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TOXICOLOGICAL REVIEW

 

OF

HEXACHLOROETHANE

(CAS No. 67-72-1) 

In Support of Summary Information on the 

Integrated Risk Information System (IRIS)

March 2010 

NOTICE 

This document is an Interagency Science Consultation draft. This information is distributed 

solely for the purpose of pre-dissemination peer review under applicable information quality 

guidelines. It has not been formally disseminated by EPA. It does not represent and should not 

be construed to represent any Agency determination or policy. It is being circulated for review 

of its technical accuracy and science policy implications.

 U.S. Environmental Protection Agency 

 Washington, DC 
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DISCLAIMER

This document is a preliminary draft for review purposes only. This information is 

distributed solely for the purpose of pre-dissemination peer review under applicable information 

quality guidelines. It has not been formally disseminated by EPA. It does not represent and 

should not be construed to represent any Agency determination or policy. Mention of trade 

names or commercial products does not constitute endorsement or recommendation for use. 
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CONTENTS—TOXICOLOGICAL REVIEW OF HEXACHLOROETHANE 

(CAS No. 67-72-1) 

LIST OF TABLES......................................................................................................................... vi 

LIST OF FIGURES ...................................................................................................................... vii 

LIST OF ABBREVIATIONS AND ACRONYMS .................................................................... viii 

FOREWORD ...................................................................................................................................x 

AUTHORS, CONTRIBUTORS, AND REVIEWERS ................................................................. xi 

1. INTRODUCTION ......................................................................................................................1 

2. CHEMICAL AND PHYSICAL INFORMATION ....................................................................3 

3. TOXICOKINETICS ...................................................................................................................5 

3.1. ABSORPTION ................................................................................................................. 5 

3.2. DISTRIBUTION............................................................................................................... 5 

3.3. METABOLISM ................................................................................................................ 8 

3.4. ELIMINATION .............................................................................................................. 15 

3.5. PHYSIOLOGICALLY BASED TOXICOKINETIC MODELS.................................... 16 

4. HAZARD IDENTIFICATION.................................................................................................17 

4.1. STUDIES IN HUMANS—EPIDEMIOLOGY, CASE REPORTS, CLINICAL 

CONTROLS .................................................................................................................. 17 

4.2. SUBCHRONIC AND CHRONIC STUDIES AND CANCER BIOASSAYS IN 

ANIMALS—ORAL AND INHALATION................................................................... 19 

4.2.1. Oral ......................................................................................................................19 

4.2.1.1. Subchronic Exposure .............................................................................. 19 

4.2.1.2. Chronic Exposure and Carcinogenicity .................................................. 23 

4.2.2. Inhalation .............................................................................................................31 

4.2.2.1. Subchronic Exposure .............................................................................. 31 

4.2.2.2. Chronic Exposure.................................................................................... 33 

4.3. REPRODUCTIVE/DEVELOPMENTAL STUDIES—ORAL AND INHALATION .. 33 

4.3.1. Oral ......................................................................................................................33 

4.3.2. Inhalation .............................................................................................................35 

4.4. OTHER DURATION- OR ENDPOINT-SPECIFIC STUDIES..................................... 36 

4.4.1. Acute Exposure Studies .......................................................................................36 

4.4.1.1. Oral ......................................................................................................... 36 

4.4.1.2. Inhalation ................................................................................................ 37 

4.4.2. Short-term Exposure Studies ...............................................................................37 

4.4.3. Neurological.........................................................................................................40 

4.4.3.1. Oral Studies............................................................................................. 41 

4.4.3.2. Inhalation Studies.................................................................................... 42 

4.4.4. Immunological .....................................................................................................42 

4.4.5. Dermatological.....................................................................................................43 

4.4.6. Eye Irritation ........................................................................................................43 

4.5. MECHANISTIC DATA AND OTHER STUDIES IN SUPPORT OF THE MODE OF 

ACTION ........................................................................................................................ 44 

4.5.1. Genotoxicity.........................................................................................................44 
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4.5.1.1. Genotoxicity of HCE Metabolites ...........Error! Bookmark not defined.

4.5.2. In Vitro and Ex Vivo Studies Using Isolated Target Tissues/Organs or Cells....49 

4.5.3. Structure Activity Relationships..........................................................................53 

4.6. SYNTHESIS OF MAJOR NONCANCER EFFECTS................................................... 54 

4.6.1. Oral ......................................................................................................................54 

4.6.1.1. Nephrotoxicity ........................................................................................ 57 

4.6.1.2. Hepatotoxicity......................................................................................... 59 

4.6.1.3. Developmental Toxicity.......................................................................... 60 

4.6.1.4. Metabolite Toxicity................................................................................. 60 

4.6.2. Inhalation .............................................................................................................61 

4.6.3. Mode-of-Action Information ...............................................................................62 

4.7. EVALUATION OF CARCINOGENICITY................................................................... 64 

4.7.1. Summary of Overall Weight of Evidence............................................................64 

4.7.2. Synthesis of Human, Animal, and Other Supporting Evidence...........................65 

4.7.3. Mode-of-Action Information for Kidney Adenomas and Carcinomas................67 

4.7.3.1. Hypothesized Mode of Action................................................................ 67 

4.7.3.2. Experimental Support for the Hypothesized Mode of Action ................ 72 

4.7.3.3. Other Possible Modes of Action............................................................. 80 

4.7.3.4. Conclusions about the Hypothesized Mode of Action ........................... 81 

4.7.4. Mode-of-Action Information for Hepatocellular Carcinomas.............................82 

4.7.4.1. Hypothesized Mode of Action..........................................................................82 

4.7.5. Mode-of-Action Information for Pheochromocytomas.......................................84 

4.7.5.1. Hypothesized Mode of Action................................................................ 84 

4.8. SUSCEPTIBLE POPULATIONS AND LIFE STAGES............................................... 84 

4.8.1. Possible Childhood Susceptibility .......................................................................84 

4.8.2. Possible Gender Differences................................................................................85 

4.8.3. Other ....................................................................................................................85 

5. DOSE-RESPONSE ASSESSMENTS......................................................................................86 

5.1. ORAL REFERENCE DOSE (RfD)................................................................................ 86 

5.1.1. Choice of Principal Study and Critical Effect—with Rationale and 

Justification.........................................................................................................86 

5.1.2. Methods of Analysis—Including Models............................................................90 

5.1.3. RfD Derivation—Including Application of Uncertainty Factors (UFs)..............92 

5.1.4. RfD Comparison Information ..............................................................................94 

5.1.5. Previous RfD Assessment....................................................................................97 

5.2. INHALATION REFERENCE CONCENTRATION (RfC) .......................................... 97 

5.2.1. Choice of Principal Study and Critical Effect—with Rationale and 

Justification.........................................................................................................97 

5.2.2. Methods of Analysis—Including Models..........................................................100 

5.2.3. RfC Derivation—Including Application of Uncertainty Factors 

(UFs) .................................................................................................................101 

5.2.4. RfC Comparison Information ............................................................................103 

5.2.5. Previous RfC Assessment..................................................................................103 

5.3. UNCERTAINTIES IN THE ORAL REFERENCE DOSE AND INHALATION 

REFERENCE CONCENTRATION............................................................................ 103

5.4. CANCER ASSESSMENT............................................................................................ 106 

5.4.1. Choice of Study/Data—with Rationale and Justification ..................................106 

5.4.2. Dose-response Data ................................................................................. 106
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5.4.3. Dose Adjustments and Extrapolation Methods........................................ 107 

5.4.4. Oral Slope Factor and Inhalation Unit Risk............................................. 110 

5.4.5. Uncertainties in Cancer Risk Values .................................................................110 

5.4.5.1. Sources of Uncertainty........................................................................112 

5.4.6. Previous Cancer Assessment .............................................................................115 

6. MAJOR CONCLUSIONS IN THE CHARACTERIZATION OF HAZARD AND 

DOSE RESPONSE ................................................................................................................116 

6.1. HUMAN HAZARD POTENTIAL............................................................................... 116 

6.2. DOSE RESPONSE ....................................................................................................... 117 

6.2.1. Oral Noncancer ..................................................................................................117 

6.2.2. Inhalation Noncancer .........................................................................................118 

6.2.3. Cancer ................................................................................................................118 

7. REFERENCES .......................................................................................................................121 

APPENDIX A: SUMMARY OF EXTERNAL PEER REVIEW AND PUBLIC 

COMMENTS AND DISPOSITION .................................................................................... A-1 

APPENDIX B: BENCHMARK DOSE MODELING OUTPUT...............................................B-1 
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LIST OF TABLES 

2-1. Physical properties of HCE .....................................................................................................3 

3-1. HCE , PERC, and pentachloroethane tissue concentrations in anaesthetized 

sheep 8.5 hours after injection of 500 mg/kg HCE..............................................................6 

3-2. Time course of HCE concentrations in male rat tissues after 57 days of dietary 

exposure to 62 mg/kg-day....................................................................................................7 

3-3. HCE concentrations in male and female rat tissues after 110 days or 111 days of 

dietary exposure...................................................................................................................8 

3-4. Disposition of HCE in rats and mice during 48 hours following administration of 

an MTD for 4 weeks ............................................................................................................9 

3-5. Metabolism amounts of HCE in rats and mice......................................................................10 

3-6. Product formation rates and relative ratios of the products formed by CYP450 

1A2 metabolism of HCE....................................................................................................14 

4-1. Body, kidney and liver weights of rats exposed to HCE in the diet for 16 weeks ................21

4-2. Histopathological results on kidney effects in rats exposed to HCE in the diet for 

16 weeks.............................................................................................................................21 

4-3. Organ weight to body weight ratios for rats exposed to HCE for 13 weeks .........................22 

4-4. Incidence and severity of nephrotoxicity in male and female rats treated with 

HCE....................................................................................................................................25 

4-5. Additional kidney effects in HCE-treated rats ......................................................................25 

4-6. Renal tubular hyperplasia and tumor incidences in HCE-treated male rats ..........................26 

4-7. Adrenal medullary lesions in HCE-treated male rats ............................................................27 

4-8. Tumor incidences in male rats gavaged with HCE ...............................................................28 

4-9. Tumor incidences in female rats gavaged with HCE ............................................................29 

4-10. Incidence of hepatocellular carcinomas in mice..................................................................30 

4-11. Summary of HCE effects on pregnant Wistar rats and their fetuses...................................34 

4-12. Summary of skeletal effects of fetuses from HCE-exposed rats.........................................35 

4-13. Summary of acute toxicity data in rats, rabbits, and guinea pigs ........................................36 

4-14. Summary of toxicity data from rats exposed to HCE for 21 days.......................................40 

4-15. Summary of genotoxicity studies of HCE...........................................................................44 

4-16. Number of enzyme-altered foci in rat liver of the promotion protocol...............................50 

4-17. In vivo covalent binding of [14C]HCE to DNA, RNA and proteins from rat and 

mouse organs .....................................................................................................................51 

4-18. In vitro binding of [14C]HCE to calf thymus DNA mediated by microsomal 

and/or cytosolic phenobarbital-induced fractions of rat and mouse organs ......................52 

4-19. Oral toxicity studies for HCE ..............................................................................................56 

4-20. Inhalation toxicity studies for HCE.....................................................................................62 

4-21. Nephrotoxic effects characteristic of α2u-globulin nephropathy observed in 

male and female rats administered HCE............................................................................70 

5-1. Incidences of non-cancerous kidney and liver effects in rats following oral 

exposure to HCE................................................................................................................89 
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5-2. Summary of the BMD modeling results for the kidney ........................................................91 

5-3. Potential points of departure (PODs) for nephrotoxicity in male rats with applied 

uncertainty factors (UF) and potential reference values ....................................................95 

5-4. Noncancerous effects observed in animals exposed to HCE via inhalation..........................99 

5-5. Summary of incidence data in rodents orally exposed to HCE for use in cancer 

dose-response assessment ................................................................................................107 

5-6. Summary of BMD modeling for oral cancer assessment of HCE.......................................109 

5-7. Summary of uncertainties in the HCE cancer risk assessment............................................111 

B-1. Dose-response modeling results using BMDS (Version 2.0) based on noncancerous kidney and liver effects in rats following oral exposure to HCE....................B-1 

B-2. Dose-response modeling results using BMDS (Version 2.0) for BMRs of 1, 5, 

and 10% based on non-cancerous kidney and liver effects in rats following 

oral exposure to HCE.......................................................................................................B-5 

LIST OF FIGURES 

2-1: Structure of HCE .................................................................................................................... 3 

3-1: Possible metabolic pathway of HCE .................................................................................... 11 

5-1. Array of potential points of departure in with applied uncertainty factors and potential 

reference values for nephrotoxic effects of studies in Table 5-3 .......................................... 96 
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LIST OF ABBREVIATIONS AND ACRONYMS

AIC Akaike’s Information Criterion 

ALD approximate lethal dosage 

ALT alanine aminotransferase 

AST aspartate aminotransferase 

BMD benchmark dose 

BMDL benchmark dose lower confidence limit 

BMDS Benchmark Dose Software 

BMR benchmark response 

BUN blood urea nitrogen 

BW body weight 

CA chromosome aberrations 

CAS Chemical Abstracts Service 

CASRN Chemical Abstracts Service Registry Number 

CBI covalent binding index 

CHO Chinese hamster ovary 

CL confidence limits 

CNS central nervous system 

CPN chronic progressive nephropathy 

CYP450 cytochrome P450 

DAF dosimetric adjustment factor 

DEN diethylnitrosamine 

DMSO dimethylsulfoxide 

FEV1.0 forced expiratory volume of 1 second 

GD gestation day 

GDH glutamate dehydrogenase 

GGT γ-glutamyl transpeptidase, γ-glutamyl transferase 

GSH glutathione 

GST glutathione-S-transferase 

Hb/g-A animal blood:gas partition coefficient 

Hb/g-H human blood:gas partition coefficient 

HCE hexachloroethane 

HEC human equivalent concentration 

HED human equivalent dose 

i.p. intraperitoneal 

IRIS Integrated Risk Information System 

IVF in vitro fertilization 

LC50 median lethal concentration 

LD50 median lethal dose 

LOAEL lowest-observed-adverse-effect level 

MN micronuclei 

MNPCE micronucleated polychromatic erythrocytes 

MTD maximum tolerated dose 

NAG N-acetyl-β-D-glucosaminidase 

NCI National Cancer Institute 

NOAEL no-observed-adverse-effect level 

NTP National Toxicology Program 
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OCT ornithine carbamoyl transferase 

PBPK physiologically based pharmacokinetic 

PBTK physiologically based toxicokinetic 

PCBs polychlorinated biphenyls 

PCNA proliferating cell nuclear antigen 

PERC tetrachloroethene, tetrachloroethylene, perchloroethylene 

POD point of departure 

POD[ADJ] duration-adjusted POD 

QSAR quantitative structure-activity relationship 

R&D research and development 

RDS replicative DNA synthesis 

REL reference exposure level 

RfC inhalation reference concentration 

RfD oral reference dose 

RGDR regional gas dose ratio 

S9 supernatant fraction from 9000 × g centrifugal spin 

SAR structure activity relationship 

SCE sister chromatid exchange 

SD standard deviation 

SDH sorbitol dehydrogenase 

SE standard error 

SGOT glutamic oxaloacetic transaminase, also known as AST 

SGPT glutamic pyruvic transaminase, also known as ALT 

SSD systemic scleroderma 

TCA trichloroacetic acid 

TCE trichloroethylene 

TCVC S-1,2,2-trichlorovinyl-L-cysteine 

TCVG S-1,2,2-trichlorovinyl glutathione 

TWA time-weighted average 

UF uncertainty factor 

UFA interspecies uncertainty factor 

UFH intraspecies uncertainty factor 

UFS subchronic-to-chronic uncertainty factor 

UFD database deficiencies uncertainty factor 

U.S. EPA United States Environmental Protection Agency 
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FOREWORD 

The purpose of this Toxicological Review is to provide scientific support and rationale 

for the hazard and dose-response assessment in IRIS pertaining to chronic exposure to

hexachloroethane. It is not intended to be a comprehensive treatise on the chemical or 

toxicological nature of hexachloroethane. 

The intent of Section 6, Major Conclusions in the Characterization of Hazard and Dose 

Response, is to present the major conclusions reached in the derivation of the reference dose, 

reference concentration and cancer assessment, where applicable, and to characterize the overall 

confidence in the quantitative and qualitative aspects of hazard and dose response by addressing 

the quality of data and related uncertainties. The discussion is intended to convey the limitations 

of the assessment and to aid and guide the risk assessor in the ensuing steps of the risk 

assessment process. 

For other general information about this assessment or other questions relating to IRIS, 

the reader is referred to EPA’s IRIS Hotline at (202) 566-1676 (phone), (202) 566-1749 (fax), or 

hotline.iris@epa.gov (email address). 
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AUTHORS, CONTRIBUTORS, AND REVIEWERS

CHEMICAL MANAGER/AUTHOR

John Cowden, Ph.D. 

National Center for Environmental Assessment 

Office of Research and Development 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 

AUTHORS

Samantha Jones, Ph.D. 

National Center for Environmental Assessment 

Office of Research and Development 

U.S. Environmental Protection Agency 

Washington, DC 

CONTRIBUTORS 

Ted Berner, M.S. 

National Center for Environmental Assessment 

Office of Research and Development 

U.S. Environmental Protection Agency 

Washington, DC 

Glinda Cooper, Ph.D. 

National Center for Environmental Assessment 

Office of Research and Development 

U.S. Environmental Protection Agency 

Washington, DC 

Andrew A. Rooney, Ph.D. 

Currently at National Toxicology Program 

Center for the Evaluation of Risks to Human Reproduction 

National Institute of Environmental Health Sciences

Research Triangle Park, NC 
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CONTRACTOR SUPPORT

James Kim, Ph.D. 

Sciences International, Inc. 

Alexandria, VA 

REVIEWERS

This document has been provided for review to EPA scientists and interagency reviewers 

from other federal agencies and White House offices. 

INTERNAL EPA REVIEWERS

Ambuja Bale, Ph.D. 

National Center for Environmental Assessment 

Office of Research and Development 

Ghazi Dannan, Ph.D. 

National Center for Environmental Assessment 

Office of Research and Development 

Kate Guyton, Ph.D. 

National Center for Environmental Assessment 

Office of Research and Development 

Maureen Gwinn, Ph.D. 

National Center for Environmental Assessment 

Office of Research and Development 

Jennifer Jinot, Ph.D. 

National Center for Environmental Assessment 

Office of Research and Development 

Channa Keshava, Ph.D. 

National Center for Environmental Assessment 

Office of Research and Development 

Allan Marcus, Ph.D. 

National Center for Environmental Assessment 

Office of Research and Development 
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D. Charles Thompson, Ph.D. 

National Center for Environmental Assessment 

Office of Research and Development 

John Whalan 

National Center for Environmental Assessment 

Office of Research and Development 
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1. INTRODUCTION

This document presents background information and justification for the Integrated Risk 

Information System (IRIS) Summary of the hazard and dose-response assessment of 

hexachloroethane (HCE). IRIS Summaries may include oral reference dose (RfD) and inhalation 

reference concentration (RfC) values for chronic and other exposure durations, and a 

carcinogenicity assessment. 

The RfD and RfC, if derived, provide quantitative information for use in risk assessments 

for health effects known or assumed to be produced through a nonlinear (presumed threshold) 

mode of action. The RfD (expressed in units of mg/kg-day) is defined as 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 inhalation RfC (expressed in units of mg/m3

) is 

analogous to the oral RfD, but provides a continuous inhalation exposure estimate. The 

inhalation RfC considers toxic effects for both the respiratory system (portal-of-entry) and for 

effects peripheral to the respiratory system (extrarespiratory or systemic effects). Reference 

values are generally derived for chronic exposures (up to a lifetime), but may also be derived for 

acute (≤24 hours), short-term (>24 hours up to 30 days), and subchronic (>30 days up to 10% of 

lifetime) exposure durations, all of which are derived based on an assumption of continuous 

exposure throughout the duration specified. Unless specified otherwise, the RfD and RfC are 

derived for chronic exposure duration. 

The carcinogenicity assessment provides information on the carcinogenic hazard 

potential of the substance in question and quantitative estimates of risk from oral and inhalation 

exposure may be derived. The information includes a weight-of-evidence judgment of the 

likelihood that the agent is a human carcinogen and the conditions under which the carcinogenic 

effects may be expressed. Quantitative risk estimates may be derived from the application of a 

low-dose extrapolation procedure. If derived, the oral slope factor is a plausible upper bound on 

the estimate of risk per mg/kg-day of oral exposure. Similarly, an inhalation unit risk is a 

plausible upper bound on the estimate of risk per μg/m3 air breathed. 

Development of these hazard identification and dose-response assessments for HCE has 

followed the general guidelines for risk assessment as set forth by the National Research Council 

(1983). EPA Guidelines and Risk Assessment Forum Technical Panel Reports that may have 

been used in the development of this assessment include the following: Guidelines for the 

Health Risk Assessment of Chemical Mixtures (U.S. EPA, 1986a), Guidelines for Mutagenicity 

Risk Assessment (U.S. EPA, 1986b), Recommendations for and Documentation of Biological

Values for Use in Risk Assessment (U.S. EPA, 1988), Guidelines for Developmental Toxicity 

Risk Assessment (U.S. EPA, 1991a), Interim Policy for Particle Size and Limit Concentration 
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Issues in Inhalation Toxicity (U.S. EPA, 1994a), Methods for Derivation of Inhalation Reference 

Concentrations and Application of Inhalation Dosimetry (U.S. EPA, 1994b), Use of the 

Benchmark Dose Approach in Health Risk Assessment (U.S. EPA, 1995), Guidelines for 

Reproductive Toxicity Risk Assessment (U.S. EPA, 1996), Guidelines for Neurotoxicity Risk 

Assessment (U.S. EPA, 1998), Science Policy Council Handbook: Risk Characterization (U.S. 

EPA, 2000a), Benchmark Dose Technical Guidance Document (U.S. EPA, 2000b), 

Supplementary Guidance for Conducting Health Risk Assessment of Chemical Mixtures (U.S. 

EPA, 2000c), A Review of the Reference Dose and Reference Concentration Processes (U.S. 

EPA, 2002), Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a), Supplemental 

Guidance for Assessing Susceptibility from Early-Life Exposure to Carcinogens (U.S. EPA, 

2005b), Science Policy Council Handbook: Peer Review (U.S. EPA, 2006a), and A Framework 

for Assessing Health Risks of Environmental Exposures to Children (U.S. EPA, 2006b). 

The literature search strategy employed for this compound was based on the Chemical 

Abstracts Service Registry Number (CASRN) and at least one common name. Any pertinent 

scientific information submitted by the public to the IRIS Submission Desk was also considered 

in the development of this document. The relevant literature was reviewed through February 

2010. 
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2. CHEMICAL AND PHYSICAL INFORMATION 

 Hexachloroethane (HCE) (CASRN 67-72-1) is a halogenated hydrocarbon consisting of 

six chlorines attached to an ethane backbone (Figure 2-1). Synonyms include 

1,1,1,2,2,2-hexachloroethane, ethane hexachloride, ethylene hexachloride, perchloroethane, 

carbon hexachloride, and carbon trichloride (ChemIDplus Advanced, 2005; ACGIH, 1991). 

Certain physical and chemical properties are shown below in Table 2-1 (ACGIH, 2001; ATSDR, 

1997a; Budavari, 1989; Howard, 1989; Weast, 1986; Spanggord et al., 1985; Verschueren, 1983; 

U.S. EPA, 1982, 1979). 

C C

Cl

Cl

Cl

Cl

Cl

Cl C C

Cl

Cl

Cl

Cl

Cl

Cl

Figure 2-1. Structure of HCE. 

Table 2-1. Physical properties of HCE. 

Name HCE 

CASRN 67-72-1 

Synonyms 1,1,1,2,2,2-hexachloroethane, ethane hexachloride, ethylene hexachloride, 

perchloroethane, carbon hexachloride, carbon trichloride 

Molecular weight 236.74 g/mol 

Molecular formula C2Cl6

Melting point Sublimes without melting 

Boiling Point 186.8°C 

Density 2.091 g/mL at 20°C 

Water solubilitya 50 mg/L at 22°C; 14 mg/L at 25°C 

Log Kow 3.82a, 3.34b, 4.14c

Log Koc 4.3 

Vapor pressure 0.5 mmHg at 20°C; 1.0 mmHg at 32.7°C

Henry’s law constant 2.8 × 10-3 atm-m3/mol at 20°C 

Conversion factor 1 ppm = 9.68 mg/m3; 1 mg/m3 = 0.10 ppm 

Sources: aHoward, 1989; bU.S. EPA, 1979; cHansch et al., 1995 

 

HCE was produced in the United States (U.S.) for commercial distribution from 1921 to 

1967 but is currently not commercially distributed (ATSDR, 1997a; IARC, 1979). In the 1970s, 

producers of HCE reported that HCE was not distributed, but used in-house or recycled 

(ATSDR, 1997a); distributors in the 1970s imported HCE from France, Spain and the United 

Kingdom (ACGIH 2001; ATSDR, 1997a). HCE and tetrachloroethane imports combined were 

1.5 million pounds in 1989 and 612,000 pounds in 2000 (NTP, 2005). HCE production in 1977 
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was between 2 and 20 million pounds; more recent information on production of HCE was not 

located (NTP, 2005; ATSDR, 1997a). HCE is produced by the chlorination of 

tetrachloroethylene (PERC) in the presence of ferric chloride at temperatures of 100–140°C 

(ATSDR, 1997a; U.S. EPA, 1991b; Fishbein, 1979; IARC, 1979). HCE is primarily used in the 

military for smoke pots, smoke grenades, and pyrotechnic devices (ACGIH, 2001; ATSDR, 

1997; U.S. EPA, 1991b; IARC, 1979). HCE was also identified in the headspace of 

chlorine-bleach-containing household products (Odabasi, 2008). In the past, HCE was used as 

an antihelminthic for the treatment of sheep flukes but is no longer used for this purpose since 

the Food and Drug Administration withdrew approval for this use in 1971 (ATSDR, 1997a). 

HCE has also been used as a polymer additive, moth repellant, plasticizer for cellulose esters, 

insecticide solvent, and in metallurgy for refining aluminum alloys (ATSDR, 1997a; U.S. EPA, 

1991b). 
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3. TOXICOKINETICS

3.1. ABSORPTION 

 There are no studies that have systematically evaluated HCE absorption in humans by the 

oral or inhalation routes of exposure. However, uptake was demonstrated by Younglai et al. 

(2002) when HCE was identified in follicular fluid during an analysis for environmental 

contaminants in 21 couples undergoing in vitro fertilization (IVF). These data identify the 

potential for HCE absorption but not the source or route of exposure. No studies have been 

reported that assess the inhalation absorption of HCE in humans. The dermal absorption rate of 

HCE has been described as limited (ATSDR, 1997a). Based on physical properties, the 

absorption of a saturated HCE solution across human skin was estimated to be 

0.023 mg/cm2

/hour (Fiserova-Bergerova et al., 1990). 

 Studies in animals via the oral route of exposure demonstrated that HCE is absorbed and 

primarily distributed to fat (Gorzinski et al., 1985; Nolan and Karbowski, 1978; Fowler, 1969). 

Fowler (1969) orally administered 500 mg/kg HCE to Scottish Blackface or Cheviot sheep and 

found maximal venous blood concentrations of HCE (10-28 μg/mL) were reached at 24 hours 

after HCE exposure, indicating slow absorption. Jondorf et al. (1957) reported that rabbits fed 

[

14C]-radiolabeled HCE at 500 mg/kg excreted only 5% of the applied radioactivity in urine over 

a period of 3 days (fecal measurements were not conducted). During this 3-day period, 14–24% 

of the applied radioactivity was detected in expired air, and the remainder was present in the 

tissues and intestinal tract. The amount of HCE absorbed by the rabbits was not determined; 

however, based on the amount of radioactivity present in urine and expired air, approximately 

19–29% of the HCE was absorbed. Studies in rats and mice (Mitoma et al., 1985) using 

[

14C]-radiolabeled HCE (500 mg/kg for rats; 1,000 mg/kg for mice) administered orally, via corn 

oil, indicated that the amount absorbed was 65–71% and 72–88%, respectively, based on the 

amount of radiolabel detected in expired air and excreta. 

3.2. DISTRIBUTION

 There are limited data on the distribution of HCE in humans (Younglai et al., 2002). The 

animal studies evaluated (Gorzinski et al., 1985; Nolan and Karbowski, 1978; Fowler, 1969) 

consistently demonstrated that HCE is distributed primarily to fat tissue followed by the kidney 

and to a lesser extent the liver and the blood (Gorzinski et al., 1985; Nolan and Karbowski, 

1978). 

 Younglai et al. (2002) evaluated the concentrations of various environmental 

contaminants in follicular fluid, serum, and seminal plasma of 21 couples undergoing IVF. HCE 

was one of the contaminants identified in >50% of follicular fluid samples, suggesting 

post-absorptive distribution to reproductive organs. The average HCE concentration in follicular 
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fluid was 232 ± 27 pg/mL (mean ± standard error [SE]). HCE was not detected in human female 

serum obtained during oocyte retrieval for IVF. This study focused primarily on chemicals such 

as pesticides and polychlorinated biphenyls (PCBs), and the authors could not make any 

conclusions with regards to the level of HCE in follicular fluid and its effect on fertility. 

 Fowler (1969) evaluated the tissue distribution of HCE in sheep. Two sheep were fasted 

for 24 hours and then anesthetized with pentobarbitone sodium. An HCE solution (15% w/v in 

olive oil) was injected for a total dose of 500 mg/kg directly into the rumen and lower duodenum 

(dose was divided). Anesthesia was maintained for 8.5 hours, after which time the sheep were 

sacrificed and tissues were taken within 10 minutes of death. Tissues that were evaluated for 

HCE include the brain, fat, kidney, liver and muscle. Bile and blood were also evaluated. HCE 

was widely distributed and the highest levels were found in fat of one sheep. Fat from different 

sites did not show significant variation in HCE concentration. The second sheep had only trace 

amounts of HCE in tissue (see Table 3-1). 

 

Table 3-1. HCE, PERC, and pentachloroethane tissue concentrations in 

anesthetized sheep 8.5 hours after injection of 500 mg/kg HCE

Concentrations (μg/g) 

Sheep 1 Sheep 2 

Tissue HCE PERC Pentachloroethane HCE PERC Pentachloroethane 

Bile (4 hr) 1.7 0.3 Trace 2.2 0.5 Nil 

Blood (6 hr) 0.2 0.4 Trace 0.2 0.2 Nil 

Brain 0.2 0.9 0.02 Trace Trace Trace 

Fat 1.1 2.1 0.02 Trace 0.6 Nil 

Kidney 0.1 1.2 Trace Trace 0.6 Trace 

Liver 0.2 0.9 0.01 Trace 2.8 Trace 

Muscle 0.04 0.5 0.01 Trace Trace Trace 

Source: Fowler (1969). 

Nolan and Karbowski (1978) studied tissue clearance of HCE in rats. Male F344 rats 

were placed on an HCE-containing diet that was calculated to deliver 100 mg/kg-day (later 

determined to be 62 mg/kg-day by Gorzinski et al., 1985) for 57 days. After this exposure 

period, the rats were returned to a HCE-free control diet and sacrificed (groups of 3 or 4 rats) 0, 

3, 6, 13, 22, and 31 days after this change in exposure. Samples of fat, liver, kidney, and whole 

blood were collected for HCE analysis. The time-course related tissue HCE concentrations are 

presented in Table 3-2. The highest tissue concentrations of HCE were in fat, which were 3-fold 

greater than the concentration in the kidney and over 100-fold greater than blood and liver 

concentrations. Fat concentrations decreased from 303 ± 50 μg/g in a first-order manner with a 

half-life of 2.7 days. Concentrations in blood and kidney also decreased in a first-order manner 

with half-lives of 2.5 and 2.6 days, respectively. Liver concentrations initially increased in the 

first 3 days postexposure, but began to decrease by day 6. The half-life for liver HCE was 2.3 
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days (calculated after peak levels were reached at day 3). These same results were published in a 

follow-up study by Gorzinski et al. (1985) that included a toxicity assessment. 

Table 3-2. Time course of HCE concentrations in male rat tissues after 

57 days of dietary exposure to 62 mg/kg-day

aValues from one of the three rats was consistently low and not used to obtain the mean ± SD. 

bOne sample was lost and a mean ± SD could not be calculated. 

cND: not detected (detection limit of 0.001 μg/g) 

Sources: Gorzinski et al. (1985); Nolan and Karbowski (1978). 

Nolan and Karbowski (1978) also evaluated tissue concentrations of HCE in both male 

and female rats after an exposure period of 110–111 days (16 weeks) to doses of 3, 30, and 

100 mg/kg-day via the diet. The actual doses were approximated as 1, 15, and 62 mg/kg-day 

after factoring in volatility of the test material from the food and based on linear nighttime food 

consumption rates (Gorzinski et al., 1985). The tissue concentrations are presented in Table 3-3. 

Kidney concentrations of HCE were much higher in male rats compared with female rats, 

particularly at the highest dose (47-fold greater in males) (Nolan and Karbowski, 1978). Kidney 

concentrations of HCE proportionately increased with the doses in males, whereas the increase in 

females was dose-dependent but not proportionate. The authors noted that the HCE kidney 

concentrations and kidney toxicity were consistently different for the male and female rats. 

Consequently, they speculated that the male rats would be 10-30 times more sensitive than 

female rats to HCE toxicity, based on the relative HCE concentration measured in the rat kidney 

(assuming that toxicity is due to HCE and not a metabolite). Both sexes exhibited comparable 

levels (although levels in males were slightly greater) of HCE in blood, liver and fat; 

concentrations in fat were the highest for both sexes. Blood levels of HCE did not correlate well 

to either the exposure dose or the dose at the major target organ, the kidney, indicating that blood 

levels of HCE may not be a suitable metric for the estimation of exposure to HCE in rats. 

 HCE tissue concentrations (n = 3 or 4) 

(mean ± SD μg/g tissue) 

Days after 

cessation of 

HCE exposure Blood Liver Kidney Fat 

0 0.834 ± 0.223 0.143 ± 0.040 81.8 ± 5.3 303 ± 50 

3 0.279 ± 0.048 0.399 ± 0.188 41.0 ± 1.4 107.8 ± 10.5 

6 0.0835 ± 0.006a 0.303 ± 0.156a 18.5b

 62.45 ± 3.04a 

13 0.015 ± 0.005 0.039 ± 0.023 2.53 ± 1.02 6.56 ± 0.52 

22 0.002 ± 0.001 0.001 ± 0.001 0.194 ± 0.171 0.472 ± 0.232 

31 NDc NDc

 0.026 ± 0.006 0.125 ± 0.020 
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Table 3-3. HCE concentrations in male and female rat tissues after 110 or 

111 days of dietary exposure

HCE tissue concentration (n = 3 or 4) 

(mean ± SD, μg/g tissue)

Dose (mg/kg-day) Blood Liver Kidney Fat 

1 Male 0.079 ± 0.057 0.291 ± 0.213 1.356 ± 0.286 3.09 ± 0.33 

Female 0.067 ± 0.039 0.260 ± 0.035 0.369 ± 0.505 2.59 ± 0.72 

15 Male 0.596 ± 0.653 1.736 ± 1.100 24.33 ± 5.73 37.90 ± 6.10 

Female 0.162 ± 0.049 0.472 ± 0.204 0.688 ± 0.165 45.27 ± 11.33 

62 Male 0.742 ± 0.111 0.713 ± 0.343 95.12 ± 11.56 176.1 ± 14.5 

Female 0.613 ± 0.231 0.631 ± 0.262 2.01 ± 0.66 162.1 ± 7.1 

Sources: Gorzinski et al. (1985); Nolan and Karbowski (1978). 

3.3. METABOLISM

 In vitro studies using liver microsomes indicated that the major enzymes involved in 

HCE metabolism are phenobarbital-inducible cytochrome P450 (CYP450) isoforms (Salmon et 

al., 1985; Town and Leibman, 1984; Nastainczyk et al., 1982; Nastainczyk et al., 1981; Salmon 

et al. 1981); however, no specific (phenobarbital-inducible) isoforms have been identified. The 

isoforms induced by phenobarbital include those from the 2A, 2B, 2C, and 3A subfamilies. One 

study (Yanagita et al., 1997) found some evidence for CYP1A2 involvement in the metabolism 

of HCE, although this was not supported by the results from in vitro studies with 

3-methylcholanthrene, an inducer of the CYP450 1 subfamily (Nastainczyk et al., 1982; 

Nastainczyk et al., 1981; Van Dyke and Wineman, 1971). Information regarding the roles of 

aroclor 1254-inducible enzymes other than 1A2 (including CYP 2A6, 2E1, 2C9, 2C19, 2D6, and 

3A4) is not available for HCE. 

The metabolism data for HCE are limited because there are only three in vivo studies 

available that provide information on metabolites: Mitoma et al. (1985) in rats and mice, Jondorf 

et al. (1957) in rabbits, and Fowler (1969) in sheep. Each of these studies tends to support 

limited metabolism for HCE. The data from the in vivo and in vitro studies support a conclusion 

that metabolism of HCE is incomplete, with excretion of unmetabolized HCE in exhaled air and 

possibly in urine. A variety of intermediary metabolites have also been identified in exhaled air 

and urine (Fowler, 1969; Jondorf et al., 1957). Figure 3-1 provides a possible metabolic pathway 

for HCE derived from the in vivo and in vitro data with ordering of metabolites based on 

sequential dechlorination and oxidation state. The HCE metabolism information was 

supplemented with data on the metabolism of the PERC (ATSDR, 1997b), trichloroethylene 

(TCE; ATSDR, 1997c), and 1,1,2,2-tetrachloroethane (ATSDR, 2008) intermediary metabolites. 

Mitoma et al. (1985) examined the distribution of HCE in male Osborne-Mendel rats and 

male B6C3F1 mice to evaluate the extent to which radiolabeled compound is metabolized in the 

48 hours after administration of 125 or 500 mg/kg to the rats and 250 or 1,000 mg/kg to the 

mice. These doses were selected based on the maximum tolerated dose (MTD) and 1⁄4 MTD of 

HCE; the MTD in rats and mice is 500 mg/kg (2.11 mmol/kg) and 1,000 mg/kg (4.22 mmol/kg), 
 DRAFT – DO NOT CITE OR QUOTE 9

respectively. The animals, 4/dose, were orally administered unlabeled HCE as a solution in corn 

oil 5 days/week for 4 weeks, followed by a single dose of [14C]-radiolabeled HCE. The 48 hour 

observation period began after administration of the radiolabeled HCE. The animals were then 

sacrificed, and urine and feces were collected from the cages. Table 3-4 summarizes the 

metabolic disposition data (based on the detection of radiolabel) at the high dose in rats and 

mice. The comparable data for the lower doses were not reported. 

Table 3-4. Disposition of HCE in male rats and mice during 48 hours following 

administration of an MTD for 4 weeks 

Rat (500 mg/kg-day) Mouse (1,000 mg/kg-day) 

Percent of administered dose 

Expired air 64.55 ± 6.67 71.51 ± 5.09 

CO2 2.37 ± 0.76 1.84 ± 0.94 

Excreta 6.33 ± 2.39 16.21 ± 3.76 

Carcass 20.02 ± 3.70 5.90 ± 1.60 

Recovery 93.28 ± 6.23 95.47 ± 23.95 

Total metabolism (CO2 + excreta + 

carcass) 28.72 23.95 

Source: Mitoma et al. (1985). 

Recovery of the radiolabel was >90% for both rats and mice. Total metabolism was 

calculated by the authors as the sum of the radiolabel present in carbon dioxide, excreta, and the 

carcass. This is an assumption by the authors and is not an accurate estimate of metabolism 

since actual metabolites were not quantified. Data on the extent of metabolism for the 

radiolabeled material are presented in Table 3-5. Based on the mass balance between dose and 

the estimate for the sum of the metabolites, 30% of the parent compound was metabolized by 

both the rats and mice. This is consistent with the 60-70% of the high dose that was reported to 

be present unchanged in exhaled air. However, this assumes that all of the exhaled radiolabel 

that was not identified as carbon dioxide was the unmetabolized parent compound. The major 

urinary metabolites, determined qualitatively by high performance liquid chromatography, were 

trichloroethanol and trichloroacetic acid (TCA) for both rats and mice. Trichloroethanol and 

TCA were also qualitatively considered the major urinary metabolites for other halogenated 

hydrocarbon compounds, including PERC, that were evaluated by Mitoma et al. (1985). 
 DRAFT – DO NOT CITE OR QUOTE 10

Table 3-5. Metabolism of HCE measured in rats and mice 

Species Dose (mmol/kg) Metabolism (mmol/kg) Percent metabolizeda

0.53 0.16 30 Rat 2.11 0.60 28 

1.05 0.32 30 Mouse 4.22 1.01 24 

a

Percent metabolism was calculated from the dose and the reported sum of the metabolites. This calculation is likely 

an underestimation of metabolism since the exhaled air was likely to include some volatile metabolites based on the 

data from Jondorf et al. (1957). 

Source: Mitoma et al. (1985).

Jondorf et al. (1957) reported that rabbits fed [14C]-radiolabeled HCE at 500 mg/kg (route 

of administration not reported by study authors) excreted only 5% of the applied radioactivity in 

urine over 3 days (72 hours), indicating slow metabolism. This is consistent with the results in 

mice and rats reported by Mitoma et al. (1985) in which approximately 2-4% of the label was 

found in urine after 48 hours. During this 3-day period, 14–24% of the radioactivity was 

detected in expired air (a lower percentage than seen for rats at a comparable dose by Mitoma et 

al., 1985), and the remainder was present in tissues and the intestinal tract. However, the authors 

did not have the capability of quantifying HCE in tissues. Reported urinary metabolites include 

trichloroethanol (1.3%), dichloroethanol (0.4%), TCA (1.3%), dichloroacetic acid (0.8%), 

monochloroacetic acid (0.7%), and oxalic acid (0.1%). The expired air contained HCE, carbon 

dioxide, PERC, and 1,1,2,2-tetrachloroethane (TCE was not found). Quantitative data on the 

volatile metabolites in exhaled air were not reported. 

The only other metabolite data come from the work of Fowler (1969) in sheep. HCE was 

administered to four Scottish Blackface and six Cheviot cross sheep at three dose levels: 0 (two 

sheep), 500 (six sheep), 750 (one sheep), and 1,000 (one sheep) mg/kg. Two HCE metabolites, 

PERC and pentachloroethane, were detected in sheep blood 24 hours after oral HCE 

administration by drenching bottle. Following administration of 500 mg/kg, blood 

measurements were 10–28 μg/mL for HCE, 0.6–1.1 μg/mL for PERC, and 0.06–0.5 μg/mL for 

pentachloroethane. Blood concentrations of HCE, PERC, and pentachloroethane were 

2.3-2.6 times greater than the corresponding concentrations in erythrocytes. Data were not 

reported for the 750 and 1,000 mg/kg doses. In vitro experiments using fresh liver slices 

suspended in an olive oil emulsion confirmed the presence of the metabolites PERC and 

pentachloroethane. 

The metabolites identified in the in vivo studies (Mitoma et al., 1985; Fowler, 1969; 

Jondorf et al., 1957) along with the in vitro studies (Town and Leibman, 1984; Nastainczyk et 

al., 1982), and ATSDR (1997a) were used in derivation of Figure 3-1. The proposed metabolic 

pathway is based on limited information; therefore, it is likely that intermediate chemical 

reactions are not captured in the figure, which presents the formation of the various metabolites 

as single step reactions. 
 DRAFT – DO NOT CITE OR QUOTE 11

Figure 3-1. Possible metabolic pathway of HCE. 

Sources: Adapted from ATSDR (1997a); Mitoma et al. (1985); Town and Leibman, 1984; 

Nastainczyk et al. (1982, 1981); Bonse and Henschler (1976); Fowler (1969); Jondorf et al. 

(1957). 

C C

OH Cl

O

Cl

Cl

O

HO OH

O

Oxalic acid

Trichloroacetic acid 

Cl

Dichloroacetic acid

H Cl

HO Cl

Dichloroethanol

CO2

One carbon pool

Cl

O Cl

HO

H

O Cl

HO

Monochloroacetic acid

Cl

Cl

C C

Cl Cl

C C

Cl Cl

Trichloroethanol Trichloroethylene

CCl3 – CCl2

● Free radical reactions

C C

Cl Cl

C C

Cl Cl

C C

Cl H

C O

Cl H

C C

Cl Cl

H H

Cl Cl

C C

Cl H

Cl Cl

C C

Cl Cl

Cl Cl

Hexachloroethane

Pentachloroethane

1,1,2,2-

Tetrachloroethane

Tetrachloroethylene

Cl H

Cl

Cl

2Cl

H

Cl

2Cl

Cl
 DRAFT – DO NOT CITE OR QUOTE 12

The in vivo data on HCE metabolism are supported by in vitro studies of hepatic 

metabolism using liver microsomes. Nastainczyk et al. (1982, 1981) reported two studies that 

provide evidence that HCE is metabolized by phenobarbital-inducible CYP450 isoforms that 

catalyze their reductive dechlorination with NADPH, cytochrome b5, and NADH as electron 

donors. HCE metabolism was measured using liver microsomes from male Sprague-Dawley rats 

that were either pretreated with phenobarbital or 3-methylcholanthrene, or were not pretreated. 

Only phenobarbital-induced rat liver microsomes demonstrated an increase in HCE metabolism 

(27.0 ± 1.1 nmol/mg protein/minute [mean + standard deviation or SD] compared with 8.0 ±

1.2 nmol/mg protein/minute for controls). Oxidation of NADPH (under anaerobic conditions) 

with an oxidation rate of 35 ± 2 nmol/mg protein/minute (mean ± SD) provided support for 

reductive dehalogenation mediated by cytochrome CYP450. Carbon monoxide inhibited the 

NADPH oxidation rate, further indicating that CYP450 enzymes were involved in the reaction. 

The major HCE metabolite of this reductive process was PERC. Nastainzcyk et al. (1982) 

determined that the stoichiometry of the reaction was represented by the following equation: 

NADPH + H+

 + Cl3C—CCl3 NADP+ + Cl2C=CCl2 + 2 H+ + 2 Cl–

(HCE) (PERC) 

Nastainczyk et al. (1982, 1981) proposed that since CYP450 is a one electron donor, the 

two electrons would be transferred sequentially. The first electron reduction would result in a 

carbon radical; the second electron reduction would result in a carbanion. From the carbanion, 

three possible stabilization reactions are possible: (1) protonation by a hydrogen atom from the 

milieu, forming pentachloroethane, (2) α-elimination of chloride, to form the carbene which

could be stabilized by the reduced CYP450, or (3) β-elimination of chloride to form PERC, 

which is the major HCE metabolite. Nastainczyk et al. (1982) found that the products of 

reductive dechlorination of HCE were 99.5% PERC and 0.5% pentachloroethane at 

physiological pHs. At a higher pH (8.4–8.8), the ratio of pentachloroethane (one electron 

reduction) to PERC (two electron reduction) increased since transfer of the second electron can 

occur via cytochrome b5, which is influenced by pH. These reaction outcomes were proposed by 

the authors to also apply to other polyhalogenated hydrocarbons. 

To provide additional support for the reaction being catalyzed by CYP450, Nastainczyk 

et al. (1982, 1981) inhibited CYP450 using carbon monoxide, metyrapone (CYP450 3A 

inhibitor), or α-naphthoflavone (CYP450 1A and CYP450 1B inhibitor) (see Omiecinski et al., 

1999 for review). In vitro metabolism of HCE by phenobarbital-induced rat liver microsomes 

was inhibited >99% when carbon monoxide was added to the incubation mixture. Metyrapone at 

a concentration of 10–4 M inhibited PERC formation by 46 ± 10% (mean ± SD) and 

pentachloroethane formation by 41 ± 8%. Treatment with 10–3 M metyrapone inhibited HCE 

metabolism to a greater extent, reducing PERC and pentachloroethane formation 66 ± 8% and 
 DRAFT – DO NOT CITE OR QUOTE 13

79 ± 10%, respectively. α-Naphthoflavone (10–4 M) was not as effective inhibiting HCE 

metabolism as metyrapone, inhibiting PERC formation 13 ± 2% and inhibiting 

pentachloroethane formation by 26 ± 4%. These data indicate that CYP450 3A isoforms are 

involved in HCE metabolism and α-naphthoflavone does not inhibit the primary CYP450

involved in the metabolism of HCE. Since metyrapone did not completely inhibit HCE 

metabolism by phenobarbital-induced liver microsomes, the remainder of HCE metabolism may 

be accounted for by the CYP450 2A and 2B subfamilies whose inhibition was not evaluated in 

this study. 

 Town and Leibman (1984) prepared liver microsomes from phenobarbital-induced male 

Holtzman rats to study the rate of metabolism of HCE to PERC. The formation of PERC was 

favored in a low oxygen environment as observed metabolism rates of 50.2 ± 0.45, 1.25 ± 0.25, 

and 0 nmol/minute/mg protein in atmospheres of N2, air, and O2, respectively. When any part of 

the NADPH-generating system, such as NADP+, glucose 6-phosphate, and glucose 6-phosphate 

dehydrogenase, was omitted from the experiment, the metabolism of HCE to PERC was 

inhibited (≥91%). In addition, the use of carbon monoxide as a monooxygenase inhibitor 

arrested HCE metabolism. Enzymes responsible for metabolism of HCE to PERC were located 

in the microsomes, rather than the cytosol, of phenobarbital-treated rat livers. Formation of 

malondialdehyde and conjugated dienes was statistically, significantly increased following 

treatment with HCE (8 mM), indicating lipid peroxidation. The authors suggested the 

involvement of a free radical. The Km and Vmax for the enzymatic formation of PERC from HCE 

were 1.20 mM and 52.0 nmol/minute/mg, respectively. Phenobarbital-induced liver microsomes 

from ICR mice were also studied and yielded Km and Vmax values of 3.34 mM and 

30.2 nmol/minute/mg, respectively. PERC formation was not detected in liver microsomes from 

phenobarbital-induced New Zealand White rabbits, suggesting that HCE metabolism resulting in 

the formation of PERC did not occur. These results support the hypothesis that rat liver 

metabolism of HCE (reductive dehalogenation) occurs by CYP450. The report identifies PERC 

as a metabolite of HCE; however, the metabolite was not quantitatively measured. 

 Salmon et al. (1981) used Aroclor 1254-induced Sprague Dawley rats to quantify the 

dechlorination of HCE. In this case, dechlorination was measured by the release of radioactive 

Cl–

 from the [36Cl]-radiolabeled HCE substrate during incubation with liver microsomes from 

induced rats. The Km and Vmax were determined as 2.37 mM and 0.91 nmol/minute/mg protein, 

respectively. A control group of noninduced rats was not included. 

Salmon et al. (1985) reported a follow-up study that used liver microsomes from 

noninduced rats (Wistar-derived Alderley Park strain) and a reconstituted CYP450 system from 

noninduced and phenobarbital-induced New Zealand White rabbits. Metabolic experiments of 

HCE using liver microsomes from noninduced rats yielded a Km of 6.0 μM and a Vmax of 

3.55 nmol NADPH/minute/mg protein (2.41 nmol NADPH/minute/nmol CYP450). These 

results are not directly comparable to the previous study (Salmon et al., 1981) because of the use 
 DRAFT – DO NOT CITE OR QUOTE 14

of a different rat strain. A reconstituted CYP450 system from phenobarbital-induced New 

Zealand White rabbits yielded Km and Vmax values of 50 μM and 

2.39 nmol NADPH/minute/nmol CYP450, respectively (Salmon et al., 1985). Microsomes from 

rabbits induced with β-naphthoflavone did not metabolize HCE. These results provide further 

evidence that the reductive dechlorination of HCE is catalyzed by phenobarbital-inducible 

CYP450 isoforms. 

Yanagita et al. (1997) used recombinantly-expressed rat CYP450 1A2 in Saccharomyces 

cerevisiae to evaluate the in vitro metabolism of several chlorinated ethylenes and ethanes, 

including HCE. The metabolism of HCE by wild-type CYP450 1A2 under aerobic conditions 

resulted in the formation of PERC (3.7 nmol/2.5 nmol CYP450/hour), pentachloroethane 

(0.8 nmol/2.5 nmol CYP450/hour), and TCE (0.6 nmol/2.5 nmol CYP450/hour). CYP450 1A2 

is a major hepatic CYP450 enzyme, but it is not a phenobarbital-inducible isoform; the major 

phenobarbital-inducible CYP450 enzymes are the 2A and 2B subfamilies. A follow-up study 

(Yanagita et al., 1998) that examined NADPH oxidation rates under anaerobic conditions found 

that CYP450 1A2 wild type had a Vmax of 1.3 mol/mol CYP450/minute, a Km of 0.25 mM, and 

an NADPH oxidation rate of 1.4 mol/mol CYP450/minute. 

Table 3-6. Product formation rates and relative ratios of the products 

formed by CYP450 1A2 metabolism of HCE 

Product formation 

(nmol/nmol CYP450/minute) 

CYP450 1A2 PERC Pentachloroethane TCE 

Ratio of PERC: 

pentachloroethane + TCE 

Wild type 0.68 0.10 0.0034 6.6 

Source: Yanagita et al. (1998). 

Beurskens et al. (1991) used HCE as a reference compound to examine the metabolism of 

three hexachlorocyclohexane isomers. Liver microsomes from male Wistar rats that were 

induced with phenobarbital converted HCE to PERC and pentachloroethane at an initial 

dechlorination rate of 12.0 nmol/minute/nmol CYP450 under anaerobic conditions. 

Van Dyke (1977) and Van Dyke and Wineman (1971) evaluated the dechlorination 

mechanisms of HCE and chlorinated olefins (alkenes) by using rat liver microsomes (a source of 

CYP450 enzymes). An initial study with HCE and other chlorinated ethanes found that the 

optimal configuration for dechlorination was a dichloromethyl group. HCE demonstrated a 

considerable amount of dechlorination (3.9%) in this in vitro study; however, the authors 

determined that HCE was unstable in aqueous solution and that this dechlorination was 

nonenzymatic based on the evidence of dechlorination in the absence of NADP. 

Gargas and Andersen (1989) and Gargas et al. (1988) determined kinetic constants for 

HCE metabolism in the rat using exhalation rates and a physiologically based pharmacokinetic 

(PBPK) inhalation model described by Ramsey and Andersen (1984) for styrene. The Vmax
 DRAFT – DO NOT CITE OR QUOTE 15

(scaled to a 1.0 kg rat) was 1.97 ± 0.05 mg/hour, or 8.3 μmol/hour. The Km was 0.80 mg/L, or 

3.38 μM. 

3.4. ELIMINATION

 No studies are available that evaluated the elimination of HCE in humans. Animal 

studies indicated that the major routes of HCE elimination are either fecal or by expired air 

(Mitoma et al., 1985; Fowler, 1969; Jondorf et al., 1957). The sheep studies (Fowler, 1969) 

indicate that orally administered HCE is eliminated by the fecal route without absorption and 

metabolism while the rodent studies (Mitoma et al., 1985) provided evidence that HCE is 

absorbed and eliminated by exhalation. It is unknown why there is a discrepancy between the 

studies in sheep and rodents. 

 Rabbits fed [14C]-radiolabeled HCE at 0.5 g/kg (Jondorf et al., 1957) eliminated 14–24% 

of the radioactivity in expired air during a 3-day period following exposure. Only 5% of the 

radiolabel was detected in urine. Fecal measurements were not conducted. 

Fowler (1969) orally administered HCE to Scottish Blackface and Cheviot cross sheep. 

Two Cheviot cross sheep were administered a single dose of 0.5 g/kg HCE and were confined to 

metabolism cages; urine and feces were collected over a period of 4 days for HCE analysis. 

More than 80% of the total fecal excretion of HCE occurred in the first 24 hours, and only small 

amounts were detected in the urine. To assess bile concentrations of HCE, two Scottish 

Blackface sheep were fasted for 24 hours and anaesthetized with pentobarbitone sodium. The 

hepatic duct was cannulated to collect bile; HCE was injected at a dose of 0.5 g/kg (15% w/v in 

olive oil) into the rumen and lower duodenum. Bile was collected continuously, with 2 mL 

retained every 30 minutes for analysis. HCE was detected in bile of anaesthetized sheep at 

15 minutes, compared with 27 minutes for blood and at maximum, HCE was 8- to 10-fold 

greater in bile. 

Mitoma et al. (1985) evaluated excretion of HCE in Osborne-Mendel rats and B6C3F1

mice following 4 weeks of administration of an MTD (500 mg/kg-day in rats, 1,000 mg/kg-day 

in mice). Excretion of radiolabel was monitored for 48 hours following administration of a 

tracer dose of [14C]-HCE. The findings are presented in Table 3-4. Most of the radiolabel was 

detected in expired air, indicating this to be a major route of elimination. The authors did not 

investigate whether the exhaled material was parent compound or volatile metabolite, assuming 

that it was the parent compound. A low percentage of the exhaled radioactivity was in the form 

of CO2, with rats exhaling slightly more than mice. The amount of radioactivity in the excreta, 

on the other hand, was lower in rats than in mice (Table 3-4). The excreta contained 6.3 and 

16.2% of the radiolabel in rats and mice, respectively.
 DRAFT – DO NOT CITE OR QUOTE 16

3.5. PHYSIOLOGICALLY BASED TOXICOKINETIC MODELS 

 No physiologically based toxicokinetic (PBTK) models for HCE have been developed 

specifically for mammalian species. Models for waterborne chloroethanes have been reported 

for rainbow trout and channel catfish; however, these are outside the scope of this toxicological 

review and are not described. 

 Gargas and Andersen (1989) and Gargas et al. (1988) determined kinetic constants for 

HCE metabolism in the rat using exhalation rates and a PBPK inhalation model described by 

Ramsey and Andersen (1984) for styrene. These reports by Gargas and Andersen (1989) and 

Gargas et al. (1988) do not describe a PBTK model for HCE, only kinetic constants for 

metabolism by inhalation. During these breath chamber experiments, fur deposition (fur 

loading) was observed to occur. At an exposure concentration of 53.3 ppm HCE at 6 hours, the 

chemical mass in body tissues was 7.29 mg and the chemical mass on fur was 0.6 mg (7.6% of 

total chemical mass). 
 DRAFT – DO NOT CITE OR QUOTE 17

4. HAZARD IDENTIFICATION

4.1. STUDIES IN HUMANS—EPIDEMIOLOGY, CASE REPORTS, CLINICAL 

CONTROLS 

There are few published studies relating to the toxicology of HCE in humans. Case 

reports of pneumonitis (Allen et al., 1992) and pneumonitis with evidence of liver abnormalities 

(Loh et al., 2008, 2006) have been described in soldiers exposed to smoke bombs containing 

HCE and zinc oxide. However, the smoke produced by this incineration is primarily zinc 

oxychloride and zinc chloride and it is not likely that these effects are a result of HCE. Some 

aluminum production processes involve the use of HCE in tablet or powder form, resulting in 

exposures to fumes containing hexachlorobenzene, octachlorostyrene, dioxins, dibenzofurans 

and other organochlorinated compounds. A case report of a hepatocellular carcinoma (Seldén et 

al., 1989), and limited data concerning some clinical serologic measures (Seldén et al., 1999, 

1997) in aluminum foundry workers involved in this process are available, but these data are not 

directly relevant to the question of health effects of HCE in other settings. No epidemiologic 

studies of the carcinogenicity of HCE were included in a 1985 review of cancer epidemiology 

with respect to halogenated alkanes and alkenes (Axelson, 1985). A study of Swedish workers 

involved in smoke bomb production has provided some information pertaining to exposure levels 

and symptoms and clinical parameters relating primarily to liver and pulmonary function (Seldén 

et al., 1994, 1993). 

 Two separate studies were conducted on a small population of Swedish workers 

occupationally exposed to HCE while producing military white smoke munitions. The first 

study reported on biological exposure monitoring (Seldén et al., 1993) and the second study 

described health effects resulting from HCE exposure (Seldén et al., 1994). The smoke 

formulation was approximately 60% HCE, 30% titanium dioxide, 8% aluminum powder, 2% 

cryolite, and a trace of zinc stearate. At the time this study was conducted in 1989, no HCE dust 

was found in the air sample filters, but the integrated results of personal and stationary charcoal 

tube samples revealed approximate HCE concentrations by location of 10–30 mg/m3

(milling/mixing), 5-25 mg/m3

 (pressing), <5 mg/m3 (assembly room), and nondetectable (storage 

room) (Seldén et al., 1993). 

 In the first study (Seldén et al., 1993), the exposed group consisted of 12 people (six men 

and six women) ranging in age from 23 to 57 (mean, 31.4 years; median, 30 years) (Seldén et al., 

1993). The principal control group (n = 12) consisted of assembly line workers from the same 

company who were unexposed to chlorinated hydrocarbons, but had some exposure to glass fiber 

dust. They were matched to the exposure group by sex and age (± 5 years), except in the case of 

one exposed male subject where only a younger control could be found. This latter-exposed 

male subject was excluded from the analysis of health effects (Seldén et al., 1994). A second 
 DRAFT – DO NOT CITE OR QUOTE 18

control group of formerly HCE-exposed workers (3 males, 10 females; age range, 31–57 years; 

mean, 43.6 years) was used in the biological exposure monitoring study. 

 Blood samples were collected for analysis of HCE concentration. For the exposed group, 

samples were drawn 5 weeks into a temporary production break (the “baseline” period), and the 

second samples were drawn 5 months later, after production had been under way for 5 weeks 

(the “production” period). Analyses of blood plasma HCE indicated that values for both control 

groups (n = 25) were all below the limit of detection (<0.02 μg/L). 

Exposed subjects were stratified into three subgroups (n = 4) of perceived exposure (low, 

medium, or high) based on information pertaining to work tasks, presence at work, and use of 

protective equipment. At baseline, the HCE concentrations in 10 of the samples from exposed 

workers were in the range of <0.02–0.06 μg/L, one sample was 0.15 μg/L, and one was 

0.52 μg/L. The last sample was from an individual who had remained in an HCE-contaminated 

area during the baseline period. Plasma HCE levels in the production period increased by nearly 

100-fold over that of the baseline samples (mean 7.30 ± 6.04 μg/L compared with 

0.08 ± 0.14 μg/L in the production and baseline samples, respectively, p < 0.01). Although the 

magnitude of individual increases varied considerably, there was a significant (p < 0.05) linear 

trend for values in the low, medium, and high exposure subgroups (means of 3.99, 7.14, and 

10.75 μg/L, respectively). These results demonstrate that a considerable increase in plasma HCE 

can occur after a relatively brief occupational exposure, even though workers used fairly 

sophisticated personal protective equipment. 

 As noted above, 11 of the subjects from the first study (Seldén et al., 1993) and their 

11 age- and sex-matched controls were included in the second health effects study (Seldén et al., 

1994). Data pertaining to 15 clinical symptoms (including headaches, sleep quality, palpations, 

difficulty concentrating, tension/restlessness, frequency of coughing, watery eyes/runny nose, 

itching/other skin problems, shortness of breath/chest discomfort, general health) were obtained 

from self-administered questionnaires for the exposed workers and the company controls. 

Similar data had been obtained in a previous study of 130 metal shop workers, and these workers 

were used as a second, “historical” comparison group in the analysis of the symptom data. 

Whole blood and serum samples from the 11 exposed and 11 matched company controls were 

analyzed for routine clinical parameters. Spot urine samples were analyzed for hemoglobin, 

protein, and glucose. Lung function was assessed by measuring vital capacity and 1-second 

forced expiratory volume (FEV1). 

The matched company controls reported more symptoms of ill health than exposed 

subjects, although the differences were not statistically significant. Although not statistically 

significant, the exposed group reported a higher prevalence of “dry skin/dry mucous 

membranes” (3/11 or 27%) than the matched controls (1/9, 9%) or historical controls (13/130, 

10%), and a higher prevalence of “itching/other skin problems” (3/11, 27%) than the historical 

controls (16/130, 12%). The prevalence of “itching/other skin problems” in the matched controls 
 DRAFT – DO NOT CITE OR QUOTE 19

(3/11, 27%) was the same as that in the exposed group. These symptoms centered on the wrist 

and neck areas, and the authors suggest that this could reflect exposure to HCE through joints in 

the protective equipment, or possibly a “traumiterative effect of the equipment itself.” Clinical 

examination revealed no dermatological or respiratory mucous membrane abnormalities in either 

group. The authors noted that a previous unpublished study of the plant workers (but with 

primarily different workers) had also found dermatologic complaints in up to 90% of the exposed 

workers. 

All of the spot urine tests were normal, and there was no evidence of an effect of HCE 

exposure on pulmonary function as measured by vital capacity and FEV1. Exposed subjects had 

significantly higher levels of serum creatinine, urate, and bilirubin than controls (p < 0.05), 

although the group means were still in the normal range. One exposed subject had a marginally 

elevated level of serum alanine aminotransferase (ALT) (70.5 U/L versus ≤41.1 U/L reference), 

while one control subject displayed increased levels of serum ALT and aspartate 

aminotransferase (AST) (67.6 and 186.4 U/L, respectively; 41.1 U/L reference for each). The 

control individual’s values returned to normal after 8 months, while the exposed subject’s serum 

ALT value worsened to 87.6 U/L 4 months later (Seldén et al., 1994). Available data pertaining 

to these liver function tests from 1982, when exposure levels at the worksite were higher than in 

the current study, did not show elevations in these liver enzymes in this individual at that time. 

Within the exposed group, there was no correlation between plasma HCE concentrations and the 

clinical chemistry parameters, although the authors do not discuss the power limitations of this 

exposure-response analysis (Seldén et al., 1993). In summary, these studies demonstrated HCE 

exposure in the smoke bomb production workers, but the health effects study is too small to 

reach definitive conclusions. The interpretation of small differences in clinical parameters, 

within the normal range, is uncertain. Based on the available data, the possible 

dermatologic/mucosal effects and hepatic effects are the areas in most need of additional 

research. 

4.2. SUBCHRONIC AND CHRONIC STUDIES AND CANCER BIOASSAYS IN 

ANIMALS—ORAL AND INHALATION

4.2.1. Oral 

4.2.1.1. Subchronic Exposure 

Two subchronic toxicity assays for HCE were reported (NTP, 1989; Gorzinski et al., 

1985, 1980). The Gorzinski et al. (1985, 1980) study (16 weeks) reported histopathological 

evaluations that found kidney degeneration in males, kidney degeneration in females, and 

minimal hepatic effects. The NTP (1989) study (13 weeks) reported kidney effects in male rats 

such as degeneration and necrosis of renal tubular epithelium, hyaline droplet formation, and 

tubular regeneration and tubular casts. Female rats in this study exhibited a dose-response 

increase in the incidence of hepatocellular necrosis of the centrilobular area. The NTP (1989) 
 DRAFT – DO NOT CITE OR QUOTE 20

study suggested that male rats may be more susceptible to kidney effects whereas female rats 

may be more susceptible to liver effects. 

Gorzinski et al. (1980) conducted a 16-week toxicity study in male and female F344 rats. 

Ten rats/sex/dose were exposed via the diet, formulated to deliver doses of 3, 30, or 

100 mg HCE/kg-day (purity 99.4%). However, due to sublimation of HCE from the feed, the 

actual doses were reported as 1.3, 20, or 82 mg/kg-day and later based on feeding and diurnal 

eating patterns were determined to be 1, 15, or 62 mg/kg-day, respectively (Gorzinski et al., 

1985). Gorzinski et al. (1980) is a Research and Development (R&D) Report by Dow Chemical 

and is not publicly available. The data for this study were published in the peer-reviewed 

literature by Gorzinski et al. (1985) and are presented in detail below. 

Gorzinski et al. (1985) fed 1, 15, or 62 mg/kg-day HCE (purity 99.4%) to F344 rats 

(10 rats/sex/dose) for 16 weeks. As described in Section 3.2, HCE concentrations in male 

kidneys were proportionately increased with administered dose while the increases in females 

were not proportionate. At the high dose, male rats displayed statistically significant increases in 

absolute and relative kidney weights accompanied by macroscopically observed alterations. 

Male rats displayed slight hypertrophy and/or dilation of proximal convoluted tubules of the 

kidneys at incidences of 0/10, 1/10, 7/10, and 10/10 for the 0, 1, 15, and 62 mg/kg-day dose 

groups, respectively. The increased incidence of slight hypertrophy and/or dilation of proximal 

convoluted tubules was statistically significant in males at the 15 and 62 mg/kg-day doses. Male 

rats displayed atrophy and degeneration of renal tubules at incidences of 1/10, 2/10, 7/10, and 

10/10 for the 0, 1, 15, and 62 mg/kg-day dose groups, respectively. The increased incidence of 

atrophy and degeneration of renal tubules was statistically significant in males at the 15 and 

62 mg/kg-day doses. Female rats did not display hypertrophy and/or dilation of proximal 

convoluted tubules of the kidneys at any dose, but did exhibit atrophy and degeneration of 

proximal tubules: 1/10, 1/10, 2/10, and 6/10 at the 0, 1, 15, and 62 mg/kg-day doses, 

respectively. However, the increased incidence of atrophy and degeneration of proximal tubules 

was only statistically significant in females at the 62 mg/kg-day dose. Male rats of the 

62 mg/kg-day group exhibited statistically significant increases in absolute and relative liver 

weights; histopathology revealed a slight swelling of the hepatocytes in this group. Although 

female rats exhibited a statistically significant increase in relative liver weight at the high dose, 

there was no evidence of hepatotoxicity in the histopathological examination. The data for liver 

and kidney weights are presented in Table 4-1 and the data for the kidney effects are presented in 

Table 4-2. 
 DRAFT – DO NOT CITE OR QUOTE 21

Table 4-1. Body, kidney and liver weights of rats exposed to HCE in the diet for 16 

weeks 

Liver Kidney 

Sex 

Dose level 

(mg/kg-day) 

Fasted body 

weight (g) Absolute (g) 

Relative 

(g/100 g body 

weight) 

Absolute (g) 

Relative 

(g/100 g body 

weight) 

0 314.4 ± 12.4 8.32 ± 0.27 2.65 ± 0.06 2.28 ± 0.08 0.73 ± 0.04 

1 328.0 ± 7.2 8.46 ± 0.22 2.58 ± 0.07 2.31 ± 0.09 0.70 ± 0.02 

15 329.0 ± 24.4 8.69 ± 0.80 2.64 ± 0.09 2.40 ± 0.15 0.73 ± 0.01 

Malea

62 324.2 ± 10.0 8.98 ± 0.54b 2.77 ± 0.12b 2.51 ± 0.12b 0.77 ± 0.02b 

0 176.7 ± 6.9 4.65 ± 0.26 2.63 ± 0.06 1.40 ± 0.08 0.79 ± 0.03 

1 174.0 ± 7.9 4.74 ± 0.22 2.73 ± 0.11 1.38 ± 0.05 0.79 ± 0.03 

15 176.7 ± 4.6 4.79 ± 0.21 2.69 ± 0.09 1.39 ± 0.06 0.79 ± 0.04 

Femalea

62 170.8 ± 5.1 4.71 ± 0.23 2.76 ± 0.10b

 1.39 ± 0.05 0.81 ± 0.02 

aData are presented as means ± SD of 10 rats/sex. 

b

Statistically significant from control using Dunnett’s test (p = 0.05). 

Source: Gorzinski et al. (1985). 

Table 4-2. Histopathological results on kidney in rats exposed to HCE in the diet 

for 16 weeksa 

Dose (mg/kg-day) 

Organ Effect Sex 0 1 15 62 

Male 0 1 7c 10cSlight hypertrophy and/or dilation of proximal 

convoluted tubules Female 0 0 0 0 

Male 1 2 7c 10 Kidney c

Atrophy and degeneration of renal tubulesb

Female 1 1 2 6c

aData are presented as number of positive observations for 10 rats/sex/dose. 

bGraded as slight in 1 of 10 male control rats and very slight in 1 of 10 control female rats. Severity of nephropathy 

was not reported for HCE-exposed rats. 

cEPA determined statistical significance from control using Fisher’s Exact Test (p = 0.05). 

Source: Gorzinski et al. (1985). 

The authors concluded that the no-observed-effect level for both male and female rats 

was 1 mg/kg-day. EPA considered 1 mg/kg-day as the male no-observed-adverse-effect level 

(NOAEL) and 15 mg/kg-day as the lowest-observed-adverse-effect level (LOAEL), based on 

renal tubule toxicity in male rats. For female rats, EPA considered the NOAEL as 15 mg/kg-day 

and the LOAEL as 62 mg/kg-day, based on renal tubule toxicity. 

 NTP (1989) conducted a 13-week study of HCE oral toxicity in F344/N rats. Groups of 

10 rats/sex/dose were administered 0, 47, 94, 188, 375, or 750 mg/kg (purity >99%) by corn oil 

gavage, 5 days/week for 13 weeks. The time-weighted average (TWA) doses were 0, 34, 67, 

134, 268, and 536 mg/kg-day, respectively. In the 536 mg/kg-day group, 5/10 male rats (only 

the 5 males that died were examined microscopically) and 2/10 female rats died before the end of 

the study. Mean body weights of 536 mg/kg-day male and female rats were decreased 19 and 
 DRAFT – DO NOT CITE OR QUOTE 22

4%, respectively, compared with controls. Statistically significant increases in liver weights are 

noted at doses of ≥67 mg/kg-day (females) and ≥134 mg/kg-day (males), and in kidney weights 

at doses of ≥268 mg/kg-day (females) and ≥67 mg/kg-day (males). Organ weight to body weight 

ratios (mg/g) generally increased in a dose-related manner for both male and female rats exposed 

to HCE (Table 4-3). 

Table 4-3. Organ weight to body weight ratios for rats exposed to HCE for 

13 weeks

HCE dose by gavage (mg/kg-day) 

0 34 67 134 268 536 

Malea

Number 10 10 10 10 9 5 

Body weight 340 ± 7.6 349 ± 8.8 343 ± 5.9 348 ± 5.9 319 ± 4.0 262 ± 13.5 

Liver 35.8 ± 0.61 37.3 ± 0.37 36.0 ± 0.71 39.1 ± 0.62c 42.5 ± 0.74c 46.3 ± 0.95c 

Brain 6.0 ± 0.30 5.7 ± 0.17 5.7 ± 0.10 5.8 ± 0.23 6.3 ± 0.21 7.2 ± 0.31c 

Heart 2.8 ± 0.04 2.8 ± 0.04 2.9 ± 0.07 3.2 ± 0.17b 3.3 ± 0.18c 3.2 ± 0.10b 

Kidney 3.0 ± 0.05 3.8 ± 0.37 4.1 ± 0.27b

 4.7 ± 0.44c 5.2 ± 0.35c 4.7 ± 0.28c 

Lung 4.2 ± 0.21 4.6 ± 0.40 4.4 ± 0.48 3.9 ± 0.22 3.9 ± 0.15 4.9 ± 0.50 

Right testis 4.2 ± 0.05 4.8 ± 0.38 4.3 ± 0.10 4.4 ± 0.17 4.7 ± 0.05 5.3 ± 0.21c 

Thymus 0.8 ± 0.04 0.8 ± 0.06 0.6 ± 0.02 0.8 ± 0.10 0.7 ± 0.04 0.6 ± 0.06 

Femalea

Number 10 10 10 10 10 8 

Body weight 206 ± 3.7 210 ± 3.9 208 ± 2.6 200 ± 2.9 203 ± 4.3 189 ± 3.8 

Liver 32.2 ± 0.56 33.4 ± 0.63 34.3 ± 0.39b 36.3 ± 0.44c 42.0 ± 0.60c 52.4 ± 0.88c 

Brain 8.7 ± 0.17 8.6 ± 0.14 8.6 ± 0.10 9.0 ± 0.14 9.0 ± 0.15 9.5 ± 0.17c 

Heart 2.9 ± 0.04 3.0 ± 0.05 3.0 ± 0.03 3.0 ± 0.04 3.1 ± 0.07 3.4 ± 0.07c 

Kidney 3.1 ± 0.04 3.2 ± 0.05 3.2 ± 0.07 3.2 ± 0.06 3.6 ± 0.05c 4.1 ± 0.10c 

Lung 4.2 ± 0.09 4.1 ± 0.09 4.2 ± 0.10 4.1 ± 0.06 4.2 ± 0.08 4.5 ± 0.13 

Thymus 1.1 ± 0.05 1.1 ± 0.05 1.1 ± 0.04 1.0 ± 0.06 1.1 ± 0.07 0.8 ± 0.05c 

 

a

Data are presented as mean ± SE in mg/g, except for body weight in g 

b

Statistically different from controls, p<0.05 

c

Statistically different from controls, p<0.01 

Source: NTP (1989). 

Kidney effects (characterized by hyaline droplet formation, tubular regeneration, and 

tubular casts), similar to the toxicity noted in the 16-day study also conducted by NTP (1989), 

were observed in 90% of 34 mg/kg-day males and in males from all other HCE dose groups 

(incidence data only reported for the 34 mg/kg-day dose group). NTP (1989) reported that the 

severity of these effects increased with dose (data not presented by NTP). These kidney effects 

were not observed in any of the treated females. At the 536 mg/kg-day dose, 5/10 males died. 

Kidneys from these five animals were examined microscopically and revealed papillary necrosis, 

degeneration, and necrosis of the renal tubular epithelium. Hepatocellular necrosis of the 

centrilobular area was observed in 2/5 males and 8/10 females at the 536 mg/kg dose, 1/10 males 

and 4/10 females at the 268 mg/kg-day dose, and 2/10 females at the 134 mg/kg-day dose. 

Additionally, males of the 536 mg/kg-day dose group exhibited hemorrhagic necrosis of the 
 DRAFT – DO NOT CITE OR QUOTE 23

urinary bladder. EPA considered the female rat NOAEL as 67 mg/kg-day and the LOAEL as 

134 mg/kg-day, based on hepatocellular necrosis. A NOAEL could not be identified for male 

rats since kidney effects were observed in 90% or more of the male rats at all tested doses 

(compared to none of the controls). EPA considered the LOAEL for male rats as 34 mg/kg-day 

(lowest dose tested), based on kidney lesions. 

4.2.1.2. Chronic Exposure and Carcinogenicity

 The National Toxicology Program (NTP) and National Cancer Institute (NCI) conducted 

two chronic toxicity/carcinogenicity bioassays in rats and one in mice. Increased incidences of 

renal tubular hyperplasia, renal adenoma or carcinoma, adrenal medulla hyperplasia, 

pheochromocytomas, and malignant pheochromocytomas were noted in male F344/N rats; 

female rats did not develop HCE-related tumors (NTP, 1989). Osborne-Mendel rats of both 

sexes in the NCI (1978) study exhibited tumor types that have been previously identified as 

spontaneous lesions in this strain, and do not provide evidence of carcinogenicity. B6C3F1 mice 

of both sexes exhibited hepatocellular carcinomas, although only male mice demonstrated a dose 

response with tumor incidence (NCI, 1978). Based on the body of evidence accumulated by 

these studies, NTP and NCI concluded that there was evidence of HCE carcinogenicity in male 

F344 rats and mice of both sexes, respectively, but there was no evidence of carcinogenicity in 

female F344 or male and female Osborne-Mendel rats (NTP 1989; NCI, 1978). 

 NTP (1989) conducted a chronic toxicity/carcinogenicity bioassay in F344/N rats. 

Groups of 50 male rats/dose were administered 0, 10, or 20 mg/kg-day (TWA doses of 0, 7, or 

14 mg/kg-day, respectively, after adjusting for continuous exposure) of HCE (purity >99%) by 

corn oil gavage, 5 days/week for 103 weeks. Groups of 50 female rats/dose were administered 0, 

80, or 160 mg HCE/kg by corn oil gavage, 5 days/week for 103 weeks (TWA doses of 0, 57, or 

114 mg/kg-day, respectively, after adjusting for continuous exposure). These sex-specific doses 

were selected based on the results of the 13-week study conducted by NTP (1989) that 

demonstrated kidney lesions in male rats at the lower doses and liver lesions in female rats at the 

higher doses. All animals were necropsied. 

Mean body weights of the 14 mg/kg-day male rats were 5–6% lower than controls after 

week 81. Mean body weights of the 114 mg/kg-day female rats were 5–9% lower between 

weeks 41 and 101. Nephropathy, characterized by tubular cell degeneration and regeneration, 

tubular dilatation and atrophy, glomerulosclerosis, interstitial fibrosis, and chronic inflammation, 

was observed in both treated and control rats. Incidences of male nephropathy were 48/50 in 

controls, 48/50 in the 7 mg/kg-day dose group, and 47/50 in the 14 mg/kg-day dose group. The 

mean severity scores for nephropathy in male rats increased with dose (2.34 ± 0.14, 2.62 ± 0.15, 

and 2.68 ± 0.16 in the 0, 7, and 14 mg/kg-day groups, respectively), with the 14 mg/kg-day 

group being statistically significantly higher than the control group. While the mean severity 

scores did not show more than a 15% increase over control in the high-dose group, an 
 DRAFT – DO NOT CITE OR QUOTE 24

examination of the various grades of severity revealed more moderate and marked nephrotoxicity 

in treated male rats compared with controls, which predominantly exhibited mild nephropathy 

(Table 4-4). 

In light of these variations in severity, EPA considered the responses observed in both the 

control and treated male rats associated with more severe (moderate and marked severity) 

nephropathy to better distinguish the HCE-related effects. Incidences of male nephropathy (that 

were of moderate or marked severity) were 18/50, 24/50, and 30/50 in the control, 7, and 

14 mg/kg-day dose groups, respectively. 

Additional kidney effects were noted in the male rats (presented in Table 4-5). Linear 

mineralization of the renal papillae was increased in a dose-dependent manner: 15/50 (30%) and 

32/50 (64%) in the 7 and 14 mg/kg-day dose groups, respectively, compared with 2/50 (4%) in 

controls. Hyperplasia of the pelvic transitional epithelium was increased in treated rats (14% in 

7 and 14 mg/kg-day HCE dose groups) compared to 0% of control rats. Nonneoplastic lesions 

such as casts (4%), cytomegaly (4%), chronic inflammation (4%), and focal necrosis (2%) were 

observed in some of the male rats administered 14 mg/kg-day. An increased incidence of renal 

tubule pigmentation was noted in 4/50 (8%) of the 7 mg/kg-day dose group and 5/50 (10%) of 

the 14 mg/kg-day dose group, compared with 1/50 (2%) in the controls. Regeneration of the 

renal tubule was observed in three males administered 14 mg/kg-day HCE. 

Incidences of female nephropathy were 22/50 for controls, 42/50 in the 57 mg/kg-day 

dose group, and 44/49 in the 114 mg/kg-day dose group. The severity scores for nephropathy in 

female rats were statistically significantly increased in both treated groups: 0.72 ± 0.13 (mean ±

SE) in controls, 1.38 ± 0.11 in the 57 mg/kg-day group, and 1.69 ± 0.12 in the 114 mg/kg-day 

group. Examination of the various grades of severity shows mild and moderate nephropathy in 

treated females compared with controls which predominantly presented less than minimally 

severe nephropathy (Table 4-4). 

Similar to the male rats, the incidence of nephropathy associated with the more severe 

(mild and moderate) responses was considered in the females rats. Therefore, incidences of 

female nephropathy (that were of mild or moderate severity) were 12/50, 25/50, and 32/50 in the 

control, 57, and 114 mg/kg-day dose groups, respectively. 

Additional kidney effects in female rats included linear mineralization of the renal 

papillae, although the incidence was not dose-dependent: 14/50 (28%) in vehicle controls, 22/50 

(44%) in the 57 mg/kg-day dose, and 13/50 (26%) in the 114 m/kg-day dose. Female rats also 

exhibited casts (4% at 114 mg/kg-day) and chronic inflammation (2% at both 57 and 

114 mg/kg-day). Pigmentation of the renal tubule was present in 4, 4, and 6% of control, 57, and 

114 mg/kg-day females, respectively. Renal tubule regeneration was observed in treated females 

(but not controls); 4% of the 57 mg/kg-day dose group and 2% of the 114 mg/kg-day dose group. 

Only male rats demonstrated an increase in hyperplasia of the pelvic transitional epithelium and 

a dose-dependent increase in incidences of mineralization along the renal papillae. 
 DRAFT – DO NOT CITE OR QUOTE 25

Table 4-4. Incidence and severity of nephropathy in male and female rats 

treated with HCE

Dose (mg/kg-day) 

Severity 0 7 14 0 57 114 

Male Female 

None (0) 2 2 3 28 8 5 

Minimal (1) 4 3 4 10 17 12 

Mild (2) 26 21 13 10 23 25 

Moderate (3) 11 10 16 2 2 7 

Marked (4) 7 14 14 0 0 0 

 

Total Incidence 

(minimal to 

marked) 

48 48 47 22 42b 45b

 

 Total # of rats 50 50 50 50 50 49 

Overall severityc 2.34 ± 0.14 2.62 ± 0.15 2.68 ± 0.16a

 0.72 ± 0.13 1.38 ± 0.11b 1.69 ± 0.12b

aAuthors reported as statistically significantly different from controls, p < 0.05. 

bAuthors reported as statistically significantly different from controls, p < 0.01. 

cMean ± SE. 

Source: NTP (1989).

Table 4-5. Additional kidney effects in HCE-treated rats

HCE Dose (mg/kg-day) 

Males Females 

Vehicle 

control 7 14 Vehicle control 57 114 

Renal tubule 

pigmentation 1/50 (2%) 4/50 (8%) 5/50 (10%) 2/50 (4%) 2/50 (4%) 3/50 (6%) 

Linear 

mineralization of 

renal papillae 

2/50 (4%) 15/50 (30%)

a 32/50 (64%) a 14/50 (28%) 22/50 (44%) 13/50 (26%) 

Hyperplasia of the 

pelvic transitional 

epithelium 

0/50 (0%) 7/50 (14%) a 7/50 (14%) a Not 

observed Not observed Not observed 

aEPA determined statistical significance using Fisher’s exact test, p < 0.05. 

Source: NTP (1989). 

EPA considered the male LOAEL as 7 mg/kg-day based on increased incidence of 

moderate or marked nephropathy (Table 4-4), hyperplasia of the pelvic transitional epithelium 

(Table 4-5), increased incidence of renal tubule pigmentation (Table 4-5), and linear 

mineralization of the renal papillae (Table 4-5). EPA considered 57 mg/kg-day the female 

LOAEL, based on dose-related increases in incidence and severity (minimal to marked) 

nephropathy. The male and female NOAELs could not be established as toxic effects were 

observed at the lowest doses tested. 
 DRAFT – DO NOT CITE OR QUOTE 26

Renal tubular hyperplasia was observed at an increased incidence in treated male rats: 

4/50 (8%) in the 7 mg/kg-day dose and 11/50 (22%; statistically significantly higher than 

controls) in the 14 mg/kg-day dose, compared with 2/50 (4%) for control (Table 4-6). Only one 

female rat, administered 57 mg/kg-day, exhibited renal hyperplasia. Dose-related increases in 

the incidence of combined renal adenomas and carcinomas were observed in males rats 

administered HCE at doses of 7 (4%) and 14 mg/kg-day (14%, statistically significantly higher 

than controls) compared with controls (2%). No HCE-related tumors were observed in female 

rats. NTP concluded that these data provided evidence of carcinogenicity in male rats based on a 

comparison with the historical controls in the study laboratory (1/300; 0.3 ± 0.8%) and in NTP 

studies (10/1,943; 0.5 ± 0.9%). 

Table 4-6. Renal tubular hyperplasia and tumor incidences in HCE-treated male 

rats

Vehicle control 7 mg/kg-day HCE 14 mg/kg-day HCE 

Hyperplasia 2/50 (4%) 4/50 (8%) 11/50 (22%)a

Adenoma 1/50 (2%) 2/50 (4%) 4/50 (8%) 

Carcinoma 0/50 (0%) 0/50 (0%) 3/50 (6%) 

Adenoma or carcinoma 1/50 (2%) 2/50 (4%) 7/50 (14%)a

a

Significantly different from vehicle controls, p < 0.01. 

Source: NTP (1989). 

This study demonstrates specificity for HCE-induced renal effects in male rats. The 

males of both dose groups were administered 8 times less HCE than the corresponding females. 

However, the male rats demonstrated more severe nephropathy than female rats. Male, but not 

female rats, also exhibited renal hyperplasia and tumors. NTP (1989) indicated that the renal 

hyperplasia and tumors observed in the HCE-treated male rats represented a morphologic 

continuum. 

Effects in the adrenal gland were also noted in HCE-treated rats. Hyperplasia of the 

adrenal medulla was reported in 9 and 20% of male rats administered 7 and 14 mg/kg-day HCE, 

respectively, compared with 12% of controls. Female rats in the control (10%) and 114 mg/kgday-treated (15%) groups exhibited hyperplasia of the adrenal medulla; this effect was not 

observed in the 57 mg/kg-day dose group. 

Adrenal medullary lesions were observed in male rats, but not female rats (Table 4-7). 

Pheochromocytoma incidences were statistically significantly increased in the 7 mg/kg group 

(26/45, 58%). The increase of pheochromocytomas 14 mg/kg-day group (19/49, 39%) was not 

statistically significant compared with controls (14/50, 28%). There were no statistically 

significant differences in the incidences of malignant pheochromocytomas and complex 

pheochromocytomas (defined as pheochromocytomas containing nervous tissue in addition to 

the typical adrenal medullary cells) between controls and treated male rats. The combined 
 DRAFT – DO NOT CITE OR QUOTE 27

incidence of all three types of pheochromocytomas was statistically significantly increased in 

males treated with 7 mg/kg-day HCE (62%) but not in males treated with 14 mg/kg-day HCE 

(43%) when compared with vehicle controls (30%) and historical controls in the study laboratory 

(75/300; 25 ± 7%) and in NTP studies (543/1,937; 28 ± 11%). NTP concluded that the increased 

incidences of pheochromocytomas in male rats were possibly treatment-related. 

Table 4-7. Adrenal medullary lesions in HCE-treated male rats 

Control 7 mg/kg-day 14 mg/kg-day 

Focal hyperplasia 6/50 (12%) 4/45 (9%) 10/49 (20%) 

Pheochromocytoma 14/50 (28%) 26/45 (58%)a 19/49 (39%) 

Complex pheochromocytoma 0/50 0/45 2/49 (4%) 

Malignant pheochromocytoma 1/50 (2%) 2/45 (4%) 1/49 (2%) 

Combined pheochromocytoma 15/50 (30%) 28/45 (62%)a 21/49 (43%) 

a

Significantly different from vehicle controls, p < 0.01. 

Source: NTP (1989). 

 The NCI (1978; Weisburger, 1977) conducted a chronic toxicity/carcinogenicity bioassay 

in Osborne-Mendel rats. HCE (purity >98%) at doses of 0, 250, or 500 mg/kg-day was 

administered by corn oil gavage to 50 rats/sex/dose for 5 days/week for 78 weeks. Following 

termination of exposure, animals were observed for 33–34 weeks for a total duration of 

111-112 weeks. Twenty rats/sex were used for the untreated and vehicle controls. Starting in 

week 23, rats treated began a 5-week cyclic rotation that involved 1 week without exposure 

followed by dosing for 4 weeks. After adjustment from 5 days/week for 78 weeks, with the 

5-week cyclic rotation for part of the time, to continuous exposure over the standard 2 years for a 

chronic bioassay, the TWA doses were 113 and 227 mg/kg-day. 

Mortality was accelerated in the HCE-treated rats (NCI reported a statistically significant 

association between increased dose and mortality). The 113 and 227 mg/kg-day males exhibited 

survival rates of 24/50 (48%) and 19/50 (38%), respectively, compared with 14/20 (70%) in the 

untreated controls and 11/20 (55%) in vehicle controls (seven rats in the vehicle control group 

were sacrificed in week 60). Mortality in the treated groups occurred early in the bioassay. 

Approximately 20% of the high- and low-dose males died by weeks 15 and 45, respectively, 

compared with 90 weeks until 20% mortality for the controls. Survival rates for the female rats 

were 14/20 (70%) for both the untreated and vehicle controls, and 27/50 (54%) and 24/50 (48%) 

for the 113 and 227 mg/kg-day dose groups, respectively. Mortality also occurred early in the 

bioassay for the female rats. Approximately 20% of the high- and low-dose females died by 

weeks 25 and 30, respectively, compared with 110 weeks until 20% mortality for the controls. 

 Chronic inflammatory kidney lesions were observed in both control and treated rats: 

male rats exhibited incidences of 15/20 (75%) in untreated controls, 14/20 (70%) in vehicle 

controls, 32/49 (65%) in the 113 mg/kg-day dose group, and 25/50 (50%) in the 227 mg/kg-day 

dose group; female rats exhibited incidences of 8/20 (40%) in untreated controls, 4/20 (20%) in 
 DRAFT – DO NOT CITE OR QUOTE 28

vehicle controls, 18/50 (36%) in the 113 mg/kg-day dose group, and 20/49 (41%) in the 

227 mg/kg-day dose group. Tubular nephropathy (characterized by degeneration, necrosis, and 

the presence of large hyperchromatic regenerative epithelial cells) was observed in 45 and 66% 

of males and 18 and 59% of females in the 113 and 227 mg/kg-day dose groups, respectively. 

These effects were not observed in the untreated or vehicle controls. EPA considered the 

LOAEL as 113 mg/kg-day (lowest dose tested), based on a dose-related increase in the incidence 

of nephropathy in both males and females. The NOAEL could not be identified. 

Tumor types exhibited by male rats surviving at least 52 weeks included kidney tubular 

cell adenoma, pituitary chromophobe adenoma, thyroid follicular cell adenoma or carcinoma, 

and testicular interstitial cell tumors (Table 4-8). Due to the high mortality in the 227 mg/kg-day 

males, statistical analyses of male rat tumors were based only on those rats surviving at least 

52 weeks. Increased incidences of kidney tubular cell adenoma (4/37) and pituitary 

chromophobe adenoma (4/32) were observed in the male rats of the 113 mg/kg-day dose group 

but not in the 227 mg/kg-day group. Male vehicle controls did not exhibit kidney tubular cell 

adenomas, although 11% (2/18) exhibited pituitary chromophobe adenomas. Thyroid follicular 

cell adenoma or carcinoma were observed in 11, 8, and 18% in vehicle control, 113, and 

227 mg/kg-day males, respectively; high dose males also demonstrated the shortest time to first 

tumor of 60 weeks, compared with vehicle control (111 weeks) and low-dose males (92 weeks). 

Testicular interstitial cell tumors were not observed in vehicle control or 113 mg/kg-day males, 

but were observed in 10% of 227 mg/kg-day males. 

Table 4-8. Tumor incidencesa in male rats gavaged with HCE

Tumor type Vehicle control 113 mg/kg-day 227 mg/kg-day 

Kidney tubular cell adenoma 0/18 (0%) 4/37 (11%) 0/29 (0%) 

Weeks to first tumor – 86 – 

Pituitary chromophobe adenoma 2/18 (11%) 4/32 (13%) 0/24 (0%) 

Weeks to first tumor 105 104 – 

Thyroid follicular cell adenoma or carcinoma 2/18 (11%) 3/36 (8%) 5/28 (18%) 

Weeks to first tumor 111 92 60 

Testis interstitial cell tumor 0/18 (0%) 0/36 (0%) 3/29 (10%) 

Weeks to first tumor – – 109 

a

Due to early accelerated mortality the statistical analyses for the incidences of tumors are based on animals 

surviving at least 52 weeks 

Source: NCI (1978). 

Tumor types exhibited by female rats included kidney hamartoma (nonneoplastic 

overgrowth), pituitary chromophobe adenoma, thyroid follicular cell adenoma or carcinoma, 

mammary gland fibroadenoma, and ovary granulose cell tumors (Table 4-9). Females 

administered 227 mg/kg-day HCE had an incidence of 6% for kidney hamartoma, while none of 

these tumors were observed in the vehicle control or 113 mg/kg-day female rats. The increased 
 DRAFT – DO NOT CITE OR QUOTE 29

incidences of the remaining tumor types observed in female rats were not dose-dependent. 

Incidence of pituitary chromophobe adenomas, thyroid follicular cell adenoma or carcinomas, 

and mammary gland fibroadenomas was lower in HCE-treated animals than in controls. Ovary 

granulose cell tumors were increased in the low dose group, compared to controls, although none 

of the female rats in the high dose group exhibited this tumor. NCI (1978) noted that all of these 

tumor types had been encountered previously as spontaneous lesions in the Osborne-Mendel rat, 

and the authors reported no statistical differences in frequencies were observed between treated 

and control rats. NCI concluded that there was no evidence of carcinogenicity in this rat study. 

Table 4-9. Tumor incidences in female rats gavaged with HCE 

Tumor type Vehicle control 113 mg/kg-day 227 mg/kg-day 

Kidney hamartoma 0/20 (0%) 0/50 (0%) 3/49 (6%) 

Weeks to first tumor – – 112 

Pituitary chromophobe adenoma 7/20 (35%) 15/50 (30%) 6/46 (13%) 

Weeks to first tumor 89 89 112 

Thyroid follicular cell adenoma or carcinoma 2/20 (10%) 3/47 (6%) 3/47 (6%) 

Weeks to first tumor 111 112 109 

Mammary gland fibroadenoma 6/20 (30%) 13/50 (26%) 9/50 (18%) 

Weeks to first tumor 106 57 94 

Ovary granulosa cell tumor 1/20 (5%) 4/48 (8%) 0/49 (0%) 

Weeks to first tumor 111 111 – 

Source: NCI (1978). 

NCI (1978; Weisburger, 1977) conducted a chronic oral study in 50 B6C3F1

mice/sex/dose administered 0, 500, or 1,000 mg/kg-day HCE (purity >98%) via corn oil gavage 

for 5 days/week for 78 weeks. Following exposure termination, animals were observed for 

12-13 weeks for a total duration of 90–91 weeks. Twenty mice/sex were included as untreated 

and vehicle controls. Starting in week 9, the doses were increased to 600 and 1,200 mg/kg-day; 

no explanation was provided for this change in dose. After adjustment from 5 days/week for 

78 weeks to continuous exposure, the TWA doses were 360 and 722 mg/kg-day. Survival rates 

were unexpectedly low in males, particularly in the control and low-dose groups: 25 and 5% in 

the vehicle and untreated control groups and 14 and 58% in the 360 and 722 mg/kg-day dose 

group, respectively. NCI (1978) did not suggest a reason why more high-dose male mice 

survived compared with the low-dose and control males. Individual animal data were not 

available to make survival adjustments to the tumor incidence data discussed below. Survival 

rates in females were 80 and 85% in vehicle and untreated control groups and 80 and 68% in the 

360 and 722 mg/kg-day dose groups, respectively. As a result of the low survival rates in the 

vehicle and untreated male control groups, NCI compared tumor incidences in the dosed males 

and females to the pooled vehicle control data derived from concurrently run bioassays for 

several other chemicals. Animals were all of the same strain, housed in the same room, 
 DRAFT – DO NOT CITE OR QUOTE 30

intubated with corn oil, tested concurrently for a least 1 year, and were examined by the same 

pathologists. 

Chronic inflammation of the kidney was observed in control and treated male mice: 67, 

80, 66, and 18% of untreated controls, pooled vehicle controls, low dose, and high dose, 

respectively. Female mice in the pooled vehicle control group (15%) and 722 mg/kg-day (2%), 

but not the untreated control and 360 mg/kg-day dose groups, exhibited chronic kidney 

inflammation. Tubular nephropathy (characterized by degeneration of convoluted tubule 

epithelium at the junction of the cortex and medulla, enlarged dark staining regenerative tubular 

epithelium, and infiltration of inflammatory cells, fibrosis and calcium deposition) was not 

observed in untreated or pooled vehicle controls of either sex, but was observed in mice treated 

with HCE: 49/50 and 47/49 in males and 50/50 and 45/49 in females in the 360 and 722 mg/kgday dose groups, respectively. Information on the severity of these effects at the different dose 

levels was not presented. No other HCE-related nonneoplastic effects were observed. EPA 

considered 360 mg/kg-day as the LOAEL for this study based on tubular nephropathy. EPA 

considered that a NOAEL was not established. 

 Increases in the incidence of hepatocellular carcinomas were observed in male and 

female mice exposed to HCE (Table 4-10). Hepatocellular adenomas were not noted in the 

report. NCI (1978) reported statistically significant increases in the incidence of hepatocellular 

carcinomas in 30 and 63% of 360 and 722 mg/kg-day males, compared with 10 and 15% of 

pooled vehicle and matched vehicle controls, respectively. Female mice also demonstrated an 

increased tumor response, 40 and 31% of 360 and 722 mg/kg-day females compared with 3 and 

10% of pooled vehicle and matched vehicle controls, respectively. Although the increases in 

HCE-treated females were not dose-dependent, a higher incidence of hepatocellular carcinomas 

was observed at the low dose (20/50) compared with the high dose (15/49). NCI concluded that 

HCE was carcinogenic in both sexes of B6C3F1 mice (1978). 

Table 4-10. Incidence of hepatocellular carcinomas in mice 

Pooled vehicle controla Matched vehicle control 360 mg/kg-day 722 mg/kg-day

Males 6/60 (10%) 3/20 (15%) 15/50 (30%)b 31/49 (63%)c

Females 2/60 (3%) 2/20 (10%) 20/50 (40%)c 15/49 (31%)c

aAs a result of the exceptionally low survival rates in the vehicle and untreated control groups, NCI used the pooled 

vehicle control data derived from concurrently run bioassays for several other chemicals. Animals were all of the 

same strain and housed in the same room. Incidences reported were not adjusted for survival. 

b

Statistically significant, p = 0.008. 

c

Statistically significant, p < 0.001. 

Source: NCI (1978). 
 DRAFT – DO NOT CITE OR QUOTE 31

4.2.2. Inhalation 

4.2.2.1. Subchronic Exposure

 Median lethal concentration (LC50) values for HCE have not been reported. Only one 

study is available in the peer-reviewed literature that evaluated the subchronic (Weeks et al., 

1979) inhalation toxicity of HCE. Weeks et al. (1979) exposed Sprague-Dawley rats, beagle 

dogs, Hartley guinea pigs, and Coturnix japonica (Japanese quail) to HCE for 6 weeks. The 

effects observed in these species include neurotoxicity, reduced body weight gain, increased 

organ weights, and some evidence of respiratory tract irritation. 

 Weeks et al. (1979) exposed Sprague-Dawley rats (25/sex/concentration) to control air, 

15, 48, or 260 ppm HCE (145, 465, and 2,517 mg/m3, respectively; purity 99.8%) for 

6 hours/day, 5 days/week for 6 weeks. Postexposure observation was carried out for 12 weeks. 

An oxygen consumption test was also conducted. The authors reported that in the 2,517 mg/m3 

group, body weight gain of male, but not the nonpregnant female, rats was reduced beginning in 

the third week of exposure (although quantitative information was not reported). All rats in the 

2,517 mg/m3 group exhibited tremors, ruffled pelt, and red exudates around the eyes following 

the fourth week of exposure. The authors reported that in the male rats, relative kidney, spleen, 

and testes weights were significantly increased; in the female rats, only relative liver weights 

were significantly increased (although quantitative information was not reported). One male and 

one female rat died during the fourth week. During the observation period, treatment-related 

effects disappeared. No gross changes were evident at necropsy; however, after sacrifice, male 

and female rats of the 2,517 mg/m3group had a higher incidence and severity of 

mycoplasma-related lesions in nasal turbinates, trachea, and lung compared with controls. The 

authors concluded these lesions were related to potentiation of an endemic mycoplasia infection 

rather than a direct effect of HCE exposure. However, no data were presented demonstrating the 

presence of mycoplasia in the lung. There were no histopathological differences observed 

between control and exposed rats sacrificed 12 weeks postexposure. No treatment-related effects 

were observed in the rats exposed to 145 and 465 mg/m3 HCE. 

In the oxygen consumption test, male rats (5/concentration) were tested prior to and 

following exposure to 145, 465, or 2,517 mg/m3 HCE for 15 minutes, 3 days/week for the 

duration of the study (6 weeks). The 2,517 mg/m3 rats exhibited significantly decreased mean 

rates of consumption both prior to (15%) and after (13%) HCE exposure. The authors suggested 

that this decrease in oxygen consumption, while nonspecific, is indicative of an alteration in 

basal metabolic rate. No histopathological effects were observed at this concentration. EPA 

considered 465 mg/m3 the NOAEL and 2,517 mg/m3 the LOAEL, based on reduced body weight 

gain, and increased organ weights. 

Weeks et al. (1979) also exposed male Sprague-Dawley rats (15/concentration) exposed 

to 15, 48, or 260 ppm HCE (145, 465, or 2,517 mg/m3

) for 6 hours/day, 5 days/week for 6 weeks 

and examined them for behavioral changes related to learned and unlearned responses (described 
 DRAFT – DO NOT CITE OR QUOTE 32

in detail in Section 4.4.3.2). Similar to the other treated rats, body weight gain was reduced. 

Final mean body weight gain in male rats was reduced 2, 5, and 10% (statistically significant) in 

the 145, 465, and 2,517 mg/m3 dose groups, respectively, compared with controls. Additionally, 

relative lung, liver, kidney, and testes weights were increased (quantitative information not 

reported) compared with controls. 

Weeks et al. (1979) also exposed four male beagle dogs/concentration to control air, 15, 

48, or 260 ppm HCE (145, 465, and 2,517 mg/m3, respectively; purity 99.8%) for 6 hours/day, 

5 days/week for 6 weeks. Postexposure observation was carried out for 12 weeks. Blood 

samples were evaluated for blood chemistry parameters. In addition, the beagle dogs underwent 

pulmonary function tests prior to and following exposure. One dog died within 5 hours of 

exposure to 2,517 mg/m3. The remaining animals in the 2,517 mg/m3 group exhibited signs of 

neurotoxicity consisting of tremors, ataxia, hypersalivation, head bobbing, and facial 

fasciculations. No blood parameters were significantly affected and no exposure-related 

histopathological lesions were observed following necropsy on dogs sacrificed 12 weeks 

postexposture. Dogs evaluated for pulmonary functions while anesthetized did not display any 

significant effects. The HCE-exposed dogs did not display any treatment-related toxicity at 

12 weeks postexposure. EPA considered 465 mg/m3 the NOAEL and 2,517 mg/m3 the LOAEL, 

based on neurotoxic effects. 

Weeks et al. (1979) also exposed male Hartley guinea pigs (10/concentration) to control 

air, 15, 48, or 260 ppm HCE (145, 465, and 2,517 mg/m3

, respectively; purity 99.8%) for 

6 hours/day, 5 days/week for 6 weeks. Postexposure observation was carried out for 12 weeks. 

Guinea pigs were also evaluated for sensitization potential following inhalation exposure to 

HCE. Two guinea pigs died during each of the fourth and fifth weeks, resulting in four total 

deaths. Guinea pigs of the 2,517 mg/m3

 group displayed reductions in body weight beginning at 

the second week of exposure and significantly increased liver to body weight ratios (quantitative 

information was not reported). No treatment-related effects were observed in the other exposure 

groups. EPA considered the NOAEL as 465 mg/m3 and the LOAEL as 2,517 mg/m3

, based on 

decreased body weight and significantly increased relative liver weight. 

Weeks et al. (1979) also exposed male and female quail (C. japonica, 20/concentration) 

to control air, 15, 48, or 260 ppm HCE (145, 465, and 2,517 mg/m3

, respectively; purity 99.8%) 

for 6 hours/day, 5 days/week for 6 weeks. Postexposure observation was carried out for 12 

weeks. The only effects observed were in 2 of 10 quail in the 2,517 mg/m3 group exhibited 

excess mucus in nasal turbinates after 6 weeks. The authors considered the excess mucus to be 

transient based on the lack of any inflammation or histopathological effects. Although the study 

authors considered the excess mucus to be a transient effect, EPA notes that the lack of 

inflammation and histopathological effects does not preclude the presence of more sensitive 

indicators of immune response (e.g., antibodies or other immune signaling chemicals) unable to 

be detected with methods available to the study authors. EPA considered 2,517 mg/m3 (highest 
 DRAFT – DO NOT CITE OR QUOTE 33

exposure concentration) as the NOAEL, while the LOAEL could not be established from this 

study. 

4.2.2.2. Chronic Exposure

 No chronic exposure studies were identified. 

4.3. REPRODUCTIVE/DEVELOPMENTAL STUDIES—ORAL AND INHALATION 

4.3.1. Oral 

Weeks et al. (1979) exposed 22 pregnant Sprague-Dawley rats/dose to 50, 100, or 

500 mg/kg HCE (purity 99.8%) by gavage on gestation days (GDs) 6–16. Gavage controls 

received corn oil, and positive controls received 250 mg/kg aspirin. Dams orally administered 

500 mg/kg HCE displayed tremors on GDs 15 and 16. Body weight gain of the 500 mg/kg dams 

was significantly lower than controls beginning on GD 8. Rats in the 500 mg/kg group exhibited 

an increased incidence of mucopurulent nasal exudates compared with controls. Approximately 

70% of the orally exposed 500 mg/kg group had upper respiratory tract irritation; 20% had 

subclinical pneumonitis, compared with 10% in controls. 

 The aspirin-positive control group produced fetuses with lower body weights, and 

malformations such as hydrocephalus, spina bifida, and cranioschesis. None of the fetuses 

exhibited any significant skeletal or soft tissue anomalies, although fetuses from dams gavaged 

with 500 mg/kg HCE displayed significantly lower gestation indices, a lower number of viable 

fetuses/dam, and higher fetal resorption rates compared with controls (data not shown). This 

study concluded that HCE did not produce teratogenic effects at doses that were not maternally 

toxic. 

Shimizu et al. (1992) evaluated the teratogenicity of HCE (purity not specified) in 

pregnant Wistar rats at doses of 0, 56, 167, or 500 mg/kg administered by gavage during 

GD 7-17 (20–21 rats/dose). The dams of the 500 mg/kg dose group exhibited significantly 

decreased weight gain after the second day of HCE treatment (8th day of pregnancy); dams in the 

167 mg/kg dose group displayed significantly decreased weight gain after the fourth day of 

treatment (10th day of pregnancy) but not after the treatment ended on the 18th day of pregnancy. 

Food intake was also significantly decreased in the 500 and 167 mg/kg dose groups after the 

second and third days, respectively, of HCE treatment; however, intake was normal when 

treatment ended. Dams in both the 167 and 500 mg/kg dose groups exhibited decreased motor 

activity (incidence and method of analysis not reported); dams in the 500 mg/kg dose group also 

exhibited piloerection and subcutaneous hemorrhage. These effects decreased or disappeared 

when HCE exposure ended. An autopsy performed on dams on GD 20, 3 days post-HCE 

exposure, revealed three rats with whitening of the liver in the 500 mg/kg dose group. The 

significance of this observation is unknown. No deaths occurred in any of the dose groups. 
 DRAFT – DO NOT CITE OR QUOTE 34

 There were no significant differences between the HCE treatment and control groups 

with respect to the number of corpora lutea, number of implants, and number of live fetuses 

(Table 4-11). There was no significant difference in the incidence of dead or resorbed fetuses 

except for a significant increase during the late stage of pregnancy in the 500 mg/kg dose group 

(6.4% versus none in the control). Fetuses in the 500 mg/kg dose group also displayed 

significantly decreased body weight; 2.5 ± 0.57 (mean ± SD) and 2.3 ± 0.45 g in male and 

female fetuses, compared with 3.3 ± 0.20 and 3.1 ± 0.24 g in male and female controls, 

respectively. 

Table 4-11. Summary of HCE effects on pregnant Wistar rats and their fetuses

Dose (mg/kg) 

0 56 167 500 

Number of dams 20 20 20 21 

% of dead or resorbed fetuses 

 Early stage 

 Late stage 

8.7 

8.7 

9.2 

8.8 

0.4 

7.0 

6.1 

0.9 

14.7 

13.1 

 6.4a 

Body weight of live fetuses (g)b 

 Male 

 Female 

3.3 ± 0.20 

3.1 ± 0.24 

3.3 ± 0.17 

3.0 ± 0.20 

3.2 ± 0.21 

2.9 ± 0.17 

2.5 ± 0.57a 

2.3 ± 0.45a 

a

Significantly different from control, p < 0.01. 

b

Values are mean ± SD. 

Source: Shimizu et al. (1992). 

Investigators examined the fetuses for external anomalies and found one case of acaudate 

in the 500 mg/kg dose group. Other anomalies included two fetuses with subcutaneous 

hemorrhage in the 167 and 500 mg/kg dose groups, and one case of hyposarca in the 500 mg/kg 

dose group. No skeletal malformations were observed in any group although a statistically 

significant increase in skeletal variations was observed in the 500 mg/kg (60.3%) group 

compared with controls (1.3%). Skeletal variations were significantly increased in the 500 

mg/kg group (2 cases in the lumbar rib and 78 cases in the rudimentary lumbar rib) and 

nonsignificantly increased in the 167 mg/kg group (6 cases in the rudimentary lumbar rib) 

compared with controls (2 cases in the rudimentary lumbar rib) (Table 4-12). The degree of 

ossification (including number of sternebrae, number of proximal and middle phalanges, and 

number of sacral and caudal vertebrae) was significantly decreased in the 500 mg/kg dose group. 

No visceral malformations were observed and no significant differences in visceral anomalies 

were noted. The authors of this study concluded that there was no indication of teratological 

effects in rats for dose levels of HCE below 500 mg/kg. Shimizu et al. (1992) established a 

NOAEL of 56 mg/kg for dams and 167 mg/kg for fetuses. EPA considered the LOAEL for dams 

as 167 mg/kg-day, based on decreased motor activity and significantly decreased body weight. 
 DRAFT – DO NOT CITE OR QUOTE 35

EPA considered the LOAEL for fetuses as 500 mg/kg, based on significantly increased skeletal 

variations and significantly decreased ossification and fetal body weight. 

Table 4-12. Summary of skeletal effects on fetuses from HCE-exposed rats

Dose (mg/kg)

0 56 167 500

Number of fetuses examined 136 136 136 137 

Percent of fetal variations 1.3 0 3.8 60.3a

Number of fetuses with variations 

 Lumbar rib 0 0 0 2 

 Rudimentary lumbar rib 2 0 6 78 

Ossificationb 

 Number of sternebrae 4.7 ± 0.07 4.5 ± 0.08 4.5 ± 0.08 3.4 ± 0.27a

 Number of proximal and middle 

phalanges 

 

 Fore limb 3.2 ± 0.05 3.1 ± 0.04 3.1 ± 0.04 2.9 ± 0.11a

 Hind limb 4.0 ± 0.01 4.0 ± 0.01 4.0 ± 0.01 3.4 ± 0.23a

 Number of sacral and caudal vertebrae 6.9 ± 0.06 6.9 ± 0.08 7.0 ± 0.04 5.7 ± 0.37a

a

Significantly different from control, p < 0.01. 

b

As reported by Shimizu et al., the litter was used as the statistical unit for calculation of fetal values, thus these 

values represent the means ± SD of litter means within each group. 

Source: Shimizu et al. (1992). 

4.3.2. Inhalation 

Weeks et al. (1979) exposed 22 pregnant Sprague-Dawley rats/concentration to control 

air, 15, 48, or 260 ppm HCE (145, 465, and 2,517 mg/m3, respectively; purity 99.8%) by 

inhalation on GDs 6-16. Dams in the 2,517 mg/m3 group displayed tremors during GDs 12–16. 

Body weight gain of the dams was significantly lower than controls beginning on GD 8 for the 

2,517 mg/m3

 group, and beginning on GD 14 for the 465 mg/m3 group. Rats in the 465 and 

2,517 mg/m3 groups exhibited an increased incidence of mucopurulent nasal exudates compared 

with controls. Inflammatory exudate was observed in the lumen of the nasal turbinates of 85% 

of the 465 mg/m3

 group and 100% of the 2,517 mg/m3 group. The authors attributed the 

increased exudate to an endemic mycoplasia infection. 

 Fetuses of HCE-treated dams did not exhibit any significant skeletal or soft tissue 

anomalies. This study concluded that HCE did not produce teratogenic effects at concentrations 

that were not maternally toxic. EPA considered the NOAEL for the dams as 465 mg/m3 and the 

LOAEL as 2,517 mg/m3

, based on neurological effects (tremors). EPA considered 2,517 mg/m3

(highest concentration tested) as a fetal NOAEL, based on the lack of treatment-related effects, 

while a fetal LOAEL could not be established from this study. 

 
 DRAFT – DO NOT CITE OR QUOTE 36

4.4. OTHER DURATION- OR ENDPOINT-SPECIFIC STUDIES

4.4.1. Acute Exposure Studies 

4.4.1.1. Oral 

Several studies evaluated acute toxicity of HCE in animal species and reported lethal 

dose concentrations. Oral lethal doses ranged from 2,332 to 8,640 mg/kg in rats, >1,000 mg/kg 

in male rabbits, and 4,970 mg/kg in guinea pigs (Kinkead and Wolfe, 1992; Weeks et al., 1979). 

According to the Hodge and Sterner Scale, these lethal doses place HCE in low toxicity range 

(Hodge and Sterner, 1949). Reynolds (1972) administered a single dose of HCE (purity not 

specified) at 26 mmol/kg (6,155 mg/kg) by gavage in mineral oil to male rats and reported that 

liver function (assessed by microsomal protein concentration, antipyrine demethylase activity, 

NADP-neotetrazolium reductase activity, glucose 6-phosphatase activity, and conjugated diene 

concentration in microsomal lipids) was unaffected 2 hours after exposure. Kinkead and Wolfe 

(1992) determined the oral median lethal dose (LD50) for HCE (purity not specified) in male and 

female Sprague-Dawley rats (5 rats/sex/dose) was 4,489 mg/kg (95% confidence limits [CLs], 

2,332–8,640 mg/kg). A study in sheep that was conducted at high doses (500-1,000 mg/kg) 

found reduced hepatic function (Fowler, 1969). 

 Weeks et al. (1979) and Weeks and Thomasino (1978) determined acute oral toxicity 

values for Sprague-Dawley rats, New Zealand White rabbits, and Hartley guinea pigs by 

administering a single dose of HCE (99.8% purity) dissolved in corn oil (50% w/v) or 

methylcellulose (5% w/v) via stomach tube. Approximate lethal dosages (ALD) or LD50 values 

were calculated after a 14-day observation period (Table 4-13). All LD50 values were 

> 1,000 mg/kg. 

Table 4-13. Summary of acute exposure data in rats, rabbits, and guinea pigs 

Lethal value 

Species Treatment Diluent mg/kg 95% CL Slope 

Rabbit, male Oral ALD Methylcellulose >1,000 

Rat, male Intraperitoneal 

(i.p.) ALD 

Corn oil 2,900 

Rat, male Oral ALD Corn oil 4,900 

Rat, female Oral LD50 Corn oil 

Methylcellulose 

4,460 

7,080 

3,900–5,110 

6,240–8,040 

9.3 

19.9 

Rat, male Oral LD50 Corn oil 

Methylcellulose 

5,160 

7,690 

4,250–6,270 

6,380–9,250 

6.1 

8.5 

Guinea pig, male Oral LD50 Corn oil 4,970 4,030–6,150 4.7 

Rabbit, male Dermal LD50 Water paste ≥32,000 

Sources: Weeks et al. (1979); Weeks and Thomasino (1978). 

Fowler (1969) orally administered a single dose of HCE (purity not specified) through a 

drenching bottle to Scottish Blackface and Cheviot cross sheep at three dose levels: 500 (six 

sheep), 750 (one sheep), and 1,000 mg/kg (one sheep). Hepatotoxicity was assessed by 
 DRAFT – DO NOT CITE OR QUOTE 37

measurement of plasma enzyme activities and bromsulphthalein dye clearance tests, which are 

widely-used indices of hepatic function in sheep. Plasma activities of glutamate dehydrogenase 

(GDH), sorbitol dehydrogenase (SDH), ornithine carbamoyl transferase (OCT), and AST were 

determined daily until they reached stable levels. Increases in these enzymes are indicative of 

hepatic damage. HCE exposure resulted in a 3-6-fold increase in GDH, with the exception of 

one sheep that exhibited a 55-fold increase. SDH was increased 3-6-fold, and OCT was 

increased 2-10-fold. GDH, SDH, and OCT levels peaked at 48 hours and returned to normal 

within 4–5 days. AST increased only slightly. Bromosulphophthalein dye clearance tests found 

a reduction in transfer from liver cells to bile at 72 hours after HCE exposure, indicating reduced 

hepatic function. 

4.4.1.2. Inhalation 

One study is available in the peer-reviewed literature that evaluated acute (Weeks and 

Thomasino, 1978) inhalation exposure to HCE. Weeks and Thomasino (1978) exposed six male 

rats/concentration (strain not specified, although one table in the report indicated strain as 

Sprague-Dawley) to 260 and 5,900 ppm HCE (2,500 or 57,000 mg/m3) for 8 hours and to 

1,000 ppm HCE (17,000 mg/m3

) for 6 hours. Postexposure observation was carried out for 14 

days. Male rats exposed for 8 hours to 2,500 mg/m3 HCE displayed no toxic signs during 

exposure or for 14 days thereafter. Body weight gain was slightly, but not statistically 

significantly, reduced over the 14-day exposure period. Male rats exposed for 8 hours to 

57,000 mg/m3

 HCE displayed severe toxic signs including death. At 6 hours, one rat had a 

staggered gait. At 8 hours, 2/6 rats were dead. The surviving rats showed statistically significant 

reductions in mean body weight on exposure day 0 (7%), 1 (21%), 3 (19%), 7 (15%), and 14 

(15%), compared with controls. Necropsy did not reveal any gross exposure-related lesions. 

Microscopy revealed that two of the four surviving rats had minimally to moderately severe 

subacute diffuse interstitial pneumonitis and vascular congestion. Additionally, a purulent 

exudate of the nasal turbinates was observed in one control and one treated rat. The authors 

concluded that this effect was not exposure-related but rather indicative of a low-grade endemic 

upper respiratory disease. The male rats exposed for 6 hours to 17,000 mg/m3 showed slight 

reductions in body weight gain on postexposure days 1 (5%) and 3 (4%) and body weights 

similar to controls for the remaining 11 days of the postexposure period. Two of the six rats 

demonstrated a staggered gait. No exposure-related gross or histopathological changes were 

observed in tissues and organs. 

4.4.2. Short-term Exposure Studies

Several studies evaluated short-term toxicity of HCE in animal species. A 12-day study 

in male New Zealand White rabbits found liver degeneration and necrosis, as well as tubular 

nephrosis in the kidney, indicating that both the liver and kidney are potential target tissues for 
 DRAFT – DO NOT CITE OR QUOTE 38

HCE-induced toxicity (Weeks et al., 1979). Short-term toxicity assays in rats (16 and 21 days) 

demonstrated kidney effects in males (NTP, 1996, 1989) but not females (NTP, 1989). 

Weeks et al. (1979) conducted a 12-day study of HCE in male New Zealand White 

rabbits. Five rabbits/dose were administered a daily oral dose via a stomach tube of 100, 320, or 

1,000 mg/kg HCE (purity 99.8%) suspended in 5% aqueous methylcellulose. Blood was drawn 

from the central ear artery of the rabbits on treatment days 1, 4, 8, and 12, and on day 4 

following termination of dosing. Serum was analyzed for the following parameters: glutamic 

oxaloacetic transaminase (SGOT; also known as AST), glutamic pyruvic transaminase (SGPT; 

also known as ALT), blood urea nitrogen (BUN), alkaline phosphatase, bilirubin, total protein, 

potassium and sodium. On the 4th day following the termination of dosing, rabbits were 

necropsied and the following tissues were examined: eye, brain, lung, kidney, liver, spleen, 

heart, stomach, pancreas, large intestine, skeletal muscle, bone, urinary bladder, small intestine, 

and testes. 

The 1,000 mg/kg dose group exhibited significantly reduced body weight (beginning on 

treatment day 7) and increased relative liver and kidney weights. The 320 mg/kg dose group 

exhibited significantly reduced body weight beginning on day 10. The 100 mg/kg dose group 

did not display any changes. The 320 and 1,000 mg/kg dose groups displayed liver degeneration 

and necrosis, including fatty degeneration, coagulation necrosis, hemorrhage, ballooning 

degeneration, eosinophilic changes, and hemosiderin-laden macrophages and giant cells. These 

effects were not observed in controls or rabbits of the 100 mg/kg dose group. Liver lesions 

increased in severity in a dose-related manner in which the effects were more severe in the 

1,000 mg/kg group compared with the 320 mg/kg group. Tubular nephrosis of the convoluted 

tubules in the corticomedullary region of the kidney was also observed in the rabbits of the 320 

and 1,000 mg/kg dose groups. These animals also exhibited tubular nephrocalcinosis of a 

minimal degree. The only blood chemistry parameters that were affected were significantly 

decreased potassium and glucose levels in the 320 and 1,000 mg/kg groups. EPA considered the 

NOAEL as 100 mg/kg and the LOAEL as 320 mg/kg, based on dose-related increases in severity 

of liver and kidney lesions. 

The NTP (1989) conducted a 16-day study of oral HCE toxicity in F344/N rats. Groups 

of five rats/sex/dose were administered 0, 187, 375, 750, 1,500 or 3,000 mg HCE/kg (purity 

>99%) for 12 doses over 16 days by corn oil gavage. TWA doses were 0, 140, 281, 563, 1,125, 

and 2,250 mg/kg-day, respectively. Necropsy was performed on all rats; all organs and tissues 

were examined for grossly visible lesions and histopathology. All rats of the 1,125 and 

2,250 mg/kg-day dose groups and 1/5 males and 2/5 females from the 563 mg/kg-day dose group 

died before the end of the study. Final mean body weights (statistical analyses were not 

reported) were decreased by 25% in males of the 563 mg/kg-day dose group; female body 

weights were decreased by 37% in the 563 mg/kg-day dose group. Microscopic observations of 

the kidneys revealed hyaline droplet formation in the cytoplasm of renal tubular epithelium in all 
 DRAFT – DO NOT CITE OR QUOTE 39

the treated males, and tubular cell regeneration and eosinophilic granular casts of cell debris in 

tubule lumina of male rats administered 140 and 281 mg/kg-day. EPA considered 

140 mg/kg-day (lowest dose tested) a male rat LOAEL based on kidney tubule lesions, while a 

NOAEL could not be established for male rats. EPA considered the female rat LOAEL as 

563 mg/kg-day, based on a dose-related decrease in body weight, and the female rat NOAEL as 

281 mg/kg-day. 

 NTP (1996) conducted a 21-day study of oral HCE toxicity in male F344/N rats. Groups 

of five rats/dose were administered 0.62 or 1.24 mmol HCE/kg-day (146 or 293 mg/kg-day, 

respectively; purity 100%) by corn oil gavage. Necropsies were performed on all rats; the right 

kidney, liver, and right testis were weighed and underwent histopathological evaluation. Urine 

samples were collected during an overnight period that began 4 days before the end of the study. 

Urinalysis included measurements of volume, specific gravity, creatinine, glucose, total protein, 

AST, γ-glutamyl transferase (GGT), and N-acetyl-β-D-glucosaminidase (NAG). A 

Mallory-Heidenhain stain was used for kidney sections to evaluate protein droplets, particularly 

hyaline droplet formation. Cell proliferation analyses were performed on kidney sections and 

were scored by a labeling index indicating the percentage of proximal and distal tubule epithelial 

cells in S-phase. 

Results from the measured endpoints/parameters are summarized in Table 4-14. 

Absolute and relative kidney weights were significantly increased in both dose groups; absolute 

and relative (significant at high dose) liver weights were increased in both dose groups. Rats of 

the 293 mg/kg-day group also exhibited significantly lower urinary creatinine and specific 

gravity, while glucose and urine volume were greater than controls. AST and NAG activities 

were significantly higher than in controls. Nephropathy, consisting of hyaline droplet 

accumulation, was observed in the male rats in addition to increased incidences of tubule 

regeneration (3/5 and 4/5 for 146 and 293 mg/kg-day, respectively) and granular casts (4/5 and 

3/5 for 146 and 293 mg/kg-day, respectively). The mean proliferating cell nuclear antigen 

(PCNA) labeling index was significantly increased 5.7- and 9.2-fold, compared with controls, in 

the 146 and 293 mg/kg-day dose groups. EPA did not identify a NOAEL because effects 

(including increased absolute and relative kidney weight, increased AST and NAG activity, 

increased PCNA labeling index, and nephropathy) were observed at the low dose level. EPA 

considered 146 mg/kg-day a LOAEL based on statistically significant increases in kidney lesions 

and urinalysis parameters. 
 DRAFT – DO NOT CITE OR QUOTE 40

Table 4-14. Summary of toxicity data from male rats exposed to HCE for 21 days 

Vehicle control 146 mg/kg-day HCE 293 mg/kg-day HCE

Right kidney weighta 

 Absolute (g) 1.009 ± 0.025 1.157 ± 0.011b

 1.250 ± 0.022b

 Relative (mg/g) 3.19 ± 0.04 3.77 ± 0.06b

 4.07 ± 0.05b

Liver weighta 

 Absolute (g) 11.041 ± 0.291 11.959 ± 0.178 13.479 ± 0.390 

 Relative (mg/g) 34.82 ± 0.60 39.01 ± 0.92 43.84 ± 0.64b

Right testis weighta 

 Absolute (g) 1.412 ± 0.037 1.409 ± 0.023 1.430 ± 0.016 

 Relative (mg/g) 4.47 ± 0.09 4.60 ± 0.11 4.66 ± 0.05 

Urinalysis 

 Creatinine (mg/dL) 143.22 ± 18.12 79.56 ± 11.01 54.48 ± 3.06b

 Glucose (μg/mg creatinine) 169 ± 3 344 ± 30 446 ± 23b

 Protein (mU/mg creatinine) 1,322 ± 59 1,748 ± 257 2,980 ± 103 

 AST (mU/mg creatinine) 6 ± 1 40 ± 6c

 66 ± 5b

 GGT (mU/mg creatinine) 1,456 ± 47 1,547 ± 66 1,897 ± 73 

 NAG (mU/mg creatinine) 11 ± 0 23 ± 2c

 36 ± 1b

 Volume (mL/16 h) 4.2 ± 0.8 7.5 ± 9 10.6 ± 1.1b

 Specific gravity (g/mL) 1.038 ± 0.005 1.024 ± 0.003 1.020 ± 0.001b

PCNA labeling index (mean ± SE) 0.13 ± 0.02 0.74 ± 0.19c 1.2 ± 0.2c

aData are mean ± SE. 

b

Significantly different from control (p ≤ 0.01). 

c

Significantly different from control (p ≤ 0.05). 

Source: NTP (1996). 

4.4.3. Neurological 

 Neurological endpoints for HCE toxicity have been evaluated in several HCE toxicity 

studies. The studies listed below provide limited evidence that HCE produces neurological 

effects; however, it is unknown if the central nervous system (CNS) effects are due to the parent 

compound or the metabolites. Although there are few studies on the neurological effects 

associated with HCE exposure, the database is extensive for two of its proposed metabolites, 

PERC and TCE. Studies have shown that PERC and TCE readily cross the blood brain barrier 

resulting in CNS depressive effects. Sheep exposed to high doses of HCE (500-1,000 mg/kg) 

developed facial muscle tremors (Fowler, 1969; Southcott, 1951), and a staggering 

uncoordinated gait (Southcott, 1951). Sprague-Dawley rats evaluated for HCE-induced effects 

on avoidance latency (i.e., learned behavior) and spontaneous motor activity (i.e., unlearned 

behavior) exhibited slight, but not statistically significant, behavioral effects at 2,517 mg/m3. 

Male and female rats exhibited tremors and ruffled pelt at 2,517 mg/m3 as well (Weeks et al., 

1979). Beagle dogs developed signs of neurotoxicity such as tremors, ataxia, hypersalivation, 

and head bobbing, following exposure to 2,517 mg/m3 HCE. Dogs showed similar signs of 
 DRAFT – DO NOT CITE OR QUOTE 41

neurotoxicity intermittently throughout the HCE exposures, with signs disappearing overnight. 

During an observation period of 12 weeks following exposure, these symptoms were not 

observed (Weeks et al., 1979). 

4.4.3.1. Oral Studies 

Fowler (1969) orally administered a single dose of HCE (purity not specified) to Scottish 

Blackface and Cheviot cross sheep at three dose levels: 500 (10 sheep), 750 (1 sheep), and 

1,000 mg/kg (1 sheep). Slight facial muscle tremors were noted in three sheep between 1 and 

4 hours after dosages of 500-1,000 mg/kg HCE. The HCE dose level for the individual sheep 

exhibiting facial tremors was not specified by the authors. Fowler (1969) also examined two 

sheep administered 0.3 mL/kg PERC and two sheep administered 0.3 mL/kg pentachloroethane, 

two proposed major metabolites of HCE. The sheep exposed to PERC exhibited no effects 

following exposure, while the sheep exposed to pentachloroethane exhibited narcosis One 

pentachloroethane-exposed animal was recumbent within 30 minutes of exposure, exhibiting 

flaccid limbs, depression of normal reflexes, and labial tremors. The sheep regained normal 

posture 9 hours postexposure and appeared normal 72 hours postexposure. The second 

pentachloroethane-treated sheep became recumbent within 20 minutes of exposure and exhibited 

labial tremors. However, unlike the first sheep, this animal appeared normal 1.5 hours 

postexposure. EPA considered the LOAEL as 500 mg/kg (lowest dose tested), based on 

neurotoxic effects (tremors), while a NOAEL could not be established from this data. 

 Southcott (1951) treated 30 Merino Wethers sheep suffering from liver fluke infections 

with 15 g HCE-bentonite dispersible powder (13.5 g HCE, 445 mg/kg; 15 sheep) or 30 g 

HCE-bentonite (27 g HCE, 906 mg/kg; 15 sheep). The purity of the HCE was not specified. 

Two sheep died a day after treatment and nine others were unable to rise and stand. One of the 

severely affected sheep (i.e., unable to rise and stand) was from the 445 mg/kg HCE group and 

the other eight were from the 906 mg/kg group. Some severely affected animals (two from the 

445 mg/kg group) could walk if placed on their feet, but displayed a staggering, uncoordinated 

gait and fell again. The lips, face, neck, and forelegs were afflicted by fine muscular tremors that 

were observed in most of the animals. EPA considered the LOAEL as 445 mg/kg (lowest dose 

tested), based on neurological effects consisting of tremors, staggering, uncoordinated gait, and 

inability to stand, while a NOAEL could not be established from this study. 

 As described in Section 4.3.1, Shimizu et al. (1992) reported decreased motor activity 

(incidence and method of analysis not reported) in pregnant Wistar rats (20–21 rats/dose) at 

doses of 167 and 500 mg/kg administered by gavage during GD 7–17. These effects decreased 

or disappeared when HCE exposure ended. Weeks et al. (1979) exposed 22 pregnant 

Sprague-Dawley rats/dose to 50, 100, or 500 mg/kg HCE by corn oil gavage on GDs 6–16. 

Dams orally administered 500 mg/kg HCE displayed tremors on GDs 15 and 16. 
 DRAFT – DO NOT CITE OR QUOTE 42

4.4.3.2. Inhalation Studies 

 Weeks et al. (1979) exposed male Sprague-Dawley rats (15/concentration) to air, 15, 48, 

or 260 ppm HCE (145, 465, or 2,517 mg/m3, respectively; purity 99.8%) for 6 hours/day, 

5 days/week for 6 weeks. Learned behavior endpoints evaluated using an avoidance latency task 

by measuring the time it took the rats to avoid foot shock by escaping into a safe compartment 

and unlearned behavior endpoints (spontaneous motor activity; evaluated by photobeam 

interruptions) were measured in the animals. The avoidance latency task was conducted prior to 

exposure, 1 day into exposure, after 3 weeks of exposure and after 6 weeks of exposure. 

Spontaneous motor activity was tested after 3 and 6 weeks of exposure. 

 Avoidance latency was slightly but not significantly increased in the 465 and 

2,517 mg/m3 groups at 6 weeks (median 3.9 and 3.3 seconds, respectively) compared with 

control (median 2.2 seconds). Spontaneous motor activity counts were slightly, but not 

significantly increased in the HCE-treated rats (mean ± SD): 231 ± 77 for 145 mg/m3, 183 ± 109 

for 465 mg/m3

, and 201 ± 102 for 2,517 mg/m3, compared with control rats (163 ± 74). Weeks 

et al. (1979) concluded that the rats did not display obvious signs of behavioral toxicity. 

However, tremors and a ruffled pelt were noted in a separate experiment in male and female rats 

exposed to 2,517 mg/m3 HCE during the fourth week of exposure. Tremors and lack of 

grooming are indicators of neurobehavioral effects (Kulig et al., 1996). The investigators 

sacrificed the rats 12 weeks after the last exposure and reported that all measurable changes (e.g., 

brain histopathology, body weights) were comparable to controls. 

 Weeks et al. (1979) also exposed 22 pregnant Sprague-Dawley rats/concentration and 

4 beagle dogs/concentration to 145, 465, and 2,517 mg/m3 HCE by inhalation. Rat dams in the 

2,517 mg/m3 group displayed tremors during GDs 12–16. Dogs in the 2,517 mg/m3 exposure 

group developed tremors, ataxia, hypersalivation, and displayed severe head bobbing, facial 

muscular fasciculations, and held their eyelids closed during exposure. One dog experienced 

convulsions and died within 5 hours after initial exposure. The surviving dogs exhibited less 

severe symptoms during exposure, but recovered overnight after removal from exposure. 

4.4.4. Immunological 

 Ten male Hartley guinea pigs/dose were exposed by inhalation to control air or three 

concentrations of HCE (purity 99.8%): 15, 48, or 260 ppm (145, 465, and 2,517 mg/m3

, 

respectively; Weeks et al., 1979). Exposures were conducted for 6 hours/day, 5 days/week for 

3 weeks. Following exposure, animals were allowed to rest for 2 weeks. The guinea pigs were 

then challenged with a single intradermal injection of 0.1% HCE in saline. A sensitization 

response was not produced. 
 DRAFT – DO NOT CITE OR QUOTE 43

4.4.5. Dermatological

 Yamakage and Ishikawa (1982) examined certain patients suffering from systemic 

scleroderma (SSD) for potential exposure to solvents. These patients also presented with 

localized scleroderma with bilateral distribution of multiple skin lesions reminiscent of those 

observed in several cases of occupational or agent-induced scleroderma. Of nine such patients, 

seven had had significant subchronic/chronic exposure (5–44 years), while an eighth had had a 

significant acute exposure (2 weeks). The solvents involved were reported as “variable and 

mostly unidentified.” As an experimental follow-up, groups of ddY mice received 

intraperitoneal (i.p.) injections daily for 17 days with 1 of 10 experimental solvents, as well as 

with 0.9% saline to mitigate treatment lethality. For HCE, 17 mice were injected daily with 0.01 

mL of HCE (purity not specified) and 0.1 mL of 0.9% saline. Along with naphtha (“Esso No. 

5”) and n-hexane, HCE was found by double-blind histological examination and electron 

microscopy to be a significant inducer of sclerodermatous changes in skin taken from the 

animals’ backs, near the forelimbs. HCE treatment resulted in evident dermal sclerosis in five 

mice, slight fibrosis in another, and no change in nine; two mice died. PERC, a primary 

metabolite of HCE, was similarly tested in 10 mice. Injections of 0.005 mL (+ saline) resulted in 

evident dermal sclerosis in one mouse, slight fibrosis in two, no change in six, and death in one. 

Even though the experimental route of exposure used is generally irrelevant to humans, the skin 

lesions produced by HCE were “fundamentally similar” to those produced by control reference 

solvents that have been implicated in human occupational SSD. Thus, this study provides 

indirect evidence that suggests HCE may be capable of inducing SSD-type conditions in humans. 

 Weeks and Thomasino (1978) conducted two dermal studies in male New Zealand White 

rabbits. A single 24-hour application of 500 mg of technical dry HCE to intact and abraded skin 

of six rabbits did not result in primary irritation of intact or abraded skin when assessed at 

24 hours, 72 hours, or 7 days after exposure. HCE was placed in Irritation Category IV (no 

irritation). In the second study, HCE was applied as a paste in 0.5 mL of distilled water. Intact 

skin displayed no edema and barely perceptible erythema at 24 hours. Abraded skin displayed 

barely perceptible erythema in one rabbit with moderate to slight erythema reactions. HCE was 

placed in Irritation Category III (mild or slight irritation). 

4.4.6. Eye Irritation

 Weeks and Thomasino (1978) applied a single, 24-hour dose of 100 mg dry HCE to one 

eye of six male New Zealand White rabbits. Moderate corneal damage, iritis, and conjunctivitis 

was observed in 5/6 rabbits 24, 48, and 72 hours after exposure. No effects were observed 

7 days after exposure. HCE was placed in Irritation Category II for eye effects (corneal opacity 

reversible within 7 days or persisting for 7 days).
 DRAFT – DO NOT CITE OR QUOTE 44

4.5. MECHANISTIC DATA AND OTHER STUDIES IN SUPPORT OF THE MODE OF 

ACTION 

4.5.1. Genotoxicity 

 In vivo genotoxicity studies have not been performed in humans exposed to HCE. In 

vivo exposure to animals resulted in predominantly nonpositive results. Similarly, in vitro 

genotoxicity studies conducted in microorganisms, cultured mammalian cells, and insects 

(Table 4-15) were largely nonpositive both in the presence and absence of exogenous metabolic 

activation. Ashby and Tennant (1988) examined genotoxic carcinogenesis in a set of 

222 chemicals tested in rodents by NCI/NTP; HCE did not induce mutagenicity in Salmonella 

typhimurium reverse mutation tester strains. NTP’s technical report on the toxicity and 

carcinogenicity of HCE in F344/N rats concluded that HCE (purity >99%) was not significantly 

genotoxic, and that the increased incidence of tumors occurred through a mechanism other than 

one involving the induction of mutations (NTP, 1989). In an examination of putative 

“nongenotoxic” carcinogens on the basis of their reported mutagenicity per se (the ability to 

induce alterations in DNA structure or content, i.e., gene mutation, chromosomal aberrations 

[CAs], or aneuploidy), HCE was categorized as having insufficient data for evaluation (Jackson 

et al., 1993). Studies conducted by Lohman and Lohman (2000) considering DNA damage, 

recombination, gene mutation, sister chromatid exchange (SCE), micronuclei (MN), CA, 

aneuploidy, and cell transformation as endpoints indicate that the genetic activity profile for 

HCE is predominantly nonpositive. However, some positive findings have been reported in 

assays for gene conversion, somatic mutation/recombination, DNA adducts, and SCEs. 

Table 4-15. Summary of genotoxicity studies of HCE

Test system Genetic 

endpoint 

Strain/cells Results Reference Comments 

In vitro tests 

TA98; TA100; 

TA1535; TA1537; 

TA1538 

– (±S9)a

Simmon and 

Kauhanen 

(1978) 

TA98; TA100; 

TA1535; TA1537; 

TA1538 

– (±S9)a Weeks et al. 

(1979) 

TA98; TA100; 

TA1535; TA1537 – (±S9)a Haworth et al. (1983) 

Liquid 

preincubation 

protocol 

Gene reversion/

S. typhimurium 

TA98; TA100; 

TA1535; TA1537 – (±S9)a Milman et al. 

(1988) 

Forward 

mutations BA13 – (±S9)a Roldán-Arjona 

et.al. (1991) 

Liquid 

preincubation 

protocol 

Bacterial 

SOS test TA1535/pSK1002 – (±S9)a Nakamura et 

al. (1987) 

umu test; 

Liquid 

preincubation 

protocol 

Mammalian CAs Chinese hamster ovary 

(CHO) – (±S9)a

 Galloway et 

al. (1987) 
 DRAFT – DO NOT CITE OR QUOTE 45

Table 4-15. Summary of genotoxicity studies of HCE

Test system Genetic 

endpoint 

Strain/cells Results Reference Comments 

SCEs CHO – (–S9)a

, 

+ (+S9)a

Galloway et al. 

(1987) 

HCE 

precipitation 

at doses 

causing 

positive 

results 

MN AHH-1 – Doherty et al. 

(1996) 

Human cell 

line 

 MCL-5 – Doherty et al. 

(1996) 

Human cell 

line 

 h2E1 – Doherty et al. 

(1996) 

Human cell 

line 

Cell 

transformation BALB/c-3T3 – Milman et al. (1988) 

DNA adduct 

formation 

(nonhuman) 

Wistar rats, calf thymus 

DNA 

+ DNA 

binding in 

liver, kidney, 

lung, and 

stomach 

Lattanzi et al. 

(1988) 

DNA adducts 

not identified 

BALB/c mice, calf 

thymus DNA 

+ DNA 

binding in 

liver, kidney, 

lung, and 

stomach 

Lattanzi et al. 

(1988) 

DNA adducts 

not identified 

S. cerevisiae D3 – (±S9) a

Simmon and 

Kauhanen 

(1978) 

S. cerevisiae D4 – (±S9) a Weeks et al. 

(1979) 

Mitotic 

recombination 

S. cerevisiae D7 – (±S9) a Bronzetti et al. 

(1990, 1989) 

Fungi 

Aneuploidy Aspergillus nidulans P1 

diploid – 

Crebelli et al. 

(1995,1992, 

1988) 

In vivo tests 

Rat liver foci Osborne-Mendel – (initiation) 

+(promotion) 

Milman et al. 

(1988) 

Initiation or 

promotion 

protocols Rat DNA adduct 

formation 

(Nonhuman) 

Wistar rats 

Weakly + 

DNA binding 

in liver 

Lattanzi et al. 

(1988) 

Adducts not 

identified 

Micronucleus 

induction CD-1 mice – Crebelli et al. (1999) 

Replicative DNA 

synthesis B6C3F1 mice + 

Yoshikawa 

(1996); 

Miyagawa et 

al. (1995) 

Hepatic cell 

Mice proliferation 

 BALB/c mice 

Moderately + 

DNA binding 

in liver 

Lattanzi et al. 

(1988) 

Adducts not 

identified 

Isolated human 

lymphocytes + (±S9) Tafazoli et al. 

Human (1998) 

Lymphocytes DNA strand 

breaks 

Human lymphocyte 

cultures – 

Tafazoli et al. 

(1998) Comet assay 
 DRAFT – DO NOT CITE OR QUOTE 46

Table 4-15. Summary of genotoxicity studies of HCE

Test system Genetic 

endpoint 

Strain/cells Results Reference Comments 

Drosophila Mitotic 

recombination Drosophila Weakly + Vogel and Nivard (1993) 

Eye mosaic 

assay 

a

S9 is a supernatant fraction from 9000 ×g centrifugal spin. 

 Using the standard Ames assay for reversion of S. typhimurium histidine tester strains 

(TA1535, TA1537, TA1538, TA98, and TA100), Simmon and Kauhanen (1978) found HCE to 

be nonmutagenic at concentrations of 5,000 or 10,000 μg HCE/plate (purity not specified), both 

in the absence and presence of an exogenous Aroclor 1254-stimulated rat liver supernatant 

fraction from 9,000 × g centrifugal spin (S9) metabolic activation system. HCE was reported to 

be slightly toxic at the 10,000 μg/plate concentration in the absence of the S9 mix. Weeks et al. 

(1979) also reported nonpositive results using the same tester strains, test protocol, solvent, and 

metabolic activation system over a concentration range of 0.1–500 μg HCE/plate (purity 99.8%). 

Further, as a part of NTP’s mutagenicity screening program, HCE was dissolved in 

dimethylsulfoxide (DMSO) and tested in two independent trials in two separate laboratories over 

a collective concentration range of 1–10,000 μg/plate. HCE was nonpositive for induction of 

reverse mutation in S. typhimurium (tester strains TA1535, TA1537, TA98, and TA100), with 

and without S9 metabolic activation (NTP, 1989; Haworth et al., 1983). Finally, HCE (purity 

>97%) was reported to be nonpositive in several Ames tester strains, both with and without S9 

from the Aroclor 1254-induced livers of both sexes of Osborne Mendel rats and B6C3F1 mice 

(Milman et al., 1988). 

 Using a different S. typhimurium indicator strain, BA13, in a liquid preincubation 

protocol of the Ara test, Roldán-Arjona et al. (1991) found HCE to be nonpositive. This 

bacterial assay examines the ability of an agent to induce forward mutations from L-arabinose 

sensitivity to resistance, and theoretically might be expected to detect a broader range of 

mutagens than reverse-mutation assays. HCE (purity 98%) was dissolved in DMSO and tested 

over a concentration range of 1.5–30.0 μmol/plate (355–7,102 μg/plate), both with and without 

rat liver S9 metabolic activation. Of the 16 chemicals tested in this study, HCE was the only one 

that did not demonstrate any toxicity, which the authors speculated was probably related to its 

low solubility in water. HCE (purity not specified) was nonpositive when assayed in the umu

test using S. typhimurium tester strain TA1535/pSK1002 (Nakamura et al., 1987). This study 

also employed a liquid preincubation protocol, and was conducted both with and without rat liver 

S9 metabolic activation, up to a concentration of 42 μg/mL (solvent, water or DMSO, was not 

specified for individual test agents). Although the available data indicate that HCE is not 

mutagenic to Salmonella, Legator and Harper (1988) suggested that this may be related to 

inadequate reductive dechlorination (i.e., if HCE is activated by metabolic pathways not present 

in the in vitro system used). 
 DRAFT – DO NOT CITE OR QUOTE 47

 HCE was assayed for its ability to induce mitotic recombination in tester strain D3 of the 

yeast S. cerevisiae (Simmon and Kauhanen, 1978). No significant activity over a concentration 

range of 0.1-5.0% HCE (1–50 mg/mL; purity not specified), with or without exogenous rat liver 

S9 metabolic activation was observed. In addition, nonpositive findings for HCE were reported 

by Weeks et al. (1979) using the S. cerevisiae D4 strain. 

Bronzetti et al. (1989) evaluated HCE (purity not specified) for mitotic gene conversion 

at the trp locus and reverse point mutation at the ilv locus in the S. cerevisiae D7 tester strain. 

Two-hour liquid suspension exposures were conducted both on a logarithmic growth phase 

culture having high levels of CYP450 metabolizing enzymes, and on stationary growth phase 

cultures either with or without exogenous liver S9 mix. Exposures were from 5 to 12.5 mM 

(1.2–3.0 mg/mL) and were reportedly limited by solubility. HCE was inactive for both gene 

conversion and reverse mutation in stationary cultures with or without S9, and for reverse 

mutation in the logarithmic culture. However, statistically significant (p ≤ 0.05–0.001) increases 

in revertant frequency of more than twofold over background were observed at every 

concentration (Bronzetti et al., 1989). 

 The ability of various halogenated hydrocarbons to induce aneuploidy in the P1 diploid 

strain of the mold Aspergillus nidulans has been reported (Crebelli et al., 1995, 1992, 1988). 

Liquid suspension exposures (3 hours) to concentrations of 0.0025–0.04% HCE 

(0.005-0.84 mg/mL; purity >98%) resulted in survival rates of 100-48%. Exposure to these 

concentrations did not induce mitotic malsegregation of chromosomes. 

 A number of studies have evaluated the effects of in vivo and in vitro HCE exposures on 

various cytogenetic endpoints in higher organisms (Crebelli et al., 1999; Tafazoli et al., 1998; 

Doherty et al., 1996; Vogel and Nivard, 1993; NTP, 1989; Galloway et al., 1987). Crebelli et al. 

(1999) utilized the mouse bone marrow micronucleus test to investigate the in vivo induction of 

micronucleated polychromatic erythrocytes (MNPCE) by 10 aliphatic halogenated hydrocarbons, 

including HCE. CD-1 mice (5/sex/concentration) were injected i.p. with HCE doses of 2,000 or 

4,000 mg/kg (purity >98%), representing approximately 40 and 70-80% of the LD50, 

respectively. Animals were sacrificed and bone marrow cells harvested at 24 and 48 hours post 

treatment. At least 5,000 polychromatic erythrocytes/animal were analyzed. HCE treatment 

induced clinical signs of general toxicity, but no significant increases in the frequency of 

MNPCE in any treated group. 

 Vogel and Nivard (1993) utilized a Drosophila eye mosaic assay to monitor genetic 

damage in somatic cells, predominantly interchromosomal mitotic recombination, caused by the 

exposure of larvae to various chemicals. In the case of HCE (3% ethanol solvent; purity not 

specified), adult flies of the C-1 cross were permitted to lay eggs for 3 days on food 

supplemented with 10 mM HCE. Examination for light spots in the normally colored eyes of the 

resulting flies revealed what the authors classified as a weak positive response for HCE—a 

reproducible increase of not more than a doubling of the spontaneous frequency at a dose 
 DRAFT – DO NOT CITE OR QUOTE 48

associated with toxicity. The authors suggested that the effect was unspecific and likely not 

related to genotoxicity. 

 HCE was evaluated for its ability to induce MN and DNA damage in isolated human 

lymphocytes from two donors (Tafazoli et al., 1998). Lymphocytes were exposed for 3 hours in 

the presence of exogenous metabolic activation (S9 mix), or for 48 hours in the absence of S9. 

Results using cells from one donor (“A”) were reported for HCE (purity >99%) for exposures of 

0.05–1.00 mM (0.012–0.24 mg/mL) in the presence of S9. Neither toxicity nor MN induction 

was evident. Cells from the other donor (“D”) were exposed to higher HCE concentrations of 

1-16 mM (a saturating concentration; 0.24–3.79 mg/mL), both with and without S9. Toxicity 

(measured as a significant decrease in the relative division index) was still not observed, but 

statistically positive results for percent cells with MN were recorded at HCE concentrations of 

1 and 8 mM (0.24 and 1.89 mg/mL, respectively) in the absence of S9 (12 and 11%, respectively, 

versus a control value of 5.5%, p < 0.05), and at 1 mM (0.24 mg/mL) in the presence of S9 

(19.8% versus a control value of 9%, p < 0.01). In the second part of the study, lymphocyte 

cultures exposed to test agents for 3 hours with and without S9 were assessed for DNA damage 

(breaks, alkali-labile sites) using the Comet assay. HCE did not affect the measured DNA 

damage parameters (tail length, fraction of total cellular DNA in the tail, and tail moment). 

 Doherty et al. (1996) examined in vitro induction of MN by HCE in three human cells 

lines with metabolic competence; lymphoblastoid AHH-1 (native CYP1A1 activity), MCL-5 

(transfected with cDNAs encoding human CYP1A2, 2A6, 3A4, 2E1, and microsomal epoxide 

hydrolase), and h2E1 (with cDNA for human CYP2E1). Exponentially growing cultures were 

exposed for approximately one cell cycle (18 hours for AHH-1, 24 hours for MCL-5 and h2E1) 

to 0, 0.01, 0.05, or 0.1 mM HCE (purity not specified; 0, 0.002, 0.012, or 0.024 mg/mL, 

respectively), and then processed for scoring of kinetochore-positive and kinetochore-negative 

MN. No MN formation was observed in any of the three cell lines in response to HCE exposure. 

However, MN induction was enhanced by exposure to an HCE metabolite, PERC, in h2E1 and 

MCL-5 cells. 

 Induction of CAs and SCEs in cultured Chinese hamster ovary (CHO) cells exposed to 

HCE was investigated as part of an NTP screening program for genotoxicity (NTP, 1989; 

Galloway et al., 1987). Concentrations for analysis were selected based on observations of cell 

confluence and mitotic cell availability. HCE concentrations (purity >99%) ranged from 10 to 

1,000 μg/mL (0.01–1.0 mg/mL). For both endpoints, linear regression was used to test for 

dose-response trends. For individual doses, induction of CA was considered significant if 

p values (adjusted by Dunnett’s method to correct for multiple dose comparisons) relative to 

controls were ≤0.05, while increases of SCEs/chromosome ≥20% over control values were 

considered significant. For CA, duration of exposure was 8–10 hours in the absence of S9 

metabolic activation and 2 hours in the presence of S9. For induction of SCEs, exposure was 26 

hours without S9 and 2 hours with S9 (followed by 24-hour incubation without HCE). CAs were 
 DRAFT – DO NOT CITE OR QUOTE 49

not observed in response to HCE exposure without S9. In the presence of S9, the first study 

(0.15–0.50 mg/mL HCE) did not induce CAs; however, the second study (0.20–0.40 mg/mL 

HCE) was judged equivocal due to a positive response at the low dose (15.0% cells with CA, 

versus 5.0% for the DMSO control). HCE (0.010–0.33 mg/mL) did not induce SCE in the 

absence of S9; however, positive results were obtained in the presence of S9 (0.10–1.0 and 

0.40-1.0 mg/mL HCE). 

 In vitro cell transformation studies were conducted to understand the effect of HCE in the 

process of chemical carcinogenesis. In the absence of exogenous metabolic activation, a 3-day 

exposure to concentrations of HCE (purity >97%) from 0.16 to 100.0 μg/mL 

(0.00016-0.100 mg/mL) failed to induce morphological cell transformation in BALB/c-3T3 

cells, as measured by the incidence of Type III foci (characterized by the authors as an 

aggregation of multilayered, densely stained cells that are randomly oriented and exhibiting a 

criss-cross array at the edge of the focus) (Milman et al., 1988; Tu et al., 1985). Milman et al. 

(1988) also examined the capacity of HCE to initiate and promote tumors in a rat liver foci assay. 

To assess initiation potential, 24 hours after partial hepatectomy, 10 young adult male OsborneMendel rats received the MTD of HCE in corn oil by gavage. Six days later, the animals 

received a 0.05% dietary exposure to the tumor promoter phenobarbital for 7 weeks. Following 

sacrifice, livers were examined histopathologically for foci containing GGT, a putative 

preneoplastic indicator. To assess promotion potential, animals were initiated 24 hours after 

partial hepatectomy with an i.p. injection of 30 mg of the tumor initiator, diethylnitrosamine 

(DEN). Six days later, the animals received the MTD of HCE in corn oil by gavage, 

5 days/week for 7 weeks. The animals were sacrificed and their livers examined for the presence 

of GGT-positive foci. In these assays, HCE failed to demonstrate any initiating activity, but did 

show significant (p < 0.05) promoting capability (4.38 ± 1.04 GGT+ foci/cm2, versus 1.77 ± 0.49 

for the corn oil control). 

 Yoshikawa and colleagues reported on the activity of HCE and other putative 

nongenotoxic (i.e., Ames-nonpositive) mouse hepatocarcinogens in an in vivo–in vitro 

hepatocyte replicative DNA synthesis (RDS) assay (Yoshikawa, 1996; Miyagawa et al., 1995). 

Groups of 4–5 male B6C3F1 mice were exposed to single gavage doses of 0, 1,000, or 2,000 

mg/kg HCE (purity not specified). The hepatocytes were prepared at 24, 39, or 48 hours after 

exposure. The 1,000 mg/kg HCE-treated hepatocytes prepared 39 hours after exposure yielded a 

positive mean RDS response of 1.21 ± 0.46% (investigators noted that an RDS incidence rate of 

0.4% for any dose group was considered a positive response for the chemical). The remaining 

HCE groups were nonpositive with mean responses of 0.15–0.35%, while the solvent control 

mean was 0.26 ± 0.17%. 

4.5.2. In Vitro and Ex Vivo Studies Using Isolated Target Tissues/Organs or Cells 
 DRAFT – DO NOT CITE OR QUOTE 50

 A study using a rat liver foci assay (Milman et al., 1988, Story et al., 1986) found that 

HCE was a tumor promoter rather than an initiator. In vitro and in vivo assays were conducted 

to assess the ability of HCE to bind to DNA, RNA, and protein in several mouse and rat tissues 

(Lattanzi et al., 1988). This study reported binding of radiolabeled carbon to DNA, RNA, and 

protein was observed following [14C]HCE administration in both in vitro and in vivo assays in 

mice and rats (Lattanzi et al., 1988), suggesting that either HCE or its metabolites may bind to 

these macromolecules. The role of this binding in mediating HCE-induced toxicity was not 

further evaluated. 

Story et al. (1986) and Milman et al. (1988) conducted a rat liver foci assay to assess the 

initiation and promotion potential of HCE, along with eight other chlorinated aliphatics. Male 

Osborne-Mendel rats (10 rats/group) were given partial hepatectomies and then administered the 

initiation protocol or the promotion protocol. In the initiation protocol, the rats were 

administered by gavage the MTD of 2.1 mmol/kg (497 mg/kg) HCE (purity 98%), followed 

6 days later with 7 weeks of phenobarbital in the diet at 0.05%. Control rats were administered 

by gavage either corn oil (negative control) or 30 mg/kg DEN (positive control), followed by the 

phenobarbital treatment. In the promotion protocol, rats were dosed with 30 mg/kg DEN by i.p. 

injection, followed 6 days later with the MTD of 497 mg/kg HCE, 5 days/week for 7 weeks. 

Phenobarbital was administered (in the same manner as HCE) as a positive control. Control rats 

were either administered DEN or water, followed by corn oil for the promotion phase. Livers 

were removed and stained for GGT activity. Results from the initiation protocol were 

nonpositive, with only a small number of GGT+ foci (1.0 foci/cm2 at most). However, initiation 

with DEN followed by HCE or phenobarbital resulted in statistically significant increases in 

GGT+ foci (Table 4-16). Absolute and relative liver weights were increased by HCE in the 

promotion protocol. These results indicate that HCE is not an initiator in the rat liver foci assay, 

but is capable of promotion. 

Table 4-16. Number of enzyme-altered foci in rat liver of the promotion 

protocol

Total number of foci/cm2

Promotion treatment + DEN initiation - DEN initiation 

HCE 4.4 ± 1.0a

 0.1 ± 0.2 

Phenobarbital 3.9 ± 1.0a 0.3 ± 0.2 

Corn oil 1.7 ± 0.5 0.2 ± 0.2 

a

Statistically different from DEN + corn oil control group, p < 0.05 

Sources: Milman et al. (1988); Story et al. (1986). 

 Lattanzi et al. (1988) conducted in vivo and in vitro assays to assess the binding of 

[

14C] HCE (specific activity 14.6 mCi/mmol, radiochemical purity 98%) to nucleic acids in 

various organs from mice and rats following metabolic activation. For the in vivo studies, 
 DRAFT – DO NOT CITE OR QUOTE 51

6 male Wistar rats and 12 male BALB/c mice were injected i.p. with 127 μCi/kg HCE (purity 

98%). The animals were fasted and sacrificed 22 hours after injection. The organs (liver, 

kidney, lung, and stomach) were removed, pooled, and processed to obtain DNA, RNA, and 

proteins. The in vitro studies examined microsomal and cytosolic fractions from these same 

organs. The incubation mixture included 2.5 μCi [14C] HCE, 1.5 mg calf thymus DNA or 

polyribonucleotide, 2 mg microsomal proteins (plus 2 mg NADPH), and/or 6 mg of cytosolic 

proteins (plus 9.2 mg GSH). Coenzymes were not utilized with the controls. Measures for 

binding to macromolecules were determined by the presence of radiolabeled carbon from 

[

14C] HCE in the DNA, RNA, and protein. The presence of radiolabeled carbon may indicate 

HCE binding directly to the macromolecules or incorporation of radiolabeled carbon from 

intermediate metabolites into these macromolecules. 

In vivo binding data for HCE are presented in Table 4-17. Binding to macromolecules 

was interpreted by the presence of radiolabeled carbon; however, HCE-specific metabolites were 

not measured. In both rats and mice, binding values (in pmol HCE/mg) for RNA were 

consistently much greater than those for DNA or protein. Greater binding to RNA was observed 

in the kidneys of rats and mice (5-28 times greater) compared with the binding measured in the 

livers, lungs, and stomachs. DNA exhibited the lowest amount of HCE binding. Species 

differences were evident for all three macromolecule types (DNA, RNA, and protein) with the 

mouse exhibiting much higher levels (9 times greater) of covalent binding for DNA in the liver 

than the rat. The binding was 2 and 3 times greater for mice than rats with RNA and protein, 

respectively, from the liver. The binding to DNA was similar between species, but slightly 

greater in mice, for the kidney, lung, and stomach analyses. According to classifications 

reported by Lutz (1986, 1979), the covalent binding index values calculated on rat and mouse 

liver indicate weak (rat) to moderate (mice) oncogenic potency in HCE-treated rodents. 

Table 4-17. In vivo covalent binding of [14C]HCE to DNA, RNA, and proteins from 

rat and mouse organs 

Livera Kidneya Lunga Stomacha

(pmol/mg) Rat Mouse Rat Mouse Rat Mouse Rat Mouse 

DNA 

(CBIb) 

0.43 ± 0.05c 

(15.1)b 

3.92 ± 0.20d 

(140)b 0.42 0.50 0.14 0.35 0.26 0.37 

RNA 46.59 ± 7.23c 108.08 ± 21.57d 232.94 564.98 15.55 60.10 8.33 21.04 

Protein 4.94 ± 1.14c 14.99 ± 0.83d 2.59 4.91 0.89 3.42 0.80 2.41 

aData are from pooled organs from 6 male Wistar rats or 12 male BALB/c mice, except for liver (see indices). 

b

Covalent binding index (CBI) calculated according to Lutz (1986, 1979) as cited in Lattanzi et al. (1988). 

Classification of CBI values for oncogenic potency: strong, in the thousands; moderate, in the hundreds; weak, in 

the tens; and below one for nongenotoxic oncogens. 

cMean ± SE of six individual values. 

dMean ± SE of four values, each obtained from three pooled livers. 

Source: Lattanzi et al. (1988). 
 DRAFT – DO NOT CITE OR QUOTE 52

In vitro binding data for HCE are presented in Table 4-18. Liver microsomes from rats 

and mice catalyzed HCE binding to DNA at comparable levels. Kidney microsomes from rats 

and mice produced statistically significantly greater amounts of HCE binding to DNA when 

compared with controls. Kidney microsomes from mice had a threefold increase in HCE binding 

to DNA when compared to controls, while kidney microsomes from rats had a twofold increase 

in HCE binding to DNA when compared to controls. Microsomes from lung and stomach in 

both species did not display increased DNA binding activity over corresponding controls in the 

absence of coenzymes. Cytosolic fractions from all organs in mice and rats exhibited higher 

levels of HCE binding to DNA than microsomal fractions. Mouse liver cytosols produced much 

greater levels of HCE binding to DNA than rat liver cytosols. When both microsomal and 

cytosolic fractions were in the incubation mixture, HCE binding to DNA was decreased for liver 

and kidney. SKF 525-A, a nonspecific CYP450 inhibitor, caused a 50.5% decrease in HCE 

binding to DNA (data not included in report). The addition of GSH to the microsomal fractions 

also resulted in inhibition of HCE binding to DNA (data not included in report). When 

microsomal and cytosolic fractions were heat-inactivated, HCE binding to DNA was similar to 

control, providing further support that HCE binding to DNA is enzymatically catalyzed. This 

study provided evidence that HCE is metabolized by microsomal CYP450 enzymes and 

cytosolic glutathione transferases, and that DNA binding may be increased following HCE 

metabolism. 

Table 4-18. In vitro binding of [14C]HCE to calf thymus DNA mediated by 

microsomal and/or cytosolic phenobarbital-induced fractions of rat and 

mouse organs

Microsomes + NADPH Cytosol + GSH Microsomes + cytosol 

(+ NADPH, + GSH) 

Rat Mouse Rat Mouse Rat Mouse 

Liver 

Standarda

90.83 ± 5.31b 105.39 ± 7.80b 195.51 ± 21.44c 346.17 ± 18.91b 95.06 ± 6.29c 133.44 ± 2.42a 

Controlsa

55.19 ± 4.90 46.96 ± 4.19 92.96 ± 26.07 128.56 ± 8.92 52.85 ± 12.93 99.84 ± 8.06 

Kidney 

Standard 395.84 ± 78.58c 78.86 ± 6.85c 246.85 ± 35.39c 251.42 ± 45.38c 247.99 ± 3.40b ND 

Controls 136.26 ± 9.04 39.12 ± 5.34 88.82 ± 30.91 81.91 ± 9.93 144.61 ± 12.86 ND 

Lung 

Standard 125.60 ± 22.37 87.37 ± 7.90 126.65 ± 16.84b 168.52 ± 19.41b 234.26 ± 28.35b ND 

Controls 121.13 ± 16.54 86.10 ± 3.27 40.23 ± 7.34 60.44 ± 21.90 56.27 ± 5.32 ND 

Stomach 

Standard 94.41 ± 14.38 47.67 ± 17.00 289.58 ± 31.19b 228.74 ± 20.42b 76.79 ± 5.34b ND 

Controls 93.20 ± 15.24 47.12 ± 11.20 130.51 ± 4.01 51.52 ± 6.20 44.77 ± 2.28 ND 

aData (total DNA binding in pmol/mg) are reported as mean ± SE of three values; ND, not determined. Controls 

were conducted in the absence of coenzymes. 

b

Statistically different from control p < 0.01 

c

Statistically different from control p < 0.05 

Source: Lattanzi et al. (1988). 
 DRAFT – DO NOT CITE OR QUOTE 53

4.5.3. Structure Activity Relationships

 Several studies were conducted with the objective of defining structure activity 

relationships (SARs) of halogenated hydrocarbons and toxicity. NTP (1996) defined a group of 

chlorinated ethanes that resulted in hyaline droplet nephropathy in male F344/N rats and a group 

of halogenated ethanes that resulted in renal toxicity in the absence of hyaline droplet 

nephropathy. In a series of studies, Crebelli et al. (1995, 1992, 1988) evaluated chlorinated and 

halogenated hydrocarbons for their ability to induce chromosome malsegregation, lethality, and 

mitotic growth arrest in the mold A. nidulans. 

NTP (1996) conducted a 21-day oral toxicity study with halogenated ethanes in male 

F344/N rats. Chemicals under investigation were 1,1,1,2-tetrachloroethane, 

1,1,2,2-tetrachloroethane, pentachloroethane, 1,1,2,2-tetrachloro-1,2-difluoroethane, 

1,1,1-trichloro-2,2,2-trifluoroethane, 1,2-dichloro-1,1-difluoroethane, 1,1,1-trichloroethane, 

1,1,1,2-tetrabromoethane, 1,1,2,2-tetrabromoethane, pentabromoethane, and HCE (purity >98%). 

Groups of five male rats/dose were administered 0.62 or 1.24 mmol/kg-day of the halogenated 

ethane (for HCE, 146 and 293 mg/kg-day, respectively). Increased kidney weights and evidence 

of renal toxicity were observed in many of the rats administered halogenated ethanes; however, 

this was not always coincident with hyaline droplet nephropathy. Hyaline droplet nephropathy 

(assessed by Mallory-Heidenhain staining, which allows for greater sensitivity in evaluating 

hyaline droplets within the tubules of the kidney) was only observed in rats administered 

pentachloroethane, 1,1,1,2-tetrachloroethane, and HCE. RDS, indicated by PCNA labeling 

index, was increased in male rats administered chemicals that induced hyaline droplet 

nephropathy (pentachloroethane, 1,1,1,2-tetrachloroethane, and HCE) as well as 

pentabromoethane and 1,1,2,2-tetrachloroethane, compared with control rats. The increase in 

cell proliferation in the kidneys (as measured by the PCNA labeling index) observed with some 

of the halogenated ethanes that did not induce hyaline droplet nephropathy suggests the 

contribution of another toxic mechanism. NTP (1996) concluded that the capacity to induce 

hyaline droplet nephropathy in male rats was restricted to ethanes with four or more halogens, 

and only the chlorinated (compared with the fluorinated and brominated) ethanes were active. 

This study also predicted that if hyaline droplet nephropathy is the determining factor in the 

induction of renal tubule cell neoplasia, then chemicals such as bromo- or chlorofluoroethanes 

would be nonpositive for kidney neoplasia in 2-year cancer bioassays of male rats. 

 Crebelli et al. (1988) evaluated three chloromethanes and eight chlorinated ethanes 

(including HCE) for the induction of chromosome malsegregation in A. nidulans. Although 8 of 

the 11 compounds tested provided positive results including the 3 chloromethanes and 5 out of 

8 chlorinated ethanes, HCE was nonpositive for chromosome malsegregation induction. 

Analyses of relationships between biological and chemical variables indicate that the ability of a 

chemical to induce chromosome malsegregation was not related to any of the chemical 

descriptors examined, including molecular weight, melting point, boiling point, refractive index, 
 DRAFT – DO NOT CITE OR QUOTE 54

octanol/water partition coefficient, and the free energy of binding to biological receptors. 

Because of the similarity of the chemical descriptors between the positive chlorinated ethanes, 

aside from 1,1,1-trichloroethane which was nonpositive, the authors argue against a previous 

hypothesis that nonspecific interactions with hydrophobic cellular structures is the mechanism of 

aneuploidy induction (Onfelt, 1987). 

 Crebelli et al. (1992) evaluated the ability of 24 chlorinated aliphatic hydrocarbons to 

induce chromosome malsegregation, lethality, and mitotic growth arrest in the mold, A. nidulans. 

Data were combined with previous data on 11 related compounds (Crebelli et al., 1988) to 

generate a database for quantitative structure-activity relationship (QSAR) analysis. Physicochemical descriptors and electronic parameters for each chemical were included in the analysis. 

Out of the 24 chemicals, 19 were nonpositive for the induction of chromosome malsegregation; 

5 chemicals produced reproducible increases in the frequency of euploid whole chromosome 

segregants. HCE was nonpositive for the induction of chromosome malsegregation. QSAR 

analyses on these 35 chlorinated aliphatic hydrocarbons indicate that toxicity, such as the 

induction of lethality, is primarily related to steric factors (the spatial orientation of reactive 

centers within a molecule) and measures of the volume occupied by an atom or functional group 

(molar refractivity). Measures of molar refractivity are a function of temperature, index of 

refraction, and atmospheric pressure. Mitotic growth arrest was also primarily related to molar 

refractivity. However, aneugenic activity was related to both molar refractivity and electronic 

factors, such as the ease in accepting electrons (described by density and the energy of the lowest 

unoccupied molecular orbital). 

 These QSAR studies (Crebelli et al., 1992, 1988) were expanded to include 20 additional 

halogenated hydrocarbons (Crebelli et al., 1995). Chemicals in this study were also assayed for 

lipid peroxidation in rat liver microsomes, and the authors reported a partial coincidence was 

found between the ability of a chemical to initiate lipid peroxidation and to disturb chromosome 

segregation at mitosis. This updated study concluded that electronic and structural parameters 

that determine the ease of homolitic cleavage of the carbon-halogen bond play a primary role in 

the peroxidative properties of haloalkanes. 

4.6. SYNTHESIS OF MAJOR NONCANCER EFFECTS

4.6.1. Oral 

 Table 4-19 summarizes the oral toxicity studies that have been reported in laboratory 

animals. The primary noncancer effects observed in these studies include decreased body weight 

or body weight gain, increased absolute and relative kidney weights, increased absolute and 

relative liver weights, various effects associated with renal tubule toxicity in the kidney, and 

hepatocellular necrosis. Developmental studies in rats did not consistently demonstrate fetal 

effects, especially in those cases where maternal toxicity was absent. 
 DRAFT – DO NOT CITE OR QUOTE 55

 Acute and short-term toxicity tests in animals reported liver necrosis and tubular 

nephrosis in male rabbits (Weeks et al., 1979), and evidence of kidney effects such as 

nephropathy with hyaline droplet formation and tubular cell regeneration in male rats (NTP, 

1996, 1989). Female rats in short-term toxicity tests displayed only decreased body weights at 

the LOAEL of 563 mg/kg-day with a NOAEL of 281 mg/kg-day (NTP, 1989). Oral LD50 values 

in rats ranged from 4,460 to 7,690 mg/kg (Weeks et al., 1979). 
 DRAFT – DO NOT CITE OR QUOTE 56

Table 4-19. Oral toxicity studies for HCE 

Species 

Dose 

(mg/kg-day)/ 

Duration 

NOAEL 

(mg/kg-day) 

LOAEL 

(mg/kg-day) Effect Reference 

New Zealand 

White rabbits, 

male (5/dose) 

0, 100, 320 or 

1,000 by oral; 

12 days 

100 320 

Increased liver and kidney 

weights; liver degeneration 

and necrosis; tubular 

nephrosis and 

nephrocalcinosis 

Weeks et 

al. (1979) 

F344/N rats 

(5/sex/dose) 

0, 140, 281, 

563, 1,125, or 

2,250 by 

gavage; 16 days 

Male: not 

established 

Female: 281 

Male: 140 

Female: 563 

Male: kidney effects 

(hyaline droplets, tubular 

cell regeneration) 

Female: decreased body 

weight 

NTP (1989) 

F344/N rats, 

male (5/dose) 

0, 146 or 293 by 

gavage; 21 days 

Not 

established 146 

Increased kidney weight, 

nephropathy (hyaline 

droplets, tubule 

regeneration, granular 

casts), effects on urinalysis 

parameters 

NTP (1996) 

F344/N rats 

(10/sex/dose) 

0, 34, 67, 134, 

268, or 536 by 

gavage; 

13 weeks 

Male: not 

established 

Female: 67 

Male: 34 

Female: 134 

Male: kidney effects in all 

dose groups 

Female: hepatocellular 

necrosis 

NTP (1989) 

F344 rats 

(10/sex/dose) 

0, 1, 15, or 62 

by diet; 16 

weeks 

Male: 1 

Female: 15 

Male: 15 

Female: 62 

Male: kidney atrophy, 

proximal tubule 

degeneration 

Female: kidney atrophy, 

tubule degeneration 

Gorzinski 

et al. 

(1985) 

OsborneMendel rats 

(50/sex/dose) 

0, 113, or 227 

by gavage; 78 

weeks 

Not 

established 113 Tubular nephropathy in both sexes 

NCI 

(1978), 

Weisburger 

(1977) 

B6C3F1 mice 

(50/sex/dose) 

0, 360, or 722 

by gavage; 91 

weeks 

Not 

established 360 Tubular nephropathy in both sexes 

NCI 

(1978); 

Weisburger 

(1977) 

F344/N rats 

(50/sex/dose) 

Male: 0, 7, or 14 

Female: 0, 57, 

or 114 by 

gavage; 

103 weeks 

Not 

established 

Male: 1 

Female: 57 

Male: Tubular nephropathy, 

renal tubular hyperplasia 

Female: Tubular 

nephropathy 

NTP (1989) 

Pregnant 

SpragueDawley rats 

(22/dose) 

0, 50, 100, or 

500 by gavage 

on GDs 6–16 

Maternal: 100 Maternal: 500 

Maternal: body weight 

decreased, increased mucus 

in nasal turbinates, 

subclinical pneumonitis 

Fetal: no effects 

Weeks et 

al. (1979) 

Pregnant 

Wistar rats 

(21/dose) 

0, 56, 167, or 

500 by gavage 

on GDs 7–17 

Maternal: 56 

Fetal: 167 

Maternal: 167 

Fetal: 500 

Maternal: decreased weight 

gain and motor activity 

Fetal: reduced body weight 

increased incidence of 

skeletal variations, 

decreased ossification 

Shimizu et 

al. (1992) 
 DRAFT – DO NOT CITE OR QUOTE 57

4.6.1.1. Nephrotoxicity 

Two short-term studies in F344 rats (NTP, 1996; NTP, 1989, 16-day study) reported 

nephrotoxic effects at all administered doses in male rats. The formation of hyaline droplets 

accompanied by cell regeneration and eosinophilic granular casts was observed in the renal 

tubules of male rats administered 140-563 mg/kg-day HCE (NTP, 1989). Female rats did not 

exhibit any renal toxicity. In a 21- day study by NTP (1996), male rats exhibited increased 

absolute and relative kidney weights, tubular regeneration and granular casts, and increased 

labeling index in kidneys at doses of 146 and 293 mg/kg-day HCE. Tubular nephrosis, and to a 

minimal degree, tubular nephrocalcinosis were observed in the kidney of male New Zealand 

White rabbits administered 320 and 1,000 mg/kg-day (but not 100 mg/kg-day) HCE (Weeks et 

al., 1979). Compared with rabbits, the rats were more sensitive to renal effects induced by HCE. 

A gender-specific response was demonstrated in the male rats (NTP, 1989). However, the use of 

only male rats (NTP, 1996) and male rabbits (Weeks et al., 1979) in the other two studies makes 

it difficult to evaluate if the renal effects observed were gender specific. 

Subchronic exposure in male F344/N rats demonstrated kidney effects including hyaline 

droplet formation, tubular regeneration, and tubular casts in male rats administered HCE ranging 

from 34 to 536 mg/kg-day in the NTP (1989) 13-week study. Males in the 536 mg/kg-day dose 

group also exhibited renal papillary necrosis and degeneration and necrosis of renal tubule 

epithelium (NTP, 1989). Female rats did not display these kidney effects. These results suggest 

a sex-specific difference in HCE toxicity. Another study (Gorzinski et al., 1985) in F344 rats 

reported slight hypertrophy and dilation of the renal tubules in males and renal tubule atrophy 

and degeneration in male and female rats. Evidence of kidney effects in female rats consisted of 

very slight renal tubular atrophy and degeneration observed histopathologically at the highest 

dose tested. EPA considered the NOAEL and LOAEL for male rats as 1 and 15 mg/kg-day, 

respectively, while the corresponding values in the females were 15 and 62 mg/kg-day, 

indicating greater sensitivity of the males to the renal effects of HCE. These data and tissue 

distribution information (see Section 3.2 Distribution) show that the male kidney accumulated 

higher HCE concentrations than the female kidney, indicating that the kidney is the primary 

target organ following oral exposure to HCE and there are potential gender differences in the 

distribution and metabolism of HCE. Consequently, male rats are likely more sensitive to the 

nephrotoxicity of HCE than female rats. Additionally, Gorzinski et al. (1985) is the only study 

of either short-term or subchronic duration to report renal effects in female rats. 

Chronic toxicity tests were conducted by the NTP on F344/N rats and by NCI on 

Osborne-Mendel rats and B6C3F1 mice (NTP, 1989; NCI, 1978). NTP (1989) administered 

much lower doses of HCE (7 and 14 mg/kg-day in males; 57 and 114 mg/kg-day in females) to 

the F344 rats compared with the Osborne-Mendel rats (113 and 227 mg/kg-day) in the NCI 

(1978) study. In the NTP (1989) chronic study, nephropathy (characterized as tubular cell 

degeneration and regeneration, dilation and atrophy, glomerulosclerosis, interstitial fibrosis, and 
 DRAFT – DO NOT CITE OR QUOTE 58

chronic inflammation) was observed in both male and female rats. In the case of the male rats, 

the response was roughly equivalent across the control and treated groups with nephropathy in 

more than 94% of animals. The high incidence of nephropathy observed in control rats could 

result of a spontaneous syndrome known as chronic progressive nephropathy (CPN) that is 

associated with aged rats, especially F344 and Osborne-Mendel strains (see Section 4.7.3.2.1 for 

additional discussion). To examine the effects of chronic HCE exposure separate from CPN, the 

nephropathy incidence in terms of severity was considered. The severity was increased in the 

treated male rats compared with the controls. In considering severity, the increases in incidence 

of male nephropathy (that was of moderate or marked severity) were 18/50 (36%), 24/50 (48%), 

and 30/50 (60%) in the control, 7, and 14 mg/kg-day dose groups, respectively. In females, both 

the incidence (44% of controls and approximately 84% of treated) and severity of nephropathy 

were dose-related. When considering the severity, incidences of female nephropathy (that were 

of mild or moderate severity) were 12/50 (24%), 25/50 (50%), and 32/50 (64%) in the control, 

57, and 114 mg/kg-day dose groups, respectively. 

Dose-related increases (30 and 64% in 7 and 14 mg/kg-day, respectively) in linear 

mineralization of the renal papillae and treatment-related increases (14% in 7 and 14 mg/kg-day) 

in hyperplasia of pelvic transitional epithelium in the kidney were observed in the male rats. In 

females, an increased incidence of mineralization was only noted at the low dose (44% at 

57 mg/kg-day compared with 28% in controls). The low dose for the females was 8 times 

greater than that for the males yet the signs of nephropathy were more severe in the males. 

In the NCI (1978) study, Osborne-Mendel rats of both sexes displayed chronic 

inflammatory kidney lesions in both control and treated groups, although tubular nephropathy 

(characterized by degeneration, necrosis, and the presence of large hyperchromatic regenerative 

epithelial cells) was observed only in the HCE-exposed male and female rats. There were 

dose-related increases in incidences of nephropathy in males (45 and 66%, respectively) and 

females (15 and 59%, respectively) administered 113 and 227 mg/kg-day HCE. The chronic 

toxicity test in B6C3F1 mice (NCI, 1978) is the only study conducted in this species. Male mice 

experienced low survival in the control and 360 mg/kg-day (low-dose) groups. Chronic kidney 

inflammation was observed in 67 and 80% of males in the vehicle and untreated control groups, 

respectively, as well as in 66 and 18% of the 360 and 722 mg/kg-day HCE males. The report did 

not provide an explanation for the large response in the control and low dose mice and the 

relatively small response in the high dose group. Female mice exhibited chronic kidney 

inflammation only in vehicle controls (15%) and the high dose group (2%). Tubular 

nephropathy was observed in both dose groups of both sexes at high incidences (92-100%), and 

was characterized by degeneration of convoluted tubule epithelium with some hyaline casts. 

Enlarged dark staining regenerative tubular epithelium was also observed, with the kidney 

exhibiting infiltration of inflammatory cells, fibrosis, and calcium deposition. The response in 
 DRAFT – DO NOT CITE OR QUOTE 59

the treated male and female mice compared with the absence of nephropathy in the controls 

suggests that the doses used in this study were too high. 

The available information for HCE-induced nephropathy in rats, mice, male rabbits, and 

sheep indicates that the male rat is the most sensitive sex/species to the renal toxicity of HCE. 

Limited, if any, information is available for species other than the rat; however, the doses that 

elicited toxic responses in mice (NCI, 1978), male rabbits (Weeks et al., 1979), and sheep 

(Fowler, 1969) were at least 45-fold greater than the lowest dose (7 mg/kg-day; NTP, 1989) that 

induced a statistically significant response in rats. 

4.6.1.2. Hepatotoxicity

Short-term studies in rats (NTP, 1996), male rabbits (Weeks et al., 1979), and sheep 

(Fowler, 1969) reported hepatotoxicity at doses approaching ≥300 mg/kg-day. Male F344 rats 

exhibited significantly increased relative liver weights at the highest dose of 293 mg/kg-day. 

AST and NAG serum activities were also significantly higher than in controls. These effects 

were not observed at 146 mg/kg-day HCE (NTP, 1996). Liver degeneration and necrosis, 

including fatty degeneration, coagulation necrosis, hemorrhage, ballooning degeneration, 

eosinophilic changes, and hemosiderin-laden macrophages and giant cells were observed in male 

New Zealand White rabbits administered 320 and 1,000 mg/kg-day HCE (but not 

100 mg/kg-day); increasing in severity with increasing dose. Sheep given single oral doses of 

500-1,000 mg/kg of HCE exhibited plasma levels of GDH, SDH, and OCT that were increased 

twofold or more than levels in controls; indicating reduced hepatic function. 

Effects in the liver of animals treated with HCE were observed in male and female rats in 

two subchronic studies (NTP, 1989; Gorzinski et al., 1985). Liver weight increased in a 

dose-related fashion from the lowest dose (34 mg/kg-day) to the highest (536 mg/kg-day). 

Females were more sensitive than males; severity and statistical significance increased in 

females at doses lower than those eliciting toxicity in male rats. Hepatocellular necrosis was 

noted in females at doses ranging from 134 to 156 mg/kg-day and in males at the two highest 

doses, 268 and 536 mg/kg-day (NTP, 1989). Gorzinski et al. (1985) reported slight swelling of 

hepatocytes in control and treated males, although the two highest doses (15 and 62 mg/kg-day) 

exhibited dose-related increases in incidences of swelling. Other than a statistically significant 

increase (5%) in liver weight at 62 mg/kg-day HCE, the females were not affected. This is in 

contrast to the hepatocellular effects noted in female rats in the NTP study (NTP, 1989). 

However, the highest dose used by Gorzinski et al. (1985), 62 mg/kg-day, is below the 

67 mg/kg-day NOAEL for females of the NTP (1989) study, indicating that sufficient doses may 

not have been reached in Gorzinski et al. (1985) to cause hepatotoxicity in female rats. 

There were no liver effects observed in the animals administered HCE for chronic 

durations. The range of doses in the subchronic assay (0, 34, 67, 134, 268, and 536 mg/kg-day 

on F344 rats; NTP, 1989) encompassed the doses used in the chronic assays for female F344 rats 
 DRAFT – DO NOT CITE OR QUOTE 60

(57 and 114 mg/kg-day; NTP, 1989) and Osborne-Mendel rats (113 and 227 mg/kg-day; NCI, 

1978). Hepatocellular necrosis was observed in female rats in the subchronic, but not the 

chronic study. The LOAEL for female F344/N rat hepatocellular necrosis, 134 mg/kg-day, in 

the subchronic study (NTP 1989) occurred at a dose that exceeded the highest dose of the 

chronic study (NTP, 1989), suggesting that a sufficiently high dose may have not been achieved 

to elicit hepatocellular necrosis despite the longer exposure period. The NCI (1978) study in 

Osborne-Mendel rats was conducted with doses above the LOAEL for hepatocellular necrosis in 

female F344/N rats (NTP, 1989), but hepatocellular effects were not observed. Osborne-Mendel 

rats may not be as sensitive to HCE-induced hepatotoxicity as F344/N rats. The only study in 

mice (NCI, 1978; chronic) did not report any hepatotoxic effects other than the development of 

hepatocellular tumors. 

 HCE-induced liver effects were only observed in animals in short-term and subchronic 

studies. Female rats exhibited a greater sensitivity to liver effects as evidenced by the effects 

observed at lower doses compared with males (NTP, 1989). The implications of the slight 

swelling of hepatocytes in the absence of other histopathological effects at 15 and 62 mg/kg-day 

in male rats (Gorzinski et al., 1985) are unknown. Rabbits (males) and sheep demonstrated 

hepatic effects at doses at least fourfold greater than the lowest dose (67 mg/kg-day) that induced 

a statistically significant response in female rats. 

4.6.1.3. Developmental Toxicity

Two developmental studies in rats indicated that HCE-induced teratogenicity in the 

presence of maternal toxicity (Shimizu et al., 1992; Weeks et al., 1979). In the Shimizu et al. 

(1992) study, maternal rats gavaged with 167 and 500 mg/kg HCE displayed decreased motor 

activity. At the high dose, dams also exhibited piloerection and subcutaneous hemorrhage. 

Fetuses of the 500 mg/kg dose displayed decreased body weight, skeletal variations such as 

rudimentary lumbar ribs, and ossification effects, but no skeletal malformations were observed. 

The NOAEL for this study was 56 mg/kg for the dams and 167 mg/kg for the fetuses. In Weeks 

et al. (1979), maternal rats gavaged with 500 mg/kg HCE displayed pulmonary effects such as 

increased incidence of mucopurulent nasal exudates, upper respiratory tract irritation, and 

subclinical pneumonitis. The fetuses did not exhibit any skeletal or soft tissue anomalies. The 

maternal LOAEL and NOAEL were 500 and 100 mg/kg, respectively. 

4.6.1.4. Metabolite Toxicity

The potential metabolites of HCE are PERC and pentachloroethane (Fowler, 1969), 

which are subsequently metabolized to TCE, TCA, and/or trichloroethanol (see Figure 3-1). 

Exposure to these potential metabolites results in effects on the liver, kidneys, and nervous 

system, similar to effects observed following HCE exposure. Potential HCE metabolites PERC 

(Cal EPA, 2001), TCE (NTP, 1990, 1988; NCI, 1976), and TCA (Mather et al., 1990; Bull et al., 
 DRAFT – DO NOT CITE OR QUOTE 61

1990) are associated with liver effects following exposure. Kidney effects have been reported 

following exposure to putative HCE metabolites PERC (JISA, 1993; NTP, 1986; NCI, 1977), 

pentachloroethane (NTP, 1996, 1983), TCE (NTP, 1990, 1988; NCI, 1976), and TCA (Mather et 

al., 1990). Neurological effects are also reported following exposure to the putative HCE 

metabolites PERC (JISA, 1993; NTP, 1986; NCI, 1977), TCE, and pentachloroethane. 

 This qualitative comparison suggests that metabolites formed during HCE metabolism 

could contribute to the liver, kidney, and neurological effects observed in animals exposed to 

HCE. However, the available metabolism data for HCE do not allow for the conclusive 

identification of metabolites or provide quantitative information on how potential metabolites 

contribute to effects associated with HCE exposure. 

4.6.2. Inhalation

 Inhalation toxicity has only been evaluated in a single 6-week repeat exposure study in 

multiple species performed by Weeks et al. (1979). There is some uncertainty regarding the 

exposure to HCE vapor because HCE would remain a vapor only when surrounded by heated air. 

However, as soon as the hot HCE vapor was mixed with room temperature air, most (but not all) 

vapor in the airstream would condense into fine particles (a solid aerosol). The data from this 

study are summarized in Table 4-20. The study authors reported NOAELs and LOAELS for 

beagle dogs, guinea pigs, and rats of 48 ppm (465 mg/m3) and 260 ppm (2,517 mg/m3), 

respectively. Neurological effects, such as tremors and ataxia, were observed in beagle dogs and 

in pregnant and nonpregnant Sprague-Dawley rats. Rats and guinea pigs exhibited reduced body 

weight gain and increased relative liver weight. Male rats also displayed increased relative 

spleen and testes weights. Behavioral tests were conducted in male Sprague-Dawley rats at the 

same exposure concentrations, and no significant effects were observed. Overall, the 

information on the inhalation toxicity of HCE is limited. 
 DRAFT – DO NOT CITE OR QUOTE 62

Table 4-20. Inhalation toxicity studies with HCE 

Species Concentration 

(mg/m3)/durationa

NOAEL 

(mg/m3) 

LOAEL 

(mg/m3) 

Effect Reference 

Male beagle dogs 

(4/concentration) 

0, 145, 465, or 

2,517; 6 weeks 465 2,517 

Tremors, ataxia, 

hypersalivation, head 

bobbing, facial muscular 

fasciculations 

Weeks et al. 

(1979) 

Male Hartley guinea 

pigs 

(10/concentration) 

0, 145, 465, or 

2,517; 6 weeks 465 2,517 

Reduced body weight, 

increased relative liver 

weight 

Weeks et al. 

(1979) 

Sprague-Dawley rats 

(25/sex/concentration) 

0, 145, 465, or 

2,517; 6 weeks 465 2,517 

Males: reduced body 

weight gain, increased 

relative kidney, spleen, 

and testes weights 

Females: increased 

relative liver weight 

Weeks et al. 

(1979) 

C. Japonica (Japanese 

quail) 

(20/concentration) 

0, 145, 465, or 

2,517; 6 weeks 2,517 Not established No effects Weeks et al. (1979) 

Pregnant SpragueDawley rats 

(22/concentration) 

0, 145, 465, or 

2,517; GDs 6-16 

Maternal: 

465 

Maternal: 

2,517 

Maternal: tremors, 

decreased body weight 

Fetal: no effects 

Weeks et al. 

(1979) 

Male SpragueDawley rats 

(15/concentration) 

0, 145, 465, or 

2,517; 6 weeks 465 2,517 

Behavioral tests: 

avoidance latency and 

spontaneous motor 

activity 

Weeks et al. 

(1979) 

a

145, 465, and 2,517 mg/m3 correspond to concentrations reported by Weeks et al. (1979) as 15, 48, and 260 ppm, 

respectively. 

4.6.3. Mode-of-Action Information

 Reports on HCE-induced human health effects are limited and confounded by coexposure 

to multiple solvents or other toxicants (e.g., HCE-zinc oxide smoke). Studies that observed 

substantial HCE exposure in smoke bomb production workers were too small to provide 

definitive conclusions on health effects. 

Animal studies suggest that HCE is primarily metabolized to PERC and 

pentachloroethane by CYP450 enzymes of the liver, with likely subsequent metabolism to TCE. 

Metabolites identified in the urine include TCA, trichloroethanol, oxalic acid, dichloroethanol, 

dichloroacetic acid, and monochloroacetic acid. However, only 5% of a radiolabeled compound 

was measured in the urine, indicating that all of the urinary metabolites account for a small 

percentage of the dose. It is unknown whether unchanged HCE or its metabolites are responsible 

for the liver and kidney toxicities observed in animal studies. Only one study attempted to assess 

the extent of HCE metabolism in rats and mice and estimated that 24–29% of administered HCE 

is metabolized (Mitoma et al., 1985). This study did not quantify actual metabolite 

concentrations, so these estimations are of questionable accuracy. 

The mode of action for HCE-induced kidney toxicity is unknown. HCE-induced 

nephropathy has been observed in both sexes of rats and mice. Specifically, short-term assays in 
 DRAFT – DO NOT CITE OR QUOTE 63

male rats showed nephropathy characterized by hyaline droplet accumulation and increased 

incidences of tubule regeneration and granular casts (NTP, 1996, 1989). Cell proliferation of 

kidney sections using PCNA labeling analysis was also increased (NTP, 1996). Subchronic and 

chronic animal bioassays confirmed these renal effects (NTP, 1989; Gorzinski et al., 1985; NCI, 

1978). Chronic inflammatory kidney lesions and tubular nephropathy were observed in rats, and 

tubular nephropathy was also observed in mice (NCI, 1978). The mode of action for 

nephropathy is unknown. Some data suggest an α2u-globulin mode of action could contribute to 

hexachloroethane-induced nephropathy. However, there is insufficient evidence to conclude that 

the kidney effects observed following HCE exposure (NTP, 1989) are related to an α2u-globulin 

mode of action for the following reasons: (1) the lack of α2u-globulin immunohistochemical data 

for HCE-induced nephrotoxicity and carcinogenicity, (2) the hyaline droplet accumulation is 

caused by excessive protein load that may not be exclusively related to α2u-globulin 

accumulation, and (3) the existence of renal toxicity in female rats and male and female mice 

indicates that the nephrotoxic effects are not limited to an α2u-globulin-induced sequence of 

lesions. It is also possible that advanced chronic progressive nephropathy (CPN), an age-related 

renal disease of laboratory rodents that occurs spontaneously, may contribute to the observed 

nephrotoxicity following HCE exposure. However, changes in severity of the nephropathy were 

observed to be greater in male rats exposed to HCE compared to controls indicating that HCE 

exposure exacerbated effects in the kidney. Additionally, HCE-exposed male rats demonstrated 

dose-dependent increases in incidence of mineralization of the renal papillae and hyperplasia of 

pelvic transitional epithelium. Neither of these effects increased in a dose-related manner in the 

controls or the HCE-exposed female rats, suggesting that CPN is not solely responsible for the 

nephropathy observed by NTP (1989). Lattanzi et al. (1988) conducted in vivo binding studies 

concluding that HCE could bind to DNA, RNA, and protein in the kidney. 

The liver has been demonstrated to be a target organ in several animal species. Sheep 

(Fowler, 1969) and male rabbits (Weeks et al., 1979) exhibited hepatotoxicity characterized by 

clinical chemistry parameters that indicated reduced hepatic function and showed 

histopathological findings including hepatocellular necrosis. Subchronic studies showed 

statistically significant decreases in relative and absolute liver weight (Gorzinski et al., 1985) and

statistically significant increases in relative liver weight and hepatocellular necrosis (NTP, 1989) 

in female F344/N rats. Studies of TCA (a potential metabolite of HCE) indicate that free radical 

generation may play a role in mediating toxicity particularly in the liver. However, no data are 

available demonstrating generation of free radicals following exposure to HCE and it is unknown 

whether unchanged HCE or its metabolites are responsible for the liver and kidney toxicities 

observed in animal studies. Town and Leibman (1984) reported lipid peroxidation (as indicated 

by a statistically significant increase in the formation of malondialdehyde and conjugated dienes) 

following treatment with HCE (8 mM). The authors suggested the involvement of a free radical. 
 DRAFT – DO NOT CITE OR QUOTE 64

However, this mode of action has not been explored or further addressed in the literature for 

HCE. 

The presence of radiolabeled carbon measured by in vivo binding studies suggested that 

HCE can bind to DNA, RNA, and protein (Lattanzi et al., 1988). Binding to macromolecules 

was interpreted by the presence of radiolabeled carbon; however, radiolabeled carbon may have 

been incorporated into these macromolecules from intermediary HCE metabolites. In the rat, 

higher levels of DNA, RNA, and protein binding were observed in the kidney and liver 

compared with the lung and stomach. The mouse demonstrated the highest levels of DNA and 

protein binding in the liver and RNA binding in the liver and kidney. Studies using CYP450 

indicate that HCE must be metabolized to reactive intermediates prior to binding to 

macromolecules. Therefore, renal toxicity and hepatotoxicity may also involve HCE binding to 

DNA, RNA, or protein, resulting in cytotoxicity and contributing to the cytotoxic damage from 

radicals. 

The neurological effects observed in beagle dogs (Weeks et al., 1979) and sheep (Fowler, 

1969; Southcott, 1951) are commonly observed effects of chlorinated hydrocarbons. These 

effects have not been extensively studied for HCE, and data are inadequate to determine a mode 

of action. 

4.7. EVALUATION OF CARCINOGENICITY 

4.7.1. Summary of Overall Weight of Evidence

Under EPA’s Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a), HCE is 

“likely to be carcinogenic to humans” based on data from oral cancer bioassays in F344/N rats 

and B6C3F1 mice (NCI, 1978; NTP, 1989). No human data are available to assess the 

carcinogenic potential of HCE. NTP (1989) reported dose-dependent increases (statistically 

significant at the high dose) in the combined incidence of adenoma or carcinoma and increases 

(statistically significant at the low dose) in the incidence of pheochromocytomas in male F344/N 

rats. Tumors were not observed in the female F344/N rats in the NTP (1989) study. In addition, 

NCI (1978) observed statistically significant increases in the incidence of hepatocellular 

carcinomas in male and female B6C3F1 mice. The male rats demonstrated a statistically 

significantly increased tumor response for hepatocellular carcinomas that was dose-related. The 

female mice displayed a statistically significantly elevated incidence of hepatocellular 

carcinomas at both doses, although no dose-related increase in tumor response was evident. The 

Osborne-Mendel rats in the NCI (1978) study did not provide consistent evidence of 

carcinogenicity. HCE was shown to be a promoter, but not an initiator, in an Osborne-Mendel 

rat liver foci assay (Milman et al., 1988; Story et al., 1986). Binding of radiolabeled carbon to 

DNA, RNA, and protein following administration of [14C]HCE was observed in both in vitro and 

in vivo assays in mice and rats (Lattanzi et al., 1988). 
 DRAFT – DO NOT CITE OR QUOTE 65

U.S. EPA’s Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a) indicate that 

for tumors occurring at a site other than the initial point of contact the weight of evidence for 

carcinogenic potential may apply to all routes of exposure that have not been adequately tested at 

sufficient doses. An exception occurs when there is convincing toxicokinetic data that 

absorption does not occur by other routes. Information available on the carcinogenic effects of 

hexachloroethane via the oral route demonstrates that tumors occur in tissues remote from the 

site of absorption. Information on the carcinogenic effects of hexachloroethane via the 

inhalation and dermal routes in humans or animals is absent. Based on the observance of 

systemic tumors following oral exposure, and in the absence of information to indicate 

otherwise, it is assumed that an internal dose will be achieved regardless of the route of 

exposure. Therefore, hexachloroethane is “likely to be carcinogenic to humans” by all routes of 

exposure. 

4.7.2. Synthesis of Human, Animal, and Other Supporting Evidence

 There are currently no data from human studies pertaining to the carcinogenicity of HCE. 

NTP (1989) conducted a chronic toxicity/carcinogenicity bioassay in F344/N rats. Groups of 

50 male rats/dose were administered TWA doses of 7 and 14 mg/kg-day of HCE (purity >99%) 

by corn oil gavage, 5 days/week for 103 weeks. Groups of 50 female rats/dose were 

administered, by corn oil gavage, 5 days/week for 103 weeks, TWA doses of 57 and 

114 mg/kg-day. Male rats exhibited a dose-related, statistically significant increase in the 

incidence of combined renal adenomas or carcinomas at the highest dose. Combined renal 

adenomas or carcinomas were observed in 2, 4, and 14% of controls, 7, and 14 mg/kg-day males, 

respectively. No HCE-related renal tumors were observed in female rats. The combined 

incidence of all three types of pheochromocytomas (benign, malignant, and complex 

pheochromocytomas) was statistically significantly increased in males treated with 7 mg/kg-day 

HCE (62%) and increased in males treated with 14 mg/kg-day (43%) when compared with 

vehicle controls (30%) and historical controls in the study laboratory (75/300; 25 ± 7%) and in 

NTP studies (543/1,937; 28 ± 11%). No HCE-related adrenal gland tumors were observed in 

female rats. 

 The NCI (1978; Weisburger, 1977) conducted a chronic toxicity/carcinogenicity bioassay 

in Osborne-Mendel rats. HCE (purity >98%) at doses of 0, 250, or 500 mg/kg-day was 

administered by corn oil gavage to 50 rats/sex/dose for 5 days/week for 78 weeks. Following 

termination of exposure, rats were observed for 33–34 weeks for a total duration of 

111-112 weeks. Twenty rats/sex were used for the untreated and vehicle controls. Starting in 

week 23, rats in the exposure groups began a 5-week cyclic rotation that involved 1 week 

without exposure followed by dosing for 4 weeks. After adjustment from 5 days/week for 78 

weeks, with the 5-week cyclic rotation for part of the time, to continuous exposure over the 

standard 2 years for a chronic bioassay, the TWA doses were 113 and 227 mg/kg-day. Mortality 
 DRAFT – DO NOT CITE OR QUOTE 66

was increased in the 113 and 227 mg/kg-day males with survival rates of 24/50 (48%) and 19/50 

(38%), respectively, compared with 14/20 (70%) in the untreated controls. Survival rates for the 

female rats were 14/20 (70%) for both the untreated and vehicle controls, and 27/50 (54%) and 

24/50 (48%) for the 113 and 227 mg/kg-day dose groups, respectively. 

All of the tumor types observed had been encountered previously as spontaneous lesions 

in the Osborne-Mendel rat and no statistical differences in frequencies were observed between 

treated and control rats. NCI concluded that there was no evidence of carcinogenicity in this rat 

study. Notably, the doses used in the Osborne-Mendel rats of the NCI (1978) study were 

approximately 16 times greater than those doses administered to F344 male rats by NTP (1989). 

A B6C3F1 mouse study conducted by NCI (1978; Weisburger, 1977), HCE (purity 

>98%) was administered by corn oil gavage at TWA doses of 360 and 722 mg/kg-day for 

5 days/week for 78 weeks, followed by 12–13 weeks of an observation period (total 91 weeks). 

Survival rates in males were 5/20 (25%), 1/20 (5%), 7/50 (14%), and 29/50 (58%) in the vehicle 

control, untreated control and the 360 and 722 mg/kg-day dose groups, respectively. Survival 

rates in females were 80, 85, 80, and 68% in vehicle control, untreated control, 360 and 722 

mg/kg-day groups, respectively. Both male and female mice exhibited statistically significantly 

increased incidences of hepatocellular carcinomas. The treated males demonstrated an increased 

tumor response for hepatocellular carcinomas that was dose-related: 30 and 63% in the 360 and 

722 mg/kg-day dose groups, respectively, compared with 10% in pooled vehicle controls and 

15% in matched vehicle controls. Females demonstrated an increased tumor response that was 

not dose related in that a higher incidence of hepatocellular carcinomas occurred at the low dose 

(40%) compared with the high dose (31%); pooled vehicle and matched vehicle controls had 

incidences of 3 and 10%, respectively. NCI concluded that HCE was carcinogenic in both sexes 

of B6C3F1 mice. 

 Evidence of HCE’s promotion (following treatment with DEN), but not initiation, 

potential was observed in the liver of male Osborne-Mendel rats administered a single gavage 

dose of 497 mg/kg HCE (Milman et al., 1988; Story et al., 1986). Lattanzi et al. (1988) reported 

in vivo and in vitro binding of HCE to DNA, RNA, and protein in mice and rats. In both rats and 

mice administered single i.p. injections of 127 μCi/kg [14C]HCE, in vivo covalent binding of 

HCE for RNA was consistently much greater than that for DNA or protein. DNA exhibited the 

lowest amount of HCE binding. Species differences were evident for all three macromolecule 

types (DNA, RNA, and protein), with the mouse exhibiting much higher levels (9 times greater) 

of covalent binding for DNA in the liver than the rat. The binding was 2 and 3 times greater for 

mice than rats with RNA and protein, respectively, from the liver. The binding was similar 

between species, but slightly greater in mice, for the kidney, lung, and stomach analyses. In 

vitro covalent binding to DNA was observed at comparable levels in liver microsomes from both 

rats and mice following exposure to HCE. Kidney microsomes from rats and mice produced 

statistically significantly greater amounts of DNA binding compared with controls, with greater 
 DRAFT – DO NOT CITE OR QUOTE 67

amounts of DNA binding from mice (threefold increase) compared with rats (twofold increase). 

Microsomes from the lungs and stomachs in both species did not display increased DNA binding 

activity over corresponding controls. 

Carcinogenicity associated with potential HCE metabolites 

 Potential metabolites of HCE include PERC, pentachloroethane, and TCE. 

Epidemiologic studies have reported associations between exposures to PERC and TCE and 

increased risks of several cancers including cancer of the lymphoid system, esophagus, cervix, 

bladder, kidney, and lung. PERC (NTP, 1986; NCI, 1977), pentachloroethane (NTP, 1983), and 

TCE (NTP, 1990; 1988; NCI, 1976) have also been evaluated for carcinogenicity in several 

chronic bioassays. Specifically, hepatocellular carcinomas have been reported for exposure to 

PERC, pentachloroethane, and TCE. Renal tubule adenomas have also been observed with 

exposure to PERC and pentachloroethane, with equivocal evidence for TCE. An increased 

incidence of pheochromocytomas was reported in studies of pentachloroethane, but not in studies 

of PERC or TCE. In addition, data suggest that some tumors are not shared between HCE and 

these potential metabolites (e.g., there is evidence of PERC-associated mononuclear cell 

leukemia in rats, but no reports of an association with HCE exposure). Based on the available 

data, the relative roles of the parental compound (HCE) and its metabolites in the carcinogenicity 

associated with exposure to HCE are unknown. However, the carcinogenicity of these HCE 

metabolites provides support for describing HCE as a rodent carcinogen. 

4.7.3. Mode-of-Action Information 

Hepatocellular and renal adenomas and carcinomas and pheochromocytomas were 

observed in rats and mice following oral exposure to HCE (NTP, 1989; NCI, 1978). The 

mode(s) of carcinogenic action of HCE in the liver, kidney, and adrenal gland is unknown. 

There are mode-of-action data suggesting that the induction of kidney tumors in male rats and 

liver tumors in male and female mice may involve the accumulation of α2u-globulin in the 

kidney and increased cytotoxicity, inflammation, and regenerative cell proliferation in the liver, 

respectively. 

4.7.3.1. Kidney Tumors 

Description of the Hypothesized Mode of Action 

Hypothesized mode of action 

The mode of action for the carcinogenic effects of HCE in the kidney is unknown. 

Specifically, the key events leading to development of kidney tumors in male rats exposed to 

HCE have not been fully characterized. Some of the experimental data suggest that development 

of kidney tumors in male rats following exposure to HCE may involve an α2u-globulin-mediated 

mode of action. Generally, kidney tumors observed in cancer bioassays are assumed to be 
 DRAFT – DO NOT CITE OR QUOTE 68

relevant for assessment of human carcinogenic potential. However, a number of chemicals have 

been shown to induce accumulation of α2u-globulin in hyaline droplets in male rat kidney. The 

α2u-globulin accumulation in hyaline droplets initiates a sequence of events that leads to renal 

nephropathy and, eventually, to renal tubular tumor formation. The phenomenon is unique to the 

male rats since female rats and other laboratory mammals administered the same chemicals do 

not accumulate α2u-globulin in the kidney and do not develop renal tubule tumors (U.S. EPA, 

1991b). 

The lack of α2u-globulin immunohistochemical data for HCE-induced nephrotoxicity and 

carcinogenicity supports the conclusion that there is insufficient evidence to establish the role of 

α2u-globulin in HCE-induced kidney tumors. Furthermore, reported renal toxicity in female rats 

and male and female mice exposed to HCE suggest a mode of action other than α2u-globulin 

nephropathy. In the absence of minimum information demonstrating the involvement of α2uglobulin processes, male rat renal toxicity/tumors is considered relevant for risk assessment 

purposes. 

Identification of Key Events 

The U.S. EPA (1991c) Risk Assessment Forum Technical Panel report provides specific 

guidance for evaluating chemical exposure-related male rat renal tubule tumors for the purpose 

of risk assessment, based on an examination of the potential involvement of α2u-globulin 

accumulation. 

The protein α2u-globulin is a member of a large superfamily of low-molecular-weight 

proteins and was first characterized in male rat urine. It has been detected in various tissues and 

fluids of most mammals, including humans. However, the particular isoform of α2u-globulin 

commonly detected in male rat urine is considered specific for the male rat; moreover, the urine 

and kidney concentrations detected in the mature male rat are several orders of magnitude greater 

than in any other age, sex, or species tested (U.S. EPA, 1991c). 

The hypothesized mode of action ascribed to α2u-globulin-associated nephropathy is 

defined by a progressive sequence of effects in the male rat kidney, often culminating in renal 

tumors. The involvement of hyaline droplet accumulation in the early stages of nephropathy 

associated with α2u-globulin-binding chemicals is an important difference from the sequence of 

events observed with classical carcinogens. The pathological changes that precede the 

proliferative sequence for classical renal carcinogens also include early nephrotoxicity (e.g., 

cytotoxicity and cellular necrosis) but no apparent hyaline droplet accumulation. Furthermore, 

the nephrotoxicity that can ensue from hyaline droplet accumulation is novel because it is 

associated with excessive α2u-globulin accumulation. This α2u-globulin accumulation is proposed 

to result from reduced renal catabolism of the α2u-globulin chemical complex and is thought to 

initiate a sequence of events leading to chronic proliferation of the renal tubule epithelium. The 

histopathological sequence of events in mature male rats consists of the following (see Table 4-
 DRAFT – DO NOT CITE OR QUOTE 69

21 for a summary of this sequence specific for HCE): 

• Excessive accumulation of hyaline droplets containing α2u-globulin in renal proximal tubules 

• Subsequent cytotoxicity and single-cell necrosis of the tubule epithelium 

• Sustained regenerative tubule cell proliferation 

• Development of intralumenal granular casts from sloughed cellular debris associated with 

tubule dilatation and papillary mineralization 

• Foci of tubule hyperplasia in the convoluted proximal tubules 

• Renal tubule tumors 
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Table 4-21. Nephrotoxic effects characteristic of α2u-globulin nephropathy observed in male and femal

HCE 

X = presence of effect NT = not tested 

NTP, 1989 

7 or 14 mg/kg-d (M); 

57 or 114 mg/kg-d (F)

103 wks 

NCI, 1978 

113 or 227 mg/kg-d

104 wks 

Gorzinski et al., 1985

1, 15, or 62 mg/kg-d 

16 wks 

NTP, 1989 

34, 67, 134, 268, 

or 536 mg/kg-d 

13 wks 

NTP, 1996 

146 or 293 mg/kg-d 

3 wks 

Study, dose, duration, 

and sex 

Males Females Males Females Males Females Males Females Males Females 

Accumulation of hyaline droplets X X NT Accumulation of α2uglobulin in hyaline 

droplets 

NT NT NT NT NT NT NT NT NT NT 

Necrosis/degeneration X X X X X X X NT Tubular regeneration X X X X X X NT Granular casts/dilatation X X X X X X X NT Papillary mineralization X NT Tubular hyperplasia X X NT 
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In addition to this histopathological sequence, EPA (1991c) provides more specific 

guidance for evaluating chemically induced male rat renal tubule tumors for the purpose of risk 

assessment. To determine the appropriateness of the data for use in risk assessment, chemicals 

inducing renal tubule tumors in the male rat are examined in terms of three categories: 

• The α2u-globulin sequence of events accounts for the renal tumors. 

• Other potential carcinogenic processes account for the renal tumors. 

• The α2u-globulin-associated events occur in the presence of other potential carcinogenic 

processes, both of which result in renal tumors. 

Therefore, it is important to determine whether the α2u-globulin process is involved and, if so, to 

what extent α2u -globulin-associated events, rather than other processes, account for the tumor 

increase. 

Determination of these elements requires a substantial database of bioassay data not only 

from male rats but also from female rats and mice, and such toxicity studies must demonstrate 

whether or not α2u-globulin processes are operative. In the absence of minimum information 

demonstrating the involvement of α2u-globulin processes, it should be assumed that any male rat 

renal toxicity/tumors are relevant for risk assessment purposes. 

As outlined in the U.S. EPA Risk Assessment Forum Technical Panel report (U.S. EPA, 

1991c), the following information from adequately conducted studies of male rats is used for 

demonstrating that the α2u-globulin process may be a factor in any observed renal effects―an 

affirmative response in each of the three categories is required. If data are lacking for any of the 

criteria in any one category, the available renal toxicity data should be analyzed in accordance 

with standard risk assessment principles. The three categories of information and criteria are as 

follows: 

• Increased number and size of hyaline droplets in the renal proximal tubule cells of treated 

male rats. The abnormal accumulation of hyaline droplets in the P2 segment helps 

differentiate α2u-globulin inducers from chemicals that produce renal tubule tumors by 

other modes of action. 

• Accumulating protein in the hyaline droplets is α2u-globulin. Hyaline droplet accumulation 

is a nonspecific response to protein overload, and, thus, it is necessary to demonstrate that 

the protein in the droplet is, in fact, α2u-globulin. 

• Additional aspects of the pathological sequence of lesions associated with α2u-globulin 

nephropathy are present. Typical lesions include single-cell necrosis, exfoliation of 

epithelial cells into the proximal tubular lumen, formation of granular casts, linear 

mineralization of papillary tubules, and tubule hyperplasia. If the response is mild, not 

all of these lesions may be observed. However, some elements consistent with the 

pathological sequence must be demonstrated to be present. 
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Experimental Support for the Hypothesized Mode of Action 

Strength, consistency, and specificity of association 

NTP (1989)—16-day study 

In a short-term exposure study, NTP (1989) administered 140, 281, 563, 1,125, or 

2,250 mg/kg-day HCE to F344/N rats via gavage for 16 days. All of the surviving HCE-exposed 

male rats exhibited hyaline droplets in the cytoplasm of the renal tubular epithelium. 

Additionally, male rats exposed to 140 and 281 mg/kg-day HCE demonstrated tubular cell 

regeneration and eosinophilic granular casts of cell debris in the tubule lumina at the 

corticomedullary junction. NTP (1989) did not report regeneration or granular casts in the 

surviving males of the 563 mg/kg-day dose group. NTP (1989) did not report the incidence or 

severity of the lesions observed in the treated males. None of the nephrotoxic effects were 

observed at any HCE dose in the female rats or in the controls. 

NTP (1996)—21-day study 

In a second short-term exposure study, NTP (1996) administered 146 or 293 mg/kg-day 

HCE by oral gavage to male F344/N rats for 21 days. Marked hyaline droplet accumulation was 

observed and categorized by severity in relation to controls. The hyaline droplet accumulation 

exhibited by HCE-exposed male rats was characterized as two severity grades above the control 

rats. A Mallory-Heidenhain stain allowed for greater sensitivity in evaluating hyaline droplets 

within the tubules of the kidney and further supported the presence of the hyaline droplets in the 

kidney tubules. Increased incidence of tubular regeneration (60 and 100% in the 146 and 

293 mg/kg-day dose groups, respectively) was also observed in male rats following HCE 

exposure. The severity of the tubular lesions was considered mild at both doses. Eosinophilic 

granular casts, of minimal to mild severity, were identified in the outer medullary tubules in male 

rats exposed to HCE: 80 and 60% in the 146 and 293 mg/kg-day HCE, respectively. There was 

a dose-related and statistically significant increase in the PCNA labeling index in HCE-treated 

male rats. The percentage of replicating proximal and distal tubule epithelial cells was increased 

5.7-fold over controls in the 146 mg/kg-day dose group and 9.2-fold over the controls in 

293 mg/kg-day dose group. The nephrotoxic effects reported by NTP (1996) were not noted in 

the control animals. Female rats were not included in this study; therefore, gender specificity 

was of the nephrotoxic effect were not examined. 

NTP (1989)—13-week study 

In a subchronic exposure study, NTP (1989) administered 34, 67, 134, 268, or 

536 mg/kg-day HCE via gavage to F344/N rats for 13 weeks. Kidney effects were reported in 

male rats from all dose groups exposed to HCE exhibited exposure-related kidney effects, 

although incidence data was only reported for the 34 mg/kg-day dose group. These kidney 
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effects were characterized by hyaline droplet formation in the renal tubular epithelium, 

eosinophilic granular casts of cell debris in the tubular lumina at the corticomedullary region 

(with associated tubular dilatation), and tubular cell regeneration. The severity of these lesions 

increased with HCE exposure dose, though the severity grades were not reported. Furthermore, 

as the HCE exposure dose increased, the animals developed additional lesions. Renal papillary 

necrosis and renal tubule epithelium degeneration and necrosis were observed in all 536 mg/kgday males (only the 5 male rats that died before the end of the study were analyzed 

microscopically). 

Urinalysis in male rats administered HCE showed fine and course granules, cellular casts, 

and epithelial cells, findings that were consistent with the histopathological changes observed in 

the male rats. Kidney weights of HCE-exposed males were increased 27, 37, 57, 73, and 57% in 

34, 67, 134, 268, and 536 mg/kg-day males, respectively (increases were statistically significant, 

compared with control kidney weights except the low-dose group). Female kidney weight was 

increased following HCE exposure: 16 and 32% (statistically significant) in the 268 and 

536 mg/kg-day dose groups, respectively. Treated females showed no other HCE-exposure 

related kidney effects. 

Gorzinski et al. (1985)—16-week study 

Gorzinski et al. (1985) observed dose-related levels of HCE in the kidneys of male F344 

rats fed 1, 15, or 62 mg/kg-day HCE for 16 weeks. HCE was also detected in the kidneys of 

female rats, although at much lower levels and did not increase proportionally with dose. Renal 

tubular atrophy and degeneration was observed in male rats: 20, 70, and 100% in the 1, 15, and 

62 mg/kg-day dose groups, respectively. These renal degenerative effects were also noted in 

10% of the male controls, although the authors noted that these lesions were graded as slight. 

Slight hypertrophy and/or dilation of the proximal convoluted tubules were noted in 10, 70, and 

100% of the HCE-exposed male rats in the 1, 15, and 62 mg/kg-day dose groups, respectively. 

Slight hypertrophy and dilation of the proximal convoluted tubules were not observed in the 

male control rats. Peritubular fibrosis was also noted in the high-dose group males. Renal 

tubular atrophy and degeneration were observed in 10, 20, and 60% of female rats in the 1, 15, 

and 62 mg/kg-day dose groups respectively. These lesions were seen in one female control rat 

(10%), although the authors characterized the severity grade of the lesions as very slight. 

Male rat sensitivity was evident in the histopathological changes seen in the HCE 

exposed male rats compared with the female rats. Renal effects were either observed in more 

male rats than female rats (statistical analyses were not reported) or did not occur in females. 

Additionally, kidney concentrations of HCE were much higher in male rats compared with 

female rats. Gorzinski et al. (1985) noted that the differences in HCE concentrations measured 

in male rat and female rat kidneys may explain the differences observed in the kidney effects 

(i.e., male sensitivity to HCE exposure). 
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NCI (1978)—78-week study 

 NCI (1978) conducted a carcinogenicity bioassay in Osborne-Mendel rats administered 

113 and 227 mg/kg-day HCE via gavage for 5 days/week for 78 weeks. Chronic inflammatory 

kidney lesions were observed in both control and HCE-exposed rats. Male rats exhibited chronic 

inflammation in the kidney, 75, 70, 65, and 50% of untreated control, vehicle control, 113, and 

227 mg/kg-day dose groups, respectively. Similarly, female rats showed an incidence of 

inflammatory lesions in 40, 20, 36, and 41% in the untreated control, vehicle control, 113, and 

227 mg/kg-day dose groups, respectively. The control and HCE-exposed male rats exhibited 

greater sensitivity to the chronic inflammation compared with the female rats. NCI (1978) noted 

that these lesions observed in the control and HCE-exposed animals of both sexes were 

characteristic of age-related renal lesions. Some renal lesions observed in older rats could be 

related to a spontaneous syndrome known as chronic progressive nephropathy (CPN). CPN is 

associated with aged rats, especially F344, Sprague-Dawley, and Osborne-Mendel strains. CPN 

is frequently more severe in males compared with females. Hard et al. (1993) reported the 

pathologic features attributed to CPN including: 

• Thickening of tubular and glomerular basement membranes 

• Basophilic segments of proximal convoluted tubules with sporadic mitoses indicative of 

tubule cell proliferation 

• Tubular hyaline casts of proteinaceous material originating the more distal portion of the 

nephron, mainly in the medulla, and later plugging a considerable length of the tubule 

• Focal interstitial aggregations of mononuclear inflammatory cells within areas of affected 

tubules 

• Glomerular hyalinization and sclerosis 

• Interstitial fibrosis and scarring 

• Tubular atrophy involving segments of proximal tubule 

• Occasional hyperplastic foci in affected tubules (chronically in advanced cases) 

• Accumulation of protein droplets in sporadic proximal tubules (in some advanced cases) 

 Several of the CPN pathological effects are similar to and can obscure the lesions 

characteristic of α2u-globulin-related hyaline droplet nephropathy (Hard et al., 1993). 

Additionally, renal effects of α2u-globulin accumulation can exacerbate the effects associated 

with CPN (U.S. EPA, 1991c). However, Webb et al. (1989) suggested that exacerbated CPN 

was one component of the nephropathy resulting from exposure to chemicals that induce 

α2u-globulin nephropathy. Male rat sensitivity has been noted with both CPN and α2u-globulin 

nephropathy. 
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With the exception of atrophy of proximal tubule, tubular cell proliferation, and hyaline 

casts of proteinaceous material, the histopathological effects associated with CPN are distinctive 

from those of α2u-globulin nephropathy. Additionally, the urinalysis and serum chemistry of 

CPN-rats show albuminuria, hypoalbuminemia, and hypocholesterolemia as well as increased 

serum creatinine and urea nitrogen levels, whereas these changes in α2u-globulin nephropathy are 

minimal (Hard et al. 1993). 

 NCI (1978) reported tubular nephropathy in HCE-exposed rats, but not in untreated or 

vehicle controls. Increased incidence of nephropathy described as tubular degeneration and 

necrosis and the presence of large hyperchromatic regenerative epithelial cells was observed in 

45 and 66% of male rats exposed to 113 and 227 mg/kg-day HCE, respectively. Female rats also 

exhibited tubular nephropathy following HCE exposure: in 18 and 59% in the 113 and 

227 mg/kg-day dose groups, respectively. In addition to the tubular nephropathy, observed 

effects overlying these lesions included focal pyonephritis, tubular ectasia, cast formation, 

chronic interstitial nephritis and fibrosis, and focal glomerulosclerosis. Renal tubular cell 

adenomas were observed in four male rats (11% incidence rate) exposed to 113 mg/kg-day HCE. 

Similar renal tumors were not observed in males from the high-dose group, males from the 

vehicle control, males from the untreated control, or female rats. NCI (1978) concluded that 

there was no evidence of HCE-exposure related carcinogenicity in Osborne-Mendel rats based 

on the lack of statistical significance and dose-response in the tumor incidence rate. However, it 

is possible that the truncated duration of HCE-treatment (78 weeks, cyclical) and the 

significantly accelerated mortality in the male rats did not allow enough time for the renal tubule 

tumors to develop. According to Goodman et al. (1980), the incidences of spontaneous renal 

tubule tumors in control male and female Osborne-Mendel rats (as recorded in the NCI 

Carcinogenesis Testing Program) were 0.3 and 0%, respectively. The incidence of renal 

adenomas (11%, first observed at 86 weeks; 8 weeks after the treatment period ended) following 

administration of 113 mg/kg-day HCE exceeded both the concurrent (0%) and historical (0.3%) 

controls in males. 

NTP (1989)—103-week study 

NTP (1989) administered 7 or 14 mg/kg-day HCE in corn oil via gavage to male F344/N 

rats for 103 weeks. Kidney effects consisting of tubular cell degeneration and atrophy, tubular 

dilatation, tubular cell regeneration, glomerulosclerosis, interstitial fibrosis, and chronic 

inflammation were observed in ≥94% of the HCE-exposed male rats. The incidence of

nephropathy in male control rats was 96%. The mean severity of the kidney effects in male rats 

increased following HCE exposure: 2.34 ± 0.14, 2.62 ± 0.15, and 2.68 ± 0.16 (statistically 

significant) in the control, 7 mg/kg-day, and 14 mg/kg-day dose groups, respectively. Kidney 

effect severity was considered mild for the controls and mild to moderate for the HCE-exposed 

male rats. While the mean severity scores do not show more than a 15% increase over control in 
 DO NOT CITE OR QUOTE 76

the high-dose group, more moderate and marked nephropathy was observed in HCE-exposed 

male rats compared with controls. The incidence of severe (moderate or marked) nephropathy in 

males was 18/50, 24/50, and 30/50 in the control, 7, and 14 mg/kg-day dose groups, respectively. 

Additionally, the male rats exhibited increased incidence in linear mineralization of the renal 

papillae: 4, 30, and 64% in the control, 7 mg/kg-day, and 14 mg/kg-day dose groups, 

respectively. Pelvic epithelium hyperplasia was also observed in 14% of male rats exposed to 

either 7 or 14 mg/kg-day HCE. These hyperplastic effects were not observed in either the 

controls or the treated females. 

NTP (1989) administered 57 or 114 mg/kg-day HCE in corn oil via gavage to female 

F344/N rats for 103 weeks. The incidence of nephropathy in female rats following chronic HCE 

exposure was 44, 84, and 90% for the control, 57, and 114 mg/kg-day dose groups, respectively. 

The severity scores for nephrotoxicity in female rats were statistically significantly increased in 

both treated groups: 0.72 ± 0.13, 1.38 ± 0.11, and 1.69 ± 0.12 in the control, 57 mg/kg-day, and 

114 m/kg-day dose groups, respectively. The average severity of nephropathy was considered 

minimal for the controls and minimal to mild for the HCE-exposed female rats. Examination of 

the various grades of nephropathy severity shows more mild and moderate nephrotoxicity in 

HCE-exposed females compared with controls. In females, the incidence of severe (mild or 

moderate) nephropathy was 12/50, 25/50, and 32/50 in the control, 57, and 114 mg/kg-day dose 

groups, respectively (statistical analysis was not reported). Female rats also showed an increase 

in linear mineralization at 57 (44%) and 114 mg/kg-day (26%) compared with relatively high 

response in the controls (28%). This increase in linear mineralization was not dose-related. The 

HCE-exposed male rats also exhibited renal tubular hyperplasia, renal tubule adenomas, and 

renal tubule carcinomas. The combined renal adenoma or carcinoma incidence was 2, 4, and 

14% (3, 6, and 24% after adjusting for intercurrent mortality) in the control, 7, and 14 mg/kg-day 

dose groups, respectively. There were no HCE-related neoplasms observed in female rats treated 

with 57 or 114 mg/kg-day HCE. NTP (1989) noted that the hyperplasia and tumors of the renal 

tubules represented a morphologic continuum. The hyperplasia incidence was observed in 4, 8, 

and 22% of the control, 7 mg/kg-day, and 14 mg/kg-day dose groups, respectively. The 

incidence of renal tubule neoplasia in male rats also exceeded historical controls (0.5%). Female 

rats did not exhibit renal tubule hyperplasia. 

A sex difference was noted in the observed nephropathy, as males were more sensitive to 

HCE-exposure related nephropathy than females. This sex-specificity is apparent for the 

nephrotoxicity and grades of nephropathy severity in both control and HCE-treated groups. 

Although administered only one-eighth of the dose given to the female rats; the male rats 

demonstrated a greater incidence of nephropathy that was more severe and included additional 

kidney effects (i.e., increases in incidence of mineralization of the renal papillae and hyperplasia 

of pelvic transitional epithelium) compared with the female rats. 
 DO NOT CITE OR QUOTE 77

With the exceptions of glomerulosclerosis, interstitial fibrosis, and chronic inflammation, 

the observed nephrotoxic effects in the male rats are characteristic of α2u-globulin nephropathy. 

However, NTP (1989) did not report accumulation of hyaline droplets containing the 

α2u-globulin protein in the proximal tubule. It is possible that hyaline droplets were present 

considering that the 16-day and 13-week rats examined by NTP (1989) exhibited hyaline 

droplets; however, the hyaline droplets were likely obscured by the prevalence of the other 

lesions. Evidence of these effects in almost all of the control males and in treated and control 

female rats also complicates the characterization of the mode of action. Considering that 

α2u-globulin nephropathy is typically male rat-specific, the appearance of nephrotoxic effects in 

the female rats as well as the male and female controls, and the identification of other effects not 

specifically associated with α2u-globulin (i.e., glomerulosclerosis and interstitial fibrosis) 

suggests that the effects are not the result of α2u-globulin accumulation. 

Considering the strain and age of the rats in the chronic (103 weeks) NTP study (1989), it 

is also possible that the rats were affected by CPN (i.e., increased incidence of nephrotoxicity in 

the control rats). However, changes in severity of the nephropathy that are greater in the HCEexposed animals indicate some chemical-related effects. Additionally, HCE-exposed male rats 

demonstrated dose-dependent increases in incidence of mineralization of the renal papillae and 

hyperplasia of pelvic transitional epithelium. Neither of these effects increased in a dose-related 

manner in the controls or the HCE-exposed female rats. Therefore, the treatment-related effects 

in male and female rats serve as evidence that CPN is not solely responsible for the nephropathy 

observed by NTP (1989). 

Limitations in the available studies 

These studies describe the effects associated with HCE exposure using a general, 

nonspecific term: tubular nephropathy (Weeks et al., 1979; NCI, 1978). This general term does 

not provide information on the specific histopathological changes characterizing the 

nephropathy. Additionally, the reported incidences of effects were grouped and measured as 

nephropathy rather than individual effects. Effects described in this way are difficult to interpret 

with regards to α2u-globulin nephropathy. One study (NTP, 1996) was limited in its usefulness 

because the only in male rats were exposed and the experimental design sought to draw 

conclusions about SARs involved in the induction of hyaline droplet nephropathy of 

11 halogenated ethanes. The study focused predominantly on the kidneys and the purpose of the 

study was to compare chlorinated ethanes, not examine the mode of action of HCE. The 

divergence in doses used for male and females in the NTP (1989) chronic exposure experiment 

highlighted the male sensitivity to HCE-induced nephrotoxicity. However, this study design 

made it difficult to otherwise compare the sexes. Additionally, three of the six HCE exposure 

studies utilized only two dose groups, limiting the ability to characterize the dose response of 

HCE-exposure related nephropathy. 
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Summary of evidence for strength, specificity, and consistency 

Generally, kidney tumors observed in cancer bioassays are assumed to be relevant for 

assessment of human carcinogenic potential. However, when the mode-of-action evidence 

demonstrates that kidney tumors in male rats result from an α2u-globulin-related mode of action, 

the data are not suitable for use in risk assessment (U.S. EPA, 1991b). The criteria for 

demonstrating the α2u-globulin-related mode of action for risk assessment purposes have been 

defined (U.S. EPA, 1991b). Three criteria must be met: (1) increase in hyaline droplets in the 

renal proximal tubule cells; (2) determination that the accumulating protein in the droplets is 

α2u-globulin; and (3) presence of additional pathological lesions associated with α2u-globulin. 

The key event in the histopathological sequence for the α2u-globulin-related mode of action is 

excessive accumulation of hyaline droplets containing α2u-globulin in renal proximal tubules. 

None of the HCE studies performed the necessary immunohistochemical assays to confirm the 

presence of α2u-globulin protein within the hyaline droplets observed following administration of 

HCE (NTP, 1996, 1989). It is unknown if HCE is binding to α2u-globulin or to other proteins 

during the formation of hyaline droplets. This represents an important data gap considering that 

the presence of this protein is essential to identifying the α2u-globulin-related mode of action. 

Therefore, there is insufficient evidence to support an α2u-globulin-related mode of action for 

renal tumors following HCE exposure. 

In addition, the data on female rats and mice of both sexes from chronic exposure studies 

(NTP, 1989; NCI, 1978) do not support the α2u-globulin mode of action for HCE-exposure 

related nephropathy. The appearance and type of nephrotoxicity noted in control and female rats 

suggest a mode of action other than α2u-globulin. NCI (1978) reported dose-related nephropathy 

in female rats that was not apparent in the controls. The dose-responsive kidney effects observed 

in the female rats treated with HCE suggests that a mode of action other than α2u-globulin 

nephropathy was occurring. Nephropathy was also reported in male and female mice 

chronically-administered HCE (NCI, 1978). The NCI (1978) reported the appearance of renal 

tubular effects in almost all (≥92%) of the HCE-treated male and female mice following chronic 

HCE exposure, but the mice did not develop renal tubule tumors. The presence of kidney effects 

in HCE-exposed male and female mice, which generally do not accumulate the α2u-globulin 

protein, suggests that a mode of action other than α2u-globulin nephropathy. 

Dose-response concordance 

The key event in the histopathological sequence for the α2u-globulin-related mode of 

action is excessive accumulation of hyaline droplets containing α2u-globulin in renal proximal 

tubules. The accumulation of α2u-globulin in hyaline droplets must occur at lower doses than 

subsequent α2u-globulin-related effects. None of the HCE studies performed the necessary 

immunohistochemical assays to confirm the presence of α2u-globulin protein within the hyaline 
 DO NOT CITE OR QUOTE 79

droplets observed following administration of HCE (NTP, 1996, 1989). Therefore, this key 

event cannot be demonstrated from the available data. 

Most of the effects characterizing the histopathological sequence of events in epithelial 

cells of the proximal tubules leading to renal tumors (U.S. EPA, 1991c) increased in incidence 

with dose of HCE in the short-term and subchronic exposure studies. Dose-related increases in 

nephrotoxicity and renal carcinogenicity were noted in the two chronic HCE exposure studies. 

The short-term and subchronic exposure studies did not report evidence of carcinogenicity in rats 

administered HCE. In the NTP (1989) study, male rats administered 7 or 14 mg/kg-day HCE for 

2 years exhibited a dose-related increased incidence of renal tubule adenomas and carcinomas. 

Typical histopathological effects associated with α2u-globulin nephropathy (tubular cell 

degeneration and atrophy, tubular dilatation, and tubular cell regeneration) were noted in almost 

all of the treated and untreated animals. A dose-response relationship was difficult to detect 

considering the number of animals affected by nephrotoxicity. However, dose-related increases 

over controls for toxic kidney effects such as linear mineralization, severity of nephrotoxicity, 

and renal tubule hyperplasia were observed. NTP (1989) did not report interim data; therefore, 

examinations were performed at study termination. Consequently, the nephrotoxicity (that is 

generally attributed to leading up to the formation of renal tubular tumors associated with α2uglobulin) is reportedly increased at doses similar to those that induce tumor formation. 

 Overall, dose-related kidney effects were noted for almost all of the male rats 

administered HCE at doses ranging from 1 to 563 mg/kg-day. Even at the lowest HCE dose 

administered in the studies, renal effects were observed in male rats. Animals treated with 

greater amounts of HCE exhibited dose-related increases in incidence and severity of effect 

when compared with those of the lower dose groups. It is difficult to establish dose-response 

concordance between the noncancer nephropathy and the renal tubule tumors reported by NTP 

(1989). Renal tubule tumors were observed at 7 mg/kg-day HCE, the lowest dose administered 

for a chronic duration, which also induced significant nephropathy in HCE-exposed animals. 

The other studies that administered doses within an order of magnitude of 7 mg/kg-day were the 

NTP (1989) study [34 or 67 mg/kg-day for 13 weeks] and the Gorzinski et al. (1985) study [1, 

15, or 62 mg/kg-day for 16 weeks]. Although nephropathy was noted in the shorter duration 

studies (NTP, 1996, 1989; Gorzinski et al., 1985), there only evidence of carcinogenicity was 

from the chronic exposure studies (NTP, 1989; NCI, 1978). 

Temporal relationship 

The key event in the histopathological sequence for the α2u-globulin-related mode of 

action is excessive accumulation of hyaline droplets containing α2u-globulin in renal proximal 

tubules. The accumulation of α2u-globulin in hyaline droplets must occur first in the sequela 

leading to α2u-globulin-related nephrotoxicity and tumor formation. None of the HCE studies 

performed the necessary immunohistochemical assays to confirm the presence of α2u-globulin 
 DO NOT CITE OR QUOTE 80

protein within the hyaline droplets observed following administration of HCE (NTP, 1996, 

1989). Therefore, this key event and the important temporal relationship for the accumulation of 

α2u-globulin cannot be demonstrated from the available data.

Many of the histopathological effects associated with α2u-globulin-related nephropathy 

were observed in animals treated with HCE in studies that varied in exposure duration from 

16 days to 2 years. The sequence of histopathological events characteristic of the 

α2u-globulin-related mode of action was noted in the chronic exposure study NTP (1989) that 

reported renal tubule adenomas and carcinomas. All of the studies (NTP, 1996, 1989; Gorzinski 

et al., 1985; NCI, 1978) that administered HCE for shorter durations than the NTP (1989) study 

reported similar histopathological changes, although an increase in renal tubule tumors was not 

observed. It is unknown if the nephropathy observed by NTP (1989) leads to the reported renal 

tubule tumors because the animals were only examined at the end of the 103-week study period. 

A temporal relationship cannot be distinguished from reported data. 

Biological plausibility and coherence 

The sequence of events including accumulation of α2u-globulin protein in the renal 

tubules of male rats initiating a sequence of nephrotoxic events leading to renal tubule tumor 

formation is plausible ( U.S. EPA, 1991c). These effects are typically not observed in female 

rats or other species due to the absence or minimal presence of the α2u-globulin protein in these 

animals (Hard et al., 1993). Concluding that HCE is acting through an α2u-globulin-associated 

mode of carcinogenic action is precluded by evidence of nephrotoxic effects in female rats in 

two chronic studies (NTP, 1989; NCI, 1978) and in male and female mice in one chronic study 

(NCI, 1978). Nephropathy associated with α2u-globulin is generally not observed in female rats 

or other species due to the absence or minimal presence of the α2u-globulin protein in these 

animals (Hard et al., 1993). 

 

Other Possible Modes of Action

 There is insufficient evidence to support an α2u-globulin-related mode of action for renal 

tumors following HCE exposure. It is possible that advanced CPN may play a role in the 

incidence of nephrotoxicity and kidney tumors in male rats following HCE exposure. CPN is an 

age-related renal disease of laboratory rodents that occurs spontaneously. The observed renal 

lesions in male rats following exposure to HCE are effects commonly associated with CPN. 

Nephropathy (described as tubular cell degeneration and regeneration, tubular dilatation and 

atrophy, glomerulosclerosis, interstitial fibrosis, and chronic inflammation) was also observed 

following chronic HCE exposure in female rats in the NTP (1989) chronic study, as well as male 

and female mice following chronic HCE exposure (NCI, 1978). However, changes in severity of 

the nephropathy were observed to be greater in male rats exposed to HCE compared to controls 
 DO NOT CITE OR QUOTE 81

indicating that HCE exposure exacerbated effects in the kidney. Additionally, HCE-exposed 

male rats demonstrated dose-dependent increases in incidence of mineralization of the renal 

papillae and hyperplasia of pelvic transitional epithelium. Neither of these effects increased in a 

dose-related manner in the controls or the HCE-exposed female rats. The treatment-related 

effects in male and female rats serve as evidence that CPN is not solely responsible for the 

nephropathy observed by NTP (1989). 

Conclusions about the Hypothesized Mode of Action 

Support for the Hypothesized Mode of Action in Animals 

The mode of action for the carcinogenic effects of HCE in the kidney is unknown. There 

is insufficient evidence to establish the role of α2u-globulin in HCE-exposure related 

nephropathy. Studies following short-term, subchronic, and chronic exposure of male rats have 

reported renal lesions consistent with α2u-globulin nephropathy (NTP, 1996, 1989; Gorzinski et 

al., 1985; NCI, 1978). The formation of renal tubule adenomas and carcinomas (preceded by 

hyperplasia) following chronic HCE exposure (NTP, 1989) are also consistent with an 

α2u-globulin-related mode of action. However, the key event in the histopathological sequence of 

events demonstrating a α2u-globulin-related mode of action (excessive accumulation of hyaline 

droplets containing α2u-globulin in renal proximal tubules) leading to development of kidney 

tumors in male rats exposed to HCE has not been fully characterized. None of the HCE studies 

performed the necessary immunohistochemical assays to confirm the presence of α2u-globulin 

protein within the hyaline droplets observed following administration of HCE (NTP, 1996, 

1989). It is unknown if HCE is binding to α2u-globulin or to other proteins during the formation

of hyaline droplets. This represents an important data gap considering that the presence of this 

protein is essential to identifying this hypothesized mode of action. In addition, data are 

available that demonstrate kidney effects in female rats and mice of both sexes from chronic 

exposure studies (NTP, 1989; NCI, 1978). The NCI (1978) study reported dose-related 

nephropathy in female rats that was not apparent in the controls. Nephropathy was also reported 

in male and female mice chronically-administered HCE (NCI, 1978). The presence of kidney 

effects in HCE-exposed male and female mice, which generally do not accumulate the α2uglobulin protein, suggests that a mode of action other than α2u-globulin nephropathy. 

Role of metabolites 

Studies of proposed HCE metabolites such as PERC, pentachloroethane, and TCE 

revealed similar noncancer and cancer effects. PERC exposure caused toxic nephropathy and 

kidney tumors (NCI, 1977); pentachloroethane exposure resulted in chronic inflammation of the 

kidney, (NTP, 1983); and TCE exposure produced chronic nephropathy (NCI, 1976), toxic 

nephrosis, kidney cytomegaly (NTP, 1990), renal tubular adenomas (NTP, 1990), tubular cell 

cytomegaly, and toxic nephropathy (NTP, 1988). These studies consistently demonstrate that the 
 DO NOT CITE OR QUOTE 82

kidney is the principal target tissue for noncancer effects following exposure to putative HCE 

metabolites. Based on the available data, it is not possible to define the relative roles of the 

parental compound (HCE) or its metabolites in the kidney effects associated with exposure to 

HCE. However, the kidney effects of these potential HCE metabolites further support the kidney 

as the principal target organ of HCE exposure. 

Relevance of the Hypothesized Mode of Action to Humans

Generally, kidney tumors observed in cancer bioassays are assumed to be relevant for 

assessment of human carcinogenic potential. However, for male rat kidney tumors, when the 

mode-of-action evidence convincingly demonstrates that the response is secondary to 

α2u-globulin accumulation, the tumor data are not used in the cancer assessment (U.S. EPA, 

1991b). There is insufficient evidence to conclude that the renal adenomas and carcinomas 

observed in male rats administered HCE (NTP, 1989) are related to an α2u-globulin mode of 

action for the following reasons: (1) the lack of α2u-globulin immunohistochemical data for 

HCE-induced nephrotoxicity and carcinogenicity, (2) the hyaline droplet accumulation is caused 

by excessive protein load that may not be exclusively related to α2u-globulin accumulation, and 

(3) the existence of renal toxicity in female rats and male and female mice indicates that the 

nephrotoxic effects are not limited to an α2u-globulin-induced sequence of lesions. Therefore, 

the renal adenomas and carcinomas observed in male rats administered HCE (NTP, 1989) were 

considered relevant to humans. 

4.7.3.2. Liver Tumors 

Description of the Hypothesized Mode of Action 

Hypothesized mode of action 

Hepatocellular carcinomas were observed in male and female B6C3F1 mice administered 

360 or 722 mg/kg-day HCE, via gavage, in a chronic oral bioassay conducted by NCI (1978). 

Tumor incidences in males of both dose groups were statistically significantly elevated compared 

with control groups, and demonstrated a dose response. Both dose groups of female mice 

presented statistically significantly elevated incidences of hepatocellular carcinoma compared 

with control groups but a dose response was not observed. The investigators did not find 

nonneoplastic liver effects (such as organized thrombus, inflammation, fibrosis, necrosis, 

infarctions, amyloidosis, or hyperplasia) in either sex. 

The mode of action for the carcinogenic effects of HCE in the liver is unknown. 

Metabolism studies of HCE indicate that the major enzymes involved are 

phenobarbital-inducible CYP450s. These are primarily localized in the liver. Although 

tissue-specific metabolism of HCE has not been studied extensively, the majority of HCE 

metabolism is presumed to occur in the liver. HCE is proposed to metabolize to PERC and 

pentachloroethane and is likely, subsequently metabolized to TCE. It is possible that the HCE-
 DO NOT CITE OR QUOTE 83

induced hepatocellular carcinomas in mice occur as a result of the binding of HCE metabolites to 

liver macromolecules and the generation of free radicals during HCE metabolism, causing key 

events in the carcinogenic process such as cytotoxicity, inflammation and regenerative cell 

proliferation. However, these potential key events have not been systematically evaluated for 

HCE. 

In a 13-week study, hepatocellular necrosis of the centrilobular area was observed in rats 

(NTP, 1989). It is unknown if this could be considered a key event in the carcinogenic process 

because rats in the available studies (NTP, 1989; NCI, 1978) have not displayed hepatocellular 

neoplastic endpoints. Although mice demonstrated hepatocellular carcinoma, nonneoplastic 

effects such as hepatocellular necrosis were not observed (NCI, 1978). HCE-induced 

hepatocellular carcinomas in mice varied in microscopic appearance (NCI, 1978). Some 

carcinomas were characterized by well-differentiated hepatic cells with uniform cord 

arrangement, while others had anaplastic liver cells with large hyperchromatic nuclei, often with 

inclusion bodies and vacuolated pale cytoplasm. Arrangement of neoplastic liver cells also 

varied from short stubby cords to nests of cells, and occasional pseudo-acinar formations. 

Neoplasms in control mice did not vary in appearance from those in HCE-treated mice. 

In vivo binding of radiolabeled carbon to DNA, RNA and protein from liver, kidney, 

lung, and stomach following administration of [14C]HCE was consistently greater in mice 

compared with rats (Lattanzi et al., 1988). Binding to macromolecules was interpreted by the 

presence of radiolabeled carbon; however, radiolabeled carbon may have been incorporated into 

these macromolecules from intermediary HCE metabolites. In vitro binding studies using calf 

thymus DNA demonstrated that mouse liver cytosol (induced by phenobarbital) mediated more 

extensive DNA binding than rat liver cytosol (Lattanzi et al., 1988). Comparisons of HCE 

metabolism rates indicated that mice metabolize HCE at twice the rate of rats (Mitoma et al., 

1985). 

Cellular damage leading to cytotoxicity, inflammation, and regenerative cell proliferation 

is a possible consequence of this binding in the liver. The binding studies provide a line of 

evidence as to why the liver is the major carcinogenic target in the mouse, but not the rat. 

Regenerative cell proliferation has been evaluated in the kidney, but not in the liver of 

HCE-treated rats (NTP, 1996). RDS in hepatocytes was evaluated in mice treated with HCE 

(Yoshikawa, 1996; Miyagawa et al., 1995). This study reported ambiguous results; the lower 

HCE dose caused a statistically significant increase in RDS, whereas the higher dose did not 

(Yoshikawa, 1996; Miyagawa et al., 1995). Rat liver foci experiments provide support for the 

hypothesis that HCE acts as a tumor promoter, not as a tumor initiator (Milman et al., 1988; 

Story et al., 1986). 

The in vivo binding data suggest that HCE is sequestered in the liver of mice and rats and 

metabolic data suggest that mice metabolize HCE at a greater rate compared with rats. 

Considering the greater potential for metabolism in mice compared with rats and the proposed 
 DO NOT CITE OR QUOTE 84

increase in DNA binding following metabolism of HCE (Lattanzi et al., 1988), the increased 

incidence of hepatocellular carcinomas in mice, but not rats, may be related to DNA binding. 

However, the DNA binding measurements were based solely on the presence of radiolabeled 

carbon; specific HCE metabolites were not identified. Therefore, this process does not take into 

account the possibility of normal biological mechanisms in which the radiolabeled carbon can be 

incorporated into the macromolecules via anabolic processes. All together, while it is possible 

that metabolism and binding in mice are involved in the development of liver tumors, the role of 

DNA binding in the mode of action for HCE-induced hepatotoxicity and carcinogenesis is not 

known and, as such, the mode of action is not known. 

4.7.3.3. Pheochromocytomas 

Description of the Hypothesized Mode of Action 

Hypothesized mode of action 

No studies were identified to determine a mode of action for HCE-induced tumors of the 

adrenal gland. Therefore, the mode of action for pheochromocytomas observed following oral 

exposure to HCE is unknown. 

4.8. SUSCEPTIBLE POPULATIONS AND LIFE STAGES 

 No studies were located that address the susceptibility of populations or life stages to 

HCE-induced toxicity or carcinogenicity in humans. 

4.8.1. Possible Childhood Susceptibility 

 No studies were located that addressed possible childhood susceptibility to HCE-induced 

toxicity or carcinogenicity. Although it is unknown if HCE toxicity is mediated by parent 

compound or its metabolites, CYP450 enzymes of the 2A, 2B, and 3A subfamilies and CYP450 

1A2 are involved in HCE metabolism. Many drugs reportedly exhibit a higher systemic 

clearance in children than in adults (Evans et al., 1989). Blanco et al. (2000) compared liver 

microsomal CYP450 activities of humans aged <10 years with those aged >10-60 years and 

concluded that factors other than maximal CYP450 catalytic activities, such as reductions in 

hepatic blood flow, hepatic size, and oxygen supply in the elderly, may be responsible for agerelated changes in drug clearance. Studies of fetal and neonatal livers indicate that CYP450 

expression is similar to adult levels by a few months of age (Lacroix et al., 1997; Vieira et al., 

1996; Cazeneuve et al., 1994; Treluyer et al., 1991). However, Dorne (2004) reported in a 

review article that Phase I (including CYP450 activities) and Phase II enzymatic activities are 

1.3–1.5-fold higher in children (aged 1–16 years) compared with adults. Therefore, the extent to 

which variable age-related expression of CYP450 contributes to childhood susceptibility is 

unknown. Research in developmental expression of CYP450 is ongoing; conclusions regarding 

the differential expression of CYP450 are premature. Considering the substantial portion of 
 DO NOT CITE OR QUOTE 85

HCE that remains as parent compound, the impact, if any, of age on CYP450 expression and 

HCE metabolism cannot be assessed. 

4.8.2. Possible Gender Differences 

Toxicity studies in rats indicate that male rats are more sensitive to HCE-induced 

nephrotoxicity than females (NTP, 1989; Gorzinski et al., 1985; Gorzinski et al., 1980; NCI, 

1978). Evidence suggests that female rats are more sensitive to HCE-induced hepatotoxicity. 

The reasons for these sex-specific differences are unknown, but may be related to sex-specific 

differences in tissue concentrations following HCE administration (i.e., higher concentrations 

observed in male rat tissues when compared with female rats, see Table 3-3), sex hormone 

differences, and/or gender difference in CYP450 activities. No additional studies were located 

that addressed possible gender differences for HCE-induced toxicity or carcinogenicity. 

4.8.3. Other 

 CYP450 enzymes are polymorphic in the human population. Polymorphisms result in 

CYP450 enzymes with variant catalytic activity for substrates such as HCE. This could 

potentially result in decreased HCE detoxification or increased HCE bioactivation. 

Detoxification enzymes such as the GST family are also polymorphic in the human population, 

with variant catalytic activities that could affect the detoxification of HCE. There are numerous 

epidemiology studies in the literature examining CYP450 and GST polymorphisms and increases 

in cancer risks; however, none of these are specific for HCE. 
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5. DOSE-RESPONSE ASSESSMENTS

5.1. ORAL REFERENCE DOSE (RfD) 

5.1.1. Choice of Principal Study and Critical Effect—with Rationale and Justification

Data on the health effects of oral HCE exposure in humans are not available. The oral 

exposure database for HCE includes a 103-week gavage study in F344 rats (NTP, 1989), a 

78-week gavage study in Osborne-Mendel rats (NCI, 1978), a 91-week gavage study in B6C3F1

mice (NCI, 1978), a 16-week feeding study in F344 rats (Gorzinski et al., 1985), and a 13-week 

gavage study in F344 rats (NTP, 1989). The short-term study data were not considered in the 

selection of the principal study for the derivation of the RfD because the database contains 

reliable dose-response data from studies of subchronic and chronic durations. However, 

short-term studies in rats (NTP, 1996, 1989) were used to support findings in the subchronic and 

chronic studies. The available oral exposure studies identified kidney or liver effects associated 

with exposure to HCE. Reported effects include tubular nephropathy (NTP, 1989; NCI, 1978), 

atrophy and degeneration of renal tubules (NTP, 1989; Gorzinski et al., 1985), slight hypertrophy 

and/or dilation of proximal convoluted renal tubules (Gorzinski et al., 1985), linear 

mineralization of renal tubules (NTP, 1989), hyperplasia of the renal pelvic transitional 

epithelium (NTP, 1989), and hepatocellular necrosis (NTP, 1989). 

In the NTP (1989) chronic study, HCE was administered via gavage at doses of 7 and 

14 mg/kg-day in male F344 rats and 57 and 114 mg/kg-day in female F344 rats for 103 weeks. 

Nephropathy (characterized by tubular cell degeneration and regeneration, tubular dilatation and 

atrophy, glomerulosclerosis, interstitial fibrosis, and chronic inflammation) was observed in 

HCE treated rats of both sexes. Nephropathy was also reported in control rats of both sexes. 

Although a high incidence of nephropathy was observed in control rats, the study authors 

reported that the incidence of more severe nephropathy increased in dosed rats relative to 

controls (NTP, 1989). EPA considered the increase in severity of nephropathy in male rats by 

analyzing the incidence of greater than mild nephropathy. EPA determined that the increased 

incidence of moderate or marked nephropathy in males was statistically significant at the 14 

mg/kg-day dose (see Table 5-1). EPA considered the increased severity of nephropathy in 

female rats by analyzing the incidence of nephropathy that was greater than minimal 

nephropathy. EPA determined that the increased incidences of mild to marked nephropathy were 

statistically significant in females at the 57 and 114 mg/kg-day doses (see Table 5-1). Linear 

mineralization of the renal papillae and hyperplasia of the renal pelvic epithelium were increased 

in a dose-dependent, statistically significant manner in the treated male rats. EPA determined 

that the increased incidences of linear mineralization of the renal papillae and hyperplasia of the 

renal pelvic epithelium were statistically significant in males at the 7 and 14 mg/kg-day doses 

(see Table 5-1). Considering the increased severity of nephropathy following HCE exposure and 
 DO NOT CITE OR QUOTE 87

dose-dependent increases in the incidence of mineralization of the renal papillae and hyperplasia 

of renal pelvic transitional epithelium in male rats, EPA determined that the nephropathy 

observed in the NTP (1989) study was exacerbated by HCE exposure. The NTP (1989) chronic 

study did not identify NOAELs for male or female rats as kidney effects were observed at the 

lowest doses tested. EPA considered the male rat LOAEL as 7 mg/kg-day based on increased 

incidence in moderate or marked tubular nephropathy (characterized by degeneration, necrosis, 

and regenerative epithelial cells), hyperplasia of the pelvic transitional epithelium, and linear 

mineralization of the renal papillae in the NTP (1989) study. EPA considered the female rat 

LOAEL as 57 mg/kg-day, based on dose-related increases in incidence and severity of 

nephropathy in the NTP (1989) study. 

In the NCI (1978) chronic rat study, HCE was administered via gavage to groups of 50 

male and 50 female Osborne-Mendel rats for 5 days/week, cyclically for 66 of the 78 weeks, 

followed by an observation period of 33-34 weeks (total of 112 weeks). The TWA doses of 

HCE were 113 and 227 mg/kg-day. Tubular nephropathy was observed in all groups of treated 

animals, but was not observed in either untreated or vehicle controls. Statistically significant 

increases in incidence of tubular nephropathy were observed at 113 and 227 mg/kg-day HCE in 

both male and female rats (see Table 5-1). The NCI (1978) study did not identify a NOAEL for 

tubular nephropathy in rats. EPA considered the LOAEL as 113 mg/kg-day, based on a doserelated increase in incidence of nephropathy in both male and female rats. 

In the NCI (1978) chronic mouse study, HCE was administered via corn oil gavage to 

groups of 50 male and 50 female B6C3F1 mice for 5 days/ week for 78 weeks followed by an 

observation period of 12-13 weeks (total of 90 weeks). Starting in week 9, the HCE doses were 

increased, though no explanation for the increase was provided. The TWA doses of HCE were 

360 and 722 mg/kg-day. Because of low survival rates in the vehicle and untreated male control 

groups, NCI (1978) compared tumor incidences in the dosed males and females to the pooled 

vehicle control data derived from concurrently run bioassays for several other chemicals. NCI 

(1978) reported chronic kidney inflammation (i.e., tubular nephropathy characterized by 

degeneration of the convoluted tubule epithelium at the junction of the cortex and medulla and 

hyaline casts) in male and female B6C3F1 mice administered 360 and 721 mg/kg-day HCE. 

EPA considered the LOAEL for this study as 360 mg/kg-day based on tubular nephropathy, 

while a NOAEL could not established from these data. 

In Gorzinski et al. (1985), HCE was administered (in feed) to groups of 10 male and 

10 female F344 rats at doses of 0, 1, 15, or 62 mg/kg-day for a period of 16 weeks. Kidney 

effects consisted of slight hypertrophy and/or dilation of proximal convoluted renal tubules and 

atrophy and degeneration of renal tubules. Slight hypertrophy and/or dilation of the proximal 

convoluted renal tubules was not observed in the control rats of either sex or in HCE exposed 

female rats. EPA determined that increases in slight hypertrophy and/or dilation of the proximal 

convoluted renal tubules were statistically significant in male rats treated with 15 or 
 DO NOT CITE OR QUOTE 88

62 mg/kg-day HCE (see Table 5-1). Atrophy and degeneration of renal tubules was observed in 

both male and female rats. EPA determined that increases in incidences of atrophy and 

degeneration of renal tubules were statistically significant in male rats treated with 15 or 

62 mg/kg-day HCE and in female rats fed 62 mg/kg-day HCE (see Table 5-1). EPA considered 

the male rat LOAEL as 15 mg/kg-day and the male rat NOAEL as 1 mg/kg-day, based on 

increased incidence of the renal tubule effects. EPA considered the female rat LOAEL as 

62 mg/kg-day and the female rat NOAEL as 15 mg/kg-day, based on increased incidence of 

renal tubule effects. 

In the NTP (1989) subchronic study, HCE was administered via gavage to groups of 

10 male and 10 female F344 rats at TWA doses of 0, 34, 67, 134, 268, and 536 mg/kg-day for 

13 weeks. Kidney effects (i.e. hyaline droplet formation, renal tubular regeneration, and renal 

tubular casts) were observed in male rats from all HCE exposure groups, though incidence data 

was only provided for the 34 mg/kg-day dose group. NTP (1989) reported that the severity of 

kidney effects in male rats increased with dose, but no data on severity were presented. No 

kidney effects were reported in female F344 rats exposed to HCE. Liver effects were observed 

in male and female rats at higher doses of HCE and EPA determined that statistically significant 

increases in hepatocellular necrosis were observed in female rats exposed to 268 or 

536 mg/kg-day HCE (see Table 5-1). 

The short-term studies in rats were not used for subsequent dose-response assessment 

because of short study durations (12-day study [Weeks et al., 1979], 16-day study [NTP, 1989], 

21-day study [NTP, 1996], and 1-day study [Fowler, 1969]) and the availability of reliable 

dose-response data from studies of subchronic and chronic durations. The effects and the 

incidence of kidney and liver effects from the studies considered for selection as the principal 

study, serving as the basis for the derivation of the RfD, are summarized in Table 5-1. As 

incidence data on kidney effects reported in the 13-week subchronic study (NTP, 1989) were 

limited to males in the 34 mg/kg-day dose group, these data are not presented in Table 5-1. 
 DO NOT CITE OR QUOTE 89

Table 5-1. Incidences of non-cancerous kidney and liver effects in rats following 

oral exposure to HCE 

Study Duration 

(route) Strain/sex/species Endpoint Dose (mg/kg-day) Incidence 

Kidney Effects 

0 0/20 (0%) 

113 22/49a (45%) Osborne-Mendel 

Male Rat Tubular Nephropathy 

227 33/50a (66%) 

0 0/20 (0%) 

113 9/50a (18%) 

NCI (1978) 78 weeks 

(gavage) 

Osborne-Mendel 

Female Rat Tubular Nephropathy 

227 29/49a

 (59%) 

0 18/50 (36%) 

F344 Male Rat 7 24/50 (48%) 

Moderate to 

Marked Tubular 

Nephropathy 14 30/50a

 (60%) 

0 12/50 (24%) 

57 25/50a (50%) 

NTP (1989) 103 weeks 

(gavage) 

F344 Female Rat 

Mild to 

Marked Tubular 

Nephropathy 114 32/49a

 (65%) 

0 2/50 (4%) 

7 15/50 a (30%) NTP (1989) 103 weeks 

(gavage) 

F344 Male Rat Linear Mineralization 

14 32/50 a (64%) 

0 0/50 (0%) 

7 7/50a (14%) 

NTP (1989) 103 weeks 

(gavage) 

F344 Male Rat Hyperplasia of the 

Renal Pelvic 

Transitional 

Epithelium

14 7/50a (14%)

0 0/10 (0%) 

1 1/10 (10%) 

15 7/10a (70%) 

Gorzinski 

et al. (1985) 

16 weeks 

(dietary) F344 Male Rat 

Slight Hypertrophy 

and/or Dilation of 

Proximal Convoluted 

Renal Tubules 62 10/10a

 (100%) 

0 1/10 (0%) 

1 2/10 (20%) 

15 7/10a (70%) 

F344 Male Rat 

Atrophy and 

Degeneration of Renal 

Tubules 

62 10/10a (100%) 

0 1/10 (0%) 

1 1/10 (10%) 

15 2/10 (20%) 

Gorzinski 

et al. (1985) 

16 weeks 

(dietary) 

F344 Female Rat 

Atrophy and 

Degeneration of Renal 

Tubules 

62 6/10a

 (60%) 

Liver Effects 

0 0/10 (0%) 

33.5 0/10 (0%)

67.1 0/10 (0%)

134.3 0/10 (0%)

267.8 1/10 (10%) 

F344 Male Rat Hepatocellular 

Necrosis 

535.7 2/5 (40%) 

0 0/10 (0%) 

33.5 0/10 (0%) 

67.1 0/10 (0%) 

134.3 2/10 (20%) 

267.8 4/10a (40%) 

NTP (1989) 13 weeks 

(gavage) 

F344 Female Rat Hepatocellular 

Necrosis 

535.7 8/10a

 (80%) 

aEPA determined statistical significance using Fisher’s Exact Test (p < 0.05) 

These chronic and subchronic studies in rats and mice indicate that the kidney and liver 

are both target organs of HCE oral toxicity in rodents. Given the number of effects reported in 
 DO NOT CITE OR QUOTE 90

the kidney and the greater sensitivity of these effects in available studies, the kidney is the 

primary target of oral HCE exposure toxicity in rodents. HCE exposure resulted in a number of 

kidney effects: atrophy and degeneration of renal tubules in male and female F344 rats 

(Gorzinski et al., 1985), slight hypertrophy and/or dilation of proximal convoluted renal tubules 

in male F344 rats (Gorzinski et al., 1985), linear mineralization in male F344 rats (NTP, 1989), 

tubular nephropathy in male and female F344 rats (NTP, 1989), hyperplasia of the renal pelvic 

transitional epithelium in male F344 rats (NTP, 1989), and tubular nephropathy in male and 

female Osborne-Mendel rats (NCI, 1978). Further consideration was given to these endpoints as 

potential critical effects for the determination of the point of departure (POD) for derivation of 

the oral RfD. 

Although the doses associated with hepatic effects were more than 10-fold higher than 

doses associated with kidney effects, data from the NTP (1989) study on incidence of 

hepatocellular necrosis from the female rats were subject to BMD modeling for comparison 

purposes. The data on the male rat liver effects from the NTP (1989) study were not considered 

because the incidence of hepatocellular necrosis was not significantly elevated above controls at 

any HCE dose. The kidney effects reported in the 13-week subchronic study (NTP, 1989) were 

not further considered because the lack of the incidence data for the control groups made it 

uncertain if the 34 mg/kg-day HCE dose represented a LOAEL. In addition, the HCE doses 

administered were more than four-fold higher than those doses associated with kidney effects in 

other subchronic (Gorzinski et al., 1985) and chronic (NTP, 1989) studies. The chronic study in 

B6C3F1 mice (NCI, 1978) was not considered for selection as the principal study because HCE 

doses that induced kidney effects were more than 7-fold higher than doses associated with 

kidney effects in rats following subchronic (Gorzinski et al., 1985) or chronic (NTP, 1989; NCI, 

1978) exposure. The ability of the chronic NTP (1989) study to inform the effects observed at 

the lowest dose tested in the Gorzinski et al. (1985) study is limited because the lowest dose 

tested in the chronic exposure study represented a LOAEL. Table 5-2 summarizes the BMD 

modeling results of the available data and the benchmark response (BMR) levels and the 

potential PODs are identified for each effect. 

5.1.2. Methods of Analysis—Including Models

For this assessment, BMD approach (U.S. EPA, 2001) was employed to identify the 

potential POD for each of the endpoints described above. A BMR of 10% extra risk was 

considered appropriate for derivation under the assumption that it represents a minimally 

biologically significant response level (U.S. EPA, 2000b). All of the dichotomous dose-response 

models available in the EPA benchmark dose software (BMDS), version 2.0, were fit to the 

incidence data for kidney effects in male and female rats reported by NTP (1989), NCI (1978), 

and Gorzinski et al. (1985), as well as the incidence data for hepatocellular necrosis in female 

rats reported by NTP (1989). Details of the BMD dose-response modeling reported in Table 5-2 
 DO NOT CITE OR QUOTE 91

are presented in Appendix B (Table B-1). In addition, the BMD and BMDL modeling outcomes 

for a BMR of 5 and 1% are presented in Appendix B (Table B-2) for comparison with the 10% 

BMR. 

From the BMD modeling analysis results presented in Table B-1, candidate PODs were 

selected. Table 5-2 summarizes the BMD modeling results of the available data and the 

benchmark response (BMR) levels and the potential PODs are identified for each effect. 

Table 5-2. Summary of the BMD modeling results for the kidney 

Study Endpoint Sex/species 

(group size)

Duration 

(route) 

“Best-Fit” 

Model 

Goodness 

of fit 

p-value 

AIC 

BMD 

(mg/kgday) 

BMDL10

(mg/kgday) 

Gorzinski 

et al. 

(1985) 

Slight hypertrophy 

and/or dilation of 

proximal 

convoluted renal 

tubules 

Male Rats 

(n = 10) 

16 weeks 

(dietary) Gammaa

 0.99 20.88 1.22 0.710 

Male Rats 

(n = 10) Gammab Gorzinski 0.70 34.94 1.34 0.728 

et al. 

(1985) 

Atrophy and 

degeneration of 

renal tubules Female Rats 

(n = 10) 

16 weeks 

(dietary) Multistage 

1o c 0.93 40.61 8.54 4.49 

Male Rats 

(n ≈ 50) Gammad

 0.93 133.68 21.22 16.99 NCI 

(1978) 

Tubular 

Nephropathy Female Rats 

(n ≈ 50) 

78 weeks 

(gavage) Multistage 

2o 0.94 116.09 80.63 41.89 

Male Rats 

(n ≈ 50) 

Quantallineare 0.87 205.90 3.20 1.88 NTP 

(1989) 

Increased Severity 

of Tubular 

Nephropathy Female Rats 

(n ≈ 50) 

103 weeks 

(gavage) 

Gammaf 0.86 191.90 15.17 10.72 

NTP 

(1989) 

Linear 

Mineralization 

Male Rats 

(n ≈ 50) 

103 weeks 

(gavage) Probit 0.51 147.66 3.98 3.22 

NTP 

(1989) 

Hyperplasia of the 

pelvic transitional 

epithelium 

Male Rats 

(n ≈ 50) 

103 weeks 

(gavage) LogLogistic 0.48 84.42 7.05 4.48 

aGamma, Quantal-linear, and Weibull models had identical AIC, goodness of fit p-values, as well as BMD10 and 

BMDL10 values. 

bGamma, Multistage 1o, and Quantal-linear models had identical AIC, goodness of fit p-values, as well as BMD10

and BMDL10 values. 

cMultistage 1o and Quantal-linear models had identical AIC, goodness of fit p-values, as well as BMD10 and 

BMDL10 values. 

dGamma, Multistage 1o, and Weibull models had identical AIC, goodness of fit p-values, as well as BMD10 and 

BMDL10 values. 

eQuantal-linear and Multistage 1o models had identical AIC, goodness of fit p-values, as well as BMD10 and 

BMDL10 values. 

fGamma, Quantal-linear, and Weibull models had identical AIC, goodness of fit p-values, as well as BMD10 and 

BMDL10 values. 

AIC = Akaike’s Information Criteria (i.e., a measure of the deviance of the model fit that allows for comparison 

across models for a particular endpoint). 

The range of potential PODs (approximately 60-0.6 mg/kg-day) is about 100-fold. 

Kidney effects (i.e., tubular nephropathy, linear mineralization of the renal tubules, hyperplasia 

of the pelvic transitional epithelium, atrophy and degeneration of renal tubules, and slight 
 DO NOT CITE OR QUOTE 92

hypertrophy and/or dilation of the proximal convoluted renal tubules) observed in male rats 

resulted in lower candidate PODs than comparable effects in female rats. 

The most sensitive effect observed in male rats exposed to HCE is slight hypertrophy 

and/or dilation of proximal convoluted renal tubules (Gorzinski et al., 1985). However, the 

potential POD for slight hypertrophy and/or dilation of proximal convoluted renal tubules (i.e., 

0.710 mg/kg-day) is nearly identical to the candidate POD for atrophy and degeneration of renal 

tubules (i.e., 0.728 mg/kg-day). As tubular nephropathy in the chronic studies (NTP, 1989; NCI, 

1978) was characterized as atrophy and degeneration of renal tubules, this endpoint has been 

consistently observed following HCE exposure in several studies. Therefore, atrophy and 

degeneration of renal tubules was selected as the candidate POD for this subchronic exposure 

study. As shown in Appendix B, the gamma, multistage 1°, logistic, probit, Weibull models in 

BMDS (version 2.0) provided adequate fits to the incidence data for atrophy and degeneration of 

renal tubules in male rats from the Gorzinski et al. (1989) 16-week study (Table B-1), as 

assessed by a chi-square goodness-of-fit p-values, as well as BMD10 and BMDL10 values were 

identical for the gamma, multistage 1°, and quantal-linear model; therefore, the model with the 

lowest BMDL10 was selected. The gamma, multistage 1°, and quantal-linear model had identical 

BMDL10 values. Therefore, the BMD10 associated with a 10% extra risk for nephropathy in 

male rats of 1.34 mg/kg-day, and the lower 95% confidence limit on this BMD10 (BMDL10) of 

0.728 mg/kg-day was selected as the potential POD for these data. 

The tubular nephropathy in male rats observed in the chronic exposure study (NTP, 1989) 

resulted in higher PODs than the atrophy and degeneration of renal tubules in male rats observed 

following 16 weeks of HCE exposure (Gorzinski et al., 1985). The ability of the chronic NTP 

(1989) study to inform the effects observed at the lowest dose tested in the Gorzinski et al. 

(1985) study is limited because the lowest dose tested in the chronic exposure study represented 

a LOAEL. Therefore, the Gorzinski et al. (1985) study was selected as the principal study and 

atrophy and degeneration of renal tubules in male rats was selected as the critical effect. The 

BMDL10 of 0.728 mg/kg-day was selected as the POD and serves as the basis for the derivation 

of the oral RfD for HCE. This endpoint is supported by additional kidney effects associated with 

oral exposure to HCE and supports the weight of evidence for HCE-associated nephrotoxicity. 

5.1.3. RfD Derivation—Including Application of Uncertainty Factors (UFs)

The derivation of the RfD for atrophy and degeneration of renal tubules in male F344 rats 

from the Gorzinski et al. (1985) 16-week toxicity study was calculated from the BMDL10 of 

0.728 mg/kg-day. The composite UF of 3,000 was comprised of the following: 

• An interspecies UF (UFA) of 10 was applied to account for the variability in extrapolating 

from rats to humans. Although the toxicokinetics have been minimally evaluated in animals, 

the toxicokinetics of HCE have not been fully characterized in either rats or humans. 
 DO NOT CITE OR QUOTE 93

• An intraspecies UF (UFH) of 10 was applied to adjust for potentially sensitive human 

subpopulations in the absence of information on the variability of response to HCE in the 

human population. Current information is unavailable to assess human-to-human variability 

in HCE toxicokinetics and toxicodynamics. 

• The study selected as the principal study was a 16 week study by Gorzinski et al (1985), a 

study duration that is minimally past the standard subchronic (90 day) study and falls well 

short of a standard lifetime study. Kidney effects were observed in male rats in the Gorzinski 

et al. (1985) subchronic study at doses below the range of exposure tested in the available 

chronic exposure studies. In addition, the ability of the available chronic studies to inform 

the effects observed at the low dose is limited because the lowest dose tested in the NTP 

(1989) chronic exposure study represented a LOAEL. Therefore, there are no data to exclude 

the possibility that chronic exposure could increase the severity of the observed kidney 

effects or could result in similar effects at lower doses. For these reasons, 

subchronic-to-chronic UF (UFS) of 10 was used to account for the extrapolation from 

subchronic-to-chronic exposure duration. 

• An UF for a LOAEL to a NOAEL extrapolation was not used because the current approach is 

to address this extrapolation as one of the considerations in selecting a BMR for BMD 

modeling. In this case, a BMR of a 10% increase in the incidence of renal tubule atrophy and 

degeneration was selected under an assumption that it represents a minimal biologically 

significant change. 

• A UF of 3 was applied to account for deficiencies in the HCE toxicity database, including the 

lack of a multigenerational reproductive study. The database includes studies in laboratory 

animals, including chronic and subchronic dietary exposure studies and two oral 

developmental toxicity studies. One of the available oral developmental toxicity studies 

demonstrated that HCE exposure decreased gestational indices and fetal viability, and 

increased resorptions with maternal toxicity at 500 mg/kg-day (Weeks et al., 1979). The 

second oral developmental toxicity study showed maternal toxicity at both the mid- and high 

doses (167 and 500 mg/kg-day) with decreased fetal body weight and increased late stage 

resorptions and skeletal variations at the high dose (Shimizu et al., 1992). The toxic effects 

observed in the developmental toxicity studies were observed at doses higher than those 

observed to induce renal toxicity in the subchronic and chronic toxicity studies. Therefore, in 

consideration of the entire database for HCE, a database UF of 3 was applied to account for 

the lack of a two-generational reproductive study. 
 DO NOT CITE OR QUOTE 94

Given the UFs established above, the RfD for HCE was calculated employing the following 

equation: 

RfD = POD ÷ UF 

 = 0.728 mg/kg-day ÷ 3,000 

 = 2 × 10–4 mg/kg-day 

 

5.1.4. RfD Comparison Information 

 The predominant noncancer effect of acute, short-term, subchronic, and chronic oral 

exposure to HCE is renal toxicity. Table 5-3 presents the PODs for nephrotoxicity in male rats 

with applied uncertainty factors (UFs) and potential reference values. Figure 5-1 provides a 

graphical display of dose-response information from three studies that reported kidney toxicity in 

male rats following chronic and subchronic oral exposure to HCE, focusing on potential PODs 

that could be considered in deriving the oral RfD. As discussed in Sections 5.1.1 and 5.1.2, 

among those studies that demonstrated kidney toxicity, atrophy and degeneration of renal tubules 

in male F344 rats from the Gorzinski et al. (1985) study provided the POD for deriving the RfD 

(see dotted box in Figure 5-1). Potential reference values that might be derived from other 

studies are also presented. Only endpoints observed in male rats are presented because the 

database for HCE consistently showed that male rats exhibited greater sensitivity to HCE 

toxicity compared to females. 

 The nephropathy observed by NCI (1978) was similar to that reported by NTP (1989); 

however, the animals in the NTP study were exposed to and exhibited effects at a lower range of 

doses of HCE than those in the NCI study (Table 5-1). NTP (1989) described tubular 

nephropathy characterized by degeneration, necrosis, and regenerative epithelial cells in rats. 

Gorzinski et al. (1985) described similar renal effects characterized by atrophy and degeneration 

of renal tubules and slight hypertrophy and/or dilation of proximal convoluted tubules. Linear 

mineralization of the renal tubules, hyperplasia of the pelvic transitional epithelium, slight 

hypertrophy and/or dilation of the proximal convoluted tubules, increased severity of tubular 

nephropathy, and atrophy and degeneration or renal tubules were all reported in male rats 

exposed to HCE (NTP, 1989; Gorzinski et al., 1985). Additionally, nephropathy was observed 

in both male and female rats, whereas linear mineralization was only observed in male rats. 

Kidney effects were observed in male rats in the Gorzinski et al. (1985) study at doses below the 

range of exposure tested in the NTP (1989) study. In addition, the ability of the chronic studies 

to inform the effects observed at the low dose in the Gorzinski et al. (1985) study is limited 

because the lowest dose tested in the NTP (1989) chronic exposure study represented a LOAEL. 

The potential POD associated with atrophy and degeneration of renal tubules from the Gorzinski 

et al. (1985) study was lower than the POD based on increased severity of tubular nephropathy 

from NTP (1989). Therefore the POD based on atrophy and degeneration of renal tubules from 

the Gorzinski et al. (1985) study was selected to serve as the basis for the derivation of the RfD. 
 DO NOT CITE OR QUOTE 95

Table 5-3. Potential points of departure (PODs) for nephrotoxicity in male 

rats with applied uncertainty factors (UFs) and potential reference values 

 

Potential PODs 

(mg/kg-day) 

Total 

UF UFA UFH UFS UFD

Potential 

Reference Values 

 (mg/kg-day) 

Reference 

Tubular nephropathy 

BMDL 

2-year

16.99 300 10 10 1 3 0.0566 NCI (1978) 

Hyperplasia of pelvic 

transitional 

epithelium 

BMDL 

2-year

4.48 300 10 10 1 3 0.0149 

Linear mineralization 

BMDL 

2-year 

3.22 300 10 10 1 3 0.0107 

Moderate to Marked 

Tubular Nephropathy 

BMDL 

2-year 

1.88 300 10 10 1 3 0.0075 

NTP (1989) 

Slight hypertrophy 

and/or dilation of 

proximal convoluted 

renal tubules 

BMDL 

16 week

0.710 3000 10 10 10 3 0.0002 

Atrophy and 

degeneration of renal 

tubules 

BMDL 

16 week 

0.728 3000 10 10 10 3 0.0002 

Gorzinski et 

al. (1985) 
96 DO NO

Figure 5-1. Array of potential points of departure with applied uncertainty factors and potential refernephrotoxic effects of studies in Table 5-3. 

Increased 

incidence of 

atrophy and 

degeneration of 

tubules BMDL 

Gorzinski et 

al., 1985 

16-wk 

Point UFA, I

UFH, IUFS, S

 C UFD, 

Refere

Increased 

incidence of 

moderate to 

marked 

tubular 

nephropathy 

BMDL 

NTP, 1989 

2-yr 

Increased incidence 

linear 

mineralization 

BMDL 

NTP, 1989 

2-yr 

Increased 

incidence of 

tubular 

nephropathy 

BMDL 

NCI, 1978 

91-wk 

Increased 

incidence of 

hyperplasia of 

the pelvic 

transitional 

epithelium 

BMDL 

NTP, 1989 

2-yr

Increased 

incidence of 

slight 

hypertrophy 

and/or dilation 

of convoluted 

renal tubules 

BMDL 

Gorzinski et 

al., 1985 

16-wk

0.0001

0.001

0.01

0.1

1

10

100

Dose (mg/kg-day)
97 DO NOT CITE OR QUOTE

5.1.5. Previous RfD Assessment

 In the previous RfD assessment for HCE, completed in 1987, the Gorzinski et al. (1985) 

study was employed in deriving the RfD using the NOAEL/LOAEL approach. In this study, the 

identified LOAEL for atrophy and degeneration of renal tubules was 15 mg/kg-day, with a 

corresponding NOAEL of 1 mg/kg-day. A composite UF of 1,000 was employed to account for 

the following three limitations or uncertainties: (1) interspecies extrapolation (UFA = 10), 

(2) intraspecies variation (UFH = 10), and (3) subchronic-to-chronic extrapolation (UFS = 10). 

An RfD of 1 × 10–3 mg/kg-day was derived. 

5.2. INHALATION REFERENCE CONCENTRATION (RfC) 

5.2.1. Choice of Principal Study and Critical Effect—with Rationale and Justification

 The short-term inhalation toxicity studies on HCE is limited. Human studies 

demonstrated HCE exposure in smoke bomb production workers, but the sample sizes are too 

small to reach definitive conclusions regarding health effects and the exposure was likely a 

mixture of HCE and zinc oxide. There are no chronic studies available, and only a single 

subchronic inhalation study (in four species) which included a developmental toxicity 

experiment is available. Weeks et al. (1979) exposed Sprague-Dawley rats, male beagle dogs, 

male Hartley guinea pigs, and Japanese quail to HCE air concentrations of 145, 465, or 2,517 

mg/m3 for 6 hours/day, 5 days/week, for 6 weeks. Postexposure observations were carried out 

for 12 weeks. 

As discussed in Section 4.4.3.2, toxic effects observed in treated rats, dogs, and guinea 

pigs (the quail did not show signs of toxicity) were at the highest exposure level, 2,517 mg/m3

, 

except for dams in of the 465 mg/m3 exposure group in the developmental study; which 

exhibited significantly decreased body weight gain and an increased incidence (85%) of 

mucopurulent nasal exudate. This inflammatory exudate was observed in 100% of the dams 

treated with 2,517 mg/m3. Similar to the dams, male and female rats exposed to 2,517 mg/m3 

HCE for 6 weeks exhibited mucopurulent exudate in the nasal turbinates. Excess mucus in the 

nasal turbinates was also observed in 2/10 quail in the 2,517 mg/m3

 concentration group. Effects 

of this nature were not observed in the 465 or 145 mg/m3 rats and quail or in the treated guinea 

pigs and dogs. 

Weeks et al. (1979) concluded that the excess mucus in two of the 2,517 mg/m3

 quails 

was a transient effect of the HCE exposure because there was no evidence of inflammatory cells 

or tissue damage. The authors attributed the increased incidence of respiratory lesions in rats to 

an endemic mycoplasia infection as evidenced by the histopathological observation of an 

increased incidence and severity of mycoplasia-related lesions in the nasal turbinates 

(mucopurulent exudate), trachea (lymphoid hyperplasia in the lamina propria), and lung 

(pneumonitis) of 2,517 mg/m3 male and female rats. Similar lesions characteristic of respiratory 
98 DO NOT CITE OR QUOTE

mycoplasmosis in rodents were detected in an oral developmental study in rats that paralleled the 

inhalation developmental study described above (both conducted by Weeks et al., 1979). 

Irritation of the upper respiratory tract was observed in approximately 70% of the pregnant rats 

(20% diagnosed with subclinical pneumonitis) orally exposed to 500 mg/kg HCE, compared 

with 10% of controls showing irritation and pneumonitis. 

The presence of the infection in the rats in both the oral and inhalation studies and in the 

controls of the oral study suggests that respiratory tract effects are a potentiation of the 

underlying mycoplasia infection rather than a direct result of HCE exposure. Additionally, the 

reduced weight gain in the rats is likely related to the condition of the infected animals, 

considering mycoplasma-infected rodents generally gain less weight or lose weight compared 

with noninfected rodents (Xu et al., 2006; Sandstedt et al., 1997). Reduced weight gain was also 

observed in the 2,517 mg/m3 guinea pigs, but mycoplasma infection was not reported (Weeks et 

al., 1979). Like rats and mice, guinea pigs can carry mycoplasma organism; however, they are 

not clinically affected (Fox et al., 1984; Holmes, 1984). However, no data was presented 

demonstrating the presence of mycoplasma in the lungs. Therefore, the respiratory tract effects 

cannot be excluded from consideration as a potential critical effect. 

As discussed in Section 4.4.3, neurobehavioral effects were consistently observed in the 

rats and dogs exposed to 2,517 mg/m3. The male and female rats in the 6-week study exhibited 

tremors and ruffled pelt. The pregnant rats developed tremors on GDs 12-16. Similarly, the 

dams exposed to 500 mg/kg HCE in the concurrent oral developmental study by Weeks et al. 

(1979) experienced tremors on GDs 15 and 16 of the 11-day exposure period. The HCE-exposed 

dogs showed tremors, ataxia, and hypersalivation, severe head bobbing, facial muscular 

fasciculations, and closed eyelids. These effects were noted in the dogs throughout the study, 

although they disappeared overnight during nonexposure time periods. 

 Supporting data for the study was reported in an acute study by Weeks and Thomasino 

(1978), in which a single 8-hour inhalation exposure to 2,500 or 57,000 mg/m3 HCE and a single 

6-hour exposure to 17,000 mg/m3 HCE in male Sprague-Dawley rats resulted in neurological and 

lung effects. The male rats exposed to 57,000 mg/m3

 HCE had reduced body weight gain 

compared with controls over the 14 days postexposure. By 6 hours of exposure, one rat had a 

staggered gait. Necropsy did not reveal any gross exposure-related lesions, although microscopy 

revealed that two of these rats had subacute diffuse interstitial pneumonitis of minimal to 

moderate severity and vascular congestion associated with these lung effects. Following 6 hours 

of exposure to 17,000 mg/m3, the six rats in this group showed reduced weight gain compared 

with controls and two of these rats exhibited a staggered gait. No exposure-related gross or 

histopathological changes were observed in tissues and organs. These effects were not 

noticeable 14 days postexposure. 

The short-term inhalation study by Weeks et al. (1979), as the only repeated exposure 

study available, was selected as the principal study for the derivation of the RfC. This study 
99 DO NOT CITE OR QUOTE

used three concentrations and incorporated a variety of endpoints (toxicological, teratological, 

neurological, pulmonary) across a range of species. The primary limitation of Weeks et al. 

(1979) is the minimal amount of quantitative information provided characterizing the reported 

effects. Several experiments only utilized one sex, and additional exposure concentration(s) 

between the mid- and high-concentration would have allowed for better characterization of the 

exposure-response curve. However, this study identified neurotoxicity, statistically significant 

decreases in body weight gain, and upper and lower respiratory tract irritation. The responses 

were generally observed following exposure to the highest concentration, and not in the two 

lower concentrations. Considering the uncertainty surrounding the body weight changes and 

respiratory tract irritation in the presence of the mycoplasma infection, these effects were not 

considered as the critical effect for the derivation of the RfC. Therefore, neurological effects 

following inhalation exposure to HCE were selected as the critical effect. 
100 DO NOT CITE OR QUOTE

Table 5-4. Noncancerous effects observed in animals exposed to HCE via 

inhalation 

Source: Lattanzi et al. (1988). 

5.2.2. Methods of Analysis—Including Models

 The Weeks et al. (1979) study included three exposure groups (145, 465, and 

2,517 mg/m3

) plus a control. Neurological effects were observed in male and female SpragueDawley rats, male beagle dogs, and pregnant Sprague-Dawley rats only at the highest dose 

tested. Incidence data were not reported. Application of BMD modeling was precluded because 

100% of the high dose animals displayed neurological effects. Therefore, a NOAEL served as 

the POD. The NOAEL of 465 mg/m3

, identified in Weeks et al. (1979), was selected as the POD 

for the derivation of the RfC based on effects in male and female rats and male dogs exposed to 

HCE for 6 weeks and pregnant rats exposed on GDs 6-16. Although the NOAELs are the same, 

the male and female rats exposed to HCE for 6 weeks were selected as the study animals upon 

which to base the POD, as the duration of exposure for the dams in the teratology study was only 

11 days and only 4 male dogs were exposed to HCE in the 6-week study. 

The NOAEL is based on intermittent HCE inhalation exposures in male and female rats 

for 6 hours/day, 5 days/week. Thus, prior to deriving the RfC, this POD was adjusted for 

Species Dose/Duration NOAEL 

(mg/m3) 

LOAEL 

(mg/m3) 

Effect 

Sprague-Dawley 

rats (25/sex/dose) 465 mg/m3

 2,517 mg/m3

Males: neurotoxic effects (tremors and 

ruffled pelt), reduced body weight gain, 

increased relative, spleen, and testes 

weights 

Females: neurotoxic effects (tremors and 

ruffled pelt), increased relative liver weight 

Male beagle dogs 

(4/dose)Male 

beagle dogs 

(4/dose) 

465 mg/m3

 2,517 mg/m3 Tremors, ataxia, hypersalivation, head 

bobbing, facial muscular fasciculations 

Male Hartley 

guinea pigs 

(10/dose)Male 

Hartley guinea 

pigs (10/dose) 

465 mg/m3

 2,517 mg/m3 Reduced body weight, increased relative 

liver weight 

Coturnix 

Japonica quail 

(20/dose)Coturnix 

Japonica quail 

(20/dose) 

0, 145, 465, or 

2,517 mg/m3; 

6 weeks 

2517 mg/m3 Not established No effects observed 

Pregnant 

Sprague-Dawley 

rats (22/dose) 

0, 145, 465, 

or 2,517 mg/m3 

on GDs 6-16 

Maternal: 

465 mg/m3

Maternal: 

2,517 mg/m3

Maternal: tremors 

Fetal: no effects 
101 DO NOT CITE OR QUOTE

continuous exposure (24 hours/day, 7 days/week). The duration-adjusted POD (POD[ADJ]) is 

derived using the following equation (U.S. EPA, 1994b): 

POD[ADJ] = (POD) × (hours of exposure/24 hours) × (days of exposure/7 days) 

 = (465 mg/m3) × (6/24 hours) × (5/7 days) 

 = 83.0 mg/m3

. 

 The Methods for Derivation of Inhalation Reference Concentrations and Application of 

Inhalation Dosimetry (hereafter referred to as the RfC Methodology) recommends converting the 

POD[ADJ] to a human equivalent concentration (HEC) (U.S. EPA, 1994b). The RfC 

Methodology separates gases into three categories based on their water solubility and reactivity 

with tissues in the respiratory tract. Determining whether HCE is a Category 2 or 3 gas is 

difficult because data regarding the inhalation effects are very limited. HCE is slightly water 

soluble and although HCE has been observed in blood following oral exposures to HCE, it is 

unknown whether HCE accumulates in blood following inhalation exposure. Given this limited 

information, HCE is likely a Category 2 gas because it is slightly water soluble and causes 

effects distal to the site of inhalation exposure (i.e., systemic effects). For Category 2 gases, 

HEC values are calculated using methods for Category 1 gases for portal-of-entry effects and 

Category 3 methods for systemic effects (U.S. EPA, 1994b). In view of the fact that 

neurotoxicity is a systemic effect, the methods for Category 3 gases were used to derive the 

HEC. 

The RfC Methodology (U.S. EPA, 1994b) suggests that HECs be estimated by applying 

to the duration-adjusted exposure level (POD[ADJ]), a dosimetric adjustment factor (DAF) that is 

specific for the breathing characteristic of the species to be compared. The DAF for a Category 

3 gas is based on the regional gas dose ratio (RGDR), where the RGDR is the ratio of the animal 

blood:gas partition coefficient (Hb/g)A and the human blood:gas partition coefficient (Hb/g)H. 

POD[HEC] = POD[ADJ] × (Hb/g)A/(Hb/g)H 

 However, the human and animal blood partition coefficients for HCE are not known. In 

accordance with the RfC Methodology (U.S. EPA, 1994) when the partition coefficients are 

unknown a ratio of 1 is used. This results in a NOAEL[HEC] of 83.0 mg/m3. 

POD[HEC] = POD[ADJ] × (Hb/g)A/(Hb/g)H

 = 83.0 mg/m3× 1 

 = 83.0 mg/m3

5.2.3. RfC Derivation—Including Application of Uncertainty Factors (UFs)
102 DO NOT CITE OR QUOTE

 The NOAEL[HEC] value of 83 mg/m3 for evidence of neurotoxicity in Sprague-Dawley 

rats was used as the POD to derive the RfC for HCE. A composite UF of 3,000 was applied as 

follows: 

• For animal-to-human interspecies differences (UFA), an UF of 3 was applied to account 

for the uncertainty in extrapolating from laboratory animals to humans. This value is 

adopted by convention, where an adjustment from an animal-specific NOAELADJ to a 

NOAELHEC has been incorporated. Application of an UF of 10 would depend on two 

areas of uncertainty (i.e., toxicokinetic and toxicodynamic uncertainties). In this 

assessment, the toxicokinetic component associated with HCE is mostly addressed by the 

determination of an HEC as described in the RfC methodology (U.S. EPA, 1994b). The 

toxicodynamic uncertainty is also accounted for to a certain degree by the use of the 

applied dosimetry method and an UF of 3 is retained to account for uncertainty regarding 

the toxicodynamic differences between rats and humans. 

• An intraspecies UF (UFH) of 10 was applied to account for potentially sensitive human 

subpopulations in the absence of information on the variability of response to HCE in the 

human population. Information is currently unavailable to assess human-to-human 

variability in HCE toxicokinetics and toxicodynamics. 

• A subchronic-to-chronic UF (UFS) of 10 was applied to account for the use of the POD 

selected following a short-term duration of exposure to HCE to estimate a chronic 

exposure RfC. 

• An UF for a LOAEL to a NOAEL extrapolation was not applied because this assessment 

utilized a NOAEL as the POD. 

• A 10-fold UF was used to account for deficiencies in the toxicity database on inhalation 

exposure to HCE. There are no available human occupational or epidemiological studies 

of inhalation exposure to HCE. There are no chronic toxicity or multigeneration 

reproductive toxicity animal studies available for inhalation exposure to HCE. The 

toxicity data on inhalation exposure to HCE is very limited and largely restricted to one 

subchronic (6-week) inhalation study (Weeks et al., 1979) in rats, male dogs, male 

rabbits, and quail. The same investigators performed a developmental study and an acute 

study in rats. Maternal toxicity was observed at both doses. Fetuses of HCE-treated 

dams did not exhibit any significant skeletal or soft tissue anomalies. The toxic effects 

observed in the dams in the developmental study were similar to those observed in the 

rats exposed for 6 weeks, although additional effects were observed in the rats exposed 
103 DO NOT CITE OR QUOTE

for a longer duration. The absence of teratogenic effects does not abrogate concern given 

the paucity of the inhalation database for HCE. In addition, the database lacks studies of 

neurotoxicity and developmental neurotoxicity, endpoints of concern based on the 

available inhalation data. Therefore, in consideration of the inhalation database for HCE, 

a database UF of 10 was applied. 

Given the UFs established above, the RfC for HCE was calculated employing the following 

equation: 

RfC = NOAEL[HEC] ÷ UF 

 = 83 mg/m3 ÷ 3,000 

 = 0.028 mg/m3

 or 3 × 10–2 mg/m3. 

5.2.4. RfC Comparison Information 

 The predominant noncancer effect of short-term inhalation exposure to HCE is 

neurotoxicity. The other effects noted by Weeks et al. (1979) at the same dose level were 

decreases in body weight and increases in organ (liver or kidney) weights in male guinea pigs, 

male and female rats, and pregnant rats. As discussed in Sections 5.2.1 and 5.2.2, the 

neurotoxicity reported in the available inhalation study (Weeks et al., 1979) was selected for the 

RfC derivation because of the potential impact of the mycoplasma infection on the other 

endpoints. Based on the lack of alternative endpoints to be considered for the basis of the RfC, a 

graphical display of dose-response information from the short-term inhalation study was not 

provided. For the reasons discussed above and in Section 5.2.1, neurotoxic effects in male and 

female rats, pregnant rats, and male dogs reported by Weeks et al. (1979) are considered the 

most sensitive effects and were selected to serve as the basis for the derivation of the RfC for 

HCE. 

5.2.5. Previous RfC Assessment

 An RfC for HCE was not previously developed by the U.S. EPA. In the 1987 IRIS 

Summary, Weeks et al. (1979) was briefly summarized in the Additional Studies/Comments 

section for the oral RfD. The IRIS Summary (1987) stated that Weeks et al. (1979) administered 

HCE to rats by inhalation at 145, 465, or 2520 mg/mg

3

, 6 hours/day during gestation. At the two 

highest doses, maternal toxicity was observed but there was no evidence of fetoxicity or 

teratogenicity. No additional discussion was presented in the IRIS Summary (1987) describing 

why this study was not used to develop an RfC. 

5.3. UNCERTAINTIES IN THE ORAL REFERENCE DOSE AND INHALATION 

REFERENCE CONCENTRATION 
104 DO NOT CITE OR QUOTE

 The following discussion identifies uncertainties associated with the quantification of the 

RfD and RfC for HCE. Following EPA practices and guidance (U.S. EPA, 1994b, 1993), the UF 

approach was applied to the chosen PODs to derive an RfD and RfC (see sections 5.1.3 and 

5.2.3). Factors accounting for uncertainties associated with a number of steps in the analyses 

were adopted to account for extrapolating from an animal study to human exposure, a diverse 

human population of varying susceptibilities, and database deficiencies. 

The oral database includes short-term, subchronic, and chronic studies in rats and a 

chronic study in mice, and developmental studies in rats. Toxicity associated with oral exposure 

to HCE is predominantly reported as kidney toxicity, specifically, renal tubule nephropathy. The 

inhalation database includes a short-term study in rats, pregnant rats, male dogs, male guinea 

pigs, and quail. Toxicity associated with inhalation exposure to HCE in this study is mainly 

neurotoxicity. Critical data gaps have been identified in Chapter 4 and uncertainties associated 

with data deficiencies are more fully discussed below. 

 The RfD was derived from a BMDL10 of 0.728 mg/kg-day, which was based on the 

observation of atrophy and degeneration of renal tubules in male F344 rats from the Gorzinski et 

al. (1985) 16-week toxicity study. The dose-response relationships for oral exposure to HCE and 

nephropathy in other studies of rats are also available for deriving an RfD, but are associated 

with higher NOAELs/LOAELs that are less sensitive than would be protected by the selected 

critical effect and corresponding POD. After consideration of all potential PODs, the RfD of 

2 × 10–4 mg/kg-day was based on the observation of atrophy and degeneration of renal tubules in 

male F344 rats from the Gorzinski et al. (1985) 16-week toxicity study. 

 The derived RfD was quantified using a BMDL10 for the POD. The selection of the 

BMD model for the quantitation of the RfD does not lead to significant uncertainty in estimating 

the POD since benchmark effect levels were within the range of experimental data. However, 

the selected models do not represent all possible models one might fit, and other models could be 

selected to yield different results, both higher and lower than those included in this assessment. 

Uncertainty exists in the selection of the BMR level utilized in the BMD modeling of the critical 

effect (atrophy and degeneration of renal tubules in male F344 rats) to determine the POD. In 

the absence of clear information to determine the level of change in atrophy and degeneration of 

renal tubules in male F344 rats related to a biologically significant change, a benchmark 

response (BMR) of 10% was selected for the modeling of the increased incidence was selected to 

represent a minimally biologically significant change. 

 The RfC was derived from NOAEL[HEC] value of 83 mg/m3 for evidence of neurotoxicity 

in Sprague-Dawley rats from a short-term (6 week) inhalation study by Weeks et al. (1979). A 

POD based on a NOAEL or LOAEL is, in part, a reflection of the particular exposure 

concentration or dose at which a study was conducted. It lacks characterization of the doseresponse curve and for this reason is less informative than a POD obtained from benchmark 

dose-response modeling. The short-term inhalation study in rats (Weeks et al., 1979) was 
105 DO NOT CITE OR QUOTE

selected as the principal study and neurotoxicity was identified as the critical effect. A NOAEL 

of 465 mg/m3 was selected to serve as the POD and the basis for derivation of the RfC. 

 Extrapolating from animals to humans adds further uncertainty. The effect and its 

magnitude at the concentration in rats are extrapolated to human response. Pharmacokinetic 

models are useful for examining species differences in pharmacokinetic processing; however, 

dosimetric adjustment using pharmacokinetic modeling was not available for oral exposure to 

HCE. Information was unavailable to quantitatively assess toxicokinetic or toxicodynamic 

differences between animals and humans, so the 10-fold UF was used to account for uncertainty 

in extrapolating from laboratory animals to humans in the derivation of the RfD. For the chronic 

RfC, a factor of 3 was adopted by convention where an adjustment from an animal specific 

NOAELADJ to a NOAELHEC has been incorporated. Application of a full UF of 10 would depend 

on two areas of uncertainty (i.e., toxicokinetic and toxicodynamic uncertainties). In this 

assessment, the toxicokinetic component is mostly addressed by the determination of a HEC as 

described in the RfC methodology (U.S. EPA, 1994b). The toxicodynamic uncertainty is also 

accounted for to a certain degree by the use of the applied dosimetry method and an UF of 3 is 

retained to account for this component. 

 Heterogeneity among humans is another uncertainty associated with extrapolating doses 

from animals to humans. Uncertainty related to human variation needs consideration in 

extrapolating dose from a subset or smaller sized population, say of one sex or a narrow range of 

life stages typical of occupational epidemiologic studies, to a larger, more diverse population. In 

the absence of HCE-specific data on human variation, a factor of 10 was used to account for 

uncertainty associated with human variation in the derivation of both the RfD and RfC. Human 

variation may be larger or smaller; however, HCE-specific data to examine the potential 

magnitude of over- or under-estimation are unavailable. 

 Uncertainties associated with data gaps in the HCE database have been identified. Data 

more fully characterizing potential multigenerational reproductive effects associated with both 

oral and inhalation HCE exposure are lacking. The oral database includes studies in laboratory 

animals, including chronic and subchronic dietary exposure studies and two oral developmental 

toxicity studies. The developmental studies show effects at doses higher than those observed to 

induce renal toxicity in the subchronic and chronic toxicity studies. Therefore, in consideration 

of the entire oral database for HCE, a database UF of 3 was considered appropriate to account for 

the lack of a two-generational reproductive study. There are no available human occupational or 

epidemiological studies of inhalation exposure to HCE. There are no standard chronic toxicity 

or multigeneration reproductive toxicity animal studies available for inhalation exposure to HCE. 

The toxicity data on inhalation exposure to HCE is very limited and largely restricted to one 

subchronic (6-week) inhalation study (Weeks et al., 1979) in rats, male dogs, male rabbits, and 

quail. The same investigators performed a developmental study and an acute study in rats. The 

developmental study in rats did not provide any evidence of teratogenic effects. However, this 
106 DO NOT CITE OR QUOTE

data does not abrogate concern given the paucity of the inhalation database for HCE. In 

addition, the inhalation database lacks studies of developmental neurotoxicity, endpoints of 

concern based on the available inhalation data (critical effect for the RfC). Therefore, in 

consideration of the inhalation database for HCE, a database UF of 10 is was applied. 

5.4. CANCER ASSESSMENT

There are no available studies on cancer in humans associated with exposure to HCE. 

NTP (1989) provided evidence of renal adenomas and carcinomas and pheochromocytomas and 

malignant pheochromocytomas in male F344/N rats in a 2-year cancer bioassay. NCI (1978) 

provided evidence of hepatocellular carcinomas in male and female B6C3F1 mice in a 91-week 

cancer bioassay. Additionally, HCE was shown to be a promoter, but not an initiator, in an 

Osborne-Mendel rat liver foci assay (Milman et al., 1988; Story et al., 1986). Binding of 

radiolabeled carbon to DNA, RNA, and protein was observed following [14C]HCE 

administration in both in vitro and in vivo assays in mice and rats (Lattanzi et al., 1988). 

Under the Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a), HCE is “likely 

to be carcinogenic to humans” based on dose-dependent, statistically significant increases in the 

incidence of renal adenoma or carcinoma combined in male F344/N rats, statistically significant 

increases in the incidence of pheochromocytomas/malignant pheochromocytomas combined in 

male F344/N rats (NTP, 1989), and statistically significant increases in the incidence of 

hepatocellular carcinomas in male and female B6C3F1 mice (NCI, 1978). 

5.4.1. Choice of Study/Data—with Rationale and Justification

 Two animal studies were selected for BMD analysis and subsequent quantitative cancer 

assessment. In the first study, NTP (1989) reported statistically significantly elevated incidences 

of renal adenomas and carcinomas combined and pheochromocytomas, malignant 

pheochromocytomas, and complex pheochromocytomas combined in male F344 rats 

administered HCE via gavage for 2 years. Female rats in this study did not exhibit any 

HCE-related tumors. In the second study, NCI (1978) reported statistically significantly elevated 

incidences of hepatocellular carcinomas in both sexes of B6C3F1 mice administered HCE via 

gavage for 78 weeks. However, male mice in this study demonstrated a dose-response 

relationship, while female mice did not. 

5.4.2. Dose-response Data

 The NTP (1989) administered, via gavage, TWA doses of 7 and 14 mg/kg-day HCE to 

male and female F344/N rats for 103 weeks. No HCE-related tumors were observed in female 

rats. Renal adenomas and carcinomas combined were observed in 2, 4, and 14% (statistically 

significant) of male rats administered 0 (controls), 7, and 14 mg/kg-day HCE, respectively. Male 

rats also exhibited increased incidences of pheochromocytomas and malignant 
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pheochromocytomas combined; 28, 58 (statistically significant), and 39% in the controls, 7 and 

14 mg/kg-day dose groups, respectively (NTP, 1989). The NCI (1978) gavage study 

administered TWA doses of 0, 360, and 722 mg/kg-day HCE to male and female B6C3F1 mice 

for 91-weeks. Statistically significant increases in the incidence of hepatocellular carcinomas 

were observed in 15, 30, and 63% of males and 10, 40, and 31% of females in the control, 360, 

and 722 mg/kg-day dose groups, respectively 

Both NTP (1989) and NCI (1978) are well-designed studies, conducted in both sexes of 

two species with 50 animals/sex/dose. Each study utilized two dose groups of HCE and an 

untreated control group, with examination of a wide range of toxicological endpoints in both 

sexes of the rodents. Tumor incidences were elevated over controls at two sites in rats (NTP, 

1989) and at one site in mice (NCI, 1978). Some limitations associated with the NCI (1978) 

study in mice include changes to the dosing regimen 9 weeks into the study, cyclical dosing 

periods, and decreased in all study groups for the male mice. Individual animal data were 

unavailable to perform time-to-tumor modeling or adjust the tumor incidences for survival before 

BMD modeling. The cancer incidence data are summarized in Table 5-5. 

Table 5-5. Summary of incidence data in rodents orally exposed to HCE for use in 

cancer dose-response assessment

Study Sex/strain/species Endpoint HCE dose 

(mg/kg-day) Incidence 

0 1/50 (2%) 

NTP (1989) Male F344 rats 7.1 2/50 (4%) Kidney adenoma or 

carcinoma 

14.3 7/50 (14%)a

0 14/50 (28%) 

7.1 26/45 (58%)a NTP (1989) Male F344 rats 

Pheochromocytomas/ 

malignant 

pheochromocytomas 14.3 19/49 (39%) 

0 3/20 (15%)

360 15/50 (30%)a NCI (1978) Male B6C3F1 mice Hepatocellular 

carcinoma 

722 31/49 (63%)a

0 2/20 (10%) 

360 20/50 (40%)a NCI (1978) Female B6C3F1

mice 

Hepatocellular 

carcinoma 

722 15/49 (31%)a

_ 

aDenotes statistical significance 

5.4.3. Dose Adjustments and Extrapolation Methods

 The HCE doses administered to laboratory animals were scaled to human equivalent 

doses (HEDs) according to EPA guidance (U.S. EPA, 2005a, 1992). More specifically, animal 

doses were converted to HEDs by assuming that doses in animals and humans are toxicologically 

equivalent when scaled by body weight raised to the 3⁄4 power, as follows: 

( )

[ ]

( )

[ ] human

human

animal

animal

BW

Dose mg day

BW

Dose mg day

/3 4

[ ]

/3 4

/ [ ] /

=
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The body weights for the laboratory animals used in the scaled human dose conversions are the 

mean body weights reported in the studies for each dose group. The following formula was used 

for the conversion of oral animal doses to oral HEDs: 

Scaled human dose (HED) = animal dose × (animal body weight/human body weight)1⁄4

Therefore, the HCE doses of 7 and 14 mg/kg-day employed by NTP (1989) in rats were 

converted to HEDs, as follows: 

 Scaled human dose (HED) = 7 mg/kg-day × (0.483 kg/70 kg)1⁄4

 = 2.05 mg/kg-day 

 Scaled human dose (HED) = 14 mg/kg-day × (0.471 kg/70 kg)1⁄4

 = 4.10 mg/kg-day 

Similarly, the HCE doses of 360 and 722 mg/kg-day employed by NCI (1978) in mice were 

converted to HEDs, as follows: 

 Scaled human dose (HED) = 360 mg/kg-day × (0.033 kg/70 kg)1⁄4

 = 53.05 mg/kg-day 

 Scaled human dose (HED) = 722 mg/kg-day × (0.030 kg/70 kg)1⁄4

 = 103.88 mg/kg-day 

These scaled human doses are used in the dose-response modeling described below. 

The multistage model was the primary model considered for fitting the dose-response 

data and is given by: 

P(d) = 1 – exp[–(q0 + q1d + q2d

2

 + ... + qkd

k

)], 

where: 

P(d) = lifetime risk (probability) of cancer at dose d 

qi

 = parameters estimated in fitting the model, i = 1, ..., k 

And extra risk is defined as (P(d) –P(0))/(1-P(0)).
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 The multistage model in BMDS (version 2.0) (U.S. EPA, 2008) was used to fit the 

multistage dose-response model to the incidence data summarized in Table 5-5 using the 

calculated HEDs in order to derive an oral slope factor for HCE. The BMR selected was the 

default value of 10% extra risk recommended for dichotomous models (U.S. EPA, 2000b). No 

data were excluded from the BMD multistage modeling. 

As stated above, the multistage model was fit to the incidences of renal adenomas or 

carcinomas combined in male rats and hepatocellular carcinomas in male and female mice. In all 

cases, the 2° multistage model provided the best fit. The multistage model was also fit to the 

incidence of pheochromocytomas or malignant pheochromocytomas in male rats. The model 

exhibited a significant lack of fit for the pheochromocytomas (according to the χ

2 statistic with 

p < 0.01). Thus, this dataset was not useful for dose-response assessment because the tumor 

incidence is not a monotonic increasing function of dose, as demonstrated by the 

Cochran-Armitage Trend Test. Therefore, the BMD modeling results for the kidney and liver 

tumors in rats and mice, respectively, are summarized in Table 5-6, with more detailed results 

contained in Appendix B. 

Table 5-6. Summary of BMD modeling results for oral cancer assessment of HCE 

Study Sex/strain/species Endpoint “Best-fit” 

Model 

pvalue AIC BMR BMD10BMDL10

or POD 

Oralr 

Slope 

Factor 

(mg/kgday)–1

NTP 

(1989) Male F344 rats 

Renal 

adenomas/carcinomas 

combined 

2° 

Multistage 0.75 71.19 0.1 3.73 2.44 0.040984 

NCI 

(1978) Male B6C3F1 mice Hepatocellular carcinomas 

2° 

Multistage 0.83 146.47 0.1 37.03 14.44 0.006925 

NCI 

(1978) 

 Female B6C3F1

mice 

 Hepatocellular 

carcinomas 

2° 

Multistage 0.03 149.77 0.1 286.24 136.88 0.000730 

The U.S. EPA Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a) 

recommend that the method used to characterize and quantify cancer risk from a chemical is 

determined by what is known about the mode of action of the carcinogen and the shape of the 

cancer dose-response curve. The linear approach is used as a default option if the mode-ofaction of carcinogenicity is not understood (U.S. EPA, 2005a). In the case of HCE, the mode of 

carcinogenic action of HCE in the kidneys and livers of rats and mice is unknown. There is 

some data in experimental animals evaluating α2u-globulin accumulation and toxicity in the 

kidney. As described in Section 4.7.3.1., two principal factors contribute to the conclusion that 

the available data do not support an α2u-globulin mode of action for the development of renal 

tumors: (1) the lack of information identifying the α2u-globulin protein in HCE-treated rats, and 

(2) evidence of nephropathy in female rats as well as male and female mice (because the α2u-
110 DO NOT CITE OR QUOTE

globulin-related mode of action is specific for male rats). Therefore, a linear low-dose 

extrapolation approach was used to estimate human carcinogenic risk associated with HCE 

exposure. 

5.4.4. Oral Slope Factor and Inhalation Unit Risk

The oral slope factor was derived from the BMDL10 (the lower bound on the exposure 

associated with a 10% extra cancer risk) by dividing the BMR by the BMDL10 and represents an 

upper bound on cancer risk associated with a continuous lifetime exposure to HCE. In 

accordance with the U.S. EPA (2005a) guidelines, an oral slope factor for renal tumors in male 

rats of 0.04 (mg/kg-day)-1 was calculated by dividing the human equivalent BMDL10 of 2.44 

mg/kg-day into 0.1 (10%) (Appendix B). An oral slope factor for hepatocellular tumors in male 

mice of 0.007 (mg/kg-day)-1 was calculated by dividing the human equivalent BMDL10 of 14.44 

mg/kg-day into 0.1 (10%) (Appendix B). An oral slope factor for hepatocellular tumors in 

female mice of 0.0007 (mg/kg-day)-1 was calculated by dividing the human equivalent BMDL10

of 136.88 mg/kg-day into 0.1 (10%) (Appendix B).The oral slope factors were derived by linear 

extrapolation to the origin from the POD and represent upper-bound estimates. The rats 

exhibited greater sensitivity to HCE-induced carcinogenicity than the mice. Thus, the risk 

estimate associated with the male rats that developed renal adenomas or carcinomas was selected 

as the oral slope factor of 0.04 (mg/kg-day)-1 for HCE. The slope of the linear extrapolation 

from the central estimate (i.e., BMD) is 0.1/37.03 mg/kg-day or 3 x 10-3 (mg/kg-day)-1. 

 In the absence of any suitable data on the carcinogenicity of HCE via the inhalation route, 

an inhalation unit risk has not been derived in this evaluation. 

5.4.5. Uncertainties in Cancer Risk Values 

 Extrapolation of data from animals to estimate potential cancer risks to human 

populations from exposure to HCE yields uncertainty. Several types of uncertainty may be 

considered quantitatively, whereas others can only be addressed qualitatively. Thus, an overall 

integrated quantitative uncertainty analysis cannot be developed. Major sources of uncertainty in 

the cancer assessment for HCE are summarized in Section 5.4.5.1 and in Table 5-7. 
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Table 5-7. Summary of uncertainties in the HCE cancer risk assessment

Consideration/ 

approach 

Impact on oral 

slope factora Decision Justification 

Human relevance 

of male mouse 

tumor data 

Human risk could ↓

or ↑, depending on 

relative sensitivity; 

if rodent tumors 

proved not to be 

relevant to humans, 

oral cancer risk 

estimate would not 

apply, (i.e., human 

risk would ↓)

Kidney and adrenal 

gland tumors in male 

rats and liver tumors 

in male and female 

mice are relevant to 

human exposure 

There are no mode of action data to guide 

extrapolation approach for any choice. It 

was assumed that humans are as sensitive as 

the most sensitive rodent gender/species 

tested; true correspondence is unknown. The 

carcinogenic response occurs across species. 

HCE is a multi-site carcinogen although 

direct site concordance is generally not 

assumed (U.S. EPA, 2005a); consistent with 

this view, some human tumor types are not 

found in rodents. 

Bioassay 

Alternatives could ↑

or ↓ oral slope factor 

by an unknown 

extent 

NTP study 

Alternative bioassays in rats were 

unavailable. NCI (1978) bioassay in mice 

was available, although mice were less 

sensitive than rats to HCE carcinogenicity 

and were not utilized in estimating 

carcinogenic risk to humans. 

Species/gender 

choice 

Human risk could ↑

or ↓, depending on 

relative sensitivity 

Incidence of renal 

adenoma/carcinoma 

in male rats 

It was assumed that humans are as sensitive 

as the most sensitive rodent gender/species 

tested; true correspondence is unknown. 

Increased tumor incidence in mice resulted in 

lower risk estimate than rats. No increase of 

kidney tumors in female rats. 

Dose metric 

Alternatives could ↑

or ↓ oral slope factor 

by an unknown 

extent 

Used administered 

exposure 

Experimental evidence supports a role for 

metabolism in toxicity, but actual responsible 

metabolites are not identified. 

Low-dose 

extrapolation 

procedure 

Departure from 

EPA’s Guidelines 

for Carcinogen Risk 

Assessment POD

paradigm, if 

justified, could ↓ or 

↑ oral slope factor 

an unknown extent 

Multistage model to 

determine POD, 

linear low-dose 

extrapolation from 

POD (default 

approach) 

Available mode of action data do not inform 

selection of dose-response model; linear 

approach employed in absence of support for 

an alternative approach. 

Cross-species 

scaling 

Alternatives could ↓

or ↑ the oral slope 

factor (e.g., 3.5-fold 

↓ [scaling by body 

weight] or ↑ 2-fold 

[scaling by BW2/3]) 

BW3/4 (default 

approach) 

There are no data to support alternatives. 

Because the dose metric was not an AUC, 

BW3/4 scaling was used to calculate 

equivalent cumulative exposures for 

estimating equivalent human risks. 

Statistical 

uncertainty at POD 

↓ oral slope factor 

1.5-fold ifBMD 

used as the POD 

rather than lower 

bound on POD 

BMDL (preferred 

approach for 

calculating 

reasonable upper 

bound slope factor) 

Limited size of bioassay results in sampling 

variability; lower bound is 95% confidence 

interval on administered exposure. 

Human population 

variability in 

metabolism and 

response/sensitive 

subpopulations 

Low-dose risk ↑ or ↓

to an unknown 

extent 

Considered 

qualitatively 

No data to support range of human 

variability/sensitivity, including whether 

children are more sensitive. 

a

↑ = increase; ↓ = decrease
112 DO NOT CITE OR QUOTE

5.4.5.1. Sources of Uncertainty

Relevance to humans 

The modes of action for the kidney (adenomas/carcinomas) and adrenal gland tumors 

(pheochromocytomas) in male rats and liver tumors (hepatocellular carcinomas) in male and 

female mice are unknown. There is some data in experimental animals evaluating α2u-globulin 

accumulation and toxicity in the kidney. As described in Section 4.7.3, two principal factors 

contribute to the conclusion that the available data do not support an α2u-globulin mode of action 

for the development of renal tumors. First, the presence of kidney effects in HCE-exposed male 

and female mice, which generally do not accumulate the α2u-globulin protein, suggests that a 

mode of action other than α2u-globulin nephropathy. Second, none of the HCE studies 

performed the necessary immunohistochemical assays to confirm the presence of α2u-globulin 

protein within the hyaline droplets observed following administration of HCE (NTP, 1996, 

1989). This represents a data gap, as the presence of α2u-globulin is necessary to support an 

α2u-globulin mode of action. 

The relevance of the mode of action of liver tumor induction to humans was considered 

in Section 4.7.2. There is no evidence in humans for hepatic cancer associated with HCE 

exposure. The experimental animal literature, however, shows that oral exposure to HCE 

induces liver tumors in male and female mice. It is possible that the HCE-induced hepatocellular 

carcinomas in mice occur as a result of the binding of HCE metabolites to liver macromolecules 

and the generation of free radicals during HCE metabolism, causing key events in the 

carcinogenic process such as cytotoxicity, inflammation and regenerative cell proliferation. 

Limited information exists to distinguish the similarities and differences between experimental 

animals and humans in terms of HCE metabolism or toxicity. However, these potential key 

events have not been systematically evaluated for HCE. 

In humans, pheochromocytomas are rare catecholamine-producing neuroendocrine 

tumors that are usually benign, but may also present as or develop into a malignancy (Eisenhofer 

et al., 2004; Lehnert et al., 2004; Edstrom Elder et al., 2003; Goldstein et al., 1999). Hereditary 

factors in humans have been identified as important in the development of pheochromocytomas 

(Eisenhofer et al., 2004). Therefore, in the absence of information indicating otherwise, the 

kidney and adrenal gland tumors in male rats and liver tumors in male and female mice are 

considered relevant to humans. 

Bioassay selection 

 The study by NTP (1989) was used for the development of an oral slope factor. This 

study was conducted in both sexes of F344/N rats and used 50 male and 50 female rats per dose 

group. Test animals were allocated among two dose levels of HCE and an untreated control 

group. Animals were observed twice daily and examined weekly (for 14 weeks) then monthly 
113 DO NOT CITE OR QUOTE

for body weight and monthly for feed consumption. Animals were necropsied and all organs and 

tissues were examined grossly and microscopically for histopathological lesions for a 

comprehensive set of toxicological endpoints in both sexes. 

Choice of species/gender 

The oral slope factor for HCE was quantified using the tumor incidence data for male 

rats, which were found to be more sensitive than male or female mice to the carcinogenicity of 

HCE. The oral slope factor calculated from male rats was higher than the slope factors 

calculated from male and female mice. As there is no information to inform which species or 

gender of animals would be most applicable to humans, the most sensitive group was selected for 

the basis of the oral slope factor. Though the mode of action for observed kidney tumors in 

rodents is unknown, the evidence suggesting the kidney as a target organ of HCE toxicity in both 

species lends strength to the concern for human carcinogenic potential. 

Dose metric 

HCE is likely metabolized to PERC and pentachloroethane; however, it is unknown 

whether a metabolite or some combination of parent compound and metabolites is responsible 

for the observed toxicity and carcinogenicity of HCE. If the actual carcinogenic moiety(ies) 

is(are) proportional to administered exposure, then use of administered exposure as the dose 

metric provides an unbiased estimate of carcinogenicity. On the other hand, if this is not the 

most relevant dose metric, then the impact on the human equivalent slope factor is unknown; the 

low-dose cancer risk value may be higher or lower than that estimated, by an unknown amount. 

Choice of low-dose extrapolation approach 

 The mode of action is a key consideration in clarifying how risks should be estimated for 

low-dose exposure. A linear-low-dose extrapolation approach was used to estimate human 

carcinogenic risk associated with HCE exposure, in the absence of information to inform the 

dose-response at low doses. The extent to which the overall uncertainty in low-dose risk 

estimation could be reduced if the mode of action for HCE were known is of interest, but data on 

the mode of action of HCE are limited and the mode of action is not known. 

 Etiologically different tumor types were not combined across sites prior to modeling, in 

order to allow for the possibility that different tumor types can have different dose-response 

relationships because of varying time courses or other underlying mechanisms or factors. The 

human equivalent oral slope factors estimated from the tumor sites with statistically significant 

increases ranged from 0.007 to 0.04 per mg/kg-day, a range less than one order of magnitude, 

with greater risk coming from the male rat kidney data. 

 
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Choice of model 

All risk assessments involve uncertainty, as study data are extrapolated to make 

inferences about potential effects in humans from environmental exposure. The largest sources 

of uncertainty in the HCE cancer risk estimates are interspecies extrapolation and low-dose 

extrapolation. There are no human data from which to estimate human cancer risk; therefore, the 

risk estimate must rely on data from studies of rodents exposed to levels greater than would 

occur from environmental exposures. 

 Without human cancer data or better mechanistic data, the relevance of the rodent cancer 

results to humans is uncertain. The occurrence of increased incidences of kidney and adrenal 

gland tumors in male rats, and liver tumors in male and female mice exposed to HCE from the 

oral route of exposure suggests that HCE is potentially carcinogenic to humans as well. 

However, the lack of concordance in tumor sites between the two rodent species makes it more 

difficult to quantitatively estimate human cancer risk. 

 Regarding low-dose extrapolation, in the absence of mechanistic data for biologically 

based low-dose modeling or mechanistic evidence supporting a nonlinear approach (see the 

discussion at the beginning of Section 5.4.3), a linear low-dose extrapolation was carried out 

from the BMDL10. It is expected that this approach provides an upper bound on low-dose cancer 

risk for humans. The true low-dose risks cannot be known without additional data. 

 With respect to uncertainties in the dose-response modeling, the two-step approach of 

modeling only in the observable range (U.S. EPA, 2005a) and extrapolating from a POD in the 

observable range is designed in part to minimize model dependence. Furthermore, the 

multistage model used provided an adequate fit to all the datasets. The ratios of the BMD10

values to the BMDL10 values give some indication of the uncertainties in the dose-response 

modeling. These ratios did not exceed a value of 2.6, indicating that the estimated risk is not 

influenced by any unusual variability relative to other assessments. 

 

Cross-species scaling 

 An adjustment for cross-species scaling (BW3/4) was applied to address toxicological 

equivalence of internal doses between rats and humans, consistent with the Guidelines for 

Carcinogen Risk Assessment (US EPA, 2005a). It is assumed that equal risks result from 

equivalent constant lifetime exposures. 

Human population variability

 The extent of inter-individual variability or sensitivity to the potential carcinogenicity of 

HCE is unknown. There are no data exploring whether there is differential sensitivity to HCE 

carcinogenicity across life stages. In addition, neither the extent of interindividual variability in

HCE metabolism nor human variability in response to HCE has been characterized. Factors that 

could contribute to a range of human response to HCE include variations in CYP450 levels 
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because of age-related differences or other factors (e.g., exposure to other chemicals that induce 

or inhibit microsomal enzymes), nutritional status, alcohol consumption, or the presence of 

underlying disease that could alter metabolism of HCE or antioxidant protection systems. This 

lack of understanding about potential susceptibility differences across exposed human 

populations thus represents a source of uncertainty. Humans are expected to be more 

heterogenous than laboratory animals, and this variability is likely to be influenced by ongoing 

or background exposures, diseases, and biological processes. 

5.4.6. Previous Cancer Assessment 

 The previous HCE cancer assessment by U.S. EPA reported in IRIS was last revised in 

1994. The quantitative cancer assessment was based on the incidence of hepatocellular 

carcinomas in male mice in the NCI (1978) study. The current risk value is derived from the 

incidence of renal adenomas or carcinomas in male rats (NTP, 1989), resulting in an 

approximately 2.8-fold higher than the oral slope factor derived in the previous assessment using 

BMD modeling based the hepatocellular carcinomas in male mice in the NCI (1978) study. 

In addition, the scaled human doses were calculated using a slightly different formula 

than is current practice: 

 Scaled human dose = animal dose × (animal weight/human body weight)1/3 × (546/637) 

The difference in the animal-to-human dose scaling procedure is due to the fact that current 

practice bases dose equivalence on the 3⁄4 power of body weight instead of the previous 2⁄3 power 

of body weight. 
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 6. MAJOR CONCLUSIONS IN THE CHARACTERIZATION OF HAZARD 

AND DOSE RESPONSE

6.1. HUMAN HAZARD POTENTIAL

 HCE is a halogenated hydrocarbon consisting of six chlorines attached to an ethane 

backbone. HCE was produced in the U.S. from 1921 to 1967, but is currently not commercially 

distributed. HCE is primarily used in the military for smoke pots, smoke grenades, and 

pyrotechnic devices. In the past, HCE was used as antihelminthic for the treatment of sheep 

flukes but is no longer used for this purpose since the Food and Drug Administration withdrew 

approval for this use in 1971. HCE has also been used as a polymer additive, moth repellant, 

plasticizer for cellulose esters, insecticide solvent, and in metallurgy for refining aluminum 

alloys. 

There is limited information on the toxicity of HCE in humans. Current understanding of 

HCE toxicology is based on the limited database of animal studies. After absorption by oral 

exposure, HCE is primarily distributed to fat tissue. Toxicokinetic studies in animals indicated 

that HCE is also localized and metabolized in the liver and kidney. Kidney concentrations of 

HCE were higher in male rats than female rats (Gorzinski et al., 1985; Nolan and Karbowski, 

1978). Studies of HCE metabolism indicated that the major CYP450 enzymes involved are 

phenobarbital-inducible, which include the 2A, 2B, and 3A subfamilies (Salmon et al., 1985, 

1981; Town and Leibman, 1984; Nastainczyk et al., 1982, 1981). HCE is putatively metabolized 

via a pentachloroethyl free radical to PERC and pentachloroethane. Pentachloroethane is then 

metabolized to TCE. TCE and PERC are further metabolized by hepatic oxidation to several 

urinary metabolites including TCA, trichloroethanol, oxalic acid, dichloroethanol, dichloroacetic 

acid, and monochloroacetic acid (Mitoma et al., 1985; Nastainczyk et al., 1982, 1981; Bonse and 

Henschler, 1976; Fowler, 1969; Jondorf et al., 1957). Metabolism is minimal based on the few 

studies that provided quantitative data on metabolites. However, several of these metabolites 

have demonstrated liver and kidney toxicities similar to HCE. 

The kidney has consistently been shown as the target for toxicity in acute, subchronic, 

and chronic toxicity bioassays in animals (NTP, 1996, 1989; Gorzinski et al., 1985; NCI, 1978). 

Noncancer effects include kidney degeneration (tubular nephropathy, necrosis of renal tubular 

epithelium, hyaline droplet formation, tubular regeneration and tubular casts) and hepatocellular 

necrosis. Hepatotoxicity was noted in animals exposed to HCE, although endpoints of this 

nature have not been evaluated in laboratory animals as fully as the renal effects. Hepatocellular 

necrosis was reported in female rats (NTP, 1989), but was not evaluated in a chronic exposure 

study of mice (NCI, 1978). The mouse study (NCI, 1978) focused on tumorigenic endpoints and 

not noncancer effects. 

There is no information available describing the metabolism of HCE following exposure 

via inhalation. The inhalation database for HCE contains one acute (Weeks and Thomasino, 
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1978) and one short-term (Weeks et al., 1979) study. Neurological effects, such as tremors and 

ataxia, were observed in male beagle dogs, male and female rats, and pregnant rats. Other 

effects included reduced body weight gain and increased relative liver weight in rats and guinea 

pigs exposed to HCE via inhalation. Male rats also displayed increased relative spleen and testes 

weights. 

Cancer effects observed in animal studies include hepatocellular carcinomas in mice and 

renal adenomas or carcinomas and pheochromocytomas in rats. Under EPA’s Guidelines for 

Carcinogen Risk Assessment (U.S. EPA, 2005a), HCE is “likely to be carcinogenic to humans” 

because HCE induced kidney and adrenal gland tumors in male rats and liver tumors in male and 

female mice. Studies evaluating the carcinogenicity in humans exposed to HCE are unavailable. 

The carcinogenicity incidence data in male rats (NTP, 1989) were used to develop quantitative 

cancer risk assessments for HCE. The consistency of the kidney and liver as target organs in 

different species for HCE distribution and metabolism, and both noncancer and cancer endpoints, 

provides support for the evaluation of these endpoints as relevant to humans. 

6.2. DOSE RESPONSE

6.2.1. Oral Noncancer

 Subchronic and chronic bioassays in rats and mice have identified the following 

endpoints after exposure to HCE: tubular nephropathy, atrophy and degeneration of renal 

tubules, and hepatocellular necrosis. In female rats, tubular nephropathy, atrophy and 

degeneration of the renal tubules, and hepatocellular necrosis were observed in a statistically 

significant dose-response manner (NTP, 1989; Gorzinski et al., 1985; NCI, 1978). Tubular 

nephropathy, severity of nephropathy, and atrophy and degeneration of the renal tubules in male 

rats demonstrated a statistically significant dose response. Although mice were evaluated in a 

chronic exposure study (NCI, 1978), noncancer effects were not reported because this study was 

focused on tumorigenic endpoints. 

 The most sensitive endpoint identified for effects of HCE by oral exposure relate to 

kidney toxicity in the 16-week feeding study by Gorzinski et al. (1985) in male rats. Gorzinski 

et al. (1985) was selected as the principal study and atrophy and degeneration of renal tubules in 

male rats were chosen as the critical effect for the derivation of the oral RfD. This study 

included both sexes of F344 rats, 10 animals/sex/dose, and three dose groups plus controls (0, 1, 

15, and 62 mg/kg-day). Dose-response analyses of the noncancer endpoint, atrophy and 

degeneration of renal tubules (Gorzinski et al., 1985), using EPA’s BMDS, resulted in a POD of 

0.728 mg/kg-day. A composite UF of 3,000 was applied to the POD to derive an oral RfD of 2 ×

10-4 mg/kg-day. 

Confidence in the principal study, Gorzinski et al. (1985) is high. The 16-week study is a 

well-conducted study that used three dose groups plus a control. NTP (1989) also conducted 

16-day, 13-week, and 103-week studies that supported the results observed in the 16-week study. 
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Application of BMD modeling provided a POD upon which to base the derivation of the RfD. 

The critical effect on which the RfD is based is well-supported by other oral short-term, 

subchronic, and chronic studies. Confidence in the database is low to medium because the 

database includes acute, short-term, subchronic, and chronic toxicity studies and developmental 

toxicity studies in rats and chronic carcinogenicity bioassays in rats and mice. The database 

lacks a multigenerational reproductive study and studies in other species. Overall confidence in 

the RfD is low to medium. 

6.2.2. Inhalation Noncancer

 The inhalation toxicity database is limited to a single 6-week repeat-exposure study by 

Weeks et al. (1979). This study reported a NOAEL of 465 mg/m3 and a LOAEL of 2,517 mg/m3

in several species including Sprague-Dawley rats, male beagle dogs, and male Hartley guinea 

pigs. The effects described in this report include neurotoxicity, reduced body weight gain, and 

increased relative liver, spleen and testes weights. Based on neurological effects in SpragueDawley rats, the NOAEL of 465 mg/m3

 was selected to serve as the POD. Adjustments for

continuous exposure and for the HEC, resulted in the POD[HEC] of 83 mg/m3

. An UF of 3,000 

was applied to derive an inhalation RfC of 3 × 10-2 mg/m3. Confidence in this toxicity value is 

low because of the short duration of the underlying study and the lack of support from other 

inhalation studies. 

Confidence in the principal study, Weeks et al. (1979), is low. The 6-week study was 

conducted in several species (including male dogs, male and female rats, male guinea pigs, and 

quail). The study used three exposure groups (145, 465, and 2,517 mg/m3) plus a control. The 

study is limited by the relatively short exposure duration (6 weeks) and minimal reporting of 

effects, especially quantitative changes. Application of BMD modeling was precluded based on 

a 100% response in animals for the neurological effects and the lack of quantitative information. 

Therefore, a NOAEL served as the POD. The critical effect on which the RfD is based is 

supported by the oral short-term study conducted by the same investigators and two subchronic 

studies. Confidence in the database is low because the database includes one acute and one 

short-term toxicity study in multiple species and one developmental toxicity study in rats. The 

database lacks studies by another laboratory and a multigenerational reproductive study. Overall 

confidence in the RfC is low.

6.2.3. Cancer

Under EPA’s Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a), HCE is

“likely to be carcinogenic to humans” by all routes of exposure. This descriptor is based on 

evidence of carcinogenicity from animal studies. HCE induced statistically significant increases 

in the incidence of kidney and adrenal gland tumors in male rats and liver tumors in male and 

female mice. The NTP (1989) rat study was selected for dose-response assessment based on 
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statistically significant increased incidences of renal adenomas and carcinomas and adrenal 

pheochromocytomas and malignant pheochromocytomas in male rats. This study was used for 

development of an oral slope factor. This was a well-designed study, conducted in both sexes of 

F344 rats with 50 rats/sex/dose; typical of carcinogenicity bioassays. Test animals were 

allocated among two dose levels (7 and 14 mg/kg-day) and an untreated control group. Animals 

were observed twice daily and examined weekly (for 14 weeks) then monthly for body weight 

and monthly for feed consumption. Animals were necropsied and all organs and tissues were 

examined grossly and microscopically for histopathological lesions for a comprehensive set of 

toxicological endpoints in both sexes. Renal adenomas and carcinomas and pheochromocytomas 

and malignant pheochromocytomas observed in male rats (NTP, 1989) were not seen in female 

rats or other species orally-exposed to HCE. Hepatocellular carcinomas were observed in male 

and female mice, but not in the rats. The male B6C3F1 mice tumor incidence data (NCI, 1978) 

demonstrated evidence of carcinogenicity and a low-dose quantitative risk estimate was derived. 

The cancer risk associated with mice exposed to HCE was less sensitive than that of rats. Thus, 

the oral slope factor derived for HCE is based on the increased incidence of kidney tumors in 

male rats. 

A linear approach was applied in the dose-response assessment for HCE, in which the 

mode of action is unknown, consistent with U.S. EPA’s (2005a) Guidelines for Carcinogen Risk 

Assessment. The guidelines recommend the use of a linear extrapolation as a default approach 

when the available data are insufficient to establish a mode of action for a tumor site. As 

discussed in Section 4.7, the mechanism leading to the formation of the kidney and adrenal 

tumors in rats and the liver tumors in mice following oral exposure to HCE is unknown. The 

database for HCE lacks information on the mode of action and the shape of the curve in the 

region below the POD; therefore, a linear extrapolation was performed in determining the oral 

slope factor in the derivation of a quantitative estimate of cancer risk for ingested HCE. 

Increased incidence of renal adenomas and carcinomas in a 2-year rat bioassay (NTP, 

1989) served as the basis for the oral cancer dose-response analysis. A multistage model using 

linear extrapolation from the POD was performed to derive an oral slope factor of 

4 × 10-2(mg/kg-day)-1 for HCE. Extrapolation of the experimental data to estimate potential 

cancer risk in human populations introduces uncertainty in the risk estimation for HCE. 

Uncertainty can be considered quantitatively; however, some uncertainty can only be addressed 

qualitatively. For this reason, an overall integrated quantitative uncertainty analysis cannot be 

developed. However, EPA’s development of the cancer quantitative assessment for HCE 

included consideration of potential areas of uncertainty. The following summarizes these 

considerations. 

 A biologically-based model was not supported by the available data; therefore, a 

multistage model was the preferred model. The multistage model can accommodate a wide 

variety of dose-response shapes and provides consistency with previous quantitative dose-
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response assessments for cancer. Linear low-dose extrapolation from a POD determined by an 

empirical fit of tumor data has been judged to lead to plausible upper bound risk estimates at low 

doses for several reasons. However, it is unknown how well this model or the linear low-dose 

extrapolation predicts low dose risks for HCE. An adjustment for cross-species scaling (BW3/4) 

was applied to address toxicological equivalence of internal doses between rats and humans 

based on the assumption that equal risks result from equivalent constant lifetime exposures. 

 An inhalation unit risk was not derived in this assessment. Data on the carcinogenicity of 

the compound via the inhalation route are unavailable, and route-to-route extrapolation was not 

possible due to the lack of a physiologically based pharmacokinetic model. However, it is 

proposed that HCE is likely to be carcinogenic to humans by the inhalation route since the 

compound is absorbed and, in oral studies, induces tumors at sites other than the portal of entry. 
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7. REFERENCES

ACGIH (American Conference of Governmental Industrial Hygienists). (1991). Documentation of the threshold 

limit values and biological exposure indices. 6th edition. Cincinnati, OH: American Conference of Governmental 

Industrial Hygienists. 

ACGIH (2001) Hexachloroethane. In TLV chemical substances 7th edition Documentation. Cincinnati, OH: 

American Conference of Governmental Industrial Hygienists. 

Allen, MB; Crisp, A; Snook, N; et al. (1992) Smoke-bomb pneumonitis. Respir Med 86:165-166. 

Ashby, J; Tennant, RW. (1988) Chemical structure, salmonella mutagenicity and extent of carcinogenicity as 

indicators of genotoxic carcinogenesis among 222 chemicals tested in rodents by the U.S. NCI/NTP. Mutat Res 

204:17-115. 

ATSDR (Agency for Toxic Substances and Disease Registry). (1997c) Toxicological profile for hexachloroethane. 

Atlanta, GA: U.S. Department of Health and Humans Services. Available from 

http://www.atsdr.cdc.gov/toxprofiles/tp97.pdf. 

ATSDR (1997b) Toxicological profile for tetrachloroethylene. Atlanta, GA: U.S. Department of Health and Humans 

Services. Available online at http://www.atsdr.cdc.gov/. 

ATSDR (1997a) Toxicological profile for trichloroethylene. Atlanta, GA: U.S. Department of Health and Humans 

Services. Available online at http://www.atsdr.cdc.gov/. 

ATSDR (2008) Toxicological profile for 1,1,2,2-tetrachloroethane. Atlanta, GA: U.S. Department of Health and 

Humans Services. Available online at http://www.atsdr.cdc.gov/. 

Axelson, O. (1985) Halogenated alkanes and alkenes and cancer: epidemiological aspects. In Fishbein, L; O’Neill, 

IK, eds. Environmental carcinogens: selected methods of analysis Vol. 7. Lyon, France: International Agency for 

Research on Cancer; pp. 5-20. 

 

Barrett, JC; Huff, J. (1991) Cellular and molecular mechanisms of chemically induced renal carcinogenesis. Ren 

Fail 13:211-226. 

Beurskens, JE; Stams, AJ; Zehnder, AJ; et al. (1991) Relative biochemical reactivity of three 

hexachlorocyclohexane isomers. Ecotoxicol Environ Saf 21:128-36. 

Blanco, JG; Harrison, PL; Evans, W; et al. (2000) Human cytochrome P450 maximal activities in pediatric versus 

adult liver. Drug Metab Disp 28(4):379-382. 

Bonse, G; Henschler, D. (1976) Chemical reactivity, biotransformation, and toxicity of polychlorinated aliphatic 

compounds. Crit Rev Toxicol 4:395-409. 

Borghoff, SJ; Short, BG; Swenberg, JA. (1990) Biochemical mechanisms and pathobiology of α2u-globulin 

nephropathy. Annu Rev Pharmacol and Toxicol 30:349-367. 

Borghoff, SJ. (1993) α2u -Globulin-mediated male rat nephropathy and kidney cancer: relevance to human risk 

assessment. CIIT Act13(4):l-8. 

Bronzetti, G; Morichetti, E; Del Carratore, R; et al. (1989) Tetrachloroethane, pentachloroethane, and

hexachloroethane: genetic and biochemical studies. Teratog Carcinog Mutagen 9:349-357. 

Bronzetti, G; Morichetti, E; Vellosi, R; et al. (1990) Genotoxicity and effects on microsomal enzymes of three 

chlorinated ethanes. Mutat Res 234:429-430. 

Budavari, S; O’Neil, MJ; Smith, A; et al., eds. (1989) The Merck index: an encyclopedia of chemicals, drugs, and 

biologicals. 1lth edition. Rahway, NJ: Merck & Co., Inc.; p. 740. 
122 DO NOT CITE OR QUOTE

Bull, RJ; Sanchez, IM; Nelson, MA; et al. (1990) Liver tumor induction in B6C3F1 mice by dichloroacetate and 

trichloroacetate. Toxicology 63:341B359. 

Cal EPA (California Environmental Protection Agency). (2001) Public health goal for tetrachloroethylene in 

drinking water. Office of Environmental Health Hazard Assessment. Available from online at 

http://oehha.ca.gov/water/shg/83101PHG.htm. 

Cal OEHHA. (1997) Chronic toxicity summary: hexachloroethane, draft for public review. Currently unavailable. 

Cazeneuve, C; Pons, G; Rey, E; et al. (1994) Biotransformation of caffeine in human liver microsomes from 

foetuses, neonates, infants and adults. Br J Clin Pharmacol 37:405-412. 

ChemIDplus Advanced. (2005) Hexachloroethane. ChemFinder.com database & internet searching. Available

online at http://chem.sis.nlm.nih.gov/chemidplus/. 

Crebelli, R; Andreoli, C; Carere, A; et al. (1992) The induction of mitotic chromosome malsegregation in 

Aspergillus nidulans. Quantitative structure activity relationship (QSAR) analysis with chlorinated aliphatic 

hydrocarbons. Mutat Res 266:117-134. 

Crebelli, R; Andreoli, C; Carere, A; et al. (1995) Toxicology of halogenated aliphatic hydrocarbons: structural and 

molecular determinants for the disturbance of chromosome segregation and the induction of lipid peroxidation. 

Chem Biol Interact 98(2):113-129. 

Crebelli, R; Carere, A; Leopardi, P; et al. (1999) Evaluation of 10 aliphatic halogenated hydrocarbons in the mouse 

bone marrow micronucleus test. Mutagenesis 14(2):207-215. 

Doherty, AT; Ellard, S; Parry, EM; et al. (1996) An investigation into the activation and deactivation of chlorinated 

hydrocarbons to genotoxins in metabolically competent human cells. Mutagenesis 11(3):247-274. 

Dorne, JLCM. (2004) Impact of inter-individual difference in drug metabolism and pharmacokinetics on safety 

evaluation. Fundam Clin Pharmacol 18:609-620. 

Evans, WE; Relling, MF; de Graaf, S; et al. (1989) Hepatic drug clearance in children: studies with indocyanine 

green as a model substrate. J Pharmacol Exp Ther 78:452-456. 

Fiserova-Bergerova, V; Pierce, JT; Droz, PO. (1990) Dermal absorption potential of industrial chemicals: criteria for 

skin notation. Am J Ind Med 17:617-635. 

Fishbein, L. (1979) Potential halogenated industrial carcinogenic and mutagenic chemicals. II. Halogenated 

saturated hydrocarbons. Sci Total Environ 11:163-95. 

Fowler, JS. (1969) Some hepatotoxic action of hexachloroethane and its metabolites in sheep. Br J Pharmacol 

35:530-542. 

Fox, JG; Cohen, BJ; Loew, FM. (1984) Laboratory animal medicine. Academic Press, New York. 

Galloway, SM; Armstrong, MJ; Reuben, C; et al. (1987) Chromosome aberrations and sister chromatid exchanges in 

Chinese hamster ovary cells: evaluations of 108 chemicals. Environ Mol Mutagen l0:1-175. 

Gargas, ML; Andersen, ME. (1989) Determining kinetic constants of chlorinated ethane metabolism in the rat from 

rates of exhalation. Toxicol Appl Pharmacol 99:344-353. 

Gargas, ML; Seybold, PG; Andersen, ME. (1988) Modeling the tissue solubilities and metabolic rate constant Vmax

of halogenated methanes, ethanes and ethylenes: symposium on quantitative toxicology held at the 17th conference 

on toxicology; November 3-5; Dayton, Ohio, USA, Toxicol Lett 43:235-256. 

Gephart, LA; Salminen, WF; Nicolich, MJ; et al. (2001) Evaluation of subchronic toxicity data using the benchmark 

dose approach. Regul Toxicol Pharmacol 33(1):37-59. 
123 DO NOT CITE OR QUOTE

Goodman, DG., Ward, JM, Squire, RA, et al. (1980) Neoplastic and non-neoplastic lesions 

in aging Osborne-Mendel rats. Toxicol Appl Pharmacol 55:433-447. 

Gorzinski, SJ; Nolan, RJ; McCollister, SB; et al. (1985) Subchronic oral toxicity, tissue distribution and clearance of 

hexachloroethane in the rat. Drug Chem Toxicol 8:155-169. 

Gorzinski, SJ; Wade, CE; McCollister, SB; et al. (1980) Hexachloroethane: results of a 16 week toxicity study in the 

diet of CDF Fischer 344 rats. Midland, MI: Dow Chemical Company. 

Hard, GS; Rodgers, IS; Baetcke, KP; et al. (1993) Hazard evaluation of chemicals that cause accumulation of 

α2u-globulin, hyaline droplet nephropathy, and tubule neoplasia in the kidneys of male rats. Environ Health Perspect 

99:313-349. 

Harrington-Brock, K; Doerr, CL; Moore, MM. (1998) Mutagenicity of three disinfection by-products: di- and 

trichloroacetic acid and chloral hydrate in L5178Y/TK +/- (-)3.7.2C mouse lymphoma cells. Mutat Res 

413:265B276. 

Haworth, S; Lawlor, T; Mortelmans, K; et al. (1983) Salmonella mutagenicity test results for 250 chemicals. 

Environ Mutagen 5(Suppl 1):3-142. 

Hodge, HC; Sterner, JH. (1949) Tabulation of toxicity classes. Am Indust Hyg Assoc Quar10:93-96. 

Holmes, DD. (1984) Clinical Laboratory Animal Medicine. Iowa State University Press, Ames. 

Howard, PH (ed). (1989). Handbook of environmental fate and exposure data for organic chemicals. Vol. I. Large 

production and priority pollutants. Chelsea, MI: Lewis Publishers. 

IARC (International Agency for Research on Cancer). (1979) IARC monographs on the evaluation of the 

carcinogenic risk of chemicals to humans. Vol. 20. Some halogenated hydrocarbons. Lyon, France: International 

Agency for Research on Cancer; p. 467. 

Jackson, MA; Stack, HF; Waters, MD. (1993) The genetic toxicology of putative nongenotoxic carcinogens. Mutat 

Res 296:241-277. 

Jondorf, WR; Parke; DV; Williams, RT. (1957) The metabolism of [14C]hexachloroethane. Biochem J 65:14P-15P 

JISA (Japan Industrial Safety Association). (1993) Carcinogenicity study of tetrachloroethylene by inhalation in rats 

and mice. Data No. 3-1. Kanagawa, Japan. Available from: IRIS Information Desk, U.S. Environmental Protection 

Agency, Washington, DC. 

Kinkead, ER; Wolfe, RE. (1992) Single oral toxicity of various organic compounds. J Am Coll Toxicol 11(6):713. 

Kulig, B; Alleva, E; Bignami, G; et al. (1996) Animal behavioral methods in neurotoxicity assessment: SGOMSEC 

joint report. Environ Health Perspect 104.2:193-204. 

Lacroix, D; Sonnier, M; Moncion, A; et al. (1997) Expression of CYP3A in the human liver–evidence that the shift 

between CYP3A7 and CYP3A4 occurs immediately after birth. Eur J Biochem 247:625-634. 

Lattanzi, G; Colacci, A; Grilli, S; et al. (1988) Binding of hexachloroethane to biological macromolecules from rat 

and mouse organs. J Toxicol Environ Health 24:403-411. 

Legator, MS; Harper, BL. (1988) Mutagenicity screening/in vitro testing—the end of an era; animal and human 

studies—the direction for the future. In Lliving in a chemical world occupational and environmental significance of 

industrial carcinogens. C Maltoni and IJ Selikoff, ed.s Ann NY Acad Sci 534:833-844. 

Lock, EA; Hard, GC. (2004) Chemically induced renal tubule tumors in the laboratory rat and mouse: review of the 

NCI/NTP database and categorization of renal carcinogens based on mechanistic information. Crit Rev Toxicol 

34(3):211-299. 
124 DO NOT CITE OR QUOTE

Loh, C.H.; Chang, Y.W.; Liou, S.H.; et al. (2006) Case report: hexachloroethane smoke inhalation: a rare cause of 

severe hepatic injuries. Environ Health Perspect 114(5): 763-765. 

Loh, C.H.; Liou, S.H., Chang, Y.W.; et al. (2008) Hepatic injuries of hexachloroethane smoke inhalation: The first 

analytical epidemiological study. Toxicology 247(2-3):119-122. 

Lohman, PHM; Lohman, WJA. (2000) Genetic activity profiles 2000 (program version 1.3.0), data record for 

hexachloroethane. Data base and software are a joint effort of the U.S. EPA and IARC.. 

Lutz, WK. (1979) In vivo covalent binding of organic chemicals to DNA as a quantitative indicator in the process of 

chemical carcinogenesis. Mutat Res 65:289-356. 

Lutz, WK. (1986) Quantitative evaluation of DNA binding data for risk estimation and for classification of direct 

and indirect carcinogens. J Cancer Res Clin Oncol 112:85-91. 

Mather, GG, Exon, JH, Koller, LD. (1990) Subchronic 90-day toxicity of dichloroacetic and trichloroacetic acid in 

rats. Toxicology 64:71-80. 

McGregor, DB; Brown, A; Cattanach, P; et al. (1988) Responses of the L5178Y tk+/tk- mouse lymphoma cell 

forward mutation assay: III. 72 coded chemicals. Environ Mol Mutagen 12:85-154. 

Milman, HA; Story, DL; Riccio, ES; et al. (1988) Rat liver foci and in vitro assays to detect initiating and promoting 

effects of chlorinated ethanes and ethylenes. Ann NY Acad Sci 534:521-530. 

Mitoma, C; Steeger, T; Jackson, SE; et al. (1985) Metabolic disposition study of chlorinated hydrocarbons in rats 

and mice. Drug Chem Toxicol 8:183-194. 

Miyagawa, M; Takasawa, H; Sugiyama, A; et al. (1995) The in vivo-in vitro replicative DNA synthesis (RDS) test 

with hepatocytes prepared from male B6C3F1 mice as an early prediction assay for putative nongenotoxic (Amesnegative) mouse hepatocarcinogens. Mutat Res 343:157-183. 

Moore, MM; Harrington-Brock, K. (2000) Mutagenicity of trichloroethylene and Its metabolites: implications for 

the risk assessment of trichloroethylene. Environ Health Perspect 108 Suppl 2:215B223. 

Nakamura, S; Oda, Y; Shimada, T; et al. (1987) SOS-inducing activity of chemical carcinogens and mutagens in 

Salmonella typhimurium TA1535/pSK1002: examination with 151 chemicals. Mutat Res 192:239-246. 

Nastainczyk, W; Ahr, H; Uhich, V; et al. (1981) The mechanism of the reductive dehalogenation of polyhalogenated 

compounds by microsomal cytochrome P450. Adv Exp Med Biol 136(A): 799-808. 

Nastainczyk, W; Ahr, HJ; Ullrich, V. (1982) The reductive metabolism of halogenated alkanes by liver microsomal 

cytochrome P450. Biochem Pharmacol 131:391-396. 

NCI (National Cancer Institute). (1976) Carcinogenesis bioassay of trichloroethylene (CAS No. 79-01-6). Public 

Health Service, U.S. Department of Health and Human Services; NTP TR-2. Available from: National Institute of 

Environmental Health Sciences, Research Triangle Park, NC. Available online at 

http://ntp.niehs.nih.gov/index.cfm?objectid=07028C7F-AB6E-6D29-3FC1CC9D48574701. 

NCI. (1977) Bioassay of tetrachloroethylene for possible carcinogenicity. Public Health Service, U.S. Department 

of Health, Education, and Welfare; NTP TR-13. Available from: National Cancer Institute, Bethesda, MD.

Available online at http://ntp.niehs.nih.gov/index.cfm?objectid=0702B823-CEA9-1089-DFBDC6F9207C56F2. 

NCI. (1978) Bioassay of hexachloroethane for possible carcinogenicity. Public Health Service, U.S. Department of 

Health, Education, and Welfare; NTP TR-68. Available from: National Cancer Institute, Bethesda, MD. Available 

online at http://ntp.niehs.nih.gov/ntp/htdocs/LT_rpts/tr068.pdf. 

Nelson, MA; Bull, RJ. (1988) Induction of strand breaks in DNA by trichloroethylene and metabolites in rat and 

mouse liver in vivo. Toxicol Appl Pharmacol 94:45B54.
125 DO NOT CITE OR QUOTE

NRC (National Research Council). (1983) Risk assessment in the federal government: managing the process. 

Washington, DC: National Academy Press. 

Nolan, RJ; Karbowski, RJ. (1978) Hexachloroethane: tissue clearance and distribution in Fischer 344 rats. Midland, 

MI: Dow Chemical Company. 

NTP (National Toxicology Program). (1983) Carcinogenesis studies of pentachloroethane (CAS No. 76-01-7) in 

F344/N rats and B6C3F1 mice (gavage study). Public Health Service, U.S. Department of Health and Human

Services; NTP TR-232. Available from National Institute of Environmental Health Sciences, Research Triangle 

Park, NC. Available online at http://ntp.niehs.nih.gov/ntp/htdocs/LT_rpts/tr232.pdf. 

NTP. (1986) Toxicology and carcinogenesis studies of tetrachloroethylene (perchloroethylene) (CAS No. 127-18-4) 

in F344/N rats and B6C3F1 mice (inhalation studies). Public Health Service, U.S. Department of Health and Human 

Services; NTP TR-311. Available from National Institute of Environmental Health Sciences, Research Triangle 

Park, NC. Available online at http://ntp.niehs.nih.gov/ntp/htdocs/LT_rpts/tr311.pdf. 

NTP. (1988) Toxicology and carcinogenesis studies of trichloroethylene (CAS No. 79-01-6) in four strains of rats 

(ACI, August, Marshall, Osborne-Mendel)(gavage studies). Public Health Service, U.S. Department of Health and 

Human Services; NTP TR-273. Available from National Institute of Environmental Health Sciences, Research 

Triangle Park, NC. Available online at http://ntp.niehs.nih.gov/ntp/htdocs/LT_rpts/tr273.pdf. 

NTP. (1989) Toxicology and carcinogenesis studies of hexachloroethane (CAS No. 67-72-l) in F344/N rats (gavage 

studies). Public Health Service, U.S. Department of Health and Human Services; NTP TR-361. Available from 

National Institute of Environmental Health Sciences, Research Triangle Park, NC. Available online at 

http://ntp.niehs.nih.gov/ntp/htdocs/LT_rpts/tr361.pdf. 

NTP. (1990) Carcinogenesis studies of trichloroethylene (without epichlorohydrin) (CAS No. 79-01-6) in F344/N 

rats and B6C3F1 mice (gavage study). Public Health Service, U.S. Department of Health and Human Services; NTP 

TR-243. Available from National Institute of Environmental Health Sciences, Research Triangle Park, NC. 

Available online at http://ntp.niehs.nih.gov/index.cfm?objectid=07067B36-09A5-8398-7E70FB2C35377215. 

NTP. (1996) NTP technical report on renal toxicity studies of selected halogenated ethanes administered by gavage 

to F344/N rats. Public Health Service, U.S. Department of Health and Human Services; NTP TOX-45. Available 

from National Institute of Environmental Health Sciences, Research Triangle Park, NC, and 

<http://ntp.niehs.nih.gov/ntpweb/index.cfm?objectid=D1512B41-F1F6-975E-7FBA3D4A2132F1C1>. 

NTP. (2005) 11th report on carcinogens. Public Health Service, U.S. Department of Health and Human Services, 

Research Triangle Park, NC. Available online at http://ntp-server.niehs.nih.gov. 

Odabasi, M. (2008) Halogenated volatile organic compounds from the use of chlorine-bleach-containing household 

products. Environ Sci Technol 42(5): 1445-1451. 

Omiecinski, CJ; Remmel, RP; Hosagrahara, VP. (1999) Concise review of the cytochrome P450s and their roles in 

toxicology. Toxicol Sci 48:151-156. 

Onfelt, A. (1987) Spindle disturbances in mammalian cells. III. Toxicity, c-mitosis and aneuploidy with 22 different 

compounds. Specific and unspecific mechanisms. Mutat Res 182(3):135-154. 

Ramsey, JC; Andersen, ME. (1984) A physiologically based description of the inhalation pharmacokinetics of 

styrene monomer in rats and humans. Toxicol Appl Pharmacol 73:159-175. 

Reynolds, ES. (1972) Comparison of early injury to liver endoplasmic reticulum by halomethanes, 

hexachloroethane, benzene, toluene, bromobenzene, ethionine, thioacetamide and dimethylnitrosamine. Biochem 

Pharmacol 21:2555-2561. 

Rice, JM; Baan, RA; Blettner, M; et al. (1999) Rodent tumors of urinary bladder, renal cortex, and thyroid gland in 

IARC Monographs evaluations of carcinogenic risk to humans. Toxicol Sci 49:166-171. 
126 DO NOT CITE OR QUOTE

Roldán-Arjona, T; Garcia-Pedrajas, MD; Luque-Romero, FL; et al. (1991) An association between mutagenicity in 

rodents for 16 halogenated aliphatic hydrocarbons. Mutagenesis 6:199-205. 

Salmon, AG; Jones, RB; Mackrodt, WC. (1981) Microsomal dechlorination of chloroethanes: structure reactivity 

relationships. Xenobiotica 11:723-734. 

Salmon, AG; Nash, JA; Walkin, CM; et al. (1985) Dechlorination of halocarbons by microsomes and vesicular 

reconstituted cytochrome P-450 systems under reductive conditions. Br J Ind Med 42:305-311. 

Sandstedt, K; Berglöf, A; Feinstein, R; et al. (1997) Differential susceptibility to Mycoplasma pulmonis intranasal 

infection in X-linked immunodeficient (xid), severe combined immunodeficient (scid), and immunocompetent mice. 

Clin Exp Immunol 108:490-496. 

Seldén, A; Jacobson, G; Berg, P; et al. (1989) Hepatocellular carcinoma and exposure to hexachlorobenzene: a case 

report. Br J Ind Med 46:138-140. 

Seldén, A; Nygren, M; Kvamlof, A; et al. (1993) Biological monitoring of hexachloroethane. Int Ach Occup 

Environ Health 65(Suppl 1):S111–114. 

Seldén, A; Kvarnlof, A; Bodin, L; et al. (1994) Health effects of low level occupational exposure to 

hexachloroethane. J Occup Med Toxicol 3(10):73-79.

Seldén, AI; Nygren, Y; Westberg, HB; et al. (1997) Hexachlorobenzene and octachlorostyrene in plasma of 

aluminium foundry workers using hexachloroethane for degassing. Occup Environ Med 54(8):613-618. 

Seldén, AI; Floderus, Y; Bodin, LS; et al. (1999) Porphyrin status in aluminum foundry workers exposed to 

hexachlorobenzene and octachlorostyrene. Arch Environ Health 54(4):248-53. 

Shimizu, M; Noda, T; Yamano, T; et al. (1992) Safety evaluation of chemicals for use in household products 

(XVII). A teratological study on hexachloroethane in rats. Osaka City Institute of Public Health and Environmental 

Sciences 54:70-75. (Japanese) 

Simmon, VF; Kauhanen, K. (1978) In vitro microbiological mutagenicity assays of hexachloroethane. SRI 

International, Menlo Park, CA. Prepared for U.S. Environmental Protection Agency, National Environmental 

Research Center, Water Supply Research Laboratory, Cincinnati, OH. 

Southcott, WH. (1951) The toxicity and antihelmintic efficiency of hexachloroethane in sheep. Aust Vet J 27:18-21. 

Spanggord, RJ; Chou, TW; Mill, T; et al. (1985) Environmental fate of nitroguanidine, diethyleneglycol dinitrate, 

and hexachloroethane smoke. SRI International, Menlo Park, CA. Prepared for U.S. Army Medical Research and 

Development Command. 

Story, DL; Meierhenry, EF; Tyson, CA; et al. (1986) Differences in rat liver enzyme-altered foci produced by 

chlorinated aliphatics and phenobarbital. Toxicol Ind Health 2:351-362. 

Tafazoli, M; Baeten, A; Geerlings, P; et al. (1998) In vitro mutagenicity and genotoxicity study of a number of 

short-chain chlorinated hydrocarbons using the micronucleus test and the alkaline single cell gel electrophoresis 

technique (Comet assay) in human lymphocytes: a structure-activity relationship (QSAR) analysis of the genotoxic 

and cytotoxic potential. Mutagenesis 13(2):115-126. 

Town, C; Leibman, KC. (1984) The in vitro dechlorination of some polychlorinated ethanes. Drug Metab Disp 

12:4-8. 

Treluyer, JM; Jacqz-Aigrain, E; Alvarez, F; et al. (1991) Expression of CYP2D6 in developing human liver. Eur J 

Biochem 202:583-588. 

Tu, AS; Murray, TA; Hatch, KM; et al. (1985) In vitro transformation of BALB/c-3T3 cells by chlorinated ethanes 

and ethylenes. Cancer Lett 28:85-92. 
127 DO NOT CITE OR QUOTE

U.S. EPA (Environmental Protection Agency). (1979) Water-related environmental fate of 129 priority pollutants. 

Vol. II. Monitoring and Data Support Division, Washington, DC; EPA440/4-79-029b. Available from: National 

Technical Information Service, Springfield, VA; PB80-204381. 

U.S. EPA. (1982) Aquatic fate process data for organic priority pollutants. Monitoring and Data Support Division, 

Washington, DC; EPA 440/4-81-014. 

U.S. EPA. (1986a) Guidelines for the health risk assessment of chemical mixtures. Federal Register 51(185):34014-

34025. Available online at http://www.epa.gov/iris/backgr-d.htm. 

U.S. EPA. (1986b) Guidelines for mutagenicity risk assessment. Federal Register 51(185):34006-34012. Available 

online at http://www.epa.gov/iris/backgr-d.htm. 

U.S. EPA. (1987) Health effects assessment for hexachloroethane. Cincinnati, OH, Environmental Criteria and 

Assessment Office, Office of Research and Development.; EPA/600/8-88/043; PB88-178736. 

U.S. EPA. (1988) Recommendations for and documentation of biological values for use in risk assessment. Prepared 

by the Environmental Criteria and Assessment Office, Office of Health and Environmental Assessment, Cincinnati, 

OH for the Office of Solid Waste and Emergency Response, Washington, DC; EPA 600/6-87/008. Available online 

at http://www.epa.gov/iris/backgr-d.htm. 

U.S. EPA. (1991a) Guidelines for developmental toxicity risk assessment. Federal Register 56(234):63798-63826. 

Available online at http://www.epa.gov/iris/backgr-d.htm. 

U.S. EPA. (1991b) Health advisory for hexachloroethane. Office of Water, Washington, DC; EPA/625/3-91/019F;. 

PB91-159657. 

U.S. EPA. (1991c) Alpha2u-globulin: association with chemically induced renal toxicity and neoplasia in the male 

rat. Risk Assessment Forum, Washington, DC; EPA/625/3-91/019F; PB92-143668. 

U.S. EPA. (1992) Draft report: a cross-species scaling factor for carcinogen risk assessment based on equivalence of 

mg/kg3/4/day. Federal Register 57(109):24152-24173. 

U.S. EPA. (1994a) Interim policy for particle size and limit concentration issues in inhalation toxicity studies. 

Federal Register 59(206):53799. Available online at http://www.epa.gov/iris/backgr-d.htm. 

U.S. EPA. (1994b) Methods for derivation of inhalation reference concentrations and application of inhalation 

dosimetry. Office of Research and Development, Washington, DC; EPA/600/8-90/066F. Available online at

http://www.epa.gov/iris/backgr-d.htm. 

U.S. EPA. (1995) Use of the benchmark dose approach in health risk assessment. Risk Assessment Forum, 

Washington, DC; EPA/630/R-94/007. Available online at 

http://cfpub.epa.gov/ncea/raf/recordisplay.cfm?deid=42601. 

U.S. EPA. (1996) Guidelines for reproductive toxicity risk assessment. Federal Register 61(212):56274-56322. 

Available online at http://www.epa.gov/iris/backgr-d.htm. 

U.S. EPA. (1998) Guidelines for neurotoxicity risk assessment. Federal Register 63(93):26926-26954. 

U.S. EPA. (2000a) Science policy council handbook: risk characterization. Office of Science Policy, Office of 

Research and Development, Washington, DC; EPA 100-B-00-002. Available online at 

http://www.epa.gov/iris/backgr-d.htm. 

U.S. EPA. (2000b) Benchmark dose technical guidance. External review draft. Risk Assessment Forum, 

Washington, DC; EPA/630/R-00/001. Available online at http://www.epa.gov/iris/backgr-d.htm. 

U.S. EPA. (2000c) Supplementary guidance for conducting for health risk assessment of chemical mixtures. Risk 

Assessment Forum, Washington, DC; EPA/630/R-00/002. Available online at 

http://www.epa.gov/iris/backgr-d.htm. 
128 DO NOT CITE OR QUOTE

>. 

U.S. EPA. (2002) A review of the reference dose and reference concentration processes. Risk Assessment Forum, 

Washington, DC; EPA/630/P-02/0002F. Available online at http://www.epa.gov/iris/backgr-d.htm. 

U.S. EPA. (2005a) Guidelines for carcinogen risk assessment. Risk Assessment Forum, Washington, DC; 

EPA/630/P-03/001B. Available online at http://www.epa.gov/iris/backgr-d.htm. 

U.S. EPA. (2005b) Supplemental guidance for assessing susceptibility from early-life exposure to carcinogens. Risk 

Assessment Forum, Washington, DC; EPA/630/R-03/003F, Available online at http://www.epa.gov/iris/backgrd.htm. 

U.S. EPA. (2005c) Tetrachloroethylene. Integrated Risk Information System (IRIS). National Center for 

Environmental Assessment, Washington, DC. Available online at http://www.epa.gov/iriswebp/iris/subst/0106.htm. 

U.S. EPA. (2006a) Science policy council handbook: peer review. Third edition. Office of Science Policy, Office of 

Research and Development, Washington, DC; EPA/100/B-06/002. Available online at 

http://www.epa.gov/iris/backgr-d.htm. 

U.S. EPA. (2006b) A framework for assessing health risk of environmental exposures to children. National Center 

for Environmental Assessment, Washington, DC, EPA/600/R-05/093F. Available online at 

http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=158363. 

Van Dyke, RA. (1977) Dechlorination mechanisms of chlorinated olefins. Environ Health Perspect 21:121-124. 

Van Dyke, RA; Wineman, CG. (1971) Enzymatic dechlorination of chloroethanes and propanes in vitro. Biochem 

Pharmacol 20:463-470. 

Verschueren, K. (1983) Handbook of environmental data on organic chemicals. 2nd ed. New York, NY: Van 

Nostrand Reinhold Company. 

Vieira, I; Sonnier, M; Cresteil, T. (1996) Developmental expression of CYP2E1 in the human liver: 

hypermethylation control of gene expression during the neonatal period. Eur J Biochem 238:476-483. 

Vogel, EW; Nivard, MJ. (1993) Performance of 181 chemicals in a Drosophila assay predominantly monitoring 

interchromosomal mitotic recombination. Mutagenesis 8:57-81. 

Weast, RC, ed. (1986) CRC handbook of chemistry and physics. 67th ed. Boca Raton, FL: CRC Press. 

Webb, DR; Ridder, GM; Alden, CL. (1989) Acute and subchronic nephrotoxicity of d-limonene in dogs. Food 

Chem Toxicol 28:669-675. 

Weeks, MH; Thomasino, JA. (1978) Assessment of acute toxicity of hexachloroethane in laboratory animals. U.S. 

Army Environmental Hygiene Agency, Aberdeen Proving Ground, MD; Report No. 51-0075-78. 

Weisburger, EK. (1977) Carcinogenicity studies on halogenated hydrocarbons. Environ Health Perspect 21:7-16. 

Xu, X; Zhang, D; Lyubynska, N; et al. (2006) Mast cells protect mice from mycoplasma pneumonia. Am J Respir 

Crit Care Med 173:219-225. 

Yamakage, A; Ishikawa, H. (1982) Generalized morphea-like scleroderma occurring in people exposed to organic 

solvents. Dermatologica 165:186-193. 

Yanagita, K; Sagami, I; Shimizu, T. (1997) Distal site and surface mutations of cytochrome P450 1A2 markedly 

enhance dehalogenation of chlorinated hydrocarbons. Arch Biochem Biophys 346(2):269-276. 

Yanagita, K; Sagami, I; Daff, S; et al. (1998) Marked enhancement in the reductive dehalogenation of 

hexachloroethane by a Thr319Ala mutation of cytochrome P450 1A2. Biochem Biophys Res Commun 249(3):678-

682. 
129 DO NOT CITE OR QUOTE

Yoshikawa, K. (1996) Anomalous nonidentity between Salmonella genotoxicants and rodent carcinogens and

genotoxic noncarcinogens. Environ Health Perspect 104:40-46. 

Younglai, EV; Foster, WG; Hughes, EG; et al. (2002) Levels of environmental contaminants in human follicular 

fluid, serum, and seminal plasma of couples undergoing in vitro fertilization. Arch Environ Contam Toxicol 43(1): 

121-126. 
A- DRAFT – DO NOT CITE OR QUOTE 1

APPENDIX A: SUMMARY OF EXTERNAL PEER REVIEW AND PUBLIC 

COMMENTS AND DISPOSITION

PAGE INTENTIONALLY LEFT BLANK 
B-1 DRAFT – DO NOT CITE OR QUOTE 

APPENDIX B: Benchmark Dose Modeling Output

Table B-1. Dose-response modeling results using BMDS (Version 2.0) based on noncancerous kidney and liver effects in rats following oral exposure to HCE 

Study Endpoint Sex/species Fitted modela pValue AIC BMD10

(mg/kg-day) 

BMDL10

(mg/kg-day) 

Kidney effects 

Gamma 0.93 133.68 21.22 16.99 

Male rat Multistage 1° 0.93 133.66 21.25 17.01 

Weibull 0.93 133.68 21.22 16.99 

Gamma 1.00 117.47 87.24 50.63 

Multistage 2° 0.94 116.09 80.63 41.89 

Logistic 0.42 118.61 95.19 73.25 

Probit 0.53 118.14 91.25 69.20 

NCI (1978) Tubular nephropathy 

Female rat 

Weibull 1.00 117.47 84.22 48.62 

Logistic 0.99 205.88 3.84 2.62 

Multistage 1° 0.87 205.90 3.20 1.88 

Probit 0.99 205.88 3.81 2.60 

Moderate to marked 

Tubular nephropathy Male rat 

Quantal-linear 0.87 205.90 3.20 1.88 

Gamma 0.86 191.90 15.17 10.72 

Logistic 0.46 192.42 23.06 18.33 

Multistage 1° 0.78 192.96 15.91 11.14 

Probit 0.47 192.40 22.55 18.04 

Quantal-linear 0.86 191.90 15.17 10.72 

NTP (1989)

Mild to marked Tubular 

nephropathy Female rat 

Weibull 0.86 191.90 15.17 10.72 

Logistic 0.36 148.11 4.30 3.45 

NTP (1989) Linear mineralization Male rat Multistage 1° 0.20 148.90 1.75 1.40 

Probit 0.51 147.66 3.98 3.22 

Gamma 0.42 84.64 7.33 4.87 

Logistic 0.03 90.96 11.41 8.77 

LogLogistic 0.48 84.42 7.05 4.48 

LogProbit 0.07 87.89 8.38 6.51 

Multistage 2° 0.42 84.64 7.33 4.87 

Probit 0.03 90.53 10.86 8.26 

Weibull 0.42 84.64 7.33 4.87 

NTP (1989) Hyperplasia of the pelvic 

transitional epithelium Male rat 

Quantal-linear 0.42 84.64 7.33 4.87 

Gamma 0.70 34.94 1.34 0.728 

Multistage 1° 0.93 32.94 1.34 0.728 

Logistic 0.89 32.97 3.30 1.98 

Probit 0.89 32.95 3.08 1.95 

Quantal-linear 0.93 32.94 1.34 0.728 

Male rat 

Weibull 0.69 34.92 1.72 0.729 

Gamma 0.99 42.47 13.80 4.56 

Multistage 1° 0.93 40.61 8.54 4.49 

Logistic 0.98 40.51 17.40 11.07 

Probit 0.99 40.49 16.10 10.51 

Quantal-linear 0.93 40.61 8.54 4.49 

Gorzinski 

et al. 

(1985) 

Atrophy and degeneration 

of renal tubules 

Female rat 

Weibull 0.98 42.47 13.71 4.56 

Gamma 0.99 20.88 1.22 0.710 

Logistic 0.66 23.91 4.85 2.71 

LogLogistic 0.68 23.89 1.23 0.308 

LogProbit 0.54 24.26 2.11 1.01 

Multistage 2° 0.94 22.84 1.33 0.713 

Probit 0.67 23.85 4.28 2.54 

Weibull 0.99 20.88 1.22 0.710 

Gorzinski 

et al. 

(1985) 

Slight hypertrophy and/or 

dilation of proximal 

convoluted tubules 

Male rat 

Quantal-linear 0.99 20.88 1.22 0.710 

Liver effects 

Gamma 0.93 38.62 118.04 60.18 

Multistage 1° 0.68 40.56 53.82 35.19 

Logistic 0.55 41.58 156.22 107.49 

Probit 0.61 40.95 148.49 102.71 

NTP (1989) Hepatocellular necrosis Female rat 

Weibull 0.91 38.91 114.68 56.75 

a

For all models, a BMR of 0.1 was employed in deriving the estimates of the benchmark dose (BMD10) and its 95% lower confidence limit 

(BMDL10). Modeling output is provided for models that represent the POD for each of the kidney endpoints and these models are 

highlighted in bold font. 
B-2 DRAFT – DO NOT CITE OR QUOTE 

Based on the incidence of tubular nephropathy in male rats (NCI, 1978), the logistic and 

probit models exhibited significant lack-of-fit (p < 0.1), while the gamma, multistage (1°) and 

Weibull models had p-values > 0.1. All three of these models that showed adequate fit yielded 

the same AIC values, as well as nearly equivalent BMD10 and BMDL10 estimates of 21.22 and 

16.99 mg/kg-day, respectively. Therefore, the potential POD selected for this dataset is 

16.99 mg/kg-day. 

Based on the incidence of tubular nephropathy in female rats (NCI, 1978), only the 1° 

multistage model exhibited significant lack-of-fit. Of the models that did not show significant 

lack-of-fit (i.e., gamma, multistage 2°, logistic, probit and Weibull models), the BMDL10

estimates were within a factor of three of each other suggesting no appreciable model 

dependence. As the BMDL10 values did not show large variation, the model with the lowest AIC 

value was selected. Therefore, the multistage 2° model BMDL10 of 41.89 mg/kg-day was 

selected as the potential POD for this dataset. 

In fitting the available dichotomous dose-response models to the incidence of moderate to 

marked tubular nephropathy in male rats (NTP, 1989), the gamma and Weibull models exhibited 

significant lack-of-fit (p < 0.1). The models that did not show significant lack-of-fit (i.e., 

logistic, multistage 1°, quantal-linear, and probit) yielded BMDL10 estimates that were within a 

factor of three of each other, suggesting no appreciable model dependence. As the BMDL10

values did not show large variation, the model with the lowest AIC value was selected. The AIC 

values were identical; therefore, the model with the lowest BMDL10 was selected. The 

multistage 1° and quantal-linear model had identical BMDL10 values, therefore the BMDL10 of 

1.88 mg/kg-day was selected as the potential POD for this dataset. 

Based on the incidence of mild to marked tubular nephropathy in female rats (NTP, 

1989), none of the models exhibited significant lack-of-fit. These models (i.e., gamma, logistic, 

multistage 1°, probit, quantal-linear, and Weibull models) yielded BMDL10 estimates that were 

within a factor of three of each other, suggesting no appreciable model dependence. As the 

BMDL10 values did not show large variation, the model with the lowest AIC value was selected. 

The gamma, quantal-linear, and Weibull models had identical AIC values; therefore, the model 

with the lowest BMDL10 was selected. The BMDL10 values for these models were identical, 

therefore the BMDL10 of 10.72 mg/kg-day was selected as the potential POD for this dataset. 

In fitting the available dichotomous dose-response models to the incidence of linear 

mineralization in male rats (NTP, 1989), the gamma and the Weibull models exhibited 

significant lack-of-fit (p < 0.1). Of the models that did not show significant lack-of-fit (i.e., 

logistic, multistage 1°, and probit), the resulting BMDL10 estimates were within a factor of three 

of each other, suggesting no appreciable model dependence. As the BMDL10 values did not 

show large variation, the model with the lowest AIC value was selected. Therefore the probit 

model BMDL10 of 3.2 mg/kg-day was selected as the potential POD for this dataset. 
B-3 DRAFT – DO NOT CITE OR QUOTE 

In fitting the available dichotomous dose-response models to the incidence of hyperplasia 

of the pelvic transitional epithelium in male rats (NTP, 1989), the logistic, logprobit and probit 

models exhibited significant lack-of-fit (p < 0.1). Of the models that did not show significant 

lack-of-fit (i.e., gamma, loglogistic, multistage 2°, Weibull, and quantal-linear), the resulting 

BMDL10 estimates were within a factor of three of each other, suggesting no appreciable model 

dependence. As the BMDL10 values did not show large variation, the model with the lowest AIC 

value was selected. Therefore the loglogistic model BMDL10 of 4.48 mg/kg-day was selected as 

the potential POD for this dataset. 

In fitting the available dichotomous dose-response models to the incidence of atrophy and 

degeneration of renal tubules in male and female rats (Gorzinski et al., 1985), none of the models 

exhibited a significant lack-of-fit in either sex. For male rats, these models (i.e., gamma, 

multistage 1°, logistic, probit, quantal-linear, and Weibull) yielded BMDL10 estimates that were 

within a factor of three of each other, suggesting no appreciable model dependence. As the 

BMDL10 values did not show large variation, the model with the lowest AIC value was selected. 

The AIC values for the gamma, multistage 1°, and quantal-linear were identical; therefore, the 

model with the lowest BMDL10 was selected. All of the BMDL10s were identical for these 

models; therefore, the BMDL10 of 0.728 mg/kg-day was selected as the potential POD for this 

dataset. 

For female rats, these models (i.e., gamma, multistage 1°, logistic, probit, quantal-linear, 

and Weibull) yielded BMDL10 estimates that were within a factor of three of each other, 

suggesting no appreciable model dependence. As the BMDL10 values did not show large 

variation, the model with the lowest AIC value was selected. The AIC values for the 

multistage 1° and quantal-linear models were identical; therefore, the model with the lowest 

BMDL10 was selected. The multistage 1° BMDL10 of 4.49 mg/kg-day was selected as the 

potential POD for this dataset. 

In fitting the available dichotomous dose-response models to the incidence of slight 

hypertrophy and/or dilation of proximal convoluted tubules in male rats (Gorzinski et al., 1985), 

none of the models exhibited a significant lack-of-fit. For male rats, these models (i.e., gamma, 

logistic, loglogistic, logprobit, multistage 2°, probit, Weibull, and quantal-linear) yielded 

BMDL10 estimates that were within a factor of three of each other, suggesting no appreciable 

model dependence. As the BMDL10 values did not show large variation, the model with the 

lowest AIC value was selected. The AIC values for the gamma, Weibull, and quantal-linear 

models were identical; therefore, the model with the lowest BMDL10 was selected. All of the 

BMDL10s were identical for these models, therefore the BMDL10 of 0.710 mg/kg-day was 

selected as the potential POD for this dataset. 

Based on the incidence of hepatocellular necrosis in female rats (NTP, 1989), none of the 

dichotomous dose-response models exhibited a significant lack-of-fit. All of these models (i.e., 

gamma, multistage 1°, logistic, probit, and Weibull) yielded BMDL10 estimates that were within 
B-4 DRAFT – DO NOT CITE OR QUOTE 

a factor of three of each other, suggesting no appreciable model dependence. As the BMDL10

values did not show large variation, the model with the lowest AIC value was selected. 

Therefore, the gamma model BMDL10 of 60.18 mg/kg-day was selected as the potential POD for 

this dataset. 

For comparison purposes, BMD modeling for the above endpoints was also conducted 

using BMRs of 5 and 1%. The modeling results are included in Table B-2. 
B-5 DRAFT – DO NO

Table B-2. Dose-response modeling results using BMDS (Version 2.0) for BMRs of 1, 5, and 10% base

kidney and liver effects in rats following oral exposure to HCE 

Study Endpoint Sex/ 

species 

Fitted 

model

a

BMD10

(mg/kgday) 

BMDL10

(mg/kgday) 

BMD05

(mg/kg-day) 

BMDL05

(mg/kgday) 

BMD01

(mg/kgday) 

BMD

(mg/

da

Kidney effects 

Gamma 21.22 16.99 10.33 8.27 2.02 1.6Multistage 1° Male rat 21.25 17.01 10.35 8.28 2.03 1.6

Weibull 21.22 16.99 10.33 8.27 2.02 1.6

NCI (1978) 

Tubular nephropathy 

Female rat 

Multistage 2° 80.63 41.89 56.26 21.18 24.90 4.2Multistage 1° 3.20 1.88 1.56 0.91 0.30 0.1

Moderate to marked tubular nephropathy 

Male rat Quantallinear 3.20 1.88 1.56 0.91 0.30 0.1

Gamma 15.17 10.72 7.39 5.22 1.45 1.0Quantallinear 15.17 10.72 7.39 5.22 1.45 1.0

NTP (1989) Mild to 

marked tubular nephropathy 

Female rat 

Weibull 15.17 10.72 7.39 5.22 1.45 1.0

NTP (1989) 

Linear mineralization Male rat Probit 3.98 3.22 2.36 1.80 0.58 0.4

NTP (1989) 

Hyperplasia of the pelvic transitional epithelium 

Male rat LogLogistic 7.05 4.48 3.34 2.12 0.64 0.4

Gamma 1.34 0.73 0.66 0.35 0.13 0.0Multistage 1° Male rat 1.34 0.73 0.65 0.35 0.13 0.0

Quantallinear 1.34 0.73 0.65 0.35 0.13 0.0

Multistage 1° 8.54 4.49 4.16 2.19 0.82 0.4

Gorzinski et al. (1985) 

Atrophy and degeneration of renal tubules 

Female rat Quantallinear 8.54 4.49 4.16 2.19 0.82 0.4
B-6 DRAFT – DO NO

Study Endpoint Sex/ 

species 

Fitted 

model

a

BMD10

(mg/kgday) 

BMDL10

(mg/kgday) 

BMD05

(mg/kg-day) 

BMDL05

(mg/kgday) 

BMD01

(mg/kgday) 

BMD

(mg/

da

Gamma 1.22 0.71 0.60 0.35 0.12 0.0Weibull 1.22 0.71 0.60 0.35 0.12 0.0

Gorzinski et al. (1985) 

Slight hypertrophy and/or dilation of proximal convoluted tubules 

Male rat 

Quantallinear 1.22 0.71 0.60 0.35 0.12 0.0

Liver effects 

NTP (1989) 

Hepatocellular necrosis 

Female rat Gamma 118.03 60.18 84.66 33.34 41.75 8.6
B-7 DRAFT – DO NOT CITE OR QUOTE 

Modeling for Noncancer Assessment 

NCI (1978) Tubular Nephropathy in Male Rats

Gamma Model

 ==================================================================== 

 Gamma Model. (Version: 2.13; Date: 05/16/2008) 

 Input Data File: C:\USEPA\BMDS2\Temp\tmpCDF.(d) 

 Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpCDF.plt 

 Thu Apr 09 14:55:06 2009 

 ==================================================================== 

 BMDS Model Run NCI 1978 Tubular Nephropathy Male Rat - Gamma Model 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the probability function is: 

 P[response]= background+(1-background)*CumGamma[slope*dose,power], 

 where CumGamma(.) is the cummulative Gamma distribution function 

 Dependent variable = PercentPositiveModerateMarkedTubularNephropathy 

 Independent variable = ularNephropathy 

 Power parameter is restricted as power >=1 

 Total number of observations = 3 

 Total number of records with missing values = 0 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 Default Initial (and Specified) Parameter Values 

 Background = 0.0238095 

 Slope = 0.00474439 

 Power = 1.01848 

 Asymptotic Correlation Matrix of Parameter Estimates 

 ( *** The model parameter(s) -Background -Power 

 have been estimated at a boundary point, or have been specified by 

the user, 

 and do not appear in the correlation matrix ) 

 Slope 

 Slope 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 Background 0 NA 

 Slope 0.00496352 0.000693669 0.00360396 

0.00632309 

 Power 1 NA 
B-8 DRAFT – DO NOT CITE OR QUOTE 

NA - Indicates that this parameter has hit a bound 

 implied by some inequality constraint and thus

 has no standard error. 

 Analysis of Deviance Table 

 Model Log(likelihood) # Param's Deviance Test d.f. P-value 

 Full model -65.7706 3 

 Fitted model -65.8419 1 0.142715 2 0.9311 

 Reduced model -82.1514 1 32.7616 2 <.0001 

 AIC: 133.684 

 Goodness of Fit 

 Scaled 

 Dose Est._Prob. Expected Observed Size Residual 

 ------------------------------------------------------------------------ 

 0.0000 0.0000 0.000 0.000 20 0.000 

 113.0000 0.4293 21.035 22.050 49 0.293 

 227.0000 0.6759 33.795 33.000 50 -0.240 

 Chi^2 = 0.14 d.f. = 2 P-value = 0.9308

 Benchmark Dose Computation 

Specified effect = 0.1 

Risk Type = Extra risk 

Confidence level = 0.95 

 BMD = 21.227 

 BMDL = 16.9904 
B-9 DRAFT – DO NOT CITE OR QUOTE 

 0

 0.1

 0.2

 0.3

 0.4

 0.5

 0.6

 0.7

 0.8

 0 50 100 150 200

Fraction Affected

dose

Gamma Multi-Hit Model with 0.95 Confidence Level

14:55 04/09 2009

BMDL BMD

 

Gamma Multi-Hit
B-10 DRAFT – DO NOT CITE OR QUOTE 

Multistage 1°

==================================================================== 

 Multistage Model. $Revision: 2.1 $ $Date: 2000/08/21 03:38:21 $ 

 Input Data File: C:\BMDS\UNSAVED1.(d) 

 Gnuplot Plotting File: C:\BMDS\UNSAVED1.plt 

 Thu Sep 14 09:09:29 2006 

 ==================================================================== 

 BMDS Model Run NCI 1978 Tubular Nephropathy Male Rat - Multistage 1 degree Model 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the probability function is: 

 P[response] = background + (1-background)*[1-EXP( 

 -beta1*dose^1)] 

 The parameter betas are restricted to be positive 

 Dependent variable = PercentPositiveModerateMarkedTubularNephropathy 

 Independent variable = ularNephropathy 

 Total number of observations = 3 

 Total number of records with missing values = 0 

 Total number of parameters in model = 2 

 Total number of specified parameters = 0 

 Degree of polynomial = 1 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 Default Initial Parameter Values 

 Background = 0.0201528 

 Beta(1) = 0.00475168 

 Asymptotic Correlation Matrix of Parameter Estimates 

 ( *** The model parameter(s) -Background 

 have been estimated at a boundary point, or have been specified by 

the user, 

 and do not appear in the correlation matrix ) 

 Beta(1) 

 Beta(1) 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 Background 0 * * * 

 Beta(1) 0.00495719 * * * 

* - Indicates that this value is not calculated. 
B-11 DRAFT – DO NOT CITE OR QUOTE 

 Analysis of Deviance Table 

 Model Log(likelihood) # Param's Deviance Test d.f. P-value 

 Full model -65.7706 3 

 Fitted model -65.8277 1 0.114158 2 0.9445 

 Reduced model -82.1514 1 32.7616 2 <.0001 

 AIC: 133.655 

 Goodness of Fit 

 Scaled 

 Dose Est._Prob. Expected Observed Size Residual 

 ------------------------------------------------------------------------ 

 0.0000 0.0000 0.000 0.000 20 0.000 

 113.0000 0.4289 21.015 22.050 49 0.299 

 227.0000 0.6754 33.772 33.000 50 -0.233 

 Chi^2 = 0.14 d.f. = 2 P-value = 0.9307

 Benchmark Dose Computation 

Specified effect = 0.1 

Risk Type = Extra risk 

Confidence level = 0.95 

 BMD = 21.2541 

 BMDL = 17.0107 

 BMDU = 26.9612 

Taken together, (17.0107, 26.9612) is a 90 % two-sided confidence 

interval for the BMD 

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0 50 100 150 200 250 300

Fraction Affected

dose

Multistage Model with 0.95 Confidence Level

09:09 09/14 2006

BMDL BMD

 

Multistage
B-12 DRAFT – DO NOT CITE OR QUOTE 

Weibull

 ==================================================================== 

 Weibull Model $Revision: 2.2 $ $Date: 2000/03/17 22:27:16 $ 

 Input Data File: C:\BMDS\UNSAVED1.(d) 

 Gnuplot Plotting File: C:\BMDS\UNSAVED1.plt 

 Thu Sep 14 09:13:24 2006 

 ==================================================================== 

 BMDS Model Run NCI 1978 Tubular Nephropathy Male Rat - Weibull Model 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the probability function is: 

 P[response] = background + (1-background)*[1-EXP(-slope*dose^power)] 

 Dependent variable = PercentPositiveModerateMarkedTubularNephropathy 

 Independent variable = ularNephropathy 

 Power parameter is restricted as power >=1 

 Total number of observations = 3 

 Total number of records with missing values = 0 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 Default Initial (and Specified) Parameter Values 

 Background = 0.0238095 

 Slope = 0.00453277 

 Power = 1.00295 

 Asymptotic Correlation Matrix of Parameter Estimates 

 ( *** The model parameter(s) -Background -Power 

 have been estimated at a boundary point, or have been specified by 

the user, 

 and do not appear in the correlation matrix ) 

 Slope 

 Slope 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 Background 0 NA 

 Slope 0.00496352 0.000693669 0.00360396 

0.00632309 

 Power 1 NA 

NA - Indicates that this parameter has hit a bound 

 implied by some inequality constraint and thus

 has no standard error. 

 Analysis of Deviance Table 

 Model Log(likelihood) # Param's Deviance Test d.f. P-value 
B-13 DRAFT – DO NOT CITE OR QUOTE 

 Full model -65.7706 3 

 Fitted model -65.8419 1 0.142715 2 0.9311 

 Reduced model -82.1514 1 32.7616 2 <.0001 

 AIC: 133.684 

 Goodness of Fit 

 Scaled 

 Dose Est._Prob. Expected Observed Size Residual 

 ------------------------------------------------------------------------ 

 0.0000 0.0000 0.000 0.000 20 0.000 

 113.0000 0.4293 21.035 22.050 49 0.293 

 227.0000 0.6759 33.795 33.000 50 -0.240 

 Chi^2 = 0.14 d.f. = 2 P-value = 0.9308

 Benchmark Dose Computation 

Specified effect = 0.1 

Risk Type = Extra risk 

Confidence level = 0.95 

 BMD = 21.227 

 BMDL = 16.9904 
B-14 DRAFT – DO NOT CITE OR QUOTE 

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0 50 100 150 200 250 300

Fraction Affected

dose

Weibull Model with 0.95 Confidence Level

09:13 09/14 2006

BMDL BMD

 

Weibull
B-15 DRAFT – DO NOT CITE OR QUOTE 

NCI (1978) Tubular Nephropathy in Female Rats

Multistage 2°

==================================================================== 

 Multistage Model. $Revision: 2.1 $ $Date: 2000/08/21 03:38:21 $ 

 Input Data File: C:\BMDS\UNSAVED1.(d) 

 Gnuplot Plotting File: C:\BMDS\UNSAVED1.plt 

 Thu Apr 09 16:18:29 2009 

 ==================================================================== 

 BMDS Model Run - NCI 1978 Tubular Nephropathy Female Rat - Multistage 2 degree Model 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the probability function is: 

 P[response] = background + (1-background)*[1-EXP( 

 -beta1*dose^1-beta2*dose^2)] 

 The parameter betas are restricted to be positive 

 Dependent variable = PercentPositiveModerateMarkedTubularNephropathy 

 Independent variable = ularNephropathy 

 Total number of observations = 3 

 Total number of records with missing values = 0 

 Total number of parameters in model = 3 

 Total number of specified parameters = 0 

 Degree of polynomial = 2 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 Default Initial Parameter Values 

 Background = 0 

 Beta(1) = 0 

 Beta(2) = 1.74381e-005 

 Asymptotic Correlation Matrix of Parameter Estimates 

 ( *** The model parameter(s) -Background -Beta(1) 

 have been estimated at a boundary point, or have been specified by 

the user, 

 and do not appear in the correlation matrix ) 

 Beta(2) 

 Beta(2) 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 Background 0 * * * 

 Beta(1) 0 * * * 

 Beta(2) 1.62048e-005 * * * 

* - Indicates that this value is not calculated. 
B-16 DRAFT – DO NOT CITE OR QUOTE 

 Analysis of Deviance Table 

 Model Log(likelihood) # Param's Deviance Test d.f. P-value 

 Full model -56.7357 3 

 Fitted model -57.0429 1 0.614339 2 0.7355 

 Reduced model -74.4688 1 35.466 2 <.0001 

 AIC: 116.086 

 Goodness of Fit 

 Scaled 

 Dose Est._Prob. Expected Observed Size Residual 

 ------------------------------------------------------------------------ 

 0.0000 0.0000 0.000 0.000 20 0.000 

 113.0000 0.1869 9.346 9.000 50 -0.125 

 227.0000 0.5661 27.741 28.910 49 0.337 

 Chi^2 = 0.13 d.f. = 2 P-value = 0.9374

 Benchmark Dose Computation 

Specified effect = 0.1 

Risk Type = Extra risk 

Confidence level = 0.95 

 BMD = 80.6338 

 BMDL = 41.8864 

 BMDU = 93.2552 

Taken together, (41.8864, 93.2552) is a 90 % two-sided confidence 

interval for the BMD 
B-17 DRAFT – DO NOT CITE OR QUOTE 

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0 50 100 150 200 250 300

Fraction Affected

dose

Multistage Model with 0.95 Confidence Level

09:21 09/14 2006

BMDL BMD

 

Multistage
B-18 DRAFT – DO NOT CITE OR QUOTE 

NTP (1989) Male Rat Nephropathy

Multistage 1° Model 

 ==================================================================== 

 Multistage Model. (Version: 3.0; Date: 05/16/2008) 

 Input Data File: C:\USEPA\BMDS2\Temp\tmp9D5.(d) 

 Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmp9D5.plt 

 Wed Apr 08 11:27:18 2009 

 ==================================================================== 

 BMDS Model Run NTP 1989 Tubular Nephropathy Male Rat - Multistage 1 degree 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the probability function is: 

 P[response] = background + (1-background)*[1-EXP( 

 -beta1*dose^1)] 

 The parameter betas are restricted to be positive 

 Dependent variable = PercentPositiveModerateMarkedTubularNephropathy 

 Independent variable = ularNephropathy 

 Total number of observations = 3 

 Total number of records with missing values = 0 

 Total number of parameters in model = 2 

 Total number of specified parameters = 0 

 Degree of polynomial = 1 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 Default Initial Parameter Values 

 Background = 0.354136 

 Beta(1) = 0.0335717 

 Asymptotic Correlation Matrix of Parameter Estimates 

 Background Beta(1) 

Background 1 -0.72 

 Beta(1) -0.72 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 Background 0.356651 * * * 

 Beta(1) 0.0329547 * * * 

* - Indicates that this value is not calculated. 

 Analysis of Deviance Table 
B-19 DRAFT – DO NOT CITE OR QUOTE 

 Model Log(likelihood) # Param's Deviance Test d.f. P-value 

 Full model -100.939 3 

 Fitted model -100.952 2 0.0258029 1 0.8724 

 Reduced model -103.852 1 5.82641 2 0.0543 

 AIC: 205.903 

 Goodness of Fit 

 Scaled 

 Dose Est._Prob. Expected Observed Size Residual 

 ------------------------------------------------------------------------ 

 0.0000 0.3567 17.833 18.000 50 0.049 

 7.0000 0.4892 24.459 24.000 50 -0.130 

 14.0000 0.5944 29.721 30.000 50 0.080 

 Chi^2 = 0.03 d.f. = 1 P-value = 0.8724

 Benchmark Dose Computation 

Specified effect = 0.1 

Risk Type = Extra risk 

Confidence level = 0.95 

 BMD = 3.19713 

 BMDL = 1.8769 

 BMDU = 10.0721 

Taken together, (1.8769 , 10.0721) is a 90 % two-sided confidence 

interval for the BMD 
B-20 DRAFT – DO NOT CITE OR QUOTE 

 0.2

 0.3

 0.4

 0.5

 0.6

 0.7

 0 2 4 6 8 10 12 14

Fraction Affected

dose

Multistage Model with 0.95 Confidence Level

11:27 04/08 2009

BMDL BMD

 

Multistage
B-21 DRAFT – DO NOT CITE OR QUOTE 

Quantal-linear Model 

 

 ==================================================================== 

 Quantal Linear Model using Weibull Model (Version: 2.12; Date: 05/16/2008) 

 Input Data File: C:\USEPA\BMDS2\Temp\tmpA17.(d) 

 Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpA17.plt 

 Wed Apr 08 13:34:31 2009 

 ==================================================================== 

 BMDS Model Run NTP 1989 Tubular Nephropathy Male Rat - Quantal-linear Model 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the probability function is: 

 P[response] = background + (1-background)*[1-EXP(-slope*dose)] 

 Dependent variable = PercentPositiveModerateMarkedTubularNephropathy 

 Independent variable = ularNephropathy 

 Total number of observations = 3 

 Total number of records with missing values = 0 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 Default Initial (and Specified) Parameter Values 

 Background = 0.362745 

 Slope = 0.0329154 

 Power = 1 Specified 

 Asymptotic Correlation Matrix of Parameter Estimates 

 ( *** The model parameter(s) -Power 

 have been estimated at a boundary point, or have been specified by 

the user, 

 and do not appear in the correlation matrix ) 

 Background Slope 

Background 1 -0.66 

 Slope -0.66 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 Background 0.356651 0.0642145 0.230793 

0.482509 

 Slope 0.0329547 0.0137884 0.00593 

0.0599794 

 Analysis of Deviance Table 

 Model Log(likelihood) # Param's Deviance Test d.f. P-value 

 Full model -100.939 3 
B-22 DRAFT – DO NOT CITE OR QUOTE 

 Fitted model -100.952 2 0.0258029 1 0.8724 

 Reduced model -103.852 1 5.82641 2 0.0543 

 AIC: 205.903 

 Goodness of Fit 

 Scaled 

 Dose Est._Prob. Expected Observed Size Residual 

 ------------------------------------------------------------------------ 

 0.0000 0.3567 17.833 18.000 50 0.049 

 7.0000 0.4892 24.459 24.000 50 -0.130 

 14.0000 0.5944 29.721 30.000 50 0.080 

 Chi^2 = 0.03 d.f. = 1 P-value = 0.8724

 Benchmark Dose Computation 

Specified effect = 0.1 

Risk Type = Extra risk 

Confidence level = 0.95 

 BMD = 3.19713 

 BMDL = 1.8769 
B-23 DRAFT – DO NOT CITE OR QUOTE 

 0.2

 0.3

 0.4

 0.5

 0.6

 0.7

 0 2 4 6 8 10 12 14

Fraction Affected

dose

Quantal Linear Model with 0.95 Confidence Level

13:34 04/08 2009

BMDL BMD

 

Quantal Linear
B-24 DRAFT – DO NOT CITE OR QUOTE 

NTP (1989) Female Rat Nephropathy

Gamma Model

 ==================================================================== 

 Gamma Model. (Version: 2.13; Date: 05/16/2008) 

 Input Data File: C:\USEPA\BMDS2\Temp\tmpD9.(d) 

 Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpD9.plt 

 Fri Apr 10 10:19:37 2009 

 ==================================================================== 

 BMDS Model Run NTP 1989 Tubular Nephropathy Female Rat - Gamma Model 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the probability function is: 

 P[response]= background+(1-background)*CumGamma[slope*dose,power], 

 where CumGamma(.) is the cummulative Gamma distribution function 

 Dependent variable = PercentPositiveModerateMarkedTubularNephropathy 

 Independent variable = ularNephropathy 

 Power parameter is restricted as power >=1 

 Total number of observations = 3 

 Total number of records with missing values = 0 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 Default Initial (and Specified) Parameter Values 

 Background = 0.245098 

 Slope = 0.0111213 

 Power = 1.3 

 Asymptotic Correlation Matrix of Parameter Estimates 

 ( *** The model parameter(s) -Power 

 have been estimated at a boundary point, or have been specified by 

the user, 

 and do not appear in the correlation matrix ) 

 Background Slope 

Background 1 -0.55 

 Slope -0.55 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 Background 0.242452 0.0592711 0.126283 

0.358621 

 Slope 0.00694477 0.0016862 0.00363988 

0.0102497 

 Power 1 NA 

NA - Indicates that this parameter has hit a bound 
B-25 DRAFT – DO NOT CITE OR QUOTE 

 implied by some inequality constraint and thus

 has no standard error. 

 Analysis of Deviance Table 

 Model Log(likelihood) # Param's Deviance Test d.f. P-value 

 Full model -93.9362 3 

 Fitted model -93.9519 2 0.0312372 1 0.8597 

 Reduced model -102.85 1 17.8276 2 0.0001345 

 AIC: 191.904 

 Goodness of Fit 

 Scaled 

 Dose Est._Prob. Expected Observed Size Residual 

 ------------------------------------------------------------------------ 

 0.0000 0.2425 12.123 12.000 50 -0.040 

 57.0000 0.4901 24.504 25.000 50 0.140 

 114.0000 0.6568 32.182 31.850 49 -0.100 

 Chi^2 = 0.03 d.f. = 1 P-value = 0.8596

 Benchmark Dose Computation 

Specified effect = 0.1 

Risk Type = Extra risk 

Confidence level = 0.95 

 BMD = 15.1712 

 BMDL = 10.7248 
B-26 DRAFT – DO NOT CITE OR QUOTE 

 0.1

 0.2

 0.3

 0.4

 0.5

 0.6

 0.7

 0.8

 0 20 40 60 80 100

Fraction Affected

dose

Quantal Linear Model with 0.95 Confidence Level

12:21 12/16 2008

BMDL BMD

 

Quantal Linear

BMD Lower Bound
B-27 DRAFT – DO NOT CITE OR QUOTE 

Quantal-linear Model

 ==================================================================== 

 Quantal Linear Model using Weibull Model (Version: 2.12; Date: 05/16/2008) 

 Input Data File: C:\USEPA\BMDS2\Temp\tmpE4.(d) 

 Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpE4.plt 

 Fri Apr 10 10:36:29 2009 

 ==================================================================== 

 BMDS Model Run NTP 1989 Tubular Nephropathy Female Rat - Quantal-linear Model 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the probability function is: 

 P[response] = background + (1-background)*[1-EXP(-slope*dose)] 

 Dependent variable = PercentPositiveModerateMarkedTubularNephropathy 

 Independent variable = ularNephropathy 

 Total number of observations = 3 

 Total number of records with missing values = 0 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 Default Initial (and Specified) Parameter Values 

 Background = 0.245098 

 Slope = 0.00666772 

 Power = 1 Specified 

 Asymptotic Correlation Matrix of Parameter Estimates 

 ( *** The model parameter(s) -Power 

 have been estimated at a boundary point, or have been specified by 

the user, 

 and do not appear in the correlation matrix ) 

 Background Slope 

Background 1 -0.55 

 Slope -0.55 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 Background 0.242451 0.0592711 0.126282 

0.358621 

 Slope 0.00694478 0.0016862 0.00363989 

0.0102497 

 Analysis of Deviance Table 

 Model Log(likelihood) # Param's Deviance Test d.f. P-value 

 Full model -93.9362 3 
B-28 DRAFT – DO NOT CITE OR QUOTE 

 Fitted model -93.9519 2 0.0312372 1 0.8597 

 Reduced model -102.85 1 17.8276 2 0.0001345 

 AIC: 191.904 

 Goodness of Fit 

 Scaled 

 Dose Est._Prob. Expected Observed Size Residual 

 ------------------------------------------------------------------------ 

 0.0000 0.2425 12.123 12.000 50 -0.040 

 57.0000 0.4901 24.504 25.000 50 0.140 

 114.0000 0.6568 32.182 31.850 49 -0.100 

 Chi^2 = 0.03 d.f. = 1 P-value = 0.8596

 Benchmark Dose Computation 

Specified effect = 0.1 

Risk Type = Extra risk 

Confidence level = 0.95 

 BMD = 15.1712 

 BMDL = 10.7248 
B-29 DRAFT – DO NOT CITE OR QUOTE 

 0.1

 0.2

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 0.8

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Fraction Affected

dose

Quantal Linear Model with 0.95 Confidence Level

10:36 04/10 2009

BMDL BMD

 

Quantal Linear
B-30 DRAFT – DO NOT CITE OR QUOTE 

 

Weibull Model 

 ==================================================================== 

 Weibull Model using Weibull Model (Version: 2.12; Date: 05/16/2008) 

 Input Data File: C:\USEPA\BMDS2\Temp\tmpE3.(d) 

 Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpE3.plt 

 Fri Apr 10 10:34:27 2009 

 ==================================================================== 

 BMDS Model Run NTP 1989 Tubular Nephropathy Female Rat - Weibull Model 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the probability function is: 

 P[response] = background + (1-background)*[1-EXP(-slope*dose^power)] 

 Dependent variable = PercentPositiveModerateMarkedTubularNephropathy 

 Independent variable = ularNephropathy 

 Power parameter is restricted as power >=1 

 Total number of observations = 3 

 Total number of records with missing values = 0 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 Default Initial (and Specified) Parameter Values 

 Background = 0.245098 

 Slope = 0.00666772 

 Power = 1 

 Asymptotic Correlation Matrix of Parameter Estimates 

 ( *** The model parameter(s) -Power 

 have been estimated at a boundary point, or have been specified by 

the user, 

 and do not appear in the correlation matrix ) 

 Background Slope 

Background 1 -0.55 

 Slope -0.55 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 Background 0.242451 0.0592711 0.126282 

0.358621 

 Slope 0.00694478 0.0016862 0.00363989 

0.0102497 

 Power 1 NA 

NA - Indicates that this parameter has hit a bound 

 implied by some inequality constraint and thus
B-31 DRAFT – DO NOT CITE OR QUOTE 

 has no standard error. 

 Analysis of Deviance Table 

 Model Log(likelihood) # Param's Deviance Test d.f. P-value 

 Full model -93.9362 3 

 Fitted model -93.9519 2 0.0312372 1 0.8597 

 Reduced model -102.85 1 17.8276 2 0.0001345 

 AIC: 191.904 

 Goodness of Fit 

 Scaled 

 Dose Est._Prob. Expected Observed Size Residual 

 ------------------------------------------------------------------------ 

 0.0000 0.2425 12.123 12.000 50 -0.040 

 57.0000 0.4901 24.504 25.000 50 0.140 

 114.0000 0.6568 32.182 31.850 49 -0.100 

 Chi^2 = 0.03 d.f. = 1 P-value = 0.8596

 Benchmark Dose Computation 

Specified effect = 0.1 

Risk Type = Extra risk 

Confidence level = 0.95 

 BMD = 15.1712 

 BMDL = 10.7248 
B-32 DRAFT – DO NOT CITE OR QUOTE 

 0.1

 0.2

 0.3

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 0.6

 0.7

 0.8

 0 20 40 60 80 100

Fraction Affected

dose

Weibull Model with 0.95 Confidence Level

10:34 04/10 2009

BMDL BMD

 

Weibull
B-33 DRAFT – DO NOT CITE OR QUOTE 

NTP (1989) Linear Mineralization in Male Rats 

Probit Model 

 ==================================================================== 

 Probit Model. (Version: 3.1; Date: 05/16/2008) 

 Input Data File: C:\USEPA\BMDS2\Temp\tmpA33.(d) 

 Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpA33.plt 

 Wed Apr 08 14:24:02 2009 

 ==================================================================== 

 BMDS Model Run NTP 1989 Linear Mineralization Male Rat - Probit Model 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the probability function is: 

 P[response] = CumNorm(Intercept+Slope*Dose), 

 where CumNorm(.) is the cumulative normal distribution function 

 Dependent variable = PercentPositiveLinearMineralization 

 Independent variable = ion 

 Slope parameter is not restricted 

 Total number of observations = 3 

 Total number of records with missing values = 0 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 Default Initial (and Specified) Parameter Values 

 background = 0 Specified 

 intercept = -1.67551 

 slope = 0.149038 

 Asymptotic Correlation Matrix of Parameter Estimates 

 ( *** The model parameter(s) -background 

 have been estimated at a boundary point, or have been specified by 

the user, 

 and do not appear in the correlation matrix ) 

 intercept slope 

 intercept 1 -0.87 

 slope -0.87 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 intercept -1.62793 0.244257 -2.10666 -

1.14919 

 slope 0.144885 0.0238239 0.0981906 

0.191579 
B-34 DRAFT – DO NOT CITE OR QUOTE 

 Analysis of Deviance Table 

 Model Log(likelihood) # Param's Deviance Test d.f. P-value 

 Full model -71.6113 3 

 Fitted model -71.8283 2 0.433989 1 0.51 

 Reduced model -94.7689 1 46.3152 2 <.0001 

 AIC: 147.657 

 Goodness of Fit 

 Scaled 

 Dose Est._Prob. Expected Observed Size Residual 

 ------------------------------------------------------------------------ 

 0.0000 0.0518 2.589 2.000 50 -0.376 

 7.0000 0.2697 13.485 15.000 50 0.483 

 14.0000 0.6556 32.780 32.000 50 -0.232 

 Chi^2 = 0.43 d.f. = 1 P-value = 0.5129

 Benchmark Dose Computation 

Specified effect = 0.1 

Risk Type = Extra risk 

Confidence level = 0.95 

 BMD = 3.98089 

 BMDL = 3.21773 
B-35 DRAFT – DO NOT CITE OR QUOTE 

 0

 0.1

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 0.8

 0 2 4 6 8 10 12 14

Fraction Affected

dose

Probit Model with 0.95 Confidence Level

14:24 04/08 2009

BMDL BMD

 

Probit
B-36 DRAFT – DO NOT CITE OR QUOTE 

NTP (1989) Male Rat Hyperplasia of Pelvic Transitional 

Epithelium

LogLogistic Model

 ==================================================================== 

 Logistic Model. (Version: 2.12; Date: 05/16/2008) 

 Input Data File: C:\USEPA\BMDS2\Temp\tmp4D5.(d) 

 Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmp4D5.plt 

 Wed Aug 12 14:26:53 2009 

 ==================================================================== 

 BMDS Model Run – NTP 989 – Male Rat – Hyperplasia – LogLogistic Model 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the probability function is: 

 P[response] = background+(1-background)/[1+EXP(-intercept-slope*Log(dose))] 

 Dependent variable = Effect 

 Independent variable = DOSE 

 Slope parameter is restricted as slope >= 1 

 Total number of observations = 3 

 Total number of records with missing values = 0 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 User has chosen the log transformed model 

 Default Initial Parameter Values 

 background = 0 

 intercept = -3.7612 

 slope = 1 

 Asymptotic Correlation Matrix of Parameter Estimates 

 ( *** The model parameter(s) -background -slope 

 have been estimated at a boundary point, or have been specified by 

the user, 

 and do not appear in the correlation matrix ) 

 intercept 

 intercept 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 background 0 * * * 

 intercept -4.15077 * * * 

 slope 1 * * * 

* - Indicates that this value is not calculated. 
B-37 DRAFT – DO NOT CITE OR QUOTE 

 Analysis of Deviance Table 

 Model Log(likelihood) # Param's Deviance Test d.f. P-value 

 Full model -40.4963 3 

 Fitted model -41.2103 1 1.42796 2 0.4897 

 Reduced model -46.5274 1 12.0622 2 0.002403 

 AIC: 84.4207 

 Goodness of Fit 

 Scaled 

 Dose Est._Prob. Expected Observed Size Residual 

 ------------------------------------------------------------------------ 

 0.0000 0.0000 0.000 0.000 50 0.000 

 7.0000 0.0993 4.966 7.000 50 0.962 

 14.0000 0.1807 9.034 7.000 50 -0.748 

 Chi^2 = 1.48 d.f. = 2 P-value = 0.4761

 Benchmark Dose Computation 

Specified effect = 0.1 

Risk Type = Extra risk 

Confidence level = 0.95 

 BMD = 7.05365 

 BMDL = 4.48322 
B-38 DRAFT – DO NOT CITE OR QUOTE 

 0

 0.05

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 0.3

 0 2 4 6 8 10 12 14

Fraction Affected

dose

Log-Logistic Model with 0.95 Confidence Level

14:26 08/12 2009

BMDL BMD

 

Log-Logistic
B-39 DRAFT – DO NOT CITE OR QUOTE 

Gorzinski (1985) Atrophy and Degeneration of renal tubules in 

Male Rats 

Gamma Model 

==================================================================== 

 Gamma Model. (Version: 2.13; Date: 05/16/2008) 

 Input Data File: C:\USEPA\BMDS2\Temp\tmpF14.(d) 

 Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpF14.plt 

 Thu Oct 08 08:59:00 2009 

 ==================================================================== 

 ndDegenRenalTubulesDataNoSeverityMaleRat.dax 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the probability function is: 

 P[response]= background+(1-background)*CumGamma[slope*dose,power], 

 where CumGamma(.) is the cummulative Gamma distribution function 

 Dependent variable = Effect 

 Independent variable = DOSE 

 Power parameter is restricted as power >=1 

 Total number of observations = 4 

 Total number of records with missing values = 0 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 Default Initial (and Specified) Parameter Values 

 Background = 0.136364 

 Slope = 0.0871864 

 Power = 1.3 

 Asymptotic Correlation Matrix of Parameter Estimates 

 Background Slope Power 

Background 1 0.52 0.64 

 Slope 0.52 1 0.93 

 Power 0.64 0.93 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 Background 0.110626 0.107207 -0.0994949 

0.320747 

 Slope 0.0787607 0.0846932 -0.0872348 

0.244756 

 Power 1.00164 1.07041 -1.09632 

3.0996 
B-40 DRAFT – DO NOT CITE OR QUOTE 

 Analysis of Deviance Table 

 Model Log(likelihood) # Param's Deviance Test d.f. P-value 

 Full model -14.3635 4 

 Fitted model -14.4712 3 0.215359 1 0.6426 

 Reduced model -27.7259 1 26.7248 3 <.0001 

 AIC: 34.9424 

 Goodness of Fit 

 Scaled 

 Dose Est._Prob. Expected Observed Size Residual 

 ------------------------------------------------------------------------ 

 0.0000 0.1106 1.106 1.000 10 -0.107 

 1.0000 0.1777 1.777 2.000 10 0.185 

 15.0000 0.7265 7.265 7.000 10 -0.188 

 62.0000 0.9932 9.932 10.000 10 0.261 

 Chi^2 = 0.15 d.f. = 1 P-value = 0.6994

 Benchmark Dose Computation 

Specified effect = 0.1 

Risk Type = Extra risk 

Confidence level = 0.95 

 BMD = 1.34399 

 BMDL = 0.727509 

 0

 0.2

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 1

 0 10 20 30 40 50 60

Fraction Affected

dose

Gamma Multi-Hit Model with 0.95 Confidence Level

08:59 10/08 2009

BMDL BMD

 

Gamma Multi-Hit
B-41 DRAFT – DO NOT CITE OR QUOTE 

Multistage 1 degree Model 

 ==================================================================== 

 Multistage Model. (Version: 3.0; Date: 05/16/2008) 

 Input Data File: C:\USEPA\BMDS2\Temp\tmpF17.(d) 

 Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpF17.plt 

 Thu Oct 08 09:00:57 2009 

 ==================================================================== 

 ndDegenRenalTubulesDataNoSeverityMaleRat.dax 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the probability function is: 

 P[response] = background + (1-background)*[1-EXP( 

 -beta1*dose^1)] 

 The parameter betas are restricted to be positive 

 Dependent variable = Effect 

 Independent variable = DOSE 

 Total number of observations = 4 

 Total number of records with missing values = 0 

 Total number of parameters in model = 2 

 Total number of specified parameters = 0 

 Degree of polynomial = 1 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 Default Initial Parameter Values 

 Background = 0 

 Beta(1) = 1.66732e+018 

 Asymptotic Correlation Matrix of Parameter Estimates 

 Background Beta(1) 

Background 1 -0.4 

 Beta(1) -0.4 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 Background 0.11052 * * * 

 Beta(1) 0.0786399 * * * 

* - Indicates that this value is not calculated. 

 Analysis of Deviance Table 

 Model Log(likelihood) # Param's Deviance Test d.f. P-value 
B-42 DRAFT – DO NOT CITE OR QUOTE 

 Full model -14.3635 4 

 Fitted model -14.4712 2 0.215361 2 0.8979 

 Reduced model -27.7259 1 26.7248 3 <.0001 

 AIC: 32.9424 

 Goodness of Fit 

 Scaled 

 Dose Est._Prob. Expected Observed Size Residual 

 ------------------------------------------------------------------------ 

 0.0000 0.1105 1.105 1.000 10 -0.106 

 1.0000 0.1778 1.778 2.000 10 0.184 

 15.0000 0.7266 7.266 7.000 10 -0.189 

 62.0000 0.9932 9.932 10.000 10 0.261 

 Chi^2 = 0.15 d.f. = 2 P-value = 0.9283

 Benchmark Dose Computation 

Specified effect = 0.1 

Risk Type = Extra risk 

Confidence level = 0.95 

 BMD = 1.33978 

 BMDL = 0.727509 

 BMDU = 2.66189 

Taken together, (0.727509, 2.66189) is a 90 % two-sided confidence 

interval for the BMD 
B-43 DRAFT – DO NOT CITE OR QUOTE 

 0

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 0.8

 1

 0 10 20 30 40 50 60

Fraction Affected

dose

Multistage Model with 0.95 Confidence Level

09:00 10/08 2009

BMDL BMD

 

Multistage
B-44 DRAFT – DO NOT CITE OR QUOTE 

Quantal-linear Model 

 ==================================================================== 

 Quantal Linear Model using Weibull Model (Version: 2.12; Date: 05/16/2008) 

 Input Data File: C:\USEPA\BMDS2\Temp\tmpF18.(d) 

 Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpF18.plt 

 Thu Oct 08 09:02:11 2009 

 ==================================================================== 

 ndDegenRenalTubulesDataNoSeverityMaleRat.dax 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the probability function is: 

 P[response] = background + (1-background)*[1-EXP(-slope*dose)] 

 Dependent variable = Effect 

 Independent variable = DOSE 

 Total number of observations = 4 

 Total number of records with missing values = 0 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 Default Initial (and Specified) Parameter Values 

 Background = 0.136364 

 Slope = 0.047491 

 Power = 1 Specified 

 Asymptotic Correlation Matrix of Parameter Estimates 

 ( *** The model parameter(s) -Power 

 have been estimated at a boundary point, or have been specified by 

the user, 

 and do not appear in the correlation matrix ) 

 Background Slope 

Background 1 -0.29 

 Slope -0.29 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 Background 0.11052 0.0819804 -0.0501583 

0.271199 

 Slope 0.0786399 0.0310542 0.0177749 

0.139505 

 Analysis of Deviance Table 

 Model Log(likelihood) # Param's Deviance Test d.f. P-value 

 Full model -14.3635 4 

 Fitted model -14.4712 2 0.215361 2 0.8979 
B-45 DRAFT – DO NOT CITE OR QUOTE 

 Reduced model -27.7259 1 26.7248 3 <.0001 

 AIC: 32.9424 

 Goodness of Fit 

 Scaled 

 Dose Est._Prob. Expected Observed Size Residual 

 ------------------------------------------------------------------------ 

 0.0000 0.1105 1.105 1.000 10 -0.106 

 1.0000 0.1778 1.778 2.000 10 0.184 

 15.0000 0.7266 7.266 7.000 10 -0.189 

 62.0000 0.9932 9.932 10.000 10 0.261 

 Chi^2 = 0.15 d.f. = 2 P-value = 0.9283

 Benchmark Dose Computation 

Specified effect = 0.1 

Risk Type = Extra risk 

Confidence level = 0.95 

 BMD = 1.33978 

 BMDL = 0.727509 

 0

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 0.8

 1

 0 10 20 30 40 50 60

Fraction Affected

dose

Quantal Linear Model with 0.95 Confidence Level

09:02 10/08 2009

BMDL BMD

 

Quantal Linear
B-46 DRAFT – DO NOT CITE OR QUOTE 

Gorzinski (1985) Atrophy and Degeneration of renal tubules in 

Female Rats 

Multistage 1 degree Model 

==================================================================== 

 Multistage Model. (Version: 3.0; Date: 05/16/2008) 

 Input Data File: C:\USEPA\BMDS2\Temp\tmpF51.(d) 

 Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpF51.plt 

 Thu Oct 08 09:34:37 2009 

 ==================================================================== 

 ndDegenRenalTubulesDataNoSeverityFemaleRat.dax 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the probability function is: 

 P[response] = background + (1-background)*[1-EXP( 

 -beta1*dose^1)] 

 The parameter betas are restricted to be positive 

 Dependent variable = Effect 

 Independent variable = DOSE 

 Total number of observations = 4 

 Total number of records with missing values = 0 

 Total number of parameters in model = 2 

 Total number of specified parameters = 0 

 Degree of polynomial = 1 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 Default Initial Parameter Values 

 Background = 0.0749781 

 Beta(1) = 0.0133129 

 Asymptotic Correlation Matrix of Parameter Estimates 

 Background Beta(1) 

Background 1 -0.51 

 Beta(1) -0.51 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 Background 0.0885222 * * * 

 Beta(1) 0.0123308 * * * 

* - Indicates that this value is not calculated. 
B-47 DRAFT – DO NOT CITE OR QUOTE 

 Analysis of Deviance Table 

 Model Log(likelihood) # Param's Deviance Test d.f. P-value 

 Full model -18.2358 4 

 Fitted model -18.3071 2 0.142532 2 0.9312 

 Reduced model -22.4934 1 8.51521 3 0.03648 

 AIC: 40.6141 

 Goodness of Fit 

 Scaled 

 Dose Est._Prob. Expected Observed Size Residual 

 ------------------------------------------------------------------------ 

 0.0000 0.0885 0.885 1.000 10 0.128 

 1.0000 0.0997 0.997 1.000 10 0.003 

 15.0000 0.2424 2.424 2.000 10 -0.313 

 62.0000 0.5757 5.757 6.000 10 0.156 

 Chi^2 = 0.14 d.f. = 2 P-value = 0.9330

 Benchmark Dose Computation 

Specified effect = 0.1 

Risk Type = Extra risk 

Confidence level = 0.95 

 BMD = 8.54451 

 BMDL = 4.49217 

 BMDU = 23.0819 

Taken together, (4.49217, 23.0819) is a 90 % two-sided confidence 

interval for the BMD 
B-48 DRAFT – DO NOT CITE OR QUOTE 

 0

 0.2

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 0.8

 0 10 20 30 40 50 60

Fraction Affected

dose

Multistage Model with 0.95 Confidence Level

09:34 10/08 2009

BMDL BMD

 

Multistage
B-49 DRAFT – DO NOT CITE OR QUOTE 

Quantal-linear Model 

==================================================================== 

 Quantal Linear Model using Weibull Model (Version: 2.12; Date: 05/16/2008) 

 Input Data File: C:\USEPA\BMDS2\Temp\tmpF52.(d) 

 Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpF52.plt 

 Thu Oct 08 09:35:41 2009 

 ==================================================================== 

 ndDegenRenalTubulesDataNoSeverityFemaleRat.dax 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the probability function is: 

 P[response] = background + (1-background)*[1-EXP(-slope*dose)] 

 Dependent variable = Effect 

 Independent variable = DOSE 

 Total number of observations = 4 

 Total number of records with missing values = 0 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 Default Initial (and Specified) Parameter Values 

 Background = 0.136364 

 Slope = 0.0120518 

 Power = 1 Specified 

 Asymptotic Correlation Matrix of Parameter Estimates 

 ( *** The model parameter(s) -Power 

 have been estimated at a boundary point, or have been specified by 

the user, 

 and do not appear in the correlation matrix ) 

 Background Slope 

Background 1 -0.33 

 Slope -0.33 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 Background 0.0885253 0.0622082 -0.0334005 

0.210451 

 Slope 0.0123305 0.00562158 0.00131237 

0.0233486 

 Analysis of Deviance Table 

 Model Log(likelihood) # Param's Deviance Test d.f. P-value 

 Full model -18.2358 4 

 Fitted model -18.3071 2 0.142532 2 0.9312 
B-50 DRAFT – DO NOT CITE OR QUOTE 

 Reduced model -22.4934 1 8.51521 3 0.03648 

 AIC: 40.6141 

 Goodness of Fit 

 Scaled 

 Dose Est._Prob. Expected Observed Size Residual 

 ------------------------------------------------------------------------ 

 0.0000 0.0885 0.885 1.000 10 0.128 

 1.0000 0.0997 0.997 1.000 10 0.003 

 15.0000 0.2424 2.424 2.000 10 -0.313 

 62.0000 0.5756 5.756 6.000 10 0.156 

 Chi^2 = 0.14 d.f. = 2 P-value = 0.9330

 Benchmark Dose Computation 

Specified effect = 0.1 

Risk Type = Extra risk 

Confidence level = 0.95 

 BMD = 8.54473 

 BMDL = 4.49217 

 0

 0.2

 0.4

 0.6

 0.8

 0 10 20 30 40 50 60

Fraction Affected

dose

Quantal Linear Model with 0.95 Confidence Level

09:35 10/08 2009

BMDL BMD

 

Quantal Linear
B-51 DRAFT – DO NOT CITE OR QUOTE 

Gorzinski et al. (1985) Male Rat Hypertrophy and/or Dilation of 

Proximal Tubules

Gamma Model

 ==================================================================== 

 Gamma Model. (Version: 2.13; Date: 05/16/2008) 

 Input Data File: C:\USEPA\BMDS2\Temp\tmp4D6.(d) 

 Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmp4D6.plt 

 Wed Aug 12 14:31:38 2009 

 ==================================================================== 

BMDS Model Run - Gorzinski et al (1985) - Male Rat – Hypertrophy/Dilation of Proximal 

Tubules - Gamma Model 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the probability function is: 

 P[response]= background+(1-background)*CumGamma[slope*dose,power], 

 where CumGamma(.) is the cummulative Gamma distribution function 

 Dependent variable = Effect 

 Independent variable = DOSE 

 Power parameter is restricted as power >=1 

 Total number of observations = 4 

 Total number of records with missing values = 0 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 Default Initial (and Specified) Parameter Values 

 Background = 0.0454545 

 Slope = 0.0907614 

 Power = 1.3 

 Asymptotic Correlation Matrix of Parameter Estimates 

 ( *** The model parameter(s) -Background -Power 

 have been estimated at a boundary point, or have been specified by 

the user, 

 and do not appear in the correlation matrix ) 

 Slope 

 Slope 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 Background 0 NA 

 Slope 0.0860249 0.029523 0.0281609 

0.143889 

 Power 1 NA 

NA - Indicates that this parameter has hit a bound 

 implied by some inequality constraint and thus

 has no standard error. 
B-52 DRAFT – DO NOT CITE OR QUOTE 

 Analysis of Deviance Table 

 Model Log(likelihood) # Param's Deviance Test d.f. P-value 

 Full model -9.35947 4 

 Fitted model -9.44226 1 0.165576 3 0.9829 

 Reduced model -27.5256 1 36.3322 3 <.0001 

 AIC: 20.8845 

 Goodness of Fit 

 Scaled 

 Dose Est._Prob. Expected Observed Size Residual 

 ------------------------------------------------------------------------ 

 0.0000 0.0000 0.000 0.000 10 0.000 

 1.0000 0.0824 0.824 1.000 10 0.202 

 15.0000 0.7248 7.248 7.000 10 -0.176 

 62.0000 0.9952 9.952 10.000 10 0.220 

 Chi^2 = 0.12 d.f. = 3 P-value = 0.9893

 Benchmark Dose Computation 

Specified effect = 0.1 

Risk Type = Extra risk 

Confidence level = 0.95 

 BMD = 1.22477 

 BMDL = 0.710032 
B-53 DRAFT – DO NOT CITE OR QUOTE 

 0

 0.2

 0.4

 0.6

 0.8

 1

 0 10 20 30 40 50 60

Fraction Affected

dose

Gamma Multi-Hit Model with 0.95 Confidence Level

14:31 08/12 2009

BMDL BMD

 

Gamma Multi-Hit

 
B-54 DRAFT – DO NOT CITE OR QUOTE 

Weibull Model 

==================================================================== 

 Weibull Model using Weibull Model (Version: 2.12; Date: 05/16/2008) 

 Input Data File: C:\USEPA\BMDS2\Temp\tmp4D9.(d) 

 Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmp4D9.plt 

 Wed Aug 12 14:35:51 2009 

 ==================================================================== 

 BMDS Model Run - Gorzinski et al (1985) - Male rats - Hypertrophy/Dilation of 

Proximal Tubules - Weibull Model using Weibull Model (Version: 2.12; Date: 

05/16/2008) 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the probability function is: 

 P[response] = background + (1-background)*[1-EXP(-slope*dose^power)] 

 Dependent variable = Effect 

 Independent variable = DOSE 

 Power parameter is restricted as power >=1 

 Total number of observations = 4 

 Total number of records with missing values = 0 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 Default Initial (and Specified) Parameter Values 

 Background = 0.0454545 

 Slope = 0.0491052 

 Power = 1 

 Asymptotic Correlation Matrix of Parameter Estimates 

 ( *** The model parameter(s) -Background -Power 

 have been estimated at a boundary point, or have been specified by 

the user, 

 and do not appear in the correlation matrix ) 

 Slope 

 Slope 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 Background 0 NA 

 Slope 0.086025 0.0295231 0.0281608 

0.143889 

 Power 1 NA 

NA - Indicates that this parameter has hit a bound 

 implied by some inequality constraint and thus

 has no standard error. 
B-55 DRAFT – DO NOT CITE OR QUOTE 

 Analysis of Deviance Table 

 Model Log(likelihood) # Param's Deviance Test d.f. P-value 

 Full model -9.35947 4 

 Fitted model -9.44226 1 0.165576 3 0.9829 

 Reduced model -27.5256 1 36.3322 3 <.0001 

 AIC: 20.8845 

 Goodness of Fit 

 Scaled 

 Dose Est._Prob. Expected Observed Size Residual 

 ------------------------------------------------------------------------ 

 0.0000 0.0000 0.000 0.000 10 0.000 

 1.0000 0.0824 0.824 1.000 10 0.202 

 15.0000 0.7248 7.248 7.000 10 -0.176 

 62.0000 0.9952 9.952 10.000 10 0.220 

 Chi^2 = 0.12 d.f. = 3 P-value = 0.9893

 Benchmark Dose Computation 

Specified effect = 0.1 

Risk Type = Extra risk 

Confidence level = 0.95 

 BMD = 1.22477 

 BMDL = 0.710032 

 0

 0.2

 0.4

 0.6

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 1

 0 10 20 30 40 50 60

Fraction Affected

dose

Weibull Model with 0.95 Confidence Level

14:35 08/12 2009

BMDL BMD

 

Weibull
B-56 DRAFT – DO NOT CITE OR QUOTE 

Quantal-linear Model

 ==================================================================== 

 antal-linear Model using Weibull Model (Version: 2.12; Date: 05/16/2008) 

 Input Data File: C:\USEPA\BMDS2\Temp\tmp4DA.(d) 

 Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmp4DA.plt 

 Wed Aug 12 14:37:26 2009 

 ==================================================================== 

 BMDS Model Run - Gorzinski et al (1985) - Male rats - Hypertrophy/Dilation Proximal 

Tubules - Quantal-linear Model using Weibull Model (Version: 2.12; Date: 05/16/2008) 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the probability function is: 

 P[response] = background + (1-background)*[1-EXP(-slope*dose^power)] 

 Dependent variable = Effect 

 Independent variable = DOSE 

 Power parameter is set to 1 

 Total number of observations = 4 

 Total number of records with missing values = 0 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 Default Initial (and Specified) Parameter Values 

 Background = 0.0454545 

 Slope = 0.0491052 

 Power = 1 Specified 

 Asymptotic Correlation Matrix of Parameter Estimates 

 ( *** The model parameter(s) -Background -Power 

 have been estimated at a boundary point, or have been specified by 

the user, 

 and do not appear in the correlation matrix ) 

 Slope 

 Slope 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 Background 0 NA 

 Slope 0.0860249 0.029523 0.0281608 

0.143889 

NA - Indicates that this parameter has hit a bound 

 implied by some inequality constraint and thus

 has no standard error. 

 Analysis of Deviance Table 
B-57 DRAFT – DO NOT CITE OR QUOTE 

 Model Log(likelihood) # Param's Deviance Test d.f. P-value 

 Full model -9.35947 4 

 Fitted model -9.44226 1 0.165576 3 0.9829 

 Reduced model -27.5256 1 36.3322 3 <.0001 

 AIC: 20.8845 

 Goodness of Fit 

 Scaled 

 Dose Est._Prob. Expected Observed Size Residual 

 ------------------------------------------------------------------------ 

 0.0000 0.0000 0.000 0.000 10 0.000 

 1.0000 0.0824 0.824 1.000 10 0.202 

 15.0000 0.7248 7.248 7.000 10 -0.176 

 62.0000 0.9952 9.952 10.000 10 0.220 

 Chi^2 = 0.12 d.f. = 3 P-value = 0.9893

 Benchmark Dose Computation 

Specified effect = 0.1 

Risk Type = Extra risk 

Confidence level = 0.95 

 BMD = 1.22477 

 BMDL = 0.710032 

 0

 0.2

 0.4

 0.6

 0.8

 1

 0 10 20 30 40 50 60

Fraction Affected

dose

Weibull Model with 0.95 Confidence Level

14:37 08/12 2009

BMDL BMD

 

Weibull
B-58 DRAFT – DO NOT CITE OR QUOTE 

NTP (1989) Female Rat Hepatocellular Necrosis

Gamma Model

 ==================================================================== 

 Gamma Model. (Version: 2.13; Date: 05/16/2008) 

 Input Data File: C:\USEPA\BMDS2\Temp\tmpB62.(d) 

 Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpB62.plt 

 Thu Apr 09 09:14:08 2009 

 ==================================================================== 

 BMDS Model Run NTP 1989 Hepatocellular Necrosis Female Rat - Gamma Model 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the probability function is: 

 P[response]= background+(1-background)*CumGamma[slope*dose,power], 

 where CumGamma(.) is the cummulative Gamma distribution function 

 Dependent variable = PercentPositiveHepatocellularNecrosis 

 Independent variable = rosis 

 Power parameter is restricted as power >=1 

 Total number of observations = 6 

 Total number of records with missing values = 0 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 Default Initial (and Specified) Parameter Values 

 Background = 0.0454545 

 Slope = 0.00743289 

 Power = 2.82109 

 Asymptotic Correlation Matrix of Parameter Estimates 

 ( *** The model parameter(s) -Background 

 have been estimated at a boundary point, or have been specified by 

the user, 

 and do not appear in the correlation matrix ) 

 Slope Power 

 Slope 1 0.95 

 Power 0.95 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 Background 0 NA 

 Slope 0.00723384 0.00398244 -0.000571608 

0.0150393 

 Power 2.58447 1.14213 0.345944 

4.823 

NA - Indicates that this parameter has hit a bound 

 implied by some inequality constraint and thus
B-59 DRAFT – DO NOT CITE OR QUOTE 

 has no standard error. 

 Analysis of Deviance Table 

 Model Log(likelihood) # Param's Deviance Test d.f. P-value 

 Full model -16.7382 6 

 Fitted model -17.3091 2 1.14186 4 0.8876 

 Reduced model -32.5964 1 31.7164 5 <.0001 

 AIC: 38.6182 

 Goodness of Fit 

 Scaled 

 Dose Est._Prob. Expected Observed Size Residual 

 ------------------------------------------------------------------------ 

 0.0000 0.0000 0.000 0.000 10 0.000 

 33.5000 0.0059 0.059 0.000 10 -0.244 

 67.1000 0.0300 0.300 0.000 10 -0.556 

 134.3000 0.1289 1.289 2.000 10 0.671 

 267.8000 0.4095 4.095 4.000 10 -0.061 

 535.7000 0.8159 8.159 8.000 10 -0.130 

 Chi^2 = 0.84 d.f. = 4 P-value = 0.9331

 Benchmark Dose Computation 

Specified effect = 0.1 

Risk Type = Extra risk 

Confidence level = 0.95 

 BMD = 118.037 

 BMDL = 60.1812 
B-60 DRAFT – DO NOT CITE OR QUOTE 

 0

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 0.8

 1

 0 100 200 300 400 500

Fraction Affected

dose

Gamma Multi-Hit Model with 0.95 Confidence Level

09:14 04/09 2009

BMDL BMD

 

Gamma Multi-Hit
B-61 DRAFT – DO NOT CITE OR QUOTE 

Modeling for Cancer Assessment 

NTP (1989) BMD Modeling of Adenoma/Carcinoma in Male Rats

Multistage 2°Model 

==================================================================== 

 Multistage Cancer Model. (Version: 1.7; Date: 05/16/2008) 

 Input Data File: C:\USEPA\BMDS2\Temp\tmp6E8.(d) 

 Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmp6E8.plt 

 Mon Apr 13 14:38:06 2009 

 ==================================================================== 

 BMDS Model Run NTP 1989 Kidney Adenoma-Carcinoma Male Rat - Multistage Cancer 2 

degree Model 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the probability function is: 

 P[response] = background + (1-background)*[1-EXP( 

 -beta1*dose^1-beta2*dose^2)] 

 The parameter betas are restricted to be positive 

 Dependent variable = PercentAdenomaCarcinoma 

 Independent variable = DOSE 

 Total number of observations = 3 

 Total number of records with missing values = 0 

 Total number of parameters in model = 3 

 Total number of specified parameters = 0 

 Degree of polynomial = 2 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 2.22045e-016 

 Parameter Convergence has been set to: 1.49012e-008 

**** We are sorry but Relative Function and Parameter Convergence   

** ** are currently unavailable in this model. Please keep checking   

** ** the web sight for model updates which will eventually   

** ** incorporate these convergence criterion. Default values used.   

 Default Initial Parameter Values 

 Background = 0.014541 

 Beta(1) = 0 

 Beta(2) = 0.00799069 

 Asymptotic Correlation Matrix of Parameter Estimates 

 ( ** * The model parameter(s) -Beta(1) 

 have been estimated at a boundary point, or have been specified by 

the user, 

 and do not appear in the correlation matrix ) 

 Background Beta(2) 

Background 1 -0.67 

 Beta(2) -0.67 1 
B-62 DRAFT – DO NOT CITE OR QUOTE 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 Background 0.0177261 * * * 

 Beta(1) 0 * * * 

 Beta(2) 0.00751246 * * * 

* - Indicates that this value is not calculated. 

 Analysis of Deviance Table 

 Model Log(likelihood) # Param's Deviance Test d.f. P-value 

 Full model -33.5473 3 

 Fitted model -33.6008 2 0.106829 1 0.7438 

 Reduced model -36.7395 1 6.38433 2 0.04108 

 AIC: 71.2015 

 Goodness of Fit 

 Scaled 

 Dose Est._Prob. Expected Observed Size Residual 

 ----------------------------------------------------------------------- 

i: 1 

 0.0000 0.0177 0.887 1 50 0.129 

i: 2 

 2.0400 0.0481 2.407 2 50 -0.178 

i: 3 

 4.0900 0.1343 6.717 7 50 0.049 

 Chi-square = 0.10 DF= 1 P-value = 0.7510 

 Benchmark Dose Computation 

Specified effect = 0.1 

Risk Type = Extra risk 

Confidence level = 0.95 

 BMD = 3.74496 

 BMDL = 2.45283 

 BMDU = 9.24921 

Taken together, (2.45283, 9.24921) is a 90 % two-sided confidence 

interval for the BMD 

Multistage Cancer Slope Factor = 0.0407692 
B-63 DRAFT – DO NOT CITE OR QUOTE 

 0

 0.05

 0.1

 0.15

 0.2

 0.25

 0.3

 0 0.5 1 1.5 2 2.5 3 3.5 4

Fraction Affected

dose

Multistage Cancer Model with 0.95 Confidence Level

14:38 04/13 2009

BMDL BMD

 

Multistage Cancer

Linear extrapolation
B-64 DRAFT – DO NOT CITE OR QUOTE 

NCI (1978) BMD Modeling of Hepatocellular Carcinoma in Male Mice

Multistage 2°

 ==================================================================== 

 Multistage Cancer Model. (Version: 1.7; Date: 05/16/2008) 

 Input Data File: C:\USEPA\BMDS2\Temp\tmp7B8.(d) 

 Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmp7B8.plt 

 Tue Apr 14 08:30:03 2009 

 ==================================================================== 

 BMDS Model Run NCI 1978 Hepatocellular Carcinoma Male Mice - Multistage Cancer 2 

degree Model 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 The form of the probability function is: 

 P[response] = background + (1-background)*[1-EXP( 

 -beta1*dose^1-beta2*dose^2)] 

 The parameter betas are restricted to be positive 

 Dependent variable = PercentHepatocellularCarcinoma 

 Independent variable = DOSE 

 Total number of observations = 3 

 Total number of records with missing values = 0 

 Total number of parameters in model = 3 

 Total number of specified parameters = 0 

 Degree of polynomial = 2 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 2.22045e-016 

 Parameter Convergence has been set to: 1.49012e-008 

**** We are sorry but Relative Function and Parameter Convergence   

** ** are currently unavailable in this model. Please keep checking   

** ** the web sight for model updates which will eventually   

** ** incorporate these convergence criterion. Default values used.   

 Default Initial Parameter Values 

 Background = 0.141096 

 Beta(1) = 0 

 Beta(2) = 7.77012e-005 

 Asymptotic Correlation Matrix of Parameter Estimates 

 ( ** * The model parameter(s) -Beta(1) 

 have been estimated at a boundary point, or have been specified by 

the user, 

 and do not appear in the correlation matrix ) 

 Background Beta(2) 

Background 1 -0.73 

 Beta(2) -0.73 1 

 Parameter Estimates 
B-65 DRAFT – DO NOT CITE OR QUOTE 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 Background 0.146344 * * * 

 Beta(1) 0 * * * 

 Beta(2) 7.26074e-005 * * * 

* - Indicates that this value is not calculated. 

 Analysis of Deviance Table 

 Model Log(likelihood) # Param's Deviance Test d.f. P-value 

 Full model -71.2862 3 

 Fitted model -71.7199 2 0.867331 1 0.3517 

 Reduced model -80.5752 1 18.5779 2 <.0001 

 AIC: 147.44 

 Goodness of Fit 

 Scaled 

 Dose Est._Prob. Expected Observed Size Residual 

 ------------------------------------------------------------------------ 

 0.0000 0.1463 2.927 3.000 20 0.046 

 53.0500 0.3041 15.206 15.000 50 -0.063 

 103.8800 0.6101 29.892 30.870 49 0.286 

 Chi^2 = 0.09 d.f. = 1 P-value = 0.7666

 Benchmark Dose Computation 

Specified effect = 0.1 

Risk Type = Extra risk 

Confidence level = 0.95 

 BMD = 38.0933 

 BMDL = 13.8018 

 BMDU = 49.5091 

Taken together, (13.8018, 49.5091) is a 90 % two-sided confidence 

interval for the BMD 

Multistage Cancer Slope Factor = 0.00724545 
B-66 DRAFT – DO NOT CITE OR QUOTE 

 0

 0.1

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 0.5

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 0.7

 0.8

 0 20 40 60 80 100

Fraction Affected

dose

Multistage Cancer Model with 0.95 Confidence Level

08:30 04/14 2009

BMDL BMD

 

Multistage Cancer

Linear extrapolation