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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 re view 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 implic ations.

U.S. Environmental Protection Agency Washington, DCDRAFT – DO NOT CITE OR QUOTE ii DISCLAIMER

This document is a preliminary draft for review pur poses only. This information is distributed solely for the purpose of pre-dissemina tion peer review under applicable information quality guidelines. It has not been formally disse minated by EPA. It does not represent and should not be construed to represent any Agency det ermination or policy. Mention of trade names or commercial products does not constitute en dorsement or recommendation for use.DRAFT – DO NOT CITE OR QUOTE iii CONTENTS—TOXICOLOGICAL REVIEW OF HEXACHLOROETHANE (CAS No. 67-72-1)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 ............... .................................................. x i

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 BIO ASSAYS 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 I NHALATION .. 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...............................

................................................... .......................44DRAFT – DO NOT CITE OR QUOTE iv 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 Suppo rting Evidence...........................65 4.7.3. Mode-of-Action Information for Kidney Adeno mas and Carcinomas................67 4.7.3.1. Hypothesized Mode of Action.............. .................................................. 6 7 4.7.3.2. Experimental Support for the Hypothesized Mode of Action ................ 72 4.7.3.3. Other Possible Modes of Action........... .................................................. 8 0 4.7.3.4. Conclusions about the Hypothesized Mode o f Action ........................... 81 4.7.4. Mode-of-Action Information for Hepatocellul ar Carcinomas.............................82 4.7.4.1. Hypothesized Mode of Action..............

................................................... .........82 4.7.5. Mode-of-Action Information for Pheochromocy tomas.......................................84 4.7.5.1. Hypothesized Mode of Action.............. .................................................. 8 4 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 Effe ct—with Rationale and Justification......................................

................................................... ................86 5.1.2. Methods of Analysis—Including Models.......

................................................... ..90 5.1.3. RfD Derivation—Including Application of Unc ertainty 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 Effe ct—with Rationale and Justification......................................

................................................... ................97 5.2.2. Methods of Analysis—Including Models.......

................................................... 100 5.2.3. RfC Derivation—Including Application of Unc ertainty 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 Jus tification ..................................106 5.4.2

. Dose-response Data ................................. ................................................ 106DRAFT – DO NOT CITE OR QUOTE v 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 HA ZARD 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 PU BLIC COMMENTS AND DISPOSITION ...........................

................................................... ...... A-1

APPENDIX B: BENCHMARK DOSE MODELING OUTPUT........ .......................................B-1DRAFT – DO NOT CITE OR QUOTE vi LIST OF TABLES

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

................................................... ...............................3 3-1. HCE , PERC, and pentachloroethane tissue conc entrations 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 tis sues 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 o f the products formed by CYP450 1A2 metabolism of HCE..............................

................................................... ...................14 4-1. Body, kidney and liver weights of rats expose d to HCE in the diet for 16 weeks ................21 4-2. Histopathological results on kidney effects i n rats exposed to HCE in the diet for 16 weeks...........................................

................................................... ...............................21 4-3. Organ weight to body weight ratios for rats e xposed to HCE for 13 weeks ......................... 22 4-4. Incidence and severity of nephrotoxicity in m ale and female rats treated with HCE................................................

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

................................................... ...................25 4-6. Renal tubular hyperplasia and tumor incidence s 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 HC E .................................................. .............28 4-9. Tumor incidences in female rats gavaged with HCE ................................................ ............29 4-10. Incidence of hepatocellular carcinomas in mi ce................................................. .................30 4-11. Summary of HCE effects on pregnant Wistar ra ts 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, rabb its, and guinea pigs ............................... .........36 4-14. Summary of toxicity data from rats exposed t o HCE for 21 days.................................. .....40 4-15. Summary of genotoxicity studies of HCE......

................................................... ..................44 4-16. Number of enzyme-altered foci in rat liver o f the promotion protocol........................... ....50 4-17. In vivo covalent binding of [ 14 C]HCE to DNA, RNA and proteins from rat and
mouse organs .......................................

................................................... ...........................51 4-18. In vitro binding of [ 14 C]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 α 2u -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....................................

................................................... .........................89DRAFT – DO NOT CITE OR QUOTE vii 5-2. Summary of the BMD modeling results for the k idney .............................................. ..........91 5-3. Potential points of departure (PODs) for neph rotoxicity in male rats with applied uncertainty factors (UF) and potential reference va lues ............................................... .....95 5-4. Noncancerous effects observed in animals expo sed to HCE via inhalation.......................... 99 5-5. Summary of incidence data in rodents orally e xposed to HCE for use in cancer dose-response assessment ...........................

................................................... ..................107 5-6. Summary of BMD modeling for oral cancer asses sment of HCE....................................... 109 5-7. Summary of uncertainties in the HCE cancer ri sk assessment...................................... ......111 B-1. Dose-response modeling results using BMDS (Ve rsion 2.0) based on noncancerous kidney and liver effects in rats followin g oral exposure to HCE....................B-1 B-2. Dose-response modeling results using BMDS (Ve rsion 2.0) for BMRs of 1, 5, and 10% based on non-cancerous kidney and liver eff ects 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 wit h applied uncertainty factors and potential reference values for nephrotoxic effects of studies in Table 5-3 ...................................... .... 96DRAFT – DO NOT CITE OR QUOTE viii 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 FEV 1.0 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 H b/g-A animal blood:gas partition coefficient H b/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 LC 50 median lethal concentration LD 50 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 ProgramDRAFT – DO NOT CITE OR QUOTE ix OCT ornithine carbamoyl transferase PBPK physiologically based pharmacokinetic PBTK physiologically based toxicokinetic PCBs polychlorinated biphenyls PCNA proliferating cell nuclear antigen PERC tetrachloroethene, tetrachloroethylene, perchloro ethylene 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 UF A interspecies uncertainty factor UF H intraspecies uncertainty factor UF S subchronic-to-chronic uncertainty factor UF D database deficiencies uncertainty factor U.S. EPA United States Environmental Protection AgencyDRAFT – DO NOT CITE OR QUOTE x FOREWORD

The purpose of this Toxicological Review is to prov ide 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 compr ehensive 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 th e derivation of the reference dose, reference concentration and cancer assessment, wher e applicable, and to characterize the overall confidence in the quantitative and qualitative aspe cts 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 ass essor 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 (20 2) 566-1676 (phone), (202) 566-1749 (fax), or hotline.iris@epa.gov (email address).DRAFT – DO NOT CITE OR QUOTE xi 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 Reprodu ction

National Institute of Environmental Health Sciences Research Triangle Park, NCDRAFT – DO NOT CITE OR QUOTE xii CONTRACTOR SUPPORT James Kim, Ph.D. Sciences International, Inc. Alexandria, VA

REVIEWERS This document has been provided for review to EPA s cientists 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 DevelopmentDRAFT – DO NOT CITE OR QUOTE xiii 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 DevelopmentDRAFT – DO NOT CITE OR QUOTE 1 1. INTRODUCTION

This document presents background information and j ustification 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 an d other exposure durations, and a carcinogenicity assessment. The RfD and RfC, if derived, provide quantitative i nformation 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 inhalat ion RfC (expressed in units of mg/m 3 ) is analogous to the oral RfD, but provides a continuou s 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 (extra respiratory 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), an d subchronic (>30 days up to 10% of lifetime) exposure durations, all of which are deri ved 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 quantita tive 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 estima tes 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 μ g/m 3 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 T echnical Panel Reports that may have been used in the development of this assessment inc lude 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 Concentr ationDRAFT – DO NOT CITE OR QUOTE 2 Issues in Inhalation Toxicity (U.S. EPA, 1994a), Methods for Derivation of Inhalation Reference Concentrations and Application of Inhalation Dosime try (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 A ssessment of Chemical Mixtures (U.S. EPA, 2000c), A Review of the Reference Dose and Reference Concen tration Processes (U.S. EPA, 2002), Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a) , Supplemental Guidance for Assessing Susceptibility from Early-Li fe 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 Exposur es to Children (U.S. EPA, 2006b). The literature search strategy employed for this co mpound was based on the Chemical Abstracts Service Registry Number (CASRN) and at le ast one common name. Any pertinent scientific information submitted by the public to t he IRIS Submission Desk was also considered in the development of this document. The relevant literature was reviewed through February

2010.DRAFT – DO NOT CITE OR QUOTE 3 2. CHEMICAL AND PHYSICAL INFORMATION

Hexachloroethane (HCE) (CASRN 67-72-1) is a haloge nated hydrocarbon consisting of six chlorines attached to an ethane backbone (Figur e 2-1). Synonyms include 1,1,1,2,2,2-hexachloroethane, ethane hexachloride, ethylene hexachloride, perchloroethane, carbon hexachloride, and carbon trichloride (ChemID plus 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; S panggord et al., 1985; Verschueren, 1983; U.S. EPA, 1982, 1979).



### C C C l

C l C l

Cl

C l Cl 

### C C C l

C l C l

Cl

C l 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 hexac hloride, ethylene hexachloride, perchloroethane, carbon hexachloride, carbon trichl oride Molecular weight 236.74 g/mol

Molecular formula C 2 Cl 6 Melting point Sublimes without melting Boiling Point 186.8°C Density 2.091 g/mL at 20°C Water solubility a 50 mg/L at 22°C; 14 mg/L at 25°C Log K ow 3.82 a , 3.34 b , 4.14 c Log K oc 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-m 3 /mol at 20°C Conversion factor 1 ppm = 9.68 mg/m 3 ; 1 mg/m 3 = 0.10 ppm

Sources: a Howard, 1989; b U.S. EPA, 1979; c Hansch et al., 1995

HCE was produced in the United States (U.S.) for co mmercial 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 distribu ted, 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 tetrac hloroethane imports combined were 1.5 million pounds in 1989 and 612,000 pounds in 20 00 (NTP, 2005). HCE production in 1977DRAFT – DO NOT CITE OR QUOTE 4 was between 2 and 20 million pounds; more recent in formation on production of HCE was not located (NTP, 2005; ATSDR, 1997a). HCE is produced by the chlorination of tetrachloroethylene (PERC) in the presence of ferri c chloride at temperatures of 100–140°C (ATSDR, 1997a; U.S. EPA, 1991b; Fishbein, 1979; IAR C, 1979). HCE is primarily used in the military for smoke pots, smoke grenades, and pyrote chnic devices (ACGIH, 2001; ATSDR, 1997; U.S. EPA, 1991b; IARC, 1979). HCE was also i dentified in the headspace of chlorine-bleach-containing household products (Odab asi, 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).DRAFT – DO NOT CITE OR QUOTE 5 3. TOXICOKINETICS

3.1. ABSORPTION There are no studies that have systematically eval uated HCE absorption in humans by the oral or inhalation routes of exposure. However, up take 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 fert ilization (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 H CE 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/cm

2 /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., 198 5; 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 (1 0-28 μ g/mL) were reached at 24 hours after HCE exposure, indicating slow absorption. Jo ndorf et al. (1957) reported that rabbits fed [ 14 C]-radiolabeled HCE at 500 mg/kg excreted only 5% o f the applied radioactivity in urine over a period of 3 days (fecal measurements were not con ducted). During this 3-day period, 14–24% of the applied radioactivity was detected in expire d air, and the remainder was present in the

tissues and intestinal tract. The amount of HCE ab sorbed by the rabbits was not determined; however, based on the amount of radioactivity prese nt in urine and expired air, approximately

19–29% of the HCE was absorbed. Studies in rats an d mice (Mitoma et al., 1985) using [ 14 C]-radiolabeled HCE (500 mg/kg for rats; 1,000 mg/k g 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 ex creta.

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; N olan and Karbowski, 1978; Fowler, 1969) consistently demonstrated that HCE is distributed p rimarily to fat tissue followed by the kidney and to a lesser extent the liver and the blood (Gor zinski et al., 1985; Nolan and Karbowski, 1978). Younglai et al. (2002) evaluated the concentration s of various environmental contaminants in follicular fluid, serum, and semina l plasma of 21 couples undergoing IVF. HCE

was one of the contaminants identified in >50% of f ollicular fluid samples, suggesting post-absorptive distribution to reproductive organs . The average HCE concentration in follicularDRAFT – DO NOT CITE OR QUOTE 6 fluid was 232 ± 27 pg/mL (mean ± standard error [SE]). HCE was not detected in hum an female serum obtained during oocyte retrieval for IVF. Th is 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 fol licular 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 pentobarbit one 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 muscl e. 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 con centration. 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 ( μ g/g) Sheep 1 Sheep 2

Tissue HCE PERC Pentachloroethane HCE PERC Pentachl oroethane 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 calc ulated 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 contro l diet and sacrificed (groups of 3 or 4 rats) 0, 3, 6, 13, 22, and 31 days after this change in expo sure. Samples of fat, liver, kidney, and whole blood were collected for HCE analysis. The time-co urse related tissue HCE concentrations are presented in Table 3-2. The highest tissue concent rations of HCE were in fat, which were 3-fold greater than the concentration in the kidney and ov er 100-fold greater than blood and liver

concentrations. Fat concentrations decreased from 303 ± 50 μ g/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 2.3DRAFT – DO NOT CITE OR QUOTE 7 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 inc luded a toxicity assessment.

Table 3-2. Time course of HCE concentrations in ma le rat tissues after 57 days of dietary exposure to 62 mg/kg-day

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

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

c ND: not detected (detection limit of 0.001 μ g/g)

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

Nolan and Karbowski (1978) also evaluated tissue co ncentrations 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 ti ssue concentrations are presented in Table 3-3. Kidney concentrations of HCE were much higher in ma le rats compared with female rats, particularly at the highest dose (47-fold greater i n males) (Nolan and Karbowski, 1978). Kidney concentrations of HCE proportionately increased wit h 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 consistentl y different for the male and female rats. Consequently, they speculated that the male rats wo uld 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 met abolite). Both sexes exhibited comparable levels (although levels in males were slightly grea ter) of HCE in blood, liver and fat; concentrations in fat were the highest for both sex es. Blood levels of HCE did not correlate well to either the exposure dose or the dose at the majo r 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 μ g/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 ± 0.006 a 0.303 ± 0.156 a 18.5 b 62.45 ± 3.04 a 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 ND c ND c 0.026 ± 0.006 0.125 ± 0.020DRAFT – DO NOT CITE OR QUOTE 8 Table 3-3. HCE concentrations in male and female r at tissues after 110 or 111 days of dietary exposure HCE tissue concentration (n = 3 or 4) (mean ± SD, μ g/g tissue) Dose (mg/kg-day) Blood Liver Kidney Fat Male 0.079 ± 0.057 0.291 ± 0.213 1.356 ± 0.286 3.09 ± 0.33 1 Female 0.067 ± 0.039 0.260 ± 0.035 0.369 ± 0.505 2.59 ± 0.72 Male 0.596 ± 0.653 1.736 ± 1.100 24.33 ± 5.73 37.90 ± 6.10 15 Female 0.162 ± 0.049 0.472 ± 0.204 0.688 ± 0.165 45.27 ± 11.33 Male 0.742 ± 0.111 0.713 ± 0.343 95.12 ± 11.56 176.1 ± 14.5 62 Female 0.613 ± 0.231 0.631 ± 0.262 2.01 ± 0.66 162.1 ± 7.1

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

3.3. METABOLISM In vitro studies using liver microsomes indicated that the major enzymes involved in HCE metabolism are phenobarbital-inducible cytochro me P450 (CYP450) isoforms (Salmon et al., 1985; Town and Leibman, 1984; Nastainczyk et a l., 1982; Nastainczyk et al., 1981; Salmon et al. 1981); however, no specific (phenobarbital-i nducible) isoforms have been identified. The isoforms induced by phenobarbital include those fro m the 2A, 2B, 2C, and 3A subfamilies. One study (Yanagita et al., 1997) found some evidence f or CYP1A2 involvement in the metabolism of HCE, although this was not supported by the resu lts from in vitro studies with 3-methylcholanthrene, an inducer of the CYP450 1 su bfamily (Nastainczyk et al., 1982; Nastainczyk et al., 1981; Van Dyke and Wineman, 197 1). Information regarding the roles of aroclor 1254-inducible enzymes other than 1A2 (incl uding CYP 2A6, 2E1, 2C9, 2C19, 2D6, and 3A4) is not available for HCE.

The metabolism data for HCE are limited because the re 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 shee p. Each of these studies tends to support

limited metabolism for HCE. The data from the in v ivo and in vitro studies support a conclusion that metabolism of HCE is incomplete, with excretio n of unmetabolized HCE in exhaled air and possibly in urine. A variety of intermediary metab olites have also been identified in exhaled air and urine (Fowler, 1969; Jondorf et al., 1957). Fi gure 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 PER C (ATSDR, 1997b), trichloroethylene (TCE; ATSDR, 1997c), and 1,1,2,2-tetrachloroethane (ATSDR, 2008) intermediary metabolites. Mitoma et al. (1985) examined the distribution of H CE in male Osborne-Mendel rats and

male B6C3F 1 mice to evaluate the extent to which radiolabeled compound is metabolized in the 48 hours after administration of 125 or 500 mg/kg t o the rats and 250 or 1,000 mg/kg to the

mice. These doses were selected based on the maxim um tolerated dose (MTD) and 1⁄4 MTD of HCE; the MTD in rats and mice is 500 mg/kg (2.11 mm ol/kg) and 1,000 mg/kg (4.22 mmol/kg),DRAFT – DO NOT CITE OR QUOTE 9 respectively. The animals, 4/dose, were orally adm inistered unlabeled HCE as a solution in corn oil 5 days/week for 4 weeks, followed by a single d ose of [ 14 C]-radiolabeled HCE. The 48 hour observation period began after administration of th e 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 mic e 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 CO 2 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 (CO 2 + excreta + carcass) 28.72 23.95

Source: Mitoma et al. (1985).

Recovery of the radiolabel was >90% for both rats a nd mice. Total metabolism was calculated by the authors as the sum of the radiola bel 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 th e 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 u nmetabolized parent compound. The major urinary metabolites, determined qualitatively by hi gh performance liquid chromatography, were trichloroethanol and trichloroacetic acid (TCA) for both rats and mice. Trichloroethanol and TCA were also qualitatively considered the major ur inary metabolites for other halogenated hydrocarbon compounds, including PERC, that were ev aluated 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 metabolized a 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 calcula tion 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 [ 14 C]-radiolabeled HCE at 500 mg/kg (route of administration not reported by study authors) ex creted only 5% of the applied radioactivity in urine over 3 days (72 hours), indicating slow metab olism. This is consistent with the results in mice and rats reported by Mitoma et al. (1985) in w hich approximately 2-4% of the label was found in urine after 48 hours. During this 3-day p eriod, 14–24% of the radioactivity was detected in expired air (a lower percentage than se en for rats at a comparable dose by Mitoma et al., 1985), and the remainder was present in tissue s and the intestinal tract. However, the authors did not have the capability of quantifying HCE in t issues. Reported urinary metabolites include trichloroethanol (1.3%), dichloroethanol (0.4%), TC A (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 w as not found). Quantitative data on the volatile metabolites in exhaled air were not report ed. The only other metabolite data come from the work o f Fowler (1969) in sheep. HCE was

administered to four Scottish Blackface and six Che viot 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 admi nistration of 500 mg/kg, blood

measurements were 10–28 μ g/mL for HCE, 0.6–1.1 μ g/mL for PERC, and 0.06–0.5 μ g/mL for pentachloroethane. Blood concentrations of HCE, PE RC, and pentachloroethane were 2.3-2.6 times greater than the corresponding concen trations in erythrocytes. Data were not reported for the 750 and 1,000 mg/kg doses. In vit ro experiments using fresh liver slices suspended in an olive oil emulsion confirmed the pr esence 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 studi es (Town and Leibman, 1984; Nastainczyk et al., 1982), and ATSDR (1997a) were used in derivati on 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 pre sents 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 Henschle r (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 Cl H Cl HO

Dichloroethanol

CO 2

One carbon pool Cl O Cl

HO

H Cl O

HO

Monochloroacetic acid Cl Cl C C

Cl Cl C C Cl Cl

Trichloroethanol Trichloroethylene CCl 3 – CCl 2 ● Free radical reactions

C C

Cl Cl C C Cl C l

C C

H Cl O C H Cl C C

Cl Cl H H Cl Cl

C

C

H Cl Cl Cl C C

Cl Cl Cl Cl Hexachloroethane

Pentachloroethane

1,1,2,2- Tetrachloroethane Tetrachloroethylene H Cl

Cl

Cl

2Cl H Cl

2Cl

ClDRAFT – 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 phenoba rbital-inducible CYP450 isoforms that catalyze their reductive dechlorination with NADPH, cytochrome b 5 , and NADH as electron donors. HCE metabolism was measured using liver mi crosomes from male Sprague-Dawley rats that were either pretreated with phenobarbital or 3 -methylcholanthrene, or were not pretreated. Only phenobarbital-induced rat liver microsomes dem onstrated an increase in HCE metabolism (27.0 ± 1.1 nmol/mg protein/minute [mean + standard deviat ion or SD] compared with 8.0 ± 1.2 nmol/mg protein/minute for controls). Oxidatio n 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 CYP 450. Carbon monoxide inhibited the NADPH oxidation rate, further indicating that CYP45 0 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 w as represented by the following equation:

NADPH + H + + Cl 3 C—CCl 3 NADP + + Cl 2 C=CCl 2 + 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. T he 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 r atio of pentachloroethane (one electron reduction) to PERC (two electron reduction) increas ed since transfer of the second electron can occur via cytochrome b 5 , which is influenced by pH. These reaction outcom es were proposed by the authors to also apply to other polyhalogenated hydrocarbons. To provide additional support for the reaction bein g catalyzed by CYP450, Nastainczyk et al. (1982, 1981) inhibited CYP450 using carbon m onoxide, 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 ph enobarbital-induced rat liver microsomes

was inhibited >99% when carbon monoxide was added t o 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 p entachloroethane formation 66 ± 8% andDRAFT – 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 a re involved in HCE metabolism and α -naphthoflavone does not inhibit the primary CYP450 involved in the metabolism of HCE. Since metyrapon e did not completely inhibit HCE metabolism by phenobarbital-induced liver microsome s, the remainder of HCE metabolism may be accounted for by the CYP450 2A and 2B subfamilie s 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 HC E to PERC. The formation of PERC was favored in a low oxygen environment as observed met abolism rates of 50.2 ± 0.45, 1.25 ± 0.25, and 0 nmol/minute/mg protein in atmospheres of N 2 , air, and O 2 , 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 m etabolism of HCE to PERC were located in the microsomes, rather than the cytosol, of phen obarbital-treated rat livers. Formation of malondialdehyde and conjugated dienes was statistic ally, significantly increased following treatment with HCE (8 mM), indicating lipid peroxid ation. The authors suggested the

involvement of a free radical. The K m and V max 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 K m and V max 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, su ggesting 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 Spr ague Dawley rats to quantify the dechlorination of HCE. In this case, dechlorinatio n was measured by the release of radioactive

Cl

– from the [ 36 Cl]-radiolabeled HCE substrate during incubation wi th liver microsomes from

induced rats. The K m and V max were determined as 2.37 mM and 0.91 nmol/minute/mg protein, respectively. A control group of noninduced rats w as not included. Salmon et al. (1985) reported a follow-up study tha t used liver microsomes from noninduced rats (Wistar-derived Alderley Park strai n) and a reconstituted CYP450 system from noninduced and phenobarbital-induced New Zealand Wh ite rabbits. Metabolic experiments of HCE using liver microsomes from noninduced rats yie lded a K m of 6.0 μ M and a V max 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 useDRAFT – DO NOT CITE OR QUOTE 14 of a different rat strain. A reconstituted CYP450 system from phenobarbital-induced New Zealand White rabbits yielded K m and V max values of 50 μ M and 2.39 nmol NADPH/minute/nmol CYP450, respectively (S almon et al., 1985). Microsomes from

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

evidence that the reductive dechlorination of HCE i s 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 chlo rinated 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 nmo l 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 p henobarbital-inducible isoform; the major phenobarbital-inducible CYP450 enzymes are the 2A a nd 2B subfamilies. A follow-up study (Yanagita et al., 1998) that examined NADPH oxidati on rates under anaerobic conditions found that CYP450 1A2 wild type had a V max of 1.3 mol/mol CYP450/minute, a K m of 0.25 mM, and

an NADPH oxidation rate of 1.4 mol/mol CYP450/minut e.

Table 3-6. Product formation rates and relative ra tios 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 com pound to examine the metabolism of three hexachlorocyclohexane isomers. Liver microso mes from male Wistar rats that were induced with phenobarbital converted HCE to PERC an d 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) eva luated 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 oth er chlorinated ethanes found that the optimal configuration for dechlorination was a dich loromethyl group. HCE demonstrated a considerable amount of dechlorination (3.9%) in thi s in vitro study; however, the authors determined that HCE was unstable in aqueous solutio n and that this dechlorination was nonenzymatic based on the evidence of dechlorinatio n 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 an d a physiologically based pharmacokinetic (PBPK) inhalation model described by Ramsey and And ersen (1984) for styrene. The V maxDRAFT – DO NOT CITE OR QUOTE 15 (scaled to a 1.0 kg rat) was 1.97 ± 0.05 mg/hour, or 8.3 μ mol/hour. The K m was 0.80 mg/L, or

3.38 μ M.

3.4. ELIMINATION

No studies are available that evaluated the elimin ation of HCE in humans. Animal studies indicated that the major routes of HCE elim ination 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 unkno wn why there is a discrepancy between the studies in sheep and rodents. Rabbits fed [ 14 C]-radiolabeled HCE at 0.5 g/kg (Jondorf et al., 19 57) 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 measuremen ts were not conducted. Fowler (1969) orally administered HCE to Scottish B lackface 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 ov er a period of 4 days for HCE analysis.

More than 80% of the total fecal excretion of HCE o ccurred 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 anaest hetized with pentobarbitone sodium. The hepatic duct was cannulated to collect bile; HCE wa s 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 de tected 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 B6C3F 1 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 fo r 48 hours following administration of a tracer dose of [ 14 C]-HCE. The findings are presented in Table 3-4. Most of the radiolabel was detected in expired air, indicating this to be a ma jor 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 CO 2 , 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, respectiv ely.DRAFT – DO NOT CITE OR QUOTE 16 3.5. PHYSIOLOGICALLY BASED TOXICOKINETIC MODELS No physiologically based toxicokinetic (PBTK) mode ls for HCE have been developed specifically for mammalian species. Models for wat erborne chloroethanes have been reported for rainbow trout and channel catfish; however, the se 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 an d a PBPK inhalation model described by Ramsey and Andersen (1984) for styrene. These repo rts by Gargas and Andersen (1989) and Gargas et al. (1988) do not describe a PBTK model f or HCE, only kinetic constants for metabolism by inhalation. During these breath cham ber experiments, fur deposition (fur loading) was observed to occur. At an exposure con centration of 53.3 ppm HCE at 6 hours, the chemical mass in body tissues was 7.29 mg and the c hemical 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 tox icology of HCE in humans. Case reports of pneumonitis (Allen et al., 1992) and pne umonitis with evidence of liver abnormalities (Loh et al., 2008, 2006) have been described in sol diers 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 HC E in tablet or powder form, resulting in exposures to fumes containing hexachlorobenzene, oc tachlorostyrene, dioxins, dibenzofurans and other organochlorinated compounds. A case repo rt of a hepatocellular carcinoma (Seldén et al., 1989), and limited data concerning some clinic al 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 (Ax elson, 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 prima rily to liver and pulmonary function (Seldén et al., 1994, 1993). Two separate studies were conducted on a small pop ulation of Swedish workers occupationally exposed to HCE while producing milit ary white smoke munitions. The first study reported on biological exposure monitoring (S eldén et al., 1993) and the second study described health effects resulting from HCE exposur e (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 tim e this study was conducted in 1989, no HCE dust was found in the air sample filters, but the integr ated results of personal and stationary charcoal tube samples revealed approximate HCE concentration s by location of 10–30 mg/m 3 (milling/mixing), 5-25 mg/m 3 (pressing), <5 mg/m 3 (assembly room), and nondetectable (storage room) (Seldén et al., 1993). In the first study (Seldén et al., 1993), the expo sed 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) consis ted of assembly line workers from the same company who were unexposed to chlorinated hydrocarb ons, but had some exposure to glass fiber dust. They were matched to the exposure group by s ex and age (± 5 years), except in the case of one exposed male subject where only a younger contr ol could be found. This latter-exposed male subject was excluded from the analysis of heal th effects (Seldén et al., 1994). A secondDRAFT – DO NOT CITE OR QUOTE 18 control group of formerly HCE-exposed workers (3 ma les, 10 females; age range, 31–57 years; mean, 43.6 years) was used in the biological exposu re monitoring study. Blood samples were collected for analysis of HCE c oncentration. For the exposed group, samples were drawn 5 weeks into a temporary product ion break (the “baseline” period), and the second samples were drawn 5 months later, after pro duction had been under way for 5 weeks (the “production” period). Analyses of blood plasm a HCE indicated that values for both control groups (n = 25) were all below the limit of detecti on (<0.02 μ g/L). Exposed subjects were stratified into three subgrou ps (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 concent rations in 10 of the samples from exposed workers were in the range of <0.02–0.06 μ g/L, one sample was 0.15 μ g/L, and one was

0.52 μ g/L. The last sample was from an individual who ha d 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 μ g/L compared with

0.08 ± 0.14 μ g/L in the production and baseline samples, respect ively, p < 0.01). Although the magnitude of individual increases varied considerab ly, there was a significant ( p < 0.05) linear trend for values in the low, medium, and high expos ure subgroups (means of 3.99, 7.14, and

10.75 μ g/L, respectively). These results demonstrate that a considerable increase in plasma HCE can occur after a relatively brief occupational exp osure, 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 t he second health effects study (Seldén et al., 1994). Data pertaining to 15 clinical symptoms (in cluding headaches, sleep quality, palpations, difficulty concentrating, tension/restlessness, fre quency of coughing, watery eyes/runny nose, itching/other skin problems, shortness of breath/ch est discomfort, general health) were obtained from self-administered questionnaires for the expos ed 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 grou p in the analysis of the symptom data. Whole blood and serum samples from the 11 exposed a nd 11 matched company controls were analyzed for routine clinical parameters. Spot uri ne samples were analyzed for hemoglobin, protein, and glucose. Lung function was assessed b y measuring vital capacity and 1-second forced expiratory volume (FEV 1 ). The matched company controls reported more symptoms of ill health than exposed subjects, although the differences were not statist ically significant. Although not statistically significant, the exposed group reported a higher pr evalence 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 ski n problems” (3/11, 27%) than the historical controls (16/130, 12%). The prevalence of “itching /other skin problems” in the matched controlsDRAFT – DO NOT CITE OR QUOTE 19 (3/11, 27%) was the same as that in the exposed gro up. These symptoms centered on the wrist and neck areas, and the authors suggest that this c ould reflect exposure to HCE through joints in the protective equipment, or possibly a “traumitera tive effect of the equipment itself.” Clinical examination revealed no dermatological or respirato ry mucous membrane abnormalities in either group. The authors noted that a previous unpublish ed study of the plant workers (but with primarily different workers) had also found dermato logic 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 FEV 1 . Exposed subjects had significantly higher levels of serum creatinine, ur ate, and bilirubin than controls ( p < 0.05), although the group means were still in the normal r ange. One exposed subject had a marginally elevated level of serum alanine aminotransferase (A LT) (70.5 U/L versus ≤ 41.1 U/L reference), while one control subject displayed increased level s of serum ALT and aspartate aminotransferase (AST) (67.6 and 186.4 U/L, respect ively; 41.1 U/L reference for each). The

control individual’s values returned to normal afte r 8 months, while the exposed subject’s serum ALT value worsened to 87.6 U/L 4 months later (Seld én et al., 1994). Available data pertaining to these liver function tests from 1982, when expos ure 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 o f small differences in clinical parameters, within the normal range, is uncertain. Based on th e available data, the possible dermatologic/mucosal effects and hepatic effects ar e the areas in most need of additional

research.

4.2. SUBCHRONIC AND CHRONIC STUDIES AND CANCER BIO ASSAYS IN

ANIMALS—ORAL AND INHALATION

4.2.1. Oral 4.2.1.1. Subchronic Exposure Two subchronic toxicity assays for HCE were reporte d (NTP, 1989; Gorzinski et al., 1985, 1980). The Gorzinski et al. (1985, 1980) stu dy (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 rat s in this study exhibited a dose-response increase in the incidence of hepatocellular necrosi s of the centrilobular area. The NTP (1989)DRAFT – DO NOT CITE OR QUOTE 20 study suggested that male rats may be more suscepti ble to kidney effects whereas female rats may be more susceptible to liver effects. Gorzinski et al. (1980) conducted a 16-week toxicit y study in male and female F344 rats. Ten rats/sex/dose were exposed via the diet, formul ated 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 D evelopment (R&D) Report by Dow Chemical and is not publicly available. The data for this s tudy were published in the peer-reviewed literature by Gorzinski et al. (1985) and are prese nted 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 S ection 3.2, HCE concentrations in male kidneys were proportionately increased with adminis tered dose while the increases in females were not proportionate. At the high dose, male rat s displayed statistically significant increases in absolute and relative kidney weights accompanied by macroscopically observed alterations. Male rats displayed slight hypertrophy and/or dilat ion of proximal convoluted tubules of the kidneys at incidences of 0/10, 1/10, 7/10, and 10/1 0 for the 0, 1, 15, and 62 mg/kg-day dose groups, respectively. The increased incidence of s light 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 tu bules at incidences of 1/10, 2/10, 7/10, and 10/10 for the 0, 1, 15, and 62 mg/kg-day dose group s, respectively. The increased incidence of atrophy and degeneration of renal tubules was stati stically significant in males at the 15 and 62 mg/kg-day doses. Female rats did not display hy pertrophy 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 th e 62 mg/kg-day dose. Male rats of the 62 mg/kg-day group exhibited statistically signific ant 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 i ncrease in relative liver weight at the high dose, there was no evidence of hepatotoxicity in the hist opathological examination. The data for liver and kidney weights are presented in Table 4-1 and t he 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 Male a

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

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 Female a

62 170.8 ± 5.1 4.71 ± 0.23 2.76 ± 0.10 b 1.39 ± 0.05 0.81 ± 0.02

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

b Statistically significant from control using Dunnet t’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 weeks a

Dose (mg/kg-day)

Organ Effect Sex 0 1 15 62 Male 0 1 7 c 10 c Slight hypertrophy and/or dilation of proximal

convoluted tubules Female 0 0 0 0 Male 1 2 7 c 10 c Kidney Atrophy and degeneration of renal tubules b Female 1 1 2 6 c

a Data are presented as number of positive observatio ns for 10 rats/sex/dose.

b Graded as slight in 1 of 10 male control rats and v ery slight in 1 of 10 control female rats. Severit y of nephropathy was not reported for HCE-exposed rats.

c EPA determined statistical significance from contro l using Fisher’s Exact Test ( p = 0.05).

Source: Gorzinski et al. (1985).

The authors concluded that the no-observed-effect l evel 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-adv erse-effect level (LOAEL), based on renal tubule toxicity in male rats. For female rat s, EPA considered the NOAEL as 15 mg/kg-day and the LOAEL as 62 mg/kg-day, based on renal tubul e toxicity. NTP (1989) conducted a 13-week study of HCE oral t oxicity 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-weighte d 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 andDRAFT – DO NOT CITE OR QUOTE 22 4%, respectively, compared with controls. Statisti cally 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 Male a 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 ± 0.62 c 42.5 ± 0.74 c 46.3 ± 0.95 c Brain 6.0 ± 0.30 5.7 ± 0.17 5.7 ± 0.10 5.8 ± 0.23 6.3 ± 0.21 7.2 ± 0.31 c Heart 2.8 ± 0.04 2.8 ± 0.04 2.9 ± 0.07 3.2 ± 0.17 b 3.3 ± 0.18 c 3.2 ± 0.10 b Kidney 3.0 ± 0.05 3.8 ± 0.37 4.1 ± 0.27 b 4.7 ± 0.44 c 5.2 ± 0.35 c 4.7 ± 0.28 c 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 ± 0.21 c 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 Female a 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 ± 0.39 b 36.3 ± 0.44 c 42.0 ± 0.60 c 52.4 ± 0.88 c Brain 8.7 ± 0.17 8.6 ± 0.14 8.6 ± 0.10 9.0 ± 0.14 9.0 ± 0.15 9.5 ± 0.17 c Heart 2.9 ± 0.04 3.0 ± 0.05 3.0 ± 0.03 3.0 ± 0.04 3.1 ± 0.07 3.4 ± 0.07 c Kidney 3.1 ± 0.04 3.2 ± 0.05 3.2 ± 0.07 3.2 ± 0.06 3.6 ± 0.05 c 4.1 ± 0.10 c 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 ± 0.05 c

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 fo rmation, tubular regeneration, and tubular casts), similar to the toxicity noted in th e 16-day study also conducted by NTP (1989), were observed in 90% of 34 mg/kg-day males and in m ales 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. A t the 536 mg/kg-day dose, 5/10 males died. Kidneys from these five animals were examined micro scopically and revealed papillary necrosis, degeneration, and necrosis of the renal tubular epi thelium. 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 2/1 0 females at the 134 mg/kg-day dose. Additionally, males of the 536 mg/kg-day dose group exhibited hemorrhagic necrosis of theDRAFT – DO NOT CITE OR QUOTE 23 urinary bladder. EPA considered the female rat NOA EL 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 m ore 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 r ats and one in mice. Increased incidences of renal tubular hyperplasia, renal adenoma or carcino ma, 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 prov ide evidence of carcinogenicity. B6C3F 1 mice of both sexes exhibited hepatocellular carcinomas, although only male mice demonstrated a dose response with tumor incidence (NCI, 1978). Based o n 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/carcinogen icity 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 con tinuous 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/wee k for 103 weeks (TWA doses of 0, 57, or 114 mg/kg-day, respectively, after adjusting for co ntinuous exposure). These sex-specific doses

were selected based on the results of the 13-week s tudy conducted by NTP (1989) that demonstrated kidney lesions in male rats at the low er 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 wer e 5–6% lower than controls after week 81. Mean body weights of the 114 mg/kg-day fe male rats were 5–9% lower between weeks 41 and 101. Nephropathy, characterized by tu bular cell degeneration and regeneration, tubular dilatation and atrophy, glomerulosclerosis, interstitial fibrosis, and chronic inflammation,

was observed in both treated and control rats. Inc idences 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 i ncreased 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, respect ively), 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 c ontrol in the high-dose group, anDRAFT – DO NOT CITE OR QUOTE 24 examination of the various grades of severity revea led 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 consi dered 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 e ffects. 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 ra ts (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 e pithelium 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 inflam mation (4%), and focal necrosis (2%) were

observed in some of the male rats administered 14 m g/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 administer ed 14 mg/kg-day HCE. Incidences of female nephropathy were 22/50 for con trols, 42/50 in the 57 mg/kg-day dose group, and 44/49 in the 114 mg/kg-day dose gro up. The severity scores for nephropathy in female rats were statistically significantly increa sed 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 severi ty shows mild and moderate nephropathy in treated females compared with controls which predom inantly presented less than minimally severe nephropathy (Table 4-4). Similar to the male rats, the incidence of nephropa thy 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 s everity) were 12/50, 25/50, and 32/50 in the control, 57, and 114 mg/kg-day dose groups, respect ively. Additional kidney effects in female rats included l inear mineralization of the renal papillae, although the incidence was not dose-depen dent: 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 i nflammation (2% at both 57 and 114 mg/kg-day). Pigmentation of the renal tubule w as 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 gro up and 2% of the 114 mg/kg-day dose group. Only male rats demonstrated an increase in hyperpla sia of the pelvic transitional epithelium and a dose-dependent increase in incidences of minerali zation along the renal papillae.DRAFT – DO NOT CITE OR QUOTE 25 Table 4-4. Incidence and severity of nephropathy i n 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 42 b 45 b

Total # of rats 50 50 50 50 50 49 Overall severity c 2.34 ± 0.14 2.62 ± 0.15 2.68 ± 0.16 a 0.72 ± 0.13 1.38 ± 0.11 b 1.69 ± 0.12 b

a Authors reported as statistically significantly dif ferent from controls, p < 0.05.

b Authors reported as statistically significantly dif ferent from controls, p < 0.01.

c Mean ± SE.

Source: NTP (1989).

Table 4-5. Additional kidney effects in HCE-treate d 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

a EPA determined statistical significance using Fishe r’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), hyperpl asia of the pelvic transitional epithelium (Table 4-5), increased incidence of renal tubule pi gmentation (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 increa sed incidence in treated male rats: 4/50 (8%) in the 7 mg/kg-day dose and 11/50 (22%; s tatistically 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 re nal hyperplasia. Dose-related increases in

the incidence of combined renal adenomas and carcin omas were observed in males rats administered HCE at doses of 7 (4%) and 14 mg/kg-da y (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 evide nce of carcinogenicity in male rats based on a comparison with the historical controls in the stud y 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 inc idences 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 nep hropathy than female rats. Male, but not female rats, also exhibited renal hyperplasia and t umors. 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 r ats administered 7 and 14 mg/kg-day HCE, respectively, compared with 12% of controls. Femal e rats in the control (10%) and 114 mg/kgday-treated (15%) groups exhibited hyperplasia of t he adrenal medulla; this effect was not observed in the 57 mg/kg-day dose group. Adrenal medullary lesions were observed in male rat s, but not female rats (Table 4-7). Pheochromocytoma incidences were statistically sign ificantly 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 (1 4/50, 28%). There were no statistically significant differences in the incidences of malign ant pheochromocytomas and complex pheochromocytomas (defined as pheochromocytomas con taining nervous tissue in addition to the typical adrenal medullary cells) between contro ls and treated male rats. The combinedDRAFT – DO NOT CITE OR QUOTE 27 incidence of all three types of pheochromocytomas w as 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 p ossibly treatment-related.

Table 4-7. Adrenal medullary lesions in HCE-treate d 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 chron ic 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 u ntreated and vehicle controls. Starting in week 23, rats treated began a 5-week cyclic rotatio n that involved 1 week without exposure followed by dosing for 4 weeks. After adjustment f rom 5 days/week for 78 weeks, with the 5-week cyclic rotation for part of the time, to con tinuous 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%), resp ectively, compared with 14/20 (70%) in the untreated controls and 11/20 (55%) in vehicle contr ols (seven rats in the vehicle control group were sacrificed in week 60). Mortality in the trea ted groups occurred early in the bioassay. Approximately 20% of the high- and low-dose males d ied 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, respecti vely. 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 we eks 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 un treated controls, 14/20 (70%) in vehicle controls, 32/49 (65%) in the 113 mg/kg-day dose gro up, and 25/50 (50%) in the 227 mg/kg-day dose group; female rats exhibited incidences of 8/2 0 (40%) in untreated controls, 4/20 (20%) inDRAFT – 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 (cha racterized by degeneration, necrosis, and the presence of large hyperchromatic regenerative e pithelial cells) was observed in 45 and 66%

of males and 18 and 59% of females in the 113 and 2 27 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 NOAE L could not be identified. Tumor types exhibited by male rats surviving at lea st 52 weeks included kidney tubular cell adenoma, pituitary chromophobe adenoma, thyroi d 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 c ell adenoma (4/37) and pituitary chromophobe adenoma (4/32) were observed in the mal e rats of the 113 mg/kg-day dose group but not in the 227 mg/kg-day group. Male vehicle c ontrols did not exhibit kidney tubular cell adenomas, although 11% (2/18) exhibited pituitary c hromophobe adenomas. Thyroid follicular cell adenoma or carcinoma were observed in 11, 8, a nd 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 (1 11 weeks) and low-dose males (92 weeks).

Testicular interstitial cell tumors were not observ ed in vehicle control or 113 mg/kg-day males, but were observed in 10% of 227 mg/kg-day males.

Table 4-8. Tumor incidences a 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 kidne y hamartoma (nonneoplastic overgrowth), pituitary chromophobe adenoma, thyroid follicular cell adenoma or carcinoma, mammary gland fibroadenoma, and ovary granulose cel l 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 o r 113 mg/kg-day female rats. The increasedDRAFT – 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, thyroi d follicular cell adenoma or carcinomas, and mammary gland fibroadenomas was lower in HCE-tr eated animals than in controls. Ovary granulose cell tumors were increased in the low dos e 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 spon taneous 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 or al study in 50 B6C3F 1 mice/sex/dose administered 0, 500, or 1,000 mg/kg-d ay HCE (purity >98%) via corn oil gavage for 5 days/week for 78 weeks. Following exposure t ermination, animals were observed for

12-13 weeks for a total duration of 90–91 weeks. T wenty mice/sex were included as untreated and vehicle controls. Starting in week 9, the dose s 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 mal es. 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 unt reated control groups and 80 and 68% in the 360 and 722 mg/kg-day dose groups, respectively. A s a result of the low survival rates in the vehicle and untreated male control groups, NCI comp ared tumor incidences in the dosed males and females to the pooled vehicle control data deri ved from concurrently run bioassays for

several other chemicals. Animals were all of the s ame 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 vehic le controls, low dose, and high dose, respectively. Female mice in the pooled vehicle co ntrol group (15%) and 722 mg/kg-day (2%), but not the untreated control and 360 mg/kg-day dos e groups, exhibited chronic kidney inflammation. Tubular nephropathy (characterized b y degeneration of convoluted tubule epithelium at the junction of the cortex and medull a, 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 non neoplastic effects were observed. EPA considered 360 mg/kg-day as the LOAEL for this stud y based on tubular nephropathy. EPA

considered that a NOAEL was not established. Increases in the incidence of hepatocellular carci nomas were observed in male and female mice exposed to HCE (Table 4-10). Hepatocel lular adenomas were not noted in the report. NCI (1978) reported statistically signific ant increases in the incidence of hepatocellular carcinomas in 30 and 63% of 360 and 722 mg/kg-day m ales, compared with 10 and 15% of pooled vehicle and matched vehicle controls, respec tively. 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 high er 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 B6C3F 1 mice (1978).

Table 4-10. Incidence of hepatocellular carcinomas in mice

Pooled vehicle control a 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

a As a result of the exceptionally low survival rates in the vehicle and untreated control groups, NCI u sed the pooled vehicle control data derived from concurrently run bioassays for several other chemicals. Animals wer e all of the same strain and housed in the same room. Incidence s 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 (LC 50 ) 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. (19 79) 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 neurotoxi city, reduced body weight gain, increased organ weights, and some evidence of respiratory tra ct irritation. Weeks et al. (1979) exposed Sprague-Dawley rats (2 5/sex/concentration) to control air, 15, 48, or 260 ppm HCE (145, 465, and 2,517 mg/m 3 , 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/m 3 group, body weight gain of male, but not the nonpre gnant female, rats was reduced beginning in the third week of exposure (although quantitative i nformation was not reported). All rats in the 2,517 mg/m 3 group exhibited tremors, ruffled pelt, and red exud ates 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/m 3 group had a higher incidence and severity of mycoplasma-related lesions in nasal turbinates, tra chea, and lung compared with controls. The authors concluded these lesions were related to pot entiation of an endemic mycoplasia infection rather than a direct effect of HCE exposure. Howev er, 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 week s postexposure. No treatment-related effects were observed in the rats exposed to 145 and 465 mg /m 3 HCE. In the oxygen consumption test, male rats (5/concen tration) were tested prior to and following exposure to 145, 465, or 2,517 mg/m 3 HCE for 15 minutes, 3 days/week for the duration of the study (6 weeks). The 2,517 mg/m 3 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 non specific, is indicative of an alteration in basal metabolic rate. No histopathological effects were observed at this concentration. EPA considered 465 mg/m 3 the NOAEL and 2,517 mg/m 3 the LOAEL, based on reduced body weight gain, and increased organ weights. Weeks et al. (1979) also exposed male Sprague-Dawle y rats (15/concentration) exposed to 15, 48, or 260 ppm HCE (145, 465, or 2,517 mg/m 3 ) for 6 hours/day, 5 days/week for 6 weeks and examined them for behavioral changes related to learned and unlearned responses (describedDRAFT – DO NOT CITE OR QUOTE 32 in detail in Section 4.4.3.2). Similar to the othe r treated rats, body weight gain was reduced. Final mean body weight gain in male rats was reduce d 2, 5, and 10% (statistically significant) in the 145, 465, and 2,517 mg/m 3 dose groups, respectively, compared with controls. Additionally, relative lung, liver, kidney, and testes weights we re increased (quantitative information not reported) compared with controls. Weeks et al. (1979) also exposed four male beagle d ogs/concentration to control air, 15, 48, or 260 ppm HCE (145, 465, and 2,517 mg/m 3 , 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 paramete rs. In addition, the beagle dogs underwent pulmonary function tests prior to and following exp osure. One dog died within 5 hours of exposure to 2,517 mg/m 3 . The remaining animals in the 2,517 mg/m 3 group exhibited signs of neurotoxicity consisting of tremors, ataxia, hypers alivation, head bobbing, and facial fasciculations. No blood parameters were significa ntly affected and no exposure-related histopathological lesions were observed following n ecropsy on dogs sacrificed 12 weeks postexposture. Dogs evaluated for pulmonary functi ons 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/m 3 the NOAEL and 2,517 mg/m 3 the LOAEL,

based on neurotoxic effects. Weeks et al. (1979) also exposed male Hartley guine a pigs (10/concentration) to control air, 15, 48, or 260 ppm HCE (145, 465, and 2,517 mg /m 3 , 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 p otential following inhalation exposure to HCE. Two guinea pigs died during each of the fourt h and fifth weeks, resulting in four total deaths. Guinea pigs of the 2,517 mg/m 3 group displayed reductions in body weight beginnin g at the second week of exposure and significantly incre ased liver to body weight ratios (quantitative information was not reported). No treatment-relate d effects were observed in the other exposure groups. EPA considered the NOAEL as 465 mg/m 3 and the LOAEL as 2,517 mg/m 3 , based on decreased body weight and significantly increased r elative liver weight. Weeks et al. (1979) also exposed male and female qu ail ( C. japonica, 20/concentration) to control air, 15, 48, or 260 ppm HCE (145, 465, a nd 2,517 mg/m 3 , respectively; purity 99.8%) for 6 hours/day, 5 days/week for 6 weeks. Postexpo sure observation was carried out for 12 weeks. The only effects observed were in 2 of 10 q uail in the 2,517 mg/m 3 group exhibited

excess mucus in nasal turbinates after 6 weeks. Th e 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 transie nt 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 aut hors. EPA considered 2,517 mg/m 3 (highestDRAFT – DO NOT CITE OR QUOTE 33 exposure concentration) as the NOAEL, while the LOA EL 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 I NHALATION

4.3.1. Oral Weeks et al. (1979) exposed 22 pregnant Sprague-Daw ley 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 2 50 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 exudat es 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 contr ols. The aspirin-positive control group produced fetuse s with lower body weights, and malformations such as hydrocephalus, spina bifida, and cranioschesis. None of the fetuses exhibited any significant skeletal or soft tissue a nomalies, although fetuses from dams gavaged with 500 mg/kg HCE displayed significantly lower ge station indices, a lower number of viable fetuses/dam, and higher fetal resorption rates comp ared with controls (data not shown). This study concluded that HCE did not produce teratogeni c 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 t reatment (8 th day of pregnancy); dams in the 167 mg/kg dose group displayed significantly decrea sed weight gain after the fourth day of treatment (10 th day of pregnancy) but not after the treatment ende d on the 18 th 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 treatme nt; however, intake was normal when treatment ended. Dams in both the 167 and 500 mg/k g dose groups exhibited decreased motor activity (incidence and method of analysis not repo rted); 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 d ams 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 de aths 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 s tage 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 Wis tar 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 6.4 a 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 ± 0.57 a

2.3 ± 0.45 a

a Significantly different from control, p < 0.01. b Values are mean ± SD.

Source: Shimizu et al. (1992).

Investigators examined the fetuses for external ano malies and found one case of acaudate in the 500 mg/kg dose group. Other anomalies inclu ded two fetuses with subcutaneous hemorrhage in the 167 and 500 mg/kg dose groups, an d one case of hyposarca in the 500 mg/kg dose group. No skeletal malformations were observe d in any group although a statistically significant increase in skeletal variations was obs erved 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, numbe r of proximal and middle phalanges, and number of sacral and caudal vertebrae) was signific antly decreased in the 500 mg/kg dose group. No visceral malformations were observed and no sign ificant differences in visceral anomalies were noted. The authors of this study concluded th at 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 fetuse s. 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 60.3 a

Number of fetuses with variations

Lumbar rib 0 0 0 2 Rudimentary lumbar rib 2 0 6 78

Ossification b

Number of sternebrae 4.7 ± 0.07 4.5 ± 0.08 4.5 ± 0.08 3.4 ± 0.27 a Number of proximal and middle phalanges

Fore limb 3.2 ± 0.05 3.1 ± 0.04 3.1 ± 0.04 2.9 ± 0.11 a

Hind limb 4.0 ± 0.01 4.0 ± 0.01 4.0 ± 0.01 3.4 ± 0.23 a

Number of sacral and caudal vertebrae 6.9 ± 0.06 6.9 ± 0.08 7.0 ± 0.04 5.7 ± 0.37 a

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 va lues, 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-Daw ley rats/concentration to control air, 15, 48, or 260 ppm HCE (145, 465, and 2,517 mg /m 3 , respectively; purity 99.8%) by inhalation on GDs 6-16. Dams in the 2,517 mg/m 3 group displayed tremors during GDs 12–16. Body weight gain of the dams was significantly lowe r than controls beginning on GD 8 for the 2,517 mg/m 3 group, and beginning on GD 14 for the 465 mg/m 3 group. Rats in the 465 and 2,517 mg/m 3 groups exhibited an increased incidence of mucopuru lent nasal exudates compared with controls. Inflammatory exudate was observed i n the lumen of the nasal turbinates of 85% of the 465 mg/m 3 group and 100% of the 2,517 mg/m 3 group. The authors attributed the increased exudate to an endemic mycoplasia infectio n. Fetuses of HCE-treated dams did not exhibit any si gnificant skeletal or soft tissue anomalies. This study concluded that HCE did not p roduce teratogenic effects at concentrations that were not maternally toxic. EPA considered the NOAEL for the dams as 465 mg/m 3 and the LOAEL as 2,517 mg/m 3 , based on neurological effects (tremors). EPA con sidered 2,517 mg/m 3 (highest concentration tested) as a fetal NOAEL, ba sed on the lack of treatment-related effects,

while a fetal LOAEL could not be established from t his 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 (Ki nkead and Wolfe, 1992; Weeks et al., 1979). According to the Hodge and Sterner Scale, these let hal doses place HCE in low toxicity range (Hodge and Sterner, 1949). Reynolds (1972) adminis tered 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 conc entration, antipyrine demethylase activity, NADP-neotetrazolium reductase activity, glucose 6-p hosphatase activity, and conjugated diene concentration in microsomal lipids) was unaffected 2 hours after exposure. Kinkead and Wolfe (1992) determined the oral median lethal dose (LD 50 ) 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 cond ucted 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 r abbits, and Hartley guinea pigs by administering a single dose of HCE (99.8% purity) d issolved in corn oil (50% w/v) or methylcellulose (5% w/v) via stomach tube. Approxi mate lethal dosages (ALD) or LD 50 values were calculated after a 14-day observation period ( Table 4-13). All LD 50 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 LD 50 Corn oil Methylcellulose 4,460 7,080 3,900–5,110

6,240–8,040 9.3

19.9 Rat, male Oral LD 50 Corn oil Methylcellulose 5,160 7,690 4,250–6,270

6,380–9,250 6.1

8.5 Guinea pig, male Oral LD 50 Corn oil 4,970 4,030–6,150 4.7 Rabbit, male Dermal LD 50 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 byDRAFT – DO NOT CITE OR QUOTE 37 measurement of plasma enzyme activities and bromsul phthalein dye clearance tests, which are widely-used indices of hepatic function in sheep. Plasma activities of glutamate dehydrogenase (GDH), sorbitol dehydrogenase (SDH), ornithine carb amoyl 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-fol d increase in GDH, with the exception of one sheep that exhibited a 55-fold increase. SDH w as increased 3-6-fold, and OCT was increased 2-10-fold. GDH, SDH, and OCT levels peak ed at 48 hours and returned to normal within 4–5 days. AST increased only slightly. Bro mosulphophthalein 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 literat ure 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/m 3 ) for 8 hours and to 1,000 ppm HCE (17,000 mg/m 3 ) for 6 hours. Postexposure observation was carrie d out for 14 days. Male rats exposed for 8 hours to 2,500 mg/m 3 HCE displayed no toxic signs during exposure or for 14 days thereafter. Body weight ga in was slightly, but not statistically significantly, reduced over the 14-day exposure per iod. Male rats exposed for 8 hours to 57,000 mg/m 3 HCE displayed severe toxic signs including death. At 6 hours, one rat had a staggered gait. At 8 hours, 2/6 rats were dead. T he 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 re veal any gross exposure-related lesions. Microscopy revealed that two of the four surviving rats had minimally to moderately severe subacute diffuse interstitial pneumonitis and vascu lar 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 f or 6 hours to 17,000 mg/m 3 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 th e 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 HC E in animal species. A 12-day study in male New Zealand White rabbits found liver degen eration and necrosis, as well as tubular nephrosis in the kidney, indicating that both the l iver and kidney are potential target tissues forDRAFT – DO NOT CITE OR QUOTE 38 HCE-induced toxicity (Weeks et al., 1979). Short-t erm toxicity assays in rats (16 and 21 days) demonstrated kidney effects in males (NTP, 1996, 19 89) 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 dai ly oral dose via a stomach tube of 100, 320, or 1,000 mg/kg HCE (purity 99.8%) suspended in 5% aque ous methylcellulose. Blood was drawn from the central ear artery of the rabbits on treat ment days 1, 4, 8, and 12, and on day 4 following termination of dosing. Serum was analyze d for the following parameters: glutamic oxaloacetic transaminase (SGOT; also known as AST), glutamic pyruvic transaminase (SGPT; also known as ALT), blood urea nitrogen (BUN), alka line phosphatase, bilirubin, total protein, potassium and sodium. On the 4 th day following the termination of dosing, rabbits w ere 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 k idney weights. The 320 mg/kg dose group exhibited significantly reduced body weight beginni ng 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, coagula tion 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 w hich 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 kidne y was also observed in the rabbits of the 320 and 1,000 mg/kg dose groups. These animals also ex hibited tubular nephrocalcinosis of a minimal degree. The only blood chemistry parameter s that were affected were significantly decreased potassium and glucose levels in the 320 a nd 1,000 mg/kg groups. EPA considered the NOAEL as 100 mg/kg and the LOAEL as 320 mg/kg, base d 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 pe rformed on all rats; all organs and tissues were examined for grossly visible lesions and histo pathology. All rats of the 1,125 and 2,250 mg/kg-day dose groups and 1/5 males and 2/5 f emales 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 t he cytoplasm of renal tubular epithelium in allDRAFT – DO NOT CITE OR QUOTE 39 the treated males, and tubular cell regeneration an d 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 to xicity in male F344/N rats. Groups

of five rats/dose were administered 0.62 or 1.24 mm ol HCE/kg-day (146 or 293 mg/kg-day, respectively; purity 100%) by corn oil gavage. Nec ropsies were performed on all rats; the right kidney, liver, and right testis were weighed and un derwent histopathological evaluation. Urine samples were collected during an overnight period t hat began 4 days before the end of the study. Urinalysis included measurements of volume, specifi c 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 sectio ns to evaluate protein droplets, particularly hyaline droplet formation. Cell proliferation anal yses were performed on kidney sections and were scored by a labeling index indicating the perc entage 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 significa ntly increased in both dose groups; absolute and relative (significant at high dose) liver weigh ts were increased in both dose groups. Rats of the 293 mg/kg-day group also exhibited significantl y lower urinary creatinine and specific gravity, while glucose and urine volume were greate r than controls. AST and NAG activities were significantly higher than in controls. Nephro pathy, consisting of hyaline droplet accumulation, was observed in the male rats in addi tion 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 w eight, increased AST and NAG activity, increased PCNA labeling index, and nephropathy) wer e observed at the low dose level. EPA considered 146 mg/kg-day a LOAEL based on statistic ally significant increases in kidney lesions and urinalysis parameters.DRAFT – DO NOT CITE OR QUOTE 40 Table 4-14. Summary of toxicity data from male rat s exposed to HCE for 21 days Vehicle control 146 mg/kg-day HCE 293 mg/kg-day HCE Right kidney weight a Absolute (g) 1.009 ± 0.025 1.157 ± 0.011 b 1.250 ± 0.022 b Relative (mg/g) 3.19 ± 0.04 3.77 ± 0.06 b 4.07 ± 0.05 b Liver weight a 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 ± 0.64 b Right testis weight a 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 ± 3.06 b Glucose ( μ g/mg creatinine) 169 ± 3 344 ± 30 446 ± 23 b Protein (mU/mg creatinine) 1,322 ± 59 1,748 ± 257 2,980 ± 103 AST (mU/mg creatinine) 6 ± 1 40 ± 6 c 66 ± 5 b GGT (mU/mg creatinine) 1,456 ± 47 1,547 ± 66 1,897 ± 73 NAG (mU/mg creatinine) 11 ± 0 23 ± 2 c 36 ± 1 b Volume (mL/16 h) 4.2 ± 0.8 7.5 ± 9 10.6 ± 1.1 b Specific gravity (g/mL) 1.038 ± 0.005 1.024 ± 0.003 1.020 ± 0.001 b PCNA labeling index (mean ± SE) 0.13 ± 0.02 0.74 ± 0.19 c 1.2 ± 0.2 c

a Data 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 nerv ous system (CNS) effects are due to the parent compound or the metabolites. Although there are fe w studies on the neurological effects associated with HCE exposure, the database is exten sive 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; Sout hcott, 1951), and a staggering uncoordinated gait (Southcott, 1951). Sprague-Dawl ey rats evaluated for HCE-induced effects on avoidance latency (i.e., learned behavior) and s pontaneous motor activity (i.e., unlearned behavior) exhibited slight, but not statistically s ignificant, behavioral effects at 2,517 mg/m 3

. Male and female rats exhibited tremors and ruffled pelt at 2,517 mg/m 3 as well (Weeks et al., 1979). Beagle dogs developed signs of neurotoxicit y such as tremors, ataxia, hypersalivation, and head bobbing, following exposure to 2,517 mg/m 3 HCE. Dogs showed similar signs ofDRAFT – DO NOT CITE OR QUOTE 41 neurotoxicity intermittently throughout the HCE exp osures, 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 lev els: 500 (10 sheep), 750 (1 sheep), and 1,000 mg/kg (1 sheep). Slight facial muscle tremor s 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 adm inistered 0.3 mL/kg pentachloroethane, two proposed major metabolites of HCE. The sheep e xposed to PERC exhibited no effects following exposure, while the sheep exposed to pent achloroethane exhibited narcosis One pentachloroethane-exposed animal was recumbent with in 30 minutes of exposure, exhibiting flaccid limbs, depression of normal reflexes, and l abial tremors. The sheep regained normal posture 9 hours postexposure and appeared normal 72 hours postexposure. The second pentachloroethane-treated sheep became recumbent wi thin 20 minutes of exposure and exhibited labial tremors. However, unlike the first sheep, t his animal appeared normal 1.5 hours postexposure. EPA considered the LOAEL as 500 mg/k g (lowest dose tested), based on neurotoxic effects (tremors), while a NOAEL could n ot be established from this data. Southcott (1951) treated 30 Merino Wethers sheep s uffering 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 other s were unable to rise and stand. One of the severely affected sheep (i.e., unable to rise and s tand) was from the 445 mg/kg HCE group and the other eight were from the 906 mg/kg group. Som e 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 for elegs were afflicted by fine muscular tremors that

were observed in most of the animals. EPA consider ed the LOAEL as 445 mg/kg (lowest dose tested), based on neurological effects consisting o f tremors, staggering, uncoordinated gait, and inability to stand, while a NOAEL could not be esta blished from this study. As described in Section 4.3.1, Shimizu et al. (199 2) 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 d uring GD 7–17. These effects decreased or disappeared when HCE exposure ended. Weeks et a l. (1979) exposed 22 pregnant Sprague-Dawley rats/dose to 50, 100, or 500 mg/kg H CE by corn oil gavage on GDs 6–16. Dams orally administered 500 mg/kg HCE displayed tr emors 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 ra ts (15/concentration) to air, 15, 48, or 260 ppm HCE (145, 465, or 2,517 mg/m 3 , respectively; purity 99.8%) for 6 hours/day, 5 days/week for 6 weeks. Learned behavior endpoint s evaluated using an avoidance latency task by measuring the time it took the rats to avoid foo t shock by escaping into a safe compartment and unlearned behavior endpoints (spontaneous motor activity; evaluated by photobeam interruptions) were measured in the animals. The a voidance latency task was conducted prior to exposure, 1 day into exposure, after 3 weeks of exp osure and after 6 weeks of exposure. Spontaneous motor activity was tested after 3 and 6 weeks of exposure. Avoidance latency was slightly but not significant ly increased in the 465 and 2,517 mg/m 3 groups at 6 weeks (median 3.9 and 3.3 seconds, res pectively) compared with control (median 2.2 seconds). Spontaneous motor ac tivity counts were slightly, but not significantly increased in the HCE-treated rats (me an ± SD): 231 ± 77 for 145 mg/m 3 , 183 ± 109 for 465 mg/m 3 , and 201 ± 102 for 2,517 mg/m 3 , compared with control rats (163 ± 74). Weeks et al. (1979) concluded that the rats did not displ ay 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/m 3 HCE during the fourth week of exposure. Tremors an d lack of grooming are indicators of neurobehavioral effects (Kulig et al., 1996). The investigators sacrificed the rats 12 weeks after the last exposur e and reported that all measurable changes (e.g., brain histopathology, body weights) were comparable to controls. Weeks et al. (1979) also exposed 22 pregnant Sprag ue-Dawley rats/concentration and 4 beagle dogs/concentration to 145, 465, and 2,517 mg/m 3 HCE by inhalation. Rat dams in the 2,517 mg/m 3 group displayed tremors during GDs 12–16. Dogs in the 2,517 mg/m 3 exposure group developed tremors, ataxia, hypersalivation, a nd displayed severe head bobbing, facial muscular fasciculations, and held their eyelids clo sed during exposure. One dog experienced

convulsions and died within 5 hours after initial e xposure. The surviving dogs exhibited less severe symptoms during exposure, but recovered over night 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 26 0 ppm (145, 465, and 2,517 mg/m 3

, 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 pati ents suffering from systemic scleroderma (SSD) for potential exposure to solvent s. These patients also presented with

localized scleroderma with bilateral distribution o f 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 exposu re (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 his tological examination and electron microscopy to be a significant inducer of scleroder matous 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 fibro sis 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 occupat ional 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 st udies 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 eryt hema at 24 hours. Abraded skin displayed barely perceptible erythema in one rabbit with mode rate to slight erythema reactions. HCE was placed in Irritation Category III (mild or slight i rritation).

4.4.6. Eye Irritation Weeks and Thomasino (1978) applied a single, 24-ho ur dose of 100 mg dry HCE to one eye of six male New Zealand White rabbits. Moderat e corneal damage, iritis, and conjunctivitis was observed in 5/6 rabbits 24, 48, and 72 hours af ter exposure. No effects were observed 7 days after exposure. HCE was placed in Irritatio n 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 perform ed in humans exposed to HCE. In vivo exposure to animals resulted in predominantly nonpositive results. Similarly, in vitro genotoxicity studies conducted in microorganisms, c ultured mammalian cells, and insects (Table 4-15) were largely nonpositive both in the p resence and absence of exogenous metabolic activation. Ashby and Tennant (1988) examined geno toxic 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 tha t HCE (purity >99%) was not significantly genotoxic, and that the increased incidence of tumo rs 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 re ported 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 havin g insufficient data for evaluation (Jackson et al., 1993). Studies conducted by Lohman and Loh man (2000) considering DNA damage, recombination, gene mutation, sister chromatid exch ange (SCE), micronuclei (MN), CA, aneuploidy, and cell transformation as endpoints in dicate that the genetic activity profile for HCE is predominantly nonpositive. However, some po sitive findings have been reported in assays for gene conversion, somatic mutation/recomb ination, 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

proliferation Mice

BALB/c mice Moderately + DNA binding in liver Lattanzi et al.

(1988) Adducts not

identified

Isolated human lymphocytes + ( ± S9) Tafazoli et al. (1998) Human Lymphocytes DNA strand

breaks Human lymphocyte

cultures – Tafazoli et al. (1998) Comet assayDRAFT – 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 centrifug al spin.

Using the standard Ames assay for reversion of S. typhimurium histidine tester strains (TA1535, TA1537, TA1538, TA98, and TA100), Simmon a nd Kauhanen (1978) found HCE to be nonmutagenic at concentrations of 5,000 or 10,00 0 μ g 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) metab olic activation system. HCE was reported to be slightly toxic at the 10,000 μ g/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 ra nge of 0.1–500 μ g 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 independ ent trials in two separate laboratories over a collective concentration range of 1–10,000 μ g/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; Haw orth et al., 1983). Finally, HCE (purity >97%) was reported to be nonpositive in several Ame s 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. (199 1) 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 μ mol/plate (355–7,102 μ g/plate), both with and without

rat liver S9 metabolic activation. Of the 16 chemi cals tested in this study, HCE was the only one that did not demonstrate any toxicity, which the au thors 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., 198 7). 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 4 2 μ g/mL (solvent, water or DMSO, was not specified for individual test agents). Although th e 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 i s 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 activ ity over a concentration range of 0.1-5.0% HCE (1–50 mg/mL; purity not speci fied), 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 s pecified) 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 e nzymes, 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 solu bility. HCE was inactive for both gene conversion and reverse mutation in stationary cultu res with or without S9, and for reverse mutation in the logarithmic culture. However, stat istically significant ( p ≤ 0.05–0.001) increases in revertant frequency of more than twofold over ba ckground 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, 198 8). Liquid suspension exposures (3 hours) to concentrat ions of 0.0025–0.04% HCE (0.005-0.84 mg/mL; purity >98%) resulted in surviva l rates of 100-48%. Exposure to these concentrations did not induce mitotic malsegregatio n 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) b y 10 aliphatic halogenated hydrocarbons, including HCE. CD-1 mice (5/sex/concentration) wer e injected i.p. with HCE doses of 2,000 or 4,000 mg/kg (purity >98%), representing approximate ly 40 and 70-80% of the LD 50 , respectively. Animals were sacrificed and bone mar row cells harvested at 24 and 48 hours post treatment. At least 5,000 polychromatic erythrocyt es/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 interchromos omal mitotic recombination, caused by the exposure of larvae to various chemicals. In the ca se of HCE (3% ethanol solvent; purity not specified), adult flies of the C-1 cross were permi tted 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 classifie d as a weak positive response for HCE—a reproducible increase of not more than a doubling o f the spontaneous frequency at a doseDRAFT – DO NOT CITE OR QUOTE 48 associated with toxicity. The authors suggested th at 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 repor ted 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/m L), 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 wi th MN were recorded at HCE concentrations of 1 and 8 mM (0.24 and 1.89 mg/mL, respectively) in t he 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, lymphoc yte cultures exposed to test agents for 3 hours with an d 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 c ellular 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 specifi ed; 0, 0.002, 0.012, or 0.024 mg/mL, respectively), and then processed for scoring of ki netochore-positive and kinetochore-negative MN. No MN formation was observed in any of the thr ee cell lines in response to HCE exposure. However, MN induction was enhanced by exposure to a n HCE metabolite, PERC, in h2E1 and

MCL-5 cells.

Induction of CAs and SCEs in cultured Chinese hams ter ovary (CHO) cells exposed to HCE was investigated as part of an NTP screening pr ogram for genotoxicity (NTP, 1989; Galloway et al., 1987). Concentrations for analysi s were selected based on observations of cell confluence and mitotic cell availability. HCE conc entrations (purity >99%) ranged from 10 to 1,000 μ g/mL (0.01–1.0 mg/mL). For both endpoints, linear regression was used to test for dose-response trends. For individual doses, induct ion of CA was considered significant if

p values (adjusted by Dunnett’s method to correct fo r 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 exposu re 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 2 4-hour incubation without HCE). CAs wereDRAFT – 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 respons e 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 obtai ned in the presence of S9 (0.10–1.0 and 0.40-1.0 mg/mL HCE).

In vitro cell transformation studies were conducte d 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%) fro m 0.16 to 100.0 μ g/mL (0.00016-0.100 mg/mL) failed to induce morphologica l cell transformation in BALB/c-3T3 cells, as measured by the incidence of Type III foc i (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 initiat e and promote tumors in a rat liver foci assay. To assess initiation potential, 24 hours after part ial 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 prom oter phenobarbital for 7 weeks. Following sacrifice, livers were examined histopathologically for foci containing GGT, a putative preneoplastic indicator. To assess promotion poten tial, 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 MT D of HCE in corn oil by gavage, 5 days/week for 7 weeks. The animals were sacrific ed 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/cm 2 , 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 hepatoc arcinogens in an in vivo–in vitro hepatocyte replicative DNA synthesis (RDS) assay (Y oshikawa, 1996; Miyagawa et al., 1995). Groups of 4–5 male B6C3F 1 mice were exposed to single gavage doses of 0, 1,0 00, 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% (investi gators noted that an RDS incidence rate of 0.4% for any dose group was considered a positive r esponse 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 CellsDRAFT – 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, a nd protein in several mouse and rat tissues (Lattanzi et al., 1988). This study reported bindi ng of radiolabeled carbon to DNA, RNA, and protein was observed following [ 14 C]HCE administration in both in vitro and in vivo a ssays in mice and rats (Lattanzi et al., 1988), suggesting t hat 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) conduc ted a rat liver foci assay to assess the initiation and promotion potential of HCE, along wi th eight other chlorinated aliphatics. Male Osborne-Mendel rats (10 rats/group) were given part ial 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 d iet 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 49 7 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/cm 2 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 we ights were increased by HCE in the promotion protocol. These results indicate that HC E 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 l iver of the promotion protocol Total number of foci/cm 2 Promotion treatment + DEN initiation - DEN initiation HCE 4.4 ± 1.0 a 0.1 ± 0.2 Phenobarbital 3.9 ± 1.0 a 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 vi tro assays to assess the binding of [ 14 C] HCE (specific activity 14.6 mCi/mmol, radiochemi cal purity 98%) to nucleic acids in various organs from mice and rats following metabol ic activation. For the in vivo studies,DRAFT – DO NOT CITE OR QUOTE 51 6 male Wistar rats and 12 male BALB/c mice were inj ected i.p. with 127 μ Ci/kg HCE (purity 98%). The animals were fasted and sacrificed 22 ho urs after injection. The organs (liver, kidney, lung, and stomach) were removed, pooled, an d 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 μ Ci [ 14 C] 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 uti lized with the controls. Measures for binding to macromolecules were determined by the pr esence of radiolabeled carbon from [ 14 C] HCE in the DNA, RNA, and protein. The presence of radiolabeled carbon may indicate HCE binding directly to the macromolecules or incor poration 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 car bon; however, HCE-specific metabolites were not measured. In both rats and mice, binding value s (in pmol HCE/mg) for RNA were consistently much greater than those for DNA or pro tein. 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 low est amount of HCE binding. Species

differences were evident for all three macromolecul e types (DNA, RNA, and protein) with the mouse exhibiting much higher levels (9 times greate r) of covalent binding for DNA in the liver than the rat. The binding was 2 and 3 times greate r for mice than rats with RNA and protein, respectively, from the liver. The binding to DNA w as 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) oncoge nic potency in HCE-treated rodents.

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

Liver a Kidney a Lung a Stomach a (pmol/mg) Rat Mouse Rat Mouse Rat Mouse Rat Mouse DNA (CBI b ) 0.43 ± 0.05 c

(15.1) b 3.92 ± 0.20 d

(140) b 0.42 0.50 0.14 0.35 0.26 0.37

RNA 46.59 ± 7.23 c 108.08 ± 21.57 d 232.94 564.98 15.55 60.10 8.33 21.04 Protein 4.94 ± 1.14 c 14.99 ± 0.83 d 2.59 4.91 0.89 3.42 0.80 2.41

a Data are from pooled organs from 6 male Wistar rats or 12 male BALB/c mice, except for liver (see indi ces). b Covalent binding index (CBI) calculated according to Lutz (1986, 1979) as cited in Lattanzi et al. (1 988). Classification of CBI values for oncogenic potency: strong, in the thousands; moderate, in the hundred s; weak, in the tens; and below one for nongenotoxic oncogens.

c Mean ± SE of six individual values. d Mean ± 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 Tabl e 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 great er amounts of HCE binding to DNA when compared with controls. Kidney microsomes from mic e had a threefold increase in HCE binding to DNA when compared to controls, while kidney micr osomes from rats had a twofold increase in HCE binding to DNA when compared to controls. M icrosomes 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 fracti ons. 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 inhibi tor, 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-inacti vated, 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 bi nding may be increased following HCE

metabolism.

Table 4-18. In vitro binding of [ 14 C]HCE to calf thymus DNA mediated by microsomal and/or cytosolic phenobarbital-induced f ractions of rat and
mouse organs

Microsomes + NADPH Cytosol + GSH Microsomes + cytos ol (+ NADPH, + GSH) Rat Mouse Rat Mouse Rat Mouse Liver Standard a 90.83 ± 5.31 b 105.39 ± 7.80 b 195.51 ± 21.44 c 346.17 ± 18.91 b 95.06 ± 6.29 c 133.44 ± 2.42 a Controls a 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 ± 78.58 c 78.86 ± 6.85 c 246.85 ± 35.39 c 251.42 ± 45.38 c 247.99 ± 3.40 b 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 ± 16.84 b 168.52 ± 19.41 b 234.26 ± 28.35 b 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 ± 31.19 b 228.74 ± 20.42 b 76.79 ± 5.34 b ND Controls 93.20 ± 15.24 47.12 ± 11.20 130.51 ± 4.01 51.52 ± 6.20 44.77 ± 2.28 ND

a Data (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 an d toxicity. NTP (1996) defined a group of chlorinated ethanes that resulted in hyaline drople t nephropathy in male F344/N rats and a group of halogenated ethanes that resulted in renal toxic ity in the absence of hyaline droplet nephropathy. In a series of studies, Crebelli et a l. (1995, 1992, 1988) evaluated chlorinated and halogenated hydrocarbons for their ability to induc e chromosome malsegregation, lethality, and mitotic growth arrest in the mold A. nidulans

. NTP (1996) conducted a 21-day oral toxicity study w ith 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 0.6 2 or 1.24 mmol/kg-day of the halogenated ethane (for HCE, 146 and 293 mg/kg-day, respectivel y). 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 all ows 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 H CE. RDS, indicated by PCNA labeling index, was increased in male rats administered chem icals that induced hyaline droplet nephropathy (pentachloroethane, 1,1,1,2-tetrachloro ethane, and HCE) as well as pentabromoethane and 1,1,2,2-tetrachloroethane, com pared with control rats. The increase in cell proliferation in the kidneys (as measured by t he PCNA labeling index) observed with some of the halogenated ethanes that did not induce hyal ine droplet nephropathy suggests the contribution of another toxic mechanism. NTP (1996 ) concluded that the capacity to induce hyaline droplet nephropathy in male rats was restri cted to ethanes with four or more halogens, and only the chlorinated (compared with the fluorin ated and brominated) ethanes were active. This study also predicted that if hyaline droplet n ephropathy is the determining factor in the induction of renal tubule cell neoplasia, then chem icals 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 chlorometha nes and eight chlorinated ethanes (including HCE) for the induction of chromosome mal segregation in A. nidulans . Although 8 of the 11 compounds tested provided positive results i ncluding the 3 chloromethanes and 5 out of 8 chlorinated ethanes, HCE was nonpositive for chro mosome malsegregation induction. Analyses of relationships between biological and ch emical variables indicate that the ability of a chemical to induce chromosome malsegregation was no t related to any of the chemical descriptors examined, including molecular weight, m elting point, boiling point, refractive index,DRAFT – DO NOT CITE OR QUOTE 54 octanol/water partition coefficient, and the free e nergy of binding to biological receptors. Because of the similarity of the chemical descripto rs between the positive chlorinated ethanes, aside from 1,1,1-trichloroethane which was nonposit ive, the authors argue against a previous hypothesis that nonspecific interactions with hydro phobic 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 mi totic 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-acti vity 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 th e 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 hydrocar bons indicate that toxicity, such as the induction of lethality, is primarily related to ste ric factors (the spatial orientation of reactive centers within a molecule) and measures of the volu me occupied by an atom or functional group (molar refractivity). Measures of molar refractivi ty are a function of temperature, index of refraction, and atmospheric pressure. Mitotic grow th arrest was also primarily related to molar refractivity. However, aneugenic activity was rela ted to both molar refractivity and electronic factors, such as the ease in accepting electrons (d escribed by density and the energy of the lowest unoccupied molecular orbital). These QSAR studies (Crebelli et al., 1992, 1988) w ere 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 conclud ed that electronic and structural parameters that determine the ease of homolitic cleavage of th e 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 th at 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 relativ e 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 t oxicity was absent.DRAFT – DO NOT CITE OR QUOTE 55 Acute and short-term toxicity tests in animals rep orted 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 tubu lar cell regeneration in male rats (NTP, 1996, 1989). Female rats in short-term toxicity te sts displayed only decreased body weights at the LOAEL of 563 mg/kg-day with a NOAEL of 281 mg/k g-day (NTP, 1989). Oral LD 50 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)

B6C3F 1 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)

PregnantSpragueDawley 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 ma le rats. The formation of hyaline droplets accompanied by cell regeneration and eosinophilic g ranular 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 regen eration and granular casts, and increased labeling index in kidneys at doses of 146 and 293 m g/kg-day HCE. Tubular nephrosis, and to a minimal degree, tubular nephrocalcinosis were obser ved 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 m ore 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 obser ved were gender specific. Subchronic exposure in male F344/N rats demonstrate d kidney effects including hyaline droplet formation, tubular regeneration, and tubula r 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 d egeneration and necrosis of renal tubule epithelium (NTP, 1989). Female rats did not displa y 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 ren al 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 fo r 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 Distribut ion) show that the male kidney accumulated higher HCE concentrations than the female kidney, i ndicating that the kidney is the primary target organ following oral exposure to HCE and the re 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. Additional ly, Gorzinski et al. (1985) is the only study of either short-term or subchronic duration to repo rt 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 B6C3F 1 mice (NTP, 1989; NCI, 1978). NTP (1989) administe red much lower doses of HCE (7 and 14 mg/kg-day in male s; 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, nep hropathy (characterized as tubular cell degeneration and regeneration, dilation and atrophy , glomerulosclerosis, interstitial fibrosis, andDRAFT – 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 cont rol and treated groups with nephropathy in more than 94% of animals. The high incidence of ne phropathy observed in control rats could result of a spontaneous syndrome known as chronic p rogressive nephropathy (CPN) that is associated with aged rats, especially F344 and Osbo rne-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 cons idered. The severity was increased in the treated male rats compared with the controls. In c onsidering 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/k g-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 w as only noted at the low dose (44% at 57 mg/kg-day compared with 28% in controls). The l ow dose for the females was 8 times greater than that for the males yet the signs of ne phropathy were more severe in the males. In the NCI (1978) study, Osborne-Mendel rats of bot h sexes displayed chronic inflammatory kidney lesions in both control and tre ated groups, although tubular nephropathy (characterized by degeneration, necrosis, and the p resence of large hyperchromatic regenerative epithelial cells) was observed only in the HCE-expo sed 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 B6C3F 1 mice (NCI, 1978) is the only study conducted in th is 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 a nd 722 mg/kg-day HCE males. The report did not provide an explanation for the large response i n 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 bot h sexes at high incidences (92-100%), and was characterized by degeneration of convoluted tub ule epithelium with some hyaline casts. Enlarged dark staining regenerative tubular epithel ium was also observed, with the kidney exhibiting infiltration of inflammatory cells, fibr osis, and calcium deposition. The response inDRAFT – 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 nephropat hy in rats, mice, male rabbits, and sheep indicates that the male rat is the most sensi tive sex/species to the renal toxicity of HCE. Limited, if any, information is available for speci es 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 t he lowest dose (7 mg/kg-day; NTP, 1989) that induced a statistically significant response in rat s.

4.6.1.2. Hepatotoxicity Short-term studies in rats (NTP, 1996), male rabbit s (Weeks et al., 1979), and sheep (Fowler, 1969) reported hepatotoxicity at doses app roaching ≥ 300 mg/kg-day. Male F344 rats exhibited significantly increased relative liver we ights at the highest dose of 293 mg/kg-day. AST and NAG serum activities were also significantl y 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 macroph ages and giant cells were observed in male New Zealand White rabbits administered 320 and 1,00 0 mg/kg-day HCE (but not 100 mg/kg-day); increasing in severity with increas ing dose. Sheep given single oral doses of 500-1,000 mg/kg of HCE exhibited plasma levels of G DH, 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 we re 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 an d statistical significance increased in females at doses lower than those eliciting toxicit y in male rats. Hepatocellular necrosis was noted in females at doses ranging from 134 to 156 m g/kg-day and in males at the two highest doses, 268 and 536 mg/kg-day (NTP, 1989). Gorzinsk i 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 s welling. 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 fem ale 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) st udy, 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 as say (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 ratsDRAFT – DO NOT CITE OR QUOTE 60 (57 and 114 mg/kg-day; NTP, 1989) and Osborne-Mende l rats (113 and 227 mg/kg-day; NCI, 1978). Hepatocellular necrosis was observed in fem ale rats in the subchronic, but not the chronic study. The LOAEL for female F344/N rat hep atocellular 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 suffic iently high dose may have not been achieved to elicit hepatocellular necrosis despite the longe r 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 hepatot oxicity as F344/N rats. The only study in mice (NCI, 1978; chronic) did not report any hepato toxic effects other than the development of hepatocellular tumors. HCE-induced liver effects were only observed in an imals in short-term and subchronic studies. Female rats exhibited a greater sensitivi ty to liver effects as evidenced by the effects observed at lower doses compared with males (NTP, 1 989). The implications of the slight swelling of hepatocytes in the absence of other his topathological 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 HC E-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 50 0 mg/kg HCE displayed decreased motor activity. At the high dose, dams also exhibited pi loerection and subcutaneous hemorrhage. Fetuses of the 500 mg/kg dose displayed decreased b ody 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 exhib it any skeletal or soft tissue anomalies. The maternal LOAEL and NOAEL were 500 and 100 mg/kg, re spectively.

4.6.1.4. Metabolite Toxicity The potential metabolites of HCE are PERC and penta chloroethane (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 e xposure. 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 e xposure. 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 metaboli tes formed during HCE metabolism could contribute to the liver, kidney, and neurolog ical effects observed in animals exposed to HCE. However, the available metabolism data for HC E do not allow for the conclusive identification of metabolites or provide quantitati ve information on how potential metabolites contribute to effects associated with HCE exposure.

4.6.2. Inhalation Inhalation toxicity has only been evaluated in a s ingle 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 va por only when surrounded by heated air. However, as soon as the hot HCE vapor was mixed wit h room temperature air, most (but not all) vapor in the airstream would condense into fine par ticles (a solid aerosol). The data from this study are summarized in Table 4-20. The study auth ors reported NOAELs and LOAELS for beagle dogs, guinea pigs, and rats of 48 ppm (465 m g/m 3 ) and 260 ppm (2,517 mg/m 3 ), respectively. Neurological effects, such as tremor s and ataxia, were observed in beagle dogs and in pregnant and nonpregnant Sprague-Dawley rats. R ats and guinea pigs exhibited reduced body weight gain and increased relative liver weight. M ale rats also displayed increased relative spleen and testes weights. Behavioral tests were c onducted in male Sprague-Dawley rats at the same exposure concentrations, and no significant ef fects were observed. Overall, the information on the inhalation toxicity of HCE is li mited.DRAFT – DO NOT CITE OR QUOTE 62 Table 4-20. Inhalation toxicity studies with HCE Species Concentration

(mg/m 3 )/duration a NOAEL

(mg/m 3 ) LOAEL

(mg/m 3 ) 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/m 3 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 li mited 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 w orkers were too small to provide

definitive conclusions on health effects. Animal studies suggest that HCE is primarily metabo lized to PERC and pentachloroethane by CYP450 enzymes of the liver, w ith likely subsequent metabolism to TCE. Metabolites identified in the urine include TCA, tr ichloroethanol, oxalic acid, dichloroethanol, dichloroacetic acid, and monochloroacetic acid. Ho wever, only 5% of a radiolabeled compound was measured in the urine, indicating that all of t he urinary metabolites account for a small percentage of the dose. It is unknown whether unch anged HCE or its metabolites are responsible for the liver and kidney toxicities observed in ani mal studies. Only one study attempted to assess

the extent of HCE metabolism in rats and mice and e stimated that 24–29% of administered HCE is metabolized (Mitoma et al., 1985). This study d id not quantify actual metabolite concentrations, so these estimations are of questio nable 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 inDRAFT – DO NOT CITE OR QUOTE 63 male rats showed nephropathy characterized by hyali ne droplet accumulation and increased incidences of tubule regeneration and granular cast s (NTP, 1996, 1989). Cell proliferation of kidney sections using PCNA labeling analysis was al so increased (NTP, 1996). Subchronic and chronic animal bioassays confirmed these renal effe cts (NTP, 1989; Gorzinski et al., 1985; NCI, 1978). Chronic inflammatory kidney lesions and tub ular 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 α 2u -globulin mode of action could contribute to hexachloroethane-induced nephropathy. However, the re is insufficient evidence to conclude that the kidney effects observed following HCE exposure (NTP, 1989) are related to an α 2u -globulin mode of action for the following reasons: (1) the lack of α 2u -globulin immunohistochemical data for HCE-induced nephrotoxicity and carcinogenicity, (2) the hyaline droplet accumulation is caused by excessive protein load that may not be ex clusively related to α 2u -globulin accumulation, and (3) the existence of renal toxici ty in female rats and male and female mice indicates that the nephrotoxic effects are not limi ted to an α 2u -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 spo ntaneously, may contribute to the observed nephrotoxicity following HCE exposure. However, ch anges 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. Additi onally, HCE-exposed male rats demonstrated dose-dependent increases in incidence of mineraliza tion of the renal papillae and hyperplasia of pelvic transitional epithelium. Neither of these e ffects 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 st udies concluding that HCE could bind to DNA, RNA, and pro tein in the kidney. The liver has been demonstrated to be a target orga n in several animal species. Sheep (Fowler, 1969) and male rabbits (Weeks et al., 1979 ) exhibited hepatotoxicity characterized by clinical chemistry parameters that indicated reduce d 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 liv er 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 pa rticularly 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 respon sible 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 form ation of malondialdehyde and conjugated dienes) following treatment with HCE (8 mM). The authors s uggested 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 car bon; however, radiolabeled carbon may have been incorporated into these macromolecules from in termediary 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 demo nstrated 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 i ntermediates prior to binding to macromolecules. Therefore, renal toxicity and hepa totoxicity 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 (W eeks et al., 1979) and sheep (Fowler, 1969; Southcott, 1951) are commonly observed effect s 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 d ata are available to assess the carcinogenic potential of HCE. NTP (1989) reported dose-dependent increases (statistically significant at the high dose) in the combined incid ence 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 incre ases in the incidence of hepatocellular

carcinomas in male and female B6C3F 1 mice. The male rats demonstrated a statistically significantly increased tumor response for hepatoce llular 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, b ut 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 [ 14 C]HCE was observed in both in vitro and in vivo assays in mice and rats (Lattanzi et al., 1 988).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 initi al point of contact the weight of evidence for carcinogenic potential may apply to all routes of e xposure that have not been adequately tested at sufficient doses. An exception occurs when there i s convincing toxicokinetic data that absorption does not occur by other routes. Informa tion available on the carcinogenic effects of

hexachloroethane via the oral route demonstrates th at tumors occur in tissues remote from the site of absorption. Information on the carcinogeni c effects of hexachloroethane via the

inhalation and dermal routes in humans or animals i s 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 t o be carcinogenic to humans” by all routes of

exposure.

4.7.2. Synthesis of Human, Animal, and Other Suppo rting Evidence There are currently no data from human studies per taining to the carcinogenicity of HCE. NTP (1989) conducted a chronic toxicity/carcinogeni city 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. Gro ups of 50 female rats/dose were administered, by corn oil gavage, 5 days/week for 1 03 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 1 4% of controls, 7, and 14 mg/kg-day males, respectively. No HCE-related renal tumors were obs erved 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 t he 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 chron ic 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 u ntreated 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. A fter 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 do ses were 113 and 227 mg/kg-day. MortalityDRAFT – DO NOT CITE OR QUOTE 66 was increased in the 113 and 227 mg/kg-day males wi th survival rates of 24/50 (48%) and 19/50 (38%), respectively, compared with 14/20 (70%) in t he 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 grou ps, respectively. All of the tumor types observed had been encountere d previously as spontaneous lesions

in the Osborne-Mendel rat and no statistical differ ences 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-Mend el rats of the NCI (1978) study were approximately 16 times greater than those doses adm inistered to F344 male rats by NTP (1989).

A B6C3F 1 mouse study conducted by NCI (1978; Weisburger, 19 77), HCE (purity >98%) was administered by corn oil gavage at TWA do ses of 360 and 722 mg/kg-day for 5 days/week for 78 weeks, followed by 12–13 weeks o f 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/k g-day dose groups, respectively. Survival rates in females were 80, 85, 80, and 68% in vehicl e control, untreated control, 360 and 722 mg/kg-day groups, respectively. Both male and fema le mice exhibited statistically significantly increased incidences of hepatocellular carcinomas. The treated males demonstrated an increased tumor response for hepatocellular carcinomas that w as dose-related: 30 and 63% in the 360 and 722 mg/kg-day dose groups, respectively, compared w ith 10% in pooled vehicle controls and

15% in matched vehicle controls. Females demonstra ted an increased tumor response that was not dose related in that a higher incidence of hepa tocellular carcinomas occurred at the low dose (40%) compared with the high dose (31%); pooled veh icle and matched vehicle controls had incidences of 3 and 10%, respectively. NCI conclud ed that HCE was carcinogenic in both sexes

of B6C3F 1 mice. Evidence of HCE’s promotion (following treatment w ith 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 e t al., 1986). Lattanzi et al. (1988) reported in vivo and in vitro binding of HCE to DNA, RNA, an d protein in mice and rats. In both rats and mice administered single i.p. injections of 127 μ Ci/kg [ 14 C]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 exhib iting much higher levels (9 times greater) of covalent binding for DNA in the liver than the r at. 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 compa rable levels in liver microsomes from both rats and mice following exposure to HCE. Kidney mi crosomes from rats and mice produced statistically significantly greater amounts of DNA binding compared with controls, with greaterDRAFT – DO NOT CITE OR QUOTE 67 amounts of DNA binding from mice (threefold increas e) compared with rats (twofold increase). Microsomes from the lungs and stomachs in both spec ies did not display increased DNA binding activity over corresponding controls.

Carcinogenicity associated with potential HCE metab olites Potential metabolites of HCE include PERC, pentach loroethane, and TCE. Epidemiologic studies have reported associations be tween 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, 1 977), pentachloroethane (NTP, 1983), and TCE (NTP, 1990; 1988; NCI, 1976) have also been eva luated for carcinogenicity in several chronic bioassays. Specifically, hepatocellular ca rcinomas have been reported for exposure to PERC, pentachloroethane, and TCE. Renal tubule ade nomas have also been observed with exposure to PERC and pentachloroethane, with equivo cal evidence for TCE. An increased incidence of pheochromocytomas was reported in stud ies of pentachloroethane, but not in studies of PERC or TCE. In addition, data suggest that som e tumors are not shared between HCE and these potential metabolites (e.g., there is evidenc e 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. Howev er, 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 an d pheochromocytomas were

observed in rats and mice following oral exposure to HCE (NT P, 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 i nduction of kidney tumors in male rats and liver tumors in male and female mice may involve th e accumulation of α 2u -globulin in the kidney and increased cytotoxicity, inflammation, an d 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 α 2u -globulin-mediated mode of action. Generally, kidney tumors observed in cancer bioassays are assumed to beDRAFT – DO NOT CITE OR QUOTE 68 relevant for assessment of human carcinogenic poten tial. However, a number of chemicals have

been shown to induce accumulation of α 2u -globulin in hyaline droplets in male rat kidney. The

α 2u -globulin accumulation in hyaline droplets initiate s 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 ma mmals administered the same chemicals do

not accumulate α 2u -globulin in the kidney and do not develop renal tu bule tumors (U.S. EPA, 1991b).

The lack of α 2u -globulin immunohistochemical data for HCE-induced nephrotoxicity and carcinogenicity supports the conclusion that there is insufficient evidence to establish the role of

α 2u -globulin in HCE-induced kidney tumors. Furthermor e, reported renal toxicity in female rats and male and female mice exposed to HCE suggest a m ode of action other than α 2u -globulin nephropathy. In the absence of minimum information demonstrating the involvement of α 2u globulin 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 Technica l Panel report provides specific guidance for evaluating chemical exposure-related m ale rat renal tubule tumors for the purpose of risk assessment, based on an examination of the potential involvement of α 2u -globulin

accumulation. The protein α 2u -globulin is a member of a large superfamily of low -molecular-weight proteins and was first characterized in male rat ur ine. It has been detected in various tissues and fluids of most mammals, including humans. However, the particular isoform of α 2u -globulin commonly detected in male rat urine is considered s pecific for the male rat; moreover, the urine and kidney concentrations detected in the mature ma le 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 α 2u -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 accumula tion in the early stages of nephropathy

associated with α 2u -globulin-binding chemicals is an important differe nce from the sequence of events observed with classical carcinogens. The pat hological changes that precede the proliferative sequence for classical renal carcinog ens also include early nephrotoxicity (e.g., cytotoxicity and cellular necrosis) but no apparent hyaline droplet accumulation. Furthermore, the nephrotoxicity that can ensue from hyaline drop let accumulation is novel because it is

associated with excessive α 2u -globulin accumulation. This α 2u -globulin accumulation is proposed

to result from reduced renal catabolism of the α 2u -globulin chemical complex and is thought to initiate a sequence of events leading to chronic pr oliferation 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 contai ning α 2u -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 s loughed cellular debris associated with tubule dilatation and papillary mineralization • Foci of tubule hyperplasia in the convoluted prox imal tubules

• Renal tubule tumorsDRAFT – DO NOT CITE OR QUOTE70Table 4-21. Nephrotoxic effects characteristic ofα2u-globulin nephropathy observed in male and female rats administeredHCEX = presence of effectNT = not testedNTP, 19897 or 14 mg/kg-d (M);57 or 114 mg/kg-d (F)103 wksNCI, 1978113 or 227 mg/kg-d104 wksGorzinski et al., 19851, 15, or 62 mg/kg-d16 wksNTP, 198934, 67, 134, 268,or 536 mg/kg-d13 wksNTP, 1996146 or 293 mg/kg-d3 wksNTP, 1989140, 281, or563 mg/kg-d16 dStudy, dose, duration,and sexMales Females Males Females Males Females Males Females Males Females Males FemalesAccumulation ofhyaline dropletsX X NT XAccumulation ofα2uglobulin in hyalinedropletsNT NT NT NT NT NT NT NT NT NT NT NTNecrosis/degeneration X X X X X X X NTTubular regeneration X X X X X X NT XGranular casts/dilatationX X X X X X X NT XPapillary mineralizationX NTTubular hyperplasia X X NTDO NOT CITE OR QUOTE 71 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 ex amined in terms of three categories:

• The α 2u -globulin sequence of events accounts for the renal tumors. • Other potential carcinogenic processes account fo r the renal tumors.

• The α 2u -globulin-associated events occur in the presence o f other potential carcinogenic processes, both of which result in renal tumors.

Therefore, it is important to determine whether the α 2u -globulin process is involved and, if so, to

what extent α 2u -globulin-associated events, rather than other pro cesses, account for the tumor

increase. Determination of these elements requires a substant ial 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 α 2u -globulin processes are operative. In the absence of minimum information demonstrating the involvement of α 2u -globulin processes, it should be assumed that any male rat renal toxicity/tumors are relevant for risk assessm ent purposes.
As outlined in the U.S. EPA Risk Assessment Forum T echnical Panel report (U.S. EPA, 1991c), the following information from adequately c onducted studies of male rats is used for demonstrating that the α 2u -globulin process may be a factor in any observed r enal effects

― an affirmative response in each of the three categorie s is required. If data are lacking for any of the criteria in any one category, the available renal t oxicity data should be analyzed in accordance with standard risk assessment principles. The thre e categories of information and criteria are as

follows:

• Increased number and size of hyaline droplets in th e renal proximal tubule cells of treated

male rats . The abnormal accumulation of hyaline droplets in the P 2 segment helps

differentiate α 2u- globulin inducers from chemicals that produce renal tubule tumors by

other modes of action.

• Accumulating protein in the hyaline droplets is α 2u -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, α 2u -globulin.

• Additional aspects of the pathological sequence of lesions associated with α 2u - globulin nephropathy are present . Typical lesions include single-cell necrosis, ex foliation of epithelial cells into the proximal tubular lumen, f ormation of granular casts, linear mineralization of papillary tubules, and tubule hyp erplasia. If the response is mild, not all of these lesions may be observed. However, som e elements consistent with the pathological sequence must be demonstrated to be pr esent.DO NOT CITE OR QUOTE 72 Experimental Support for the Hypothesized Mode of A ction Strength, consistency, and specificity of associati on NTP (1989)—16-day study In a short-term exposure study, NTP (1989) administ ered 140, 281, 563, 1,125, or 2,250 mg/kg-day HCE to F344/N rats via gavage for 1 6 days. All of the surviving HCE-exposed male rats exhibited hyaline droplets in the cytopla sm of the renal tubular epithelium. Additionally, male rats exposed to 140 and 281 mg/k g-day HCE demonstrated tubular cell regeneration and eosinophilic granular casts of cel l debris in the tubule lumina at the corticomedullary junction. NTP (1989) did not repo rt regeneration or granular casts in the surviving males of the 563 mg/kg-day dose group. N TP (1989) did not report the incidence or severity of the lesions observed in the treated mal es. None of the nephrotoxic effects were observed at any HCE dose in the female rats or in t he controls.

NTP (1996)—21-day study In a second short-term exposure study, NTP (1996) a dministered 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 characterize d as two severity grades above the control rats. A Mallory-Heidenhain stain allowed for great er sensitivity in evaluating hyaline droplets within the tubules of the kidney and further suppor ted the presence of the hyaline droplets in the kidney tubules. Increased incidence of tubular reg eneration (60 and 100% in the 146 and 293 mg/kg-day dose groups, respectively) was also o bserved 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 i dentified 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 increa se 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 gr oup 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) administ ered 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 exhib ited exposure-related kidney effects, although incidence data was only reported for the 3 4 mg/kg-day dose group. These kidneyDO NOT CITE OR QUOTE 73 effects were characterized by hyaline droplet forma tion in the renal tubular epithelium, eosinophilic granular casts of cell debris in the t ubular lumina at the corticomedullary region (with associated tubular dilatation), and tubular c ell regeneration. The severity of these lesions increased with HCE exposure dose, though the severi ty grades were not reported. Furthermore, as the HCE exposure dose increased, the animals dev eloped additional lesions. Renal papillary necrosis and renal tubule epithelium degeneration a nd necrosis were observed in all 536 mg/kgday males (only the 5 male rats that died before th e end of the study were analyzed microscopically). Urinalysis in male rats administered HCE showed fin e 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, respecti vely (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% (stat istically significant) in the 268 and 536 mg/kg-day dose groups, respectively. Treated f emales showed no other HCE-exposure related kidney effects.

Gorzinski et al. (1985)—16-week study Gorzinski et al. (1985) observed dose-related level s 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 ma le rats: 20, 70, and 100% in the 1, 15, and 62 mg/kg-day dose groups, respectively. These rena l degenerative effects were also noted in 10% of the male controls, although the authors note d 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 con voluted tubules were not observed in the

male control rats. Peritubular fibrosis was also n oted in the high-dose group males. Renal tubular atrophy and degeneration were observed in 1 0, 20, and 60% of female rats in the 1, 15, and 62 mg/kg-day dose groups respectively. These l esions were seen in one female control rat (10%), although the authors characterized the sever ity grade of the lesions as very slight. Male rat sensitivity was evident in the histopathol ogical changes seen in the HCE exposed male rats compared with the female rats. R enal effects were either observed in more male rats than female rats (statistical analyses we re not reported) or did not occur in females. Additionally, kidney concentrations of HCE were muc h higher in male rats compared with female rats. Gorzinski et al. (1985) noted that th e 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).DO NOT CITE OR QUOTE 74 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/wee k for 78 weeks. Chronic inflammatory kidney lesions were observed in both control and HC E-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 contr ol and HCE-exposed male rats exhibited greater sensitivity to the chronic inflammation com pared 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 fe males. Hard et al. (1993) reported the pathologic features attributed to CPN including:

• Thickening of tubular and glomerular basement membr anes

• Basophilic segments of proximal convoluted tubules with sporadic mitoses indicative of tubule cell proliferation

• Tubular hyaline casts of proteinaceous material ori ginating 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 infl ammatory cells within areas of affected

tubules

• Glomerular hyalinization and sclerosis

• Interstitial fibrosis and scarring

• Tubular atrophy involving segments of proximal tubu le

• Occasional hyperplastic foci in affected tubules (c hronically in advanced cases)

• Accumulation of protein droplets in sporadic proxim al tubules (in some advanced cases)

Several of the CPN pathological effects are simila r to and can obscure the lesions

characteristic of α 2u -globulin-related hyaline droplet nephropathy (Hard et al., 1993). Additionally, renal effects of α 2u -globulin accumulation can exacerbate the effects a ssociated 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

α 2u -globulin nephropathy. Male rat sensitivity has be en noted with both CPN and α 2u -globulin nephropathy.DO NOT CITE OR QUOTE 75 With the exception of atrophy of proximal tubule, t ubular cell proliferation, and hyaline casts of proteinaceous material, the histopathologi cal effects associated with CPN are distinctive

from those of α 2u -globulin nephropathy. Additionally, the urinalysi s and serum chemistry of CPN-rats show albuminuria, hypoalbuminemia, and hyp ocholesterolemia as well as increased serum creatinine and urea nitrogen levels, whereas these changes in α 2u -globulin nephropathy are minimal (Hard et al. 1993). NCI (1978) reported tubular nephropathy in HCE-exp osed rats, but not in untreated or vehicle controls. Increased incidence of nephropat hy described as tubular degeneration and necrosis and the presence of large hyperchromatic r egenerative epithelial cells was observed in 45 and 66% of male rats exposed to 113 and 227 mg/k g-day HCE, respectively. Female rats also exhibited tubular nephropathy following HCE exposur e: in 18 and 59% in the 113 and 227 mg/kg-day dose groups, respectively. In additi on to the tubular nephropathy, observed effects overlying these lesions included focal pyon ephritis, tubular ectasia, cast formation, chronic interstitial nephritis and fibrosis, and fo cal glomerulosclerosis. Renal tubular cell adenomas were observed in four male rats (11% incid ence rate) exposed to 113 mg/kg-day HCE.

Similar renal tumors were not observed in males fro m 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 carci nogenicity in Osborne-Mendel rats based on the lack of statistical significance and dose-re sponse in the tumor incidence rate. However, it is possible that the truncated duration of HCE-trea tment (78 weeks, cyclical) and the significantly accelerated mortality in the male rat s did not allow enough time for the renal tubule tumors to develop. According to Goodman et al. (19 80), the incidences of spontaneous renal

tubule tumors in control male and female Osborne-Me ndel rats (as recorded in the NCI Carcinogenesis Testing Program) were 0.3 and 0%, re spectively. 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 t he 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 co rn oil via gavage to male F344/N rats for 103 weeks. Kidney effects consisting of t ubular cell degeneration and atrophy, tubular dilatation, tubular cell regeneration, glomeruloscl erosis, 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 control s and mild to moderate for the HCE-exposed male rats. While the mean severity scores do not s how more than a 15% increase over control inDO NOT CITE OR QUOTE 76 the high-dose group, more moderate and marked nephr opathy 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 inc idence in linear mineralization of the renal papillae: 4, 30, and 64% in the control, 7 mg/kg-d ay, and 14 mg/kg-day dose groups, respectively. Pelvic epithelium hyperplasia was al so observed in 14% of male rats exposed to either 7 or 14 mg/kg-day HCE. These hyperplastic e ffects 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 nephro pathy in female rats following chronic HCE exposure was 44, 84, and 90% for the control, 57, a nd 114 mg/kg-day dose groups, respectively. The severity scores for nephrotoxicity in female ra ts 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 averag e severity of nephropathy was considered

minimal for the controls and minimal to mild for th e HCE-exposed female rats. Examination of the various grades of nephropathy severity shows mo re mild and moderate nephrotoxicity in HCE-exposed females compared with controls. In fem ales, the incidence of severe (mild or moderate) nephropathy was 12/50, 25/50, and 32/50 i n 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 l inear 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 adenom a 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-relat ed neoplasms observed in female rats treated with 57 or 114 mg/kg-day HCE. NTP (1989) noted tha t the hyperplasia and tumors of the renal tubules represented a morphologic continuum. The h yperplasia incidence was observed in 4, 8, and 22% of the control, 7 mg/kg-day, and 14 mg/kg-d ay dose groups, respectively. The incidence of renal tubule neoplasia in male rats al so exceeded historical controls (0.5%). Female rats did not exhibit renal tubule hyperplasia. A sex difference was noted in the observed nephropa thy, as males were more sensitive to HCE-exposure related nephropathy than females. Thi s sex-specificity is apparent for the nephrotoxicity and grades of nephropathy severity i n both control and HCE-treated groups. Although administered only one-eighth of the dose g iven to the female rats; the male rats demonstrated a greater incidence of nephropathy tha t was more severe and included additional kidney effects (i.e., increases in incidence of min eralization of the renal papillae and hyperplasia of pelvic transitional epithelium) compared with th e female rats.DO NOT CITE OR QUOTE 77 With the exceptions of glomerulosclerosis, intersti tial fibrosis, and chronic inflammation, the observed nephrotoxic effects in the male rats a re characteristic of α 2u -globulin nephropathy. However, NTP (1989) did not report accumulation of hyaline droplets containing the

α 2u -globulin protein in the proximal tubule. It is po ssible that hyaline droplets were present considering that the 16-day and 13-week rats examin ed 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 o f the control males and in treated and control female rats also complicates the characterization o f the mode of action. Considering that

α 2u -globulin nephropathy is typically male rat-specifi c, the appearance of nephrotoxic effects in

the female rats as well as the male and female cont rols, and the identification of other effects not specifically associated with α 2u -globulin (i.e., glomerulosclerosis and interstitia l fibrosis) suggests that the effects are not the result of α 2u -globulin accumulation. Considering the strain and age of the rats in the c hronic (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 ofthe nephropathy that are greater in the HCEexposed animals indicate some chemical-related effe cts. Additionally, HCE-exposed male rats demonstrated dose-dependent increases in incidence of mineralization of the renal papillae and hyperplasia of pelvic transitional epithelium. Nei ther of these effects increased in a dose-related manner in the controls or the HCE-exposed female ra ts. 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 histopathol ogical changes characterizing the nephropathy. Additionally, the reported incidences of effects were grouped and measured as nephropathy rather than individual effects. Effect s described in this way are difficult to interpret with regards to α 2u -globulin nephropathy. One study (NTP, 1996) was l imited 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 predomin antly on the kidneys and the purpose of the study was to compare chlorinated ethanes, not exami ne the mode of action of HCE. The divergence in doses used for male and females in th e NTP (1989) chronic exposure experiment highlighted the male sensitivity to HCE-induced nep hrotoxicity. 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.DO NOT CITE OR QUOTE 78 Summary of evidence for strength, specificity, and consistency Generally, kidney tumors observed in cancer bioassa ys are assumed to be relevant for assessment of human carcinogenic potential. Howeve r, when the mode-of-action evidence demonstrates that kidney tumors in male rats result from an α 2u -globulin-related mode of action,

the data are not suitable for use in risk assessmen t (U.S. EPA, 1991b). The criteria for demonstrating the α 2u -globulin-related mode of action for risk assessmen t 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

α 2u -globulin; and (3) presence of additional pathologi cal lesions associated with α 2u -globulin. The key event in the histopathological sequence for the α 2u -globulin-related mode of action is excessive accumulation of hyaline droplets containi ng α 2u -globulin in renal proximal tubules. None of the HCE studies performed the necessary imm unohistochemical assays to confirm the presence of α 2u -globulin protein within the hyaline droplets obser ved following administration of HCE (NTP, 1996, 1989). It is unknown if HCE is bin ding to α 2u -globulin or to other proteins during the formation of hyaline droplets. This rep resents an important data gap considering that the presence of this protein is essential to identi fying the α 2u -globulin-related mode of action. Therefore, there is insufficient evidence to suppor t an α 2u -globulin-related mode of action for renal tumors following HCE exposure. In addition, the data on female rats and mice of bo th sexes from chronic exposure studies (NTP, 1989; NCI, 1978) do not support the α 2u- globulin mode of action for HCE-exposure related nephropathy. The appearance and type of ne phrotoxicity noted in control and female rats suggest a mode of action other than α 2u -globulin. NCI (1978) reported dose-related nephro pathy in female rats that was not apparent in the control s. The dose-responsive kidney effects observed in the female rats treated with HCE suggests that a mode of action other than α 2u -globulin nephropathy was occurring. Nephropathy was also re ported 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 follow ing chronic HCE exposure, but the mice did not develop renal tu bule tumors. The presence of kidney effects in HCE-exposed male and female mice, which generall y do not accumulate the α 2u -globulin protein, suggests that a mode of action other than α 2u -globulin nephropathy.

Dose-response concordance The key event in the histopathological sequence for the α 2u -globulin-related mode of action is excessive accumulation of hyaline droplet s containing α 2u -globulin in renal proximal

tubules. The accumulation of α 2u -globulin in hyaline droplets must occur at lower d oses than subsequent α 2u -globulin-related effects. None of the HCE studies performed the necessary immunohistochemical assays to confirm the presence of α 2u -globulin protein within the hyalineDO NOT CITE OR QUOTE 79 droplets observed following administration of HCE ( NTP, 1996, 1989). Therefore, this key

event cannot be demonstrated from the available dat a. Most of the effects characterizing the histopatholo gical sequence of events in epithelial cells of the proximal tubules leading to renal tumo rs (U.S. EPA, 1991c) increased in incidence

with dose of HCE in the short-term and subchronic e xposure 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 ra ts administered 7 or 14 mg/kg-day HCE for 2 years exhibited a dose-related increased incidenc e of renal tubule adenomas and carcinomas. Typical histopathological effects associated with α 2u -globulin nephropathy (tubular cell degeneration and atrophy, tubular dilatation, and t ubular cell regeneration) were noted in almost

all of the treated and untreated animals. A dose-r esponse relationship was difficult to detect considering the number of animals affected by nephr otoxicity. However, dose-related increases over controls for toxic kidney effects such as line ar mineralization, severity of nephrotoxicity, and renal tubule hyperplasia were observed. NTP (1 989) 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 α 2u globulin) is reportedly increased at doses similar to those that induce tumor formation. Overall, dose-related kidney effects were noted fo r 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 obs erved in male rats. Animals treated with greater amounts of HCE exhibited dose-related incre ases 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 t he 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 signific ant 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 nephro pathy 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 α 2u -globulin-related mode of action is excessive accumulation of hyaline droplet s containing α 2u -globulin in renal proximal

tubules. The accumulation of α 2u -globulin in hyaline droplets must occur first in t he sequela leading to α 2u -globulin-related nephrotoxicity and tumor formatio n. None of the HCE studies performed the necessary immunohistochemical assays to confirm the presence of α 2u -globulinDO NOT CITE OR QUOTE 80 protein within the hyaline droplets observed follow ing administration of HCE (NTP, 1996, 1989). Therefore, this key event and the important temporal relationship for the accumulation of

α 2u -globulin cannot be demonstrated from the available data. Many of the histopathological effects associated wi th α 2u -globulin-related nephropathy

were observed in animals treated with HCE in studie s that varied in exposure duration from 16 days to 2 years. The sequence of histopathologi cal events characteristic of the

α 2u -globulin-related mode of action was noted in the c hronic 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, althoug h an increase in renal tubule tumors was not observed. It is unknown if the nephropathy observe d by NTP (1989) leads to the reported renal tubule tumors because the animals were only examine d at the end of the 103-week study period. A temporal relationship cannot be distinguished fro m reported data.

Biological plausibility and coherence The sequence of events including accumulation of α 2u -globulin protein in the renal tubules of male rats initiating a sequence of nephr otoxic events leading to renal tubule tumor formation is plausible ( U.S. EPA, 1991c). These e ffects are typically not observed in female rats or other species due to the absence or minimal presence of the α 2u -globulin protein in these animals (Hard et al., 1993). Concluding that HCE i s acting through an α 2u -globulin-associated mode of carcinogenic action is precluded by evidenc e of nephrotoxic effects in female rats in two chronic studies (NTP, 1989; NCI, 1978) and in m ale and female mice in one chronic study (NCI, 1978). Nephropathy associated with α 2u -globulin is generally not observed in female rats or other species due to the absence or minimal pres ence of the α 2u -globulin protein in these animals (Hard et al., 1993).

Other Possible Modes of Action There is insufficient evidence to support an α 2u -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 ma le rats following HCE exposure. CPN is an age-related renal disease of laboratory rodents tha t 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 th e 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 controlsDO 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 epi thelium. Neither of these effects increased in a dose-related manner in the controls or the HCE-expo sed female rats. The treatment-related

effects in male and female rats serve as evidence t hat 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 Anim als The mode of action for the carcinogenic effects of HCE in the kidney is unknown. There

is insufficient evidence to establish the role of α 2u -globulin in HCE-exposure related nephropathy. Studies following short-term, subchro nic, and chronic exposure of male rats have reported renal lesions consistent with α 2u -globulin nephropathy (NTP, 1996, 1989; Gorzinski e t al., 1985; NCI, 1978). The formation of renal tubu le adenomas and carcinomas (preceded by hyperplasia) following chronic HCE exposure (NTP, 1 989) are also consistent with an

α 2u- globulin-related mode of action. However, the key event in the histopathological sequence of events demonstrating a α 2u -globulin-related mode of action (excessive accumul ation of hyaline droplets containing α 2u -globulin in renal proximal tubules) leading to dev elopment of kidney tumors in male rats exposed to HCE has not been ful ly characterized. None of the HCE studies performed the necessary immunohistochemical assays to confirm the presence of α 2u -globulin protein within the hyaline droplets observed follow ing administration of HCE (NTP, 1996, 1989). It is unknown if HCE is binding to α 2u -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 hypothesiz ed 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 HC E (NCI, 1978). The presence of kidney effects in HCE-exposed male and female mice, which generally do not accumulate the α 2u globulin protein, suggests that a mode of action ot her than α 2u -globulin nephropathy.

Role of metabolites Studies of proposed HCE metabolites such as PERC, p entachloroethane, and TCE

revealed similar noncancer and cancer effects. PER C exposure caused toxic nephropathy and kidney tumors (NCI, 1977); pentachloroethane exposu re resulted in chronic inflammation of the kidney, (NTP, 1983); and TCE exposure produced chro nic nephropathy (NCI, 1976), toxic nephrosis, kidney cytomegaly (NTP, 1990), renal tub ular adenomas (NTP, 1990), tubular cell cytomegaly, and toxic nephropathy (NTP, 1988). The se studies consistently demonstrate that theDO 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 no t possible to define the relative roles of the parental compound (HCE) or its metabolites in the k idney effects associated with exposure to HCE. However, the kidney effects of these potentia l HCE metabolites further support the kidney as the principal target organ of HCE exposure.

Relevance of the Hypothesized Mode of Action to Hum ans Generally, kidney tumors observed in cancer bioassa ys are assumed to be relevant for assessment of human carcinogenic potential. Howeve r, for male rat kidney tumors, when the mode-of-action evidence convincingly demonstrates t hat the response is secondary to

α 2u -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 α 2u -globulin mode of action for the following reasons: (1) the lack of α 2u -globulin immunohistochemical data for HCE-induced nephrotoxicity and carcinogenicity, (2) the hyaline droplet accumulation is caused by excessive protein load that may not be exclusive ly related to α 2u -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 α 2u -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 B6C3F 1 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 respon se. Both dose groups of female mice presented statistically significantly elevated inci dences of hepatocellular carcinoma compared with control groups but a dose response was not obs erved. The investigators did not find nonneoplastic liver effects (such as organized thro mbus, 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 e nzymes involved are phenobarbital-inducible CYP450s. These are primari ly localized in the liver. Although tissue-specific metabolism of HCE has not been stud ied 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 metab olized 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 rad icals during HCE metabolism, causing key events in the carcinogenic process such as cytotoxi city, 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 consid ered a key event in the carcinogenic process because rats in the available studies (NTP, 1989; N CI, 1978) have not displayed hepatocellular neoplastic endpoints. Although mice demonstrated h epatocellular carcinoma, nonneoplastic effects such as hepatocellular necrosis were not ob served (NCI, 1978). HCE-induced hepatocellular carcinomas in mice varied in microsc opic appearance (NCI, 1978). Some carcinomas were characterized by well-differentiate d hepatic cells with uniform cord arrangement, while others had anaplastic liver cell s with large hyperchromatic nuclei, often with inclusion bodies and vacuolated pale cytoplasm. Ar rangement of neoplastic liver cells also varied from short stubby cords to nests of cells, a nd occasional pseudo-acinar formations. Neoplasms in control mice did not vary in appearanc e 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 [ 14 C]HCE was consistently greater in mice compared with rats (Lattanzi et al., 1988). Bindin g to macromolecules was interpreted by the presence of radiolabeled carbon; however, radiolabe led carbon may have been incorporated into these macromolecules from intermediary HCE metaboli tes. In vitro binding studies using calf thymus DNA demonstrated that mouse liver cytosol (i nduced by phenobarbital) mediated more extensive DNA binding than rat liver cytosol (Latta nzi 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, inflammati on, and regenerative cell proliferation is a possible consequence of this binding in the li ver. The binding studies provide a line of evidence as to why the liver is the major carcinoge nic 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 w as evaluated in mice treated with HCE (Yoshikawa, 1996; Miyagawa et al., 1995). This stu dy reported ambiguous results; the lower HCE dose caused a statistically significant increas e in RDS, whereas the higher dose did not (Yoshikawa, 1996; Miyagawa et al., 1995). Rat live r foci experiments provide support for the hypothesis that HCE acts as a tumor promoter, not a s a tumor initiator (Milman et al., 1988; Story et al., 1986). The in vivo binding data suggest that HCE is seques tered 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 proposedDO 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 so lely on the presence of radiolabeled carbon; specific HCE metabolites were not identifie d. Therefore, this process does not take into account the possibility of normal biological mechan isms in which the radiolabeled carbon can be incorporated into the macromolecules via anabolic p rocesses. 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 h epatotoxicity 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 a ction for HCE-induced tumors of the adrenal gland. Therefore, the mode of action for p heochromocytomas observed following oral exposure to HCE is unknown.

4.8. SUSCEPTIBLE POPULATIONS AND LIFE STAGES No studies were located that address the susceptib ility 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 ch ildhood susceptibility to HCE-induced toxicity or carcinogenicity. Although it is unknow n 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 rep ortedly 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 yea rs with those aged >10-60 years and

concluded that factors other than maximal CYP450 ca talytic 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 feta l and neonatal livers indicate that CYP450 expression is similar to adult levels by a few mont hs 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 activ ities) and Phase II enzymatic activities are 1.3–1.5-fold higher in children (aged 1–16 years) c ompared with adults. Therefore, the extent to which variable age-related expression of CYP450 con tributes to childhood susceptibility is unknown. Research in developmental expression of C YP450 is ongoing; conclusions regarding the differential expression of CYP450 are premature . Considering the substantial portion ofDO 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 ar e more sensitive to HCE-induced nephrotoxicity than females (NTP, 1989; Gorzinski e t 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 fem ale rats, see Table 3-3), sex hormone differences, and/or gender difference in CYP450 act ivities. 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 popula tion. 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 a lso 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 CY P450 and GST polymorphisms and increases in cancer risks; however, none of these are specifi c for HCE.DO NOT CITE OR QUOTE 86 5. DOSE-RESPONSE ASSESSMENTS

5.1. ORAL REFERENCE DOSE (RfD)

5.1.1. Choice of Principal Study and Critical Effe ct—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 gavag e study in F344 rats (NTP, 1989), a 78-week gavage study in Osborne-Mendel rats (NCI, 1 978), a 91-week gavage study in B6C3F 1 mice (NCI, 1978), a 16-week feeding study in F344 r ats (Gorzinski et al., 1985), and a 13-week gavage study in F344 rats (NTP, 1989). The short-t erm 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 subchro nic and chronic durations. However, short-term studies in rats (NTP, 1996, 1989) were u sed to support findings in the subchronic and chronic studies. The available oral exposure studi es identified kidney or liver effects associated with exposure to HCE. Reported effects include tub ular nephropathy (NTP, 1989; NCI, 1978), atrophy and degeneration of renal tubules (NTP, 198 9; Gorzinski et al., 1985), slight hypertrophy and/or dilation of proximal convoluted renal tubule s (Gorzinski et al., 1985), linear mineralization of renal tubules (NTP, 1989), hyperp lasia of the renal pelvic transitional epithelium (NTP, 1989), and hepatocellular necrosis (NTP, 1989). In the NTP (1989) chronic study, HCE was administer ed 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 degenera tion and regeneration, tubular dilatation and atrophy, glomerulosclerosis, interstitial fibrosis, and chronic inflammation) was observed in HCE treated rats of both sexes. Nephropathy was al so reported in control rats of both sexes. Although a high incidence of nephropathy was observ ed in control rats, the study authors reported that the incidence of more severe nephropa thy 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 nephro pathy. EPA determined that the increased incidence of moderate or marked nephropathy in male s 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 nephropat hy that was greater than minimal nephropathy. EPA determined that the increased inc idences 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 hyperplasi a of the renal pelvic epithelium were increased in a dose-dependent, statistically significant mann er in the treated male rats. EPA determined

that the increased incidences of linear mineralizat ion of the renal papillae and hyperplasia of the renal pelvic epithelium were statistically signific ant in males at the 7 and 14 mg/kg-day doses (see Table 5-1). Considering the increased severit y of nephropathy following HCE exposure andDO NOT CITE OR QUOTE 87 dose-dependent increases in the incidence of minera lization of the renal papillae and hyperplasia of renal pelvic transitional epithelium in male rat s, 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 ra ts as kidney effects were observed at the

lowest doses tested. EPA considered the male rat L OAEL 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 (19 89) study. EPA considered the female rat LOAEL as 57 mg/kg-day, based on dose-related increa ses in incidence and severity of nephropathy in the NTP (1989) study. In the NCI (1978) chronic rat study, HCE was admini stered via gavage to groups of 50 male and 50 female Osborne-Mendel rats for 5 days/w eek, cyclically for 66 of the 78 weeks, followed by an observation period of 33-34 weeks (t otal of 112 weeks). The TWA doses of HCE were 113 and 227 mg/kg-day. Tubular nephropath y was observed in all groups of treated animals, but was not observed in either untreated o r 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 LO AEL as 113 mg/kg-day, based on a doserelated increase in incidence of nephropathy in bot h male and female rats. In the NCI (1978) chronic mouse study, HCE was admi nistered via corn oil gavage to groups of 50 male and 50 female B6C3F 1 mice for 5 days/ week for 78 weeks followed by an observation period of 12-13 weeks (total of 90 week s). Starting in week 9, the HCE doses were increased, though no explanation for the increase w as provided. The TWA doses of HCE were 360 and 722 mg/kg-day. Because of low survival rat es 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 B6C3F 1 mice administered 360 and 721 mg/kg-day HCE. EPA considered the LOAEL for this study as 360 mg/k g-day based on tubular nephropathy,

while a NOAEL could not established from these data

. In Gorzinski et al. (1985), HCE was administered (i n 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 dila tion 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 co ntrol rats of either sex or in HCE exposed female rats. EPA determined that increases in slig ht hypertrophy and/or dilation of the proximal convoluted renal tubules were statistically signifi cant in male rats treated with 15 orDO NOT CITE OR QUOTE 88 62 mg/kg-day HCE (see Table 5-1). Atrophy and dege neration of renal tubules was observed in

both male and female rats. EPA determined that inc reases in incidences of atrophy and degeneration of renal tubules were statistically si gnificant in male rats treated with 15 or 62 mg/kg-day HCE and in female rats fed 62 mg/kg-da y 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. E PA considered the female rat LOAEL as 62 mg/kg-day and the female rat NOAEL as 15 mg/kg-d ay, based on increased incidence of

renal tubule effects. In the NTP (1989) subchronic study, HCE was adminis tered 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 for mation, 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, bu t no data on severity were presented. No kidney effects were reported in female F344 rats ex posed 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 su bsequent dose-response assessment because of short study durations (12-day study [Wee ks 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 c hronic durations. The effects and the incidence of kidney and liver effects from the stud ies considered for selection as the principal study, serving as the basis for the derivation of t he 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, th ese 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 22/49 a (45%) Osborne-Mendel Male Rat Tubular Nephropathy

227 33/50 a (66%) 0 0/20 (0%) 113 9/50 a (18%) NCI (1978) 78 weeks

(gavage) Osborne-Mendel

Female Rat Tubular Nephropathy
227 29/49 a (59%) 0 18/50 (36%) 7 24/50 (48%) F344 Male Rat Moderate to

Marked Tubular Nephropathy 14 30/50 a (60%) 0 12/50 (24%) 57 25/50 a (50%) NTP (1989) 103 weeks

(gavage) F344 Female Rat Mild to Marked Tubular Nephropathy 114 32/49 a (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 7/50 a (14%) NTP (1989) 103 weeks

(gavage) F344 Male Rat Hyperplasia of the Renal Pelvic Transitional Epithelium 14 7 /50 a ( 1 4%) 0 0/10 (0%) 1 1/10 (10%) 15 7/10 a (70%) Gorzinski et al. (1985) 16 weeks

(dietary) F344 Male Rat Slight Hypertrophy and/or Dilation of Proximal Convoluted Renal Tubules 62 10/10 a (100%) 0 1/10 (0%) 1 2/10 (20%) 15 7/10 a (70%) F344 Male Rat Atrophy and

Degeneration of Renal Tubules 62 10/10 a (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 6/10 a (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 4/10 a (40%) NTP (1989) 13 weeks

(gavage)

F344 Female Rat Hepatocellular Necrosis

535.7 8/10 a (80%)

a EPA determined statistical significance using Fishe r’s Exact Test (p < 0.05)

These chronic and subchronic studies in rats and mi ce indicate that the kidney and liver are both target organs of HCE oral toxicity in rode nts. Given the number of effects reported inDO NOT CITE OR QUOTE 90 the kidney and the greater sensitivity of these eff ects in available studies, the kidney is the primary target of oral HCE exposure toxicity in rod ents. 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 (N TP, 1989), hyperplasia of the renal pelvic transitional epithelium in male F344 rats (NTP, 198 9), and tubular nephropathy in male and female Osborne-Mendel rats (NCI, 1978). Further co nsideration 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 s ubject to BMD modeling for comparison purposes. The data on the male rat liver effects f rom the NTP (1989) study were not considered because the incidence of hepatocellular necrosis wa s not significantly elevated above controls at any HCE dose. The kidney effects reported in the 1 3-week subchronic study (NTP, 1989) were

not further considered because the lack of the inci dence 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 t hose doses associated with kidney effects in other subchronic (Gorzinski et al., 1985) and chron ic (NTP, 1989) studies. The chronic study in

B6C3F 1 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 (Gorzin ski et al., 1985) or chronic (NTP, 1989; NCI, 1978) exposure. The ability of the chronic NTP (19 89) study to inform the effects observed at the lowest dose tested in the Gorzinski et al. (198 5) 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 benc hmark 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 a bove. A BMR of 10% extra risk was considered appropriate for derivation under the ass umption 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 femal e rats reported by NTP (1989), NCI (1978), and Gorzinski et al. (1985), as well as the inciden ce data for hepatocellular necrosis in female rats reported by NTP (1989). Details of the BMD do se-response modeling reported in Table 5-2DO NOT CITE OR QUOTE 91 are presented in Appendix B (Table B-1). In additi on, 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 re sults of the available data and the benchmark response (BMR) levels and the potential P ODs 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) BMDL 10

(mg/kgday)

Gorzinski et al. (1985) Slight hypertrophy

and/or dilation of proximal convoluted renal tubules Male Rats

(n = 10) 16 weeks

(dietary) Gamma a

0.99 20.88 1.22 0.710

Male Rats (n = 10) Gamma b 0.70 34.94 1.34 0.728 Gorzinski 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) Gamma d 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) Quantallinear e 0.87 205.90 3.20 1.88 NTP (1989) Increased Severity

of Tubular Nephropathy Female Rats

(n ≈ 50) 103 weeks

(gavage) Gamma f

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

a Gamma, Quantal-linear, and Weibull models had ident ical AIC, goodness of fit p-values, as well as BMD 10 and BMDL 10 values.

b Gamma, Multistage 1o, and Quantal-linear models had identical AIC, goodness of fit p-values, as well a s BMD 10 and BMDL 10 values.

c Multistage 1o and Quantal-linear models had identic al AIC, goodness of fit p-values, as well as BMD 10 and BMDL 10 values.

d Gamma, Multistage 1o, and Weibull models had identi cal AIC, goodness of fit p-values, as well as BMD 10 and BMDL 10 values.

e Quantal-linear and Multistage 1o models had identic al AIC, goodness of fit p-values, as well as BMD 10 and BMDL 10 values.

f Gamma, Quantal-linear, and Weibull models had ident ical AIC, goodness of fit p-values, as well as BMD 10 and BMDL 10 values.

AIC = Akaike’s Information Criteria (i.e., a measur e 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 m g/kg-day) is about 100-fold. Kidney effects (i.e., tubular nephropathy, linear m ineralization of the renal tubules, hyperplasia of the pelvic transitional epithelium, atrophy and degeneration of renal tubules, and slightDO NOT CITE OR QUOTE 92 hypertrophy and/or dilation of the proximal convolu ted renal tubules) observed in male rats resulted in lower candidate PODs than comparable ef fects in female rats. The most sensitive effect observed in male rats exp osed to HCE is slight hypertrophy and/or dilation of proximal convoluted renal tubule s (Gorzinski et al., 1985). However, the potential POD for slight hypertrophy and/or dilatio n of proximal convoluted renal tubules (i.e., 0.710 mg/kg-day) is nearly identical to the candida te POD for atrophy and degeneration of renal tubules (i.e., 0.728 mg/kg-day). As tubular nephro pathy 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 sev eral studies. Therefore, atrophy and degeneration of renal tubules was selected as the c andidate POD for this subchronic exposure study. As shown in Appendix B, the gamma, multista ge 1°, logistic, probit, Weibull models in BMDS (version 2.0) provided adequate fits to the in cidence 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 BMD 10 and BMDL 10 values were identical for the gamma, multistage 1°, and quantal -linear model; therefore, the model with the

lowest BMDL 10 was selected. The gamma, multistage 1°, and quant al-linear model had identical

BMDL 10 values. Therefore, the BMD 10 associated with a 10% extra risk for nephropathy i n male rats of 1.34 mg/kg-day, and the lower 95% conf idence limit on this BMD 10 (BMDL 10 ) of 0.728 mg/kg-day was selected as the potential POD f or these data. The tubular nephropathy in male rats observed in th e chronic exposure study (NTP, 1989) resulted in higher PODs than the atrophy and degene ration 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 tes ted in the chronic exposure study represented a LOAEL. Therefore, the Gorzinski et al. (1985) st udy was selected as the principal study and atrophy and degeneration of renal tubules in male r ats was selected as the critical effect. The

BMDL 10 of 0.728 mg/kg-day was selected as the POD and ser ves as the basis for the derivation of the oral RfD for HCE. This endpoint is supporte d by additional kidney effects associated with oral exposure to HCE and supports the weight of evi dence for HCE-associated nephrotoxicity.

5.1.3. RfD Derivation—Including Application of Unc ertainty Factors (UFs) The derivation of the RfD for atrophy and degenerat ion of renal tubules in male F344 rats from the Gorzinski et al. (1985) 16-week toxicity s tudy was calculated from the BMDL 10 of 0.728 mg/kg-day. The composite UF of 3,000 was com prised of the following:

• An interspecies UF (UF A ) of 10 was applied to account for the variability in extrapolating from rats to humans. Although the toxicokinetics h ave been minimally evaluated in animals, the toxicokinetics of HCE have not been fully chara cterized in either rats or humans.DO NOT CITE OR QUOTE 93 • An intraspecies UF (UF H ) of 10 was applied to adjust for potentially sensi tive human subpopulations in the absence of information on the variability of response to HCE in the human population. Current information is unavailab le 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 r ange 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 bec ause 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 increas e the severity of the observed kidney

effects or could result in similar effects at lower doses. For these reasons, subchronic-to-chronic UF (UF S ) 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 conside rations 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 i n the HCE toxicity database, including the lack of a multigenerational reproductive study. Th e database includes studies in laboratory animals, including chronic and subchronic dietary e xposure studies and two oral developmental toxicity studies. One of the availabl e oral developmental toxicity studies demonstrated that HCE exposure decreased gestationa l 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 mat ernal 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 dos e (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 dat abase 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 wa s 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-t erm, subchronic, and chronic oral exposure to HCE is renal toxicity. Table 5-3 prese nts the PODs for nephrotoxicity in male rats with applied uncertainty factors (UFs) and potentia l 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 exp osure 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 toxici ty, 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 referenc e values that might be derived from other studies are also presented. Only endpoints observe d 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 eff ects characterized by atrophy and degeneration of renal tubules and slight hypertrophy and/or dila tion of proximal convoluted tubules. Linear mineralization of the renal tubules, hyperplasia of the pelvic transitional epithelium, slight hypertrophy and/or dilation of the proximal convolu ted 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 minera lization was only observed in male rats. Kidney effects were observed in male rats in the Go rzinski 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 t he Gorzinski et al. (1985) study is limited because the lowest dose tested in the NTP (1989) ch ronic exposure study represented a LOAEL. The potential POD associated with atrophy and degen eration 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 atroph y and degeneration of renal tubules from the Gorzinski et al. (1985) study was selected to s erve as the basis for the derivation of the RfD.DO NOT CITE OR QUOTE 95 Table 5-3. Potential points of departure (PODs) fo r nephrotoxicity in male rats with applied uncertainty factors (UFs) and pot ential reference values

Potential PODs (mg/kg-day) Total UF UF A UF H UF S UF D 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)DO NOT CITE OR QUOTE96Figure 5-1. Array of potential points of departurewith applied uncertainty factors and potential reference values fornephrotoxic effects of studies in Table 5-3.Increasedincidence ofatrophy anddegeneration oftubulesBMDLGorzinski etal., 198516-wkPoint of DepartureUFA,InterspeciesUFH, IntraspeciesUFS, Subchronic toChronicUFD, DatabaseReference DoseIncreasedincidence ofmoderate tomarkedtubularnephropathyBMDLNTP, 19892-yrIncreasedincidencelinearmineralizationBMDLNTP, 19892-yrIncreasedincidence oftubularnephropathyBMDLNCI, 197891-wkIncreasedincidence ofhyperplasia ofthe pelvictransitionalepitheliumBMDLNTP, 19892-yrIncreasedincidence ofslighthypertrophyand/or dilationof convolutedrenal tubulesBMDLGorzinski etal., 198516-wk0.00010.0010.010.1110100

Dose (mg/kg-day)DO NOT CITE OR QUOTE 97 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 NO AEL/LOAEL approach. In this study, the identified LOAEL for atrophy and degeneration of re nal 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 (UF A = 10), (2) intraspecies variation (UF H = 10), and (3) subchronic-to-chronic extrapolation (UF S = 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 Effe ct—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 hea lth effects and the exposure was likely a

mixture of HCE and zinc oxide. There are no chroni c studies available, and only a single subchronic inhalation study (in four species) which included a developmental toxicity experiment is available. Weeks et al. (1979) expos ed Sprague-Dawley rats, male beagle dogs, male Hartley guinea pigs, and Japanese quail to HCE air concentrations of 145, 465, or 2,517 mg/m 3 for 6 hours/day, 5 days/week, for 6 weeks. Postex posure observations were carried out

for 12 weeks. As discussed in Section 4.4.3.2, toxic effects obse rved in treated rats, dogs, and guinea pigs (the quail did not show signs of toxicity) wer e at the highest exposure level, 2,517 mg/m 3

, except for dams in of the 465 mg/m 3 exposure group in the developmental study; which exhibited significantly decreased body weight gain and an increased incidence (85%) of mucopurulent nasal exudate. This inflammatory exud ate was observed in 100% of the dams treated with 2,517 mg/m 3 . Similar to the dams, male and female rats expose d to 2,517 mg/m 3 HCE for 6 weeks exhibited mucopurulent exudate in t he nasal turbinates. Excess mucus in the nasal turbinates was also observed in 2/10 quail in the 2,517 mg/m 3 concentration group. Effects

of this nature were not observed in the 465 or 145 mg/m 3 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/m 3 quails was a transient effect of the HCE exposure because there was no evidence of inflammatory cells or tissue damage. The authors attributed the incre ased 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-rela ted lesions in the nasal turbinates (mucopurulent exudate), trachea (lymphoid hyperplas ia in the lamina propria), and lung (pneumonitis) of 2,517 mg/m 3 male and female rats. Similar lesions characterist ic of respiratoryDO NOT CITE OR QUOTE 98 mycoplasmosis in rodents were detected in an oral d evelopmental study in rats that paralleled the inhalation developmental study described above (bot h conducted by Weeks et al., 1979). Irritation of the upper respiratory tract was obser ved 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 pneumon itis. The presence of the infection in the rats in both t he oral and inhalation studies and in the controls of the oral study suggests that respirator y tract effects are a potentiation of the underlying mycoplasia infection rather than a direc t result of HCE exposure. Additionally, the reduced weight gain in the rats is likely related t o the condition of the infected animals, considering mycoplasma-infected rodents generally g ain less weight or lose weight compared with noninfected rodents (Xu et al., 2006; Sandsted t et al., 1997). Reduced weight gain was also observed in the 2,517 mg/m 3 guinea pigs, but mycoplasma infection was not repor ted (Weeks et al., 1979). Like rats and mice, guinea pigs can ca rry 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 lun gs. Therefore, the respiratory tract effects cannot be excluded from consideration as a potentia l critical effect. As discussed in Section 4.4.3, neurobehavioral effe cts were consistently observed in the rats and dogs exposed to 2,517 mg/m 3 . The male and female rats in the 6-week study exh ibited tremors and ruffled pelt. The pregnant rats develo ped tremors on GDs 12-16. Similarly, the dams exposed to 500 mg/kg HCE in the concurrent ora l 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, s evere head bobbing, facial muscular fasciculations, and closed eyelids. These effects were noted in the dogs throughout the study, although they disappeared overnight during nonexpos ure time periods. Supporting data for the study was reported in an a cute study by Weeks and Thomasino (1978), in which a single 8-hour inhalation exposur e to 2,500 or 57,000 mg/m 3 HCE and a single 6-hour exposure to 17,000 mg/m 3 HCE in male Sprague-Dawley rats resulted in neurol ogical and lung effects. The male rats exposed to 57,000 mg/m 3 HCE had reduced body weight gain compared with controls over the 14 days postexposur e. 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 diffus e interstitial pneumonitis of minimal to moderate severity and vascular congestion associate d with these lung effects. Following 6 hours of exposure to 17,000 mg/m 3 , the six rats in this group showed reduced weight gain compared

with controls and two of these rats exhibited a sta ggered 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. (19 79), as the only repeated exposure study available, was selected as the principal stud y for the derivation of the RfC. This studyDO NOT CITE OR QUOTE 99 used three concentrations and incorporated a variet y 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 inform ation 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 ident ified neurotoxicity, statistically significant decreases in body weight gain, and upper and lower respiratory tract irritation. The responses were generally observed following exposure to the h ighest 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 derivatio n of the RfC. Therefore, neurological effects following inhalation exposure to HCE were selected as the critical effect.DO NOT CITE OR QUOTE 100 Table 5-4. Noncancerous effects observed in animal s 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 expos ure groups (145, 465, and 2,517 mg/m 3 ) plus a control. Neurological effects were observ ed 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. Applica tion of BMD modeling was precluded because 100% of the high dose animals displayed neurologica l effects. Therefore, a NOAEL served as the POD. The NOAEL of 465 mg/m 3 , identified in Weeks et al. (1979), was selected a s the POD

for the derivation of the RfC based on effects in m ale 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 e xposures in male and female rats for 6 hours/day, 5 days/week. Thus, prior to deriv ing the RfC, this POD was adjusted for Species Dose/Duration NOAEL (mg/m 3 ) LOAEL

(mg/m 3 ) Effect

Sprague-Dawley rats (25/sex/dose) 465 mg/m 3 2,517 mg/m 3 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/m 3 2,517 mg/m 3 Tremors, ataxia, hypersalivation, head bobbing, facial muscular fasciculations

Male Hartley guinea pigs (10/dose)Male Hartley guinea pigs (10/dose) 465 mg/m 3 2,517 mg/m 3 Reduced body weight, increased relative liver weight

Coturnix Japonica quail (20/dose)Coturnix Japonica quail (20/dose) 0, 145, 465, or

2,517 mg/m 3 ; 6 weeks

2517 mg/m 3 Not established No effects observed

Pregnant Sprague-Dawley rats (22/dose) 0, 145, 465,

or 2,517 mg/m 3 on GDs 6-16 Maternal:

465 mg/m 3 Maternal:

2,517 mg/m 3 Maternal: tremors

Fetal: no effectsDO NOT CITE OR QUOTE 101 continuous exposure (24 hours/day, 7 days/week). T he duration-adjusted POD (POD [ADJ] ) is derived using the following equation (U.S. EPA, 199 4b):

POD [ADJ] = (POD) × (hours of exposure/24 hours) × (days of exposure/7 days) = (465 mg/m 3 ) × (6/24 hours) × (5/7 days) = 83.0 mg/m 3

.

The Methods for Derivation of Inhalation Reference Conc entrations and Application of Inhalation Dosimetry (hereafter referred to as the RfC Methodology) rec ommends converting the

POD [ADJ] to a human equivalent concentration (HEC) (U.S. EP A, 1994b). The RfC Methodology separates gases into three categories b ased 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 eff ects 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 Categor y 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 sp ecies 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 (H b/g ) A and the human blood:gas partition coefficient (H b/g ) H .

POD [HEC] = POD [ADJ] × (H b/g ) A /(H b/g ) H

However, the human and animal blood partition coef ficients 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 NO AEL [HEC] of 83.0 mg/m 3

.

POD [HEC] = POD [ADJ] × (H b/g ) A /(H b/g ) H = 83.0 mg/m 3 × 1 = 83.0 mg/m 3

5.2.3. RfC Derivation—Including Application of Unc ertainty Factors (UFs)DO NOT CITE OR QUOTE 102 The NOAEL [HEC] value of 83 mg/m 3 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 (UF A ), an UF of 3 was applied to account for the uncertainty in extrapolating from laborator y animals to humans. This value is adopted by convention, where an adjustment from an animal-specific NOAEL ADJ to a

NOAEL HEC has been incorporated. Application of an UF of 10 would depend on two areas of uncertainty (i.e., toxicokinetic and toxic odynamic uncertainties). In this assessment, the toxicokinetic component associated with HCE is mostly addressed by the

determination of an HEC as described in the RfC met hodology (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 huma ns.

• An intraspecies UF (UF H ) of 10 was applied to account for potentially sens itive human subpopulations in the absence of information on the variability of response to HCE in the human population. Information is currently unavail able to assess human-to-human variability in HCE toxicokinetics and toxicodynamic s.

• A subchronic-to-chronic UF (UF S ) of 10 was applied to account for the use of the P OD selected following a short-term duration of exposur e 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 i n the toxicity database on inhalation exposure to HCE. There are no available human occu pational or epidemiological studies of inhalation exposure to HCE. There are no chroni c 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 perform ed a developmental study and an acute study in rats. Maternal toxicity was observed at b oth doses. Fetuses of HCE-treated dams did not exhibit any significant skeletal or so ft tissue anomalies. The toxic effects observed in the dams in the developmental study wer e similar to those observed in the rats exposed for 6 weeks, although additional effec ts were observed in the rats exposedDO NOT CITE OR QUOTE 103 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, endp oints of concern based on the available inhalation data. Therefore, in considera tion of the inhalation database for HCE, a database UF of 10 was applied.

Given the UFs established above, the RfC for HCE wa s calculated employing the following equation:

RfC = NOAEL [HEC] ÷ UF = 83 mg/m 3 ÷ 3,000 = 0.028 mg/m 3 or 3 × 10 –2 mg/m 3

.

5.2.4. RfC Comparison Information The predominant noncancer effect of short-term inh alation 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 (li ver or kidney) weights in male guinea pigs, male and female rats, and pregnant rats. As discus sed 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 t he mycoplasma infection on the other endpoints. Based on the lack of alternative endpoi nts 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 S ection 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 a s 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. (1 979) administered HCE to rats by inhalation at 145, 465, or 2520 mg/m g 3 , 6 hours/day during gestation. At the two highest doses, maternal toxicity was observed but t here was no evidence of fetoxicity or teratogenicity. No additional discussion was prese nted 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 CONCENTRATIONDO NOT CITE OR QUOTE 104 The following discussion identifies uncertainties associated with the quantification of the RfD and RfC for HCE. Following EPA practices and g uidance (U.S. EPA, 1994b, 1993), the UF approach was applied to the chosen PODs to derive a n RfD and RfC (see sections 5.1.3 and 5.2.3). Factors accounting for uncertainties assoc iated with a number of steps in the analyses were adopted to account for extrapolating from an a nimal study to human exposure, a diverse human population of varying susceptibilities, and d atabase 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 inhalati on exposure to HCE in this study is mainly neurotoxicity. Critical data gaps have been identi fied in Chapter 4 and uncertainties associated with data deficiencies are more fully discussed bel ow. The RfD was derived from a BMDL 10 of 0.728 mg/kg-day, which was based on the observation of atrophy and degeneration of renal tu bules in male F344 rats from the Gorzinski et al. (1985) 16-week toxicity study. The dose-respon se relationships for oral exposure to HCE and nephropathy in other studies of rats are also avail able for deriving an RfD, but are associated with higher NOAELs/LOAELs that are less sensitive t han would be protected by the selected critical effect and corresponding POD. After consi deration 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 BMDL 10 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 t he range of experimental data. However, the selected models do not represent all possible m odels one might fit, and other models could be selected to yield different results, both higher an d lower than those included in this assessment. Uncertainty exists in the selection of the BMR leve l utilized in the BMD modeling of the critical effect (atrophy and degeneration of renal tubules i n male F344 rats) to determine the POD. In

the absence of clear information to determine the l evel of change in atrophy and degeneration of renal tubules in male F344 rats related to a biolog ically 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 chan ge. The RfC was derived from NOAEL [HEC] value of 83 mg/m 3 for evidence of neurotoxicity in Sprague-Dawley rats from a short-term (6 week) i nhalation study by Weeks et al. (1979). A POD based on a NOAEL or LOAEL is, in part, a reflec tion of the particular exposure concentration or dose at which a study was conducte d. It lacks characterization of the doseresponse curve and for this reason is less informat ive than a POD obtained from benchmark dose-response modeling. The short-term inhalation study in rats (Weeks et al., 1979) wasDO NOT CITE OR QUOTE 105 selected as the principal study and neurotoxicity w as identified as the critical effect. A NOAEL of 465 mg/m 3 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 extrapol ated to human response. Pharmacokinetic models are useful for examining species differences in pharmacokinetic processing; however, dosimetric adjustment using pharmacokinetic modelin g 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-f old 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

NOAEL ADJ to a NOAEL HEC has been incorporated. Application of a full UF of 10 would depend on two areas of uncertainty (i.e., toxicokinetic an d toxicodynamic uncertainties). In this assessment, the toxicokinetic component is mostly a ddressed 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 hum an variation needs consideration in extrapolating dose from a subset or smaller sized p opulation, say of one sex or a narrow range of life stages typical of occupational epidemiologic s tudies, 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-sp ecific data to examine the potential magnitude of over- or under-estimation are unavaila ble. Uncertainties associated with data gaps in the HCE database have been identified. Data more fully characterizing potential multigeneration al 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 e xposure studies and two oral developmental toxicity studies. The developmental studies show ef fects 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 H CE. There are no standard chronic toxicity or multigeneration reproductive toxicity animal stu dies 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 developm ental study and an acute study in rats. The developmental study in rats did not provide any evi dence of teratogenic effects. However, thisDO NOT CITE OR QUOTE 106 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 (cri tical 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 m ale and female B6C3F 1 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 ob served following [ 14 C]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/mal ignant pheochromocytomas combined in male F344/N rats (NTP, 1989), and statistically sig nificant increases in the incidence of hepatocellular carcinomas in male and female B6C3F 1 mice (NCI, 1978).

5.4.1. Choice of Study/Data—with Rationale and Jus tification Two animal studies were selected for BMD analysis and subsequent quantitative cancer assessment. In the first study, NTP (1989) reporte d statistically significantly elevated incidences of renal adenomas and carcinomas combined and pheoc hromocytomas, malignant pheochromocytomas, and complex pheochromocytomas co mbined in male F344 rats administered HCE via gavage for 2 years. Female ra ts 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 sex es of B6C3F 1 mice administered HCE via gavage for 78 weeks. However, male mice in this st udy 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 pheochr omocytomas and malignantDO NOT CITE OR QUOTE 107 pheochromocytomas combined; 28, 58 (statistically s ignificant), 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 B6C3F 1 mice for 91-weeks. Statistically significant increases in the incidence of hepatocellular carcinomas were observed in 15, 30, and 63% of males and 10, 4 0, 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 st udies, conducted in both sexes of two species with 50 animals/sex/dose. Each study u tilized 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 elevate d over controls at two sites in rats (NTP, 1989) and at one site in mice (NCI, 1978). Some li mitations 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 ad just the tumor incidences for survival before BMD modeling. The cancer incidence data are summar ized in Table 5-5.

Table 5-5. Summary of incidence data in rodents or ally 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%) 7.1 2/50 (4%) NTP (1989) Male F344 rats 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 B6C3F 1 mice Hepatocellular carcinoma 722 31/49 (63%) a 0 2/20 (10%) 360 20/50 (40%) a NCI (1978) Female B6C3F 1 mice Hepatocellular

carcinoma 722 15/49 (31%) a

_

a Denotes statistical significance

5.4.3. Dose Adjustments and Extrapolation Methods The HCE doses administered to laboratory animals w ere scaled to human equivalent doses (HEDs) according to EPA guidance (U.S. EPA, 2 005a, 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 day

mg Dose

BW day mg Dose

4 /3 ]

[

4 /3 ]

[ / /

=DO NOT CITE OR QUOTE 108 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 HED s:

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 empl oyed 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 e mployed 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-respo nse modeling described below.

The multistage model was the primary model consider ed for fitting the dose-response data and is given by:

P(d) = 1 – exp[–(q 0 + q 1 d + q 2 d 2 + ... + q k d k )],

where:

P(d) = lifetime risk (probability) of cancer at dos e d

q i = parameters estimated in fitting the model, i = 1 , ..., k

And extra risk is defined as (P(d) –P(0))/(1-P(0)).DO NOT CITE OR QUOTE 109 The multistage model in BMDS (version 2.0) (U.S. E PA, 2008) was used to fit the multistage dose-response model to the incidence dat a summarized in Table 5-5 using the calculated HEDs in order to derive an oral slope fa ctor for HCE. The BMR selected was the

default value of 10% extra risk recommended for dic hotomous models (U.S. EPA, 2000b). No data were excluded from the BMD multistage modeling

. As stated above, the multistage model was fit to th e 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 fi t. The multistage model was also fit to the incidence of pheochromocytomas or malignant pheochr omocytomas in male rats. The model exhibited a significant lack of fit for the pheochr omocytomas (according to the χ 2 statistic with

p < 0.01). Thus, this dataset was not useful for do se-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 mo deling results for the kidney and liver tumors in rats and mice, respectively, are summariz ed in Table 5-6, with more detailed results contained in Appendix B.

Table 5-6. Summary of BMD modeling results for ora l cancer assessment of HCE

Study Sex/strain/species Endpoint “Best-fit” Model p value AIC BMR BMD 10 BMDL 10 or POD Oralr

Slope Factor(mg/kgday) –1

NTP (1989) Male F344 rats Renal adenomas/carcinomas combined 2°

Multistage 0.75 71.1

9 0.1 3.73 2.44 0.040984

NCI (1978) Male B6C3F 1 mice Hepatocellular carcinomas 2°

Multistage 0.83 146.

47 0.1 37.03 14.44 0.006925

NCI (1978) Female B6C3F 1

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 actio n 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. E PA, 2005a). In the case of HCE, the mode of carcinogenic action of HCE in the kidneys and liver s of rats and mice is unknown. There is some data in experimental animals evaluating α 2u -globulin accumulation and toxicity in the kidney. As described in Section 4.7.3.1., two prin cipal factors contribute to the conclusion that the available data do not support an α 2u -globulin mode of action for the development of ren al tumors: (1) the lack of information identifying the α 2u -globulin protein in HCE-treated rats, and (2) evidence of nephropathy in female rats as well as male and female mice (because the α 2u -DO NOT CITE OR QUOTE 110 globulin-related mode of action is specific for mal e rats). Therefore, a linear low-dose extrapolation approach was used to estimate human c arcinogenic risk associated with HCE

exposure.

5.4.4. Oral Slope Factor and Inhalation Unit Risk The oral slope factor was derived from the BMDL 10 (the lower bound on the exposure associated with a 10% extra cancer risk) by dividin g the BMR by the BMDL 10 and represents an upper bound on cancer risk associated with a contin uous 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 BM DL 10 of 2.44 mg/kg-day into 0.1 (10%) (Appendix B). An oral slo pe factor for hepatocellular tumors in male mice of 0.007 (mg/kg-day) -1 was calculated by dividing the human equivalent BM DL 10 of 14.44 mg/kg-day into 0.1 (10%) (Appendix B). An oral slo pe factor for hepatocellular tumors in female mice of 0.0007 (mg/kg-day) -1 was calculated by dividing the human equivalent BM DL 10 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 repres ent upper-bound estimates. The rats exhibited greater sensitivity to HCE-induced carcin ogenicity than the mice. Thus, the risk estimate associated with the male rats that develop ed 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 carcino genicity of HCE via the inhalation route,

an inhalation unit risk has not been derived in thi s evaluation.

5.4.5. Uncertainties in Cancer Risk Values Extrapolation of data from animals to estimate pot ential 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 Sec tion 5.4.5.1 and in Table 5-7.DO NOT CITE OR QUOTE 111 Table 5-7. Summary of uncertainties in the HCE can cer risk assessment

Consideration/ approach Impact on oral

slope factor a 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 BW 2/3 ]) BW 3/4 (default

approach) There are no data to support alternatives.

Because the dose metric was not an AUC, BW 3/4 scaling was used to calculate equivalent cumulative exposures for estimating equivalent human risks.

Statistical uncertainty at POD ↓ oral slope factor

1.5-fold if BMD 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; ↓ = decreaseDO NOT CITE OR QUOTE 112 5.4.5.1. Sources of Uncertainty

Relevance to humans The modes of action for the kidney (adenomas/carcin omas) 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 exp erimental animals evaluating α 2u -globulin accumulation and toxicity in the kidney. As descri bed in Section 4.7.3, two principal factors

contribute to the conclusion that the available dat a do not support an α 2u -globulin mode of action for the development of renal tumors. First, the pr esence of kidney effects in HCE-exposed male and female mice, which generally do not accumulate the α 2u -globulin protein, suggests that a

mode of action other than α 2u -globulin nephropathy. Second, none of the HCE stu dies performed the necessary immunohistochemical assays to confirm the presence of α 2u -globulin protein within the hyaline droplets observed follow ing administration of HCE (NTP, 1996, 1989). This represents a data gap, as the presence of α 2u -globulin is necessary to support an

α 2u -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 f or hepatic cancer associated with HCE exposure. The experimental animal literature, howe ver, shows that oral exposure to HCE

induces liver tumors in male and female mice. It i s 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 meta bolism, causing key events in the carcinogenic process such as cytotoxicity, inflamma tion and regenerative cell proliferation. Limited information exists to distinguish the simil arities and differences between experimental

animals and humans in terms of HCE metabolism or to xicity. However, these potential key events have not been systematically evaluated for H CE. In humans, pheochromocytomas are rare catecholamine -producing neuroendocrine tumors that are usually benign, but may also presen t as or develop into a malignancy (Eisenhofer et al., 2004; Lehnert et al., 2004; Edstrom Elder e t 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 absen ce of information indicating otherwise, the kidney and adrenal gland tumors in male rats and li ver tumors in male and female mice are

considered relevant to humans.

Bioassay selection The study by NTP (1989) was used for the developme nt of an oral slope factor. This study was conducted in both sexes of F344/N rats an d 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 exami ned weekly (for 14 weeks) then monthlyDO NOT CITE OR QUOTE 113 for body weight and monthly for feed consumption. Animals were necropsied and all organs and tissues were examined grossly and microscopically f or histopathological lesions for a comprehensive set of toxicological endpoints in bot h 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 ma le or female mice to the carcinogenicity of HCE. The oral slope factor calculated from male ra ts 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 human s, the most sensitive group was selected for the basis of the oral slope factor. Though the mod e of action for observed kidney tumors in rodents is unknown, the evidence suggesting the kid ney as a target organ of HCE toxicity in both species lends strength to the concern for human car cinogenic potential.

Dose metric HCE is likely metabolized to PERC and pentachloroet hane; however, it is unknown whether a metabolite or some combination of parent compound and metabolites is responsible for the observed toxicity and carcinogenicity of HC E. 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 carcinogeni city. On the other hand, if this is not the most relevant dose metric, then the impact on the h uman equivalent slope factor is unknown; the low-dose cancer risk value may be higher or lower t han that estimated, by an unknown amount.

Choice of low-dose extrapolation approach The mode of action is a key consideration in clari fying 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 th e overall uncertainty in low-dose risk

estimation could be reduced if the mode of action f or 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 combi ned across sites prior to modeling, in order to allow for the possibility that different t umor types can have different dose-response relationships because of varying time courses or ot her 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 d ata.DO NOT CITE OR QUOTE 114 Choice of model All risk assessments involve uncertainty, as study data are extrapolated to make inferences about potential effects in humans from e nvironmental 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 rod ents exposed to levels greater than would occur from environmental exposures.

Without human cancer data or better mechanistic da ta, 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 potenti ally carcinogenic to humans as well. However, the lack of concordance in tumor sites bet ween the two rodent species makes it more difficult to quantitatively estimate human cancer r isk. Regarding low-dose extrapolation, in the absence o f mechanistic data for biologically based low-dose modeling or mechanistic evidence sup porting a nonlinear approach (see the discussion at the beginning of Section 5.4.3), a li near low-dose extrapolation was carried out

from the BMDL 10 . It is expected that this approach provides an up per 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, 20 05a) and extrapolating from a POD in the observable range is designed in part to minimize mo del dependence. Furthermore, the multistage model used provided an adequate fit to a ll the datasets. The ratios of the BMD 10 values to the BMDL 10 values give some indication of the uncertainties i n 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 o ther assessments.

Cross-species scaling An adjustment for cross-species scaling (BW 3/4 ) was applied to address toxicological equivalence of internal doses between rats and huma ns, consistent with the Guidelines for Carcinogen Risk Assessment (US EPA, 2005a). It is assumed that equal risks r esult from equivalent constant lifetime exposures.

Human population variability The extent of inter-individual variability or sens itivity to the potential carcinogenicity of HCE is unknown. There are no data exploring whethe r there is differential sensitivity to HCE carcinogenicity across life stages. In addition, n either 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 HC E include variations in CYP450 levelsDO NOT CITE OR QUOTE 115 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 H CE or antioxidant protection systems. This lack of understanding about potential susceptibilit y differences across exposed human populations thus represents a source of uncertainty . Humans are expected to be more heterogenous than laboratory animals, and this vari ability is likely to be influenced by ongoing or background exposures, diseases, and biological p rocesses.

5.4.6. Previous Cancer Assessment The previous HCE cancer assessment by U.S. EPA rep orted 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. T he current risk value is derived from the

incidence of renal adenomas or carcinomas in male r ats (NTP, 1989), resulting in an approximately 2.8-fold higher than the oral slope f actor 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 b ody weight instead of the previous 2⁄3 power of body weight.DO NOT CITE OR QUOTE 116 6. MAJOR CONCLUSIONS IN THE CHARACTERIZATION OF H AZARD

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 t o 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 u sed as a polymer additive, moth repellant, plasticizer for cellulose esters, insecticide solve nt, 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 tissu e. Toxicokinetic studies in animals indicated

that HCE is also localized and metabolized in the l iver and kidney. Kidney concentrations of HCE were higher in male rats than female rats (Gorz inski et al., 1985; Nolan and Karbowski, 1978). Studies of HCE metabolism indicated that th e 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., 1 982, 1981). HCE is putatively metabolized via a pentachloroethyl free radical to PERC and pen tachloroethane. Pentachloroethane is then

metabolized to TCE. TCE and PERC are further metab olized by hepatic oxidation to several urinary metabolites including TCA, trichloroethanol , oxalic acid, dichloroethanol, dichloroacetic acid, and monochloroacetic acid (Mitoma et al., 198 5; 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 metaboli tes. However, several of these metabolites have demonstrated liver and kidney toxicities simil ar to HCE. The kidney has consistently been shown as the targe t for toxicity in acute, subchronic, and chronic toxicity bioassays in animals (NTP, 199 6, 1989; Gorzinski et al., 1985; NCI, 1978). Noncancer effects include kidney degeneration (tubu lar nephropathy, necrosis of renal tubular epithelium, hyaline droplet formation, tubular rege neration and tubular casts) and hepatocellular necrosis. Hepatotoxicity was noted in animals expo sed to HCE, although endpoints of this nature have not been evaluated in laboratory animal s as fully as the renal effects. Hepatocellular necrosis was reported in female rats (NTP, 1989), b ut was not evaluated in a chronic exposure study of mice (NCI, 1978). The mouse study (NCI, 1 978) focused on tumorigenic endpoints and

not noncancer effects. There is no information available describing the me tabolism of HCE following exposure

via inhalation. The inhalation database for HCE co ntains one acute (Weeks and Thomasino,DO NOT CITE OR QUOTE 117 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 incre ased 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 h epatocellular 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 carcinogenicit y in humans exposed to HCE are unavailable. The carcinogenicity incidence data in male rats (NT P, 1989) were used to develop quantitative cancer risk assessments for HCE. The consistency o f the kidney and liver as target organs in different species for HCE distribution and metaboli sm, and both noncancer and cancer endpoints, provides support for the evaluation of these endpoi nts 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 nephropath y, atrophy and degeneration of renal tubules, and hepatocellular necrosis. In female ra ts, tubular nephropathy, atrophy and degeneration of the renal tubules, and hepatocellul ar necrosis were observed in a statistically significant dose-response manner (NTP, 1989; Gorzin ski et al., 1985; NCI, 1978). Tubular nephropathy, severity of nephropathy, and atrophy a nd 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 effec ts 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 Gor zinski et al. (1985) in male rats. Gorzinski et al. (1985) was selected as the principal study a nd atrophy and degeneration of renal tubules in

male rats were chosen as the critical effect for th e derivation of the oral RfD. This study included both sexes of F344 rats, 10 animals/sex/do se, and three dose groups plus controls (0, 1, 15, and 62 mg/kg-day). Dose-response analyses of t he noncancer endpoint, atrophy and degeneration of renal tubules (Gorzinski et al., 19 85), using EPA’s BMDS, resulted in a POD of 0.728 mg/kg-day. A composite UF of 3,000 was appli ed 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 pl us a control. NTP (1989) also conducted 16-day, 13-week, and 103-week studies that supporte d the results observed in the 16-week study.DO NOT CITE OR QUOTE 118 Application of BMD modeling provided a POD upon whi ch to base the derivation of the RfD.

The critical effect on which the RfD is based is we ll-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, an d chronic toxicity studies and developmental toxicity studies in rats and chronic carcinogenicit y bioassays in rats and mice. The database lacks a multigenerational reproductive study and st udies 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 s ingle 6-week repeat-exposure study by Weeks et al. (1979). This study reported a NOAEL o f 465 mg/m 3 and a LOAEL of 2,517 mg/m 3 in several species including Sprague-Dawley rats, m ale 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/m 3 was selected to serve as the POD. Adjustments for continuous exposure and for the HEC, resulted in th e POD [HEC] of 83 mg/m 3 . An UF of 3,000 was applied to derive an inhalation RfC of 3 × 10 -2 mg/m 3 . 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. (19 79), 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/m 3 ) plus a control. The study is limited by the relatively short exposure d uration (6 weeks) and minimal reporting of effects, especially quantitative changes. Applicat ion of BMD modeling was precluded based on a 100% response in animals for the neurological eff ects 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 o ne 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. H CE 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 onDO NOT CITE OR QUOTE 119 statistically significant increased incidences of r enal adenomas and carcinomas and adrenal pheochromocytomas and malignant pheochromocytomas i n male rats. This study was used for development of an oral slope factor. This was a we ll-designed study, conducted in both sexes of F344 rats with 50 rats/sex/dose; typical of carcino genicity 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 nec ropsied and all organs and tissues were examined grossly and microscopically for histopatho logical lesions for a comprehensive set of toxicological endpoints in both sexes. Renal adeno mas and carcinomas and pheochromocytomas and malignant pheochromocytomas observed in male ra ts (NTP, 1989) were not seen in female rats or other species orally-exposed to HCE. Hepat ocellular carcinomas were observed in male and female mice, but not in the rats. The male B6C 3F 1 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 t he 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 ext rapolation as a default approach

when the available data are insufficient to establi sh 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 followi ng oral exposure to HCE is unknown. The

database for HCE lacks information on the mode of a ction and the shape of the curve in the region below the POD; therefore, a linear extrapola tion was performed in determining the oral slope factor in the derivation of a quantitative es timate of cancer risk for ingested HCE.

Increased incidence of renal adenomas and carcinoma s 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 t o estimate potential cancer risk in human populations introduces uncerta inty in the risk estimation for HCE. Uncertainty can be considered quantitatively; howev er, some uncertainty can only be addressed qualitatively. For this reason, an overall integra ted quantitative uncertainty analysis cannot be developed. However, EPA’s development of the cance r quantitative assessment for HCE included consideration of potential areas of uncert ainty. The following summarizes these

considerations. A biologically-based model was not supported by th e available data; therefore, a multistage model was the preferred model. The mult istage model can accommodate a wide variety of dose-response shapes and provides consis tency with previous quantitative dose-DO NOT CITE OR QUOTE 120 response assessments for cancer. Linear low-dose e xtrapolation 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 (BW 3/4 ) was applied to address toxicological equivalence of internal doses between rats and humans based on the assumption that equal risks result fro m equivalent constant lifetime exposures.

An inhalation unit risk was not derived in this as sessment. Data on the carcinogenicity of the compound via the inhalation route are unavailab le, 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 h umans by the inhalation route since the compound is absorbed and, in oral studies, induces tumors at sites other than the portal of entry.DO NOT CITE OR QUOTE 121 7. REFERENCES

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Younglai, EV; Foster, WG; Hughes, EG; et al. (2002) Levels of environmental contaminants in human foll icular fluid, serum, and seminal plasma of couples undergo ing in vitro fertilization. Arch Environ Contam To xicol 43(1): 121-126.A- DRAFT – DO NOT CITE OR QUOTE 1 APPENDIX A: SUMMARY OF EXTERNAL PEER REVIEW AND PU BLIC

COMMENTS AND DISPOSITION

PAGE INTENTIONALLY LEFT BLANKB-1 DRAFT – DO NOT CITE OR QUOTE APPENDIX B: Benchmark Dose Modeling Output

Table B-1. Dose-response modeling results using BM DS (Version 2.0) based on noncancerous kidney and liver effects in rats followin g oral exposure to HCE

Study Endpoint Sex/species Fitted model a p Value AIC BMD 10

(mg/kg-day) BMDL 10

(mg/kg-day) Kidney effects Gamma 0.93 133.68 21.22 16.99 Multistage 1° 0.93 133.66 21.25 17.01 Male rat 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 Multistage 1° 0.20 148.90 1.75 1.40 NTP (1989) Linear mineralization Male rat 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 derivi ng the estimates of the benchmark dose (BMD 10 ) and its 95% lower confidence limit (BMDL 10 ). Modeling output is provided for models that rep resent 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 ma le 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 show ed adequate fit yielded the same AIC values, as well as nearly equivalent B MD 10 and BMDL 10 estimates of 21.22 and 16.99 mg/kg-day, respectively. Therefore, the pote ntial POD selected for this dataset is 16.99 mg/kg-day. Based on the incidence of tubular nephropathy in fe male 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 BMDL 10 estimates were within a factor of three of each oth er suggesting no appreciable model dependence. As the BMDL 10 values did not show large variation, the model wit h the lowest AIC value was selected. Therefore, the multistage 2° m odel BMDL 10 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 BMDL 10 estimates that were within a factor of three of each other, suggesting no apprec iable model dependence. As the BMDL 10 values did not show large variation, the model with the lowest AIC value was selected. The AIC values were identical; therefore, the model with th e lowest BMDL 10 was selected. The multistage 1° and quantal-linear model had identica l BMDL 10 values, therefore the BMDL 10 of 1.88 mg/kg-day was selected as the potential POD fo r this dataset.

Based on the incidence of mild to marked tubular ne phropathy in female rats (NTP, 1989), none of the models exhibited significant lac k-of-fit. These models (i.e., gamma, logistic, multistage 1°, probit, quantal-linear, and Weibull models) yielded BMDL 10 estimates that were within a factor of three of each other, suggesting no appreciable model dependence. As the

BMDL 10 values did not show large variation, the model wit h the lowest AIC value was selected. The gamma, quantal-linear, and Weibull models had i dentical AIC values; therefore, the model

with the lowest BMDL 10 was selected. The BMDL 10 values for these models were identical,

therefore the BMDL 10 of 10.72 mg/kg-day was selected as the potential P OD 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 significa nt lack-of-fit (i.e., logistic, multistage 1°, and probit), the resulting BMDL 10 estimates were within a factor of three of each other, suggesting no appreciable model depe ndence. As the BMDL 10 values did not show large variation, the model with the lowest AIC value was selected. Therefore the probit

model BMDL 10 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 significa nt lack-of-fit (i.e., gamma, loglogistic, multistage 2 °, Weibull, and quantal-linear), the resulting

BMDL 10 estimates were within a factor of three of each ot her, suggesting no appreciable model dependence. As the BMDL 10 values did not show large variation, the model wit h the lowest AIC value was selected. Therefore the loglogistic mode l BMDL 10 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 ra ts (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, an d Weibull) yielded BMDL 10 estimates that were within a factor of three of each other, suggesting no appreciable model dependence. As the

BMDL 10 values did not show large variation, the model wit h the lowest AIC value was selected. The AIC values for the gamma, multistage 1°, and qu antal-linear were identical; therefore, the

model with the lowest BMDL 10 was selected. All of the BMDL 10 s were identical for these models; therefore, the BMDL 10 of 0.728 mg/kg-day was selected as the potential P OD for this

dataset. For female rats, these models (i.e., gamma, multist age 1°, logistic, probit, quantal-linear, and Weibull) yielded BMDL 10 estimates that were within a factor of three of ea ch other, suggesting no appreciable model dependence. As the BMDL 10 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 identi cal; therefore, the model with the lowest

BMDL 10 was selected. The multistage 1° BMDL 10 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°, pr obit, Weibull, and quantal-linear) yielded

BMDL 10 estimates that were within a factor of three of ea ch other, suggesting no appreciable model dependence. As the BMDL 10 values did not show large variation, the model wit h the

lowest AIC value was selected. The AIC values for the gamma, Weibull, and quantal-linear models were identical; therefore, the model with th e lowest BMDL 10 was selected. All of the

BMDL 10 s were identical for these models, therefore the BM DL 10 of 0.710 mg/kg-day was selected as the potential POD for this dataset. Based on the incidence of hepatocellular necrosis i n female rats (NTP, 1989), none of the dichotomous dose-response models exhibited a signif icant lack-of-fit. All of these models (i.e., gamma, multistage 1°, logistic, probit, and Weibull ) yielded BMDL 10 estimates that were withinB-4 DRAFT – DO NOT CITE OR QUOTE

a factor of three of each other, suggesting no appr eciable model dependence. As the BMDL 10 values did not show large variation, the model with the lowest AIC value was selected. Therefore, the gamma model BMDL 10 of 60.18 mg/kg-day was selected as the potential P OD 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 i ncluded in Table B-2.B-5 DRAFT – DO NOT CITE OR QUOTETable B-2. Dose-response modeling results using BMDS (Version 2.0) for BMRs of 1, 5, and 10% based onnon-cancerouskidney and liver effects in rats following oral exposure to HCEStudy Endpoint Sex/speciesFittedmodelaBMD10(mg/kgday)BMDL10(mg/kgday)BMD05(mg/kg-day)BMDL05(mg/kgday)BMD01(mg/kgday)BMDL01(mg/kgday)Kidney effectsGamma 21.22 16.99 10.33 8.27 2.02 1.62Multistage1°21.25 17.01 10.35 8.28 2.03 1.62Male ratWeibull 21.22 16.99 10.33 8.27 2.02 1.62NCI(1978)TubularnephropathyFemaleratMultistage2°80.63 41.89 56.26 21.18 24.90 4.28Multistage1°3.20 1.88 1.56 0.91 0.30 0.18Moderate tomarkedtubularnephropathyMale ratQuantallinear3.20 1.88 1.56 0.91 0.30 0.18Gamma15.17 10.72 7.39 5.22 1.45 1.02Quantallinear15.17 10.72 7.39 5.22 1.45 1.02NTP(1989)Mild tomarkedtubularnephropathyFemaleratWeibull15.17 10.72 7.39 5.22 1.45 1.02NTP(1989)LinearmineralizationMale rat Probit 3.98 3.22 2.36 1.80 0.58 0.40NTP(1989)Hyperplasia ofthe pelvictransitionalepitheliumMale rat LogLogistic 7.05 4.48 3.34 2.12 0.64 0.41Gamma 1.34 0.73 0.66 0.35 0.13 0.07Multistage1°1.34 0.73 0.65 0.35 0.13 0.07Male ratQuantallinear1.34 0.73 0.65 0.35 0.13 0.07Multistage1°8.54 4.49 4.16 2.19 0.82 0.43Gorzinskiet al.(1985)Atrophy anddegenerationof renaltubulesFemaleratQuantallinear8.54 4.49 4.16 2.19 0.82 0.43B-6 DRAFT – DO NOT CITE OR QUOTEStudy Endpoint Sex/speciesFittedmodelaBMD10(mg/kgday)BMDL10(mg/kgday)BMD05(mg/kg-day)BMDL05(mg/kgday)BMD01(mg/kgday)BMDL01(mg/kgday)Gamma 1.22 0.71 0.60 0.35 0.12 0.07Weibull 1.22 0.71 0.60 0.35 0.12 0.07Gorzinskiet al.(1985)Slighthypertrophyand/or dilationof proximalconvolutedtubulesMale ratQuantallinear1.22 0.71 0.60 0.35 0.12 0.07Liver effectsNTP(1989)HepatocellularnecrosisFemaleratGamma 118.03 60.18 84.66 33.34 41.75 8.60B-7 DRAFT – DO NOT CITE OR QUOTE Modeling for Noncancer Assessment

NCI (1978) Tubular Nephropathy in Male Rats Gamma Model

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BMDS Model Run NCI 1978 Tubular Nephropathy Male R at - Gamma Model

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ ~~~~~~~~~~~~~~~~~~

The form of the probability function is:

P[response]= background+(1-background)*CumGamma[ slope*dose,power], where CumGamma(.) is the cummulative Gamma distr ibution function

Dependent variable = PercentPositiveModerateMark edTubularNephropathy 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: 1 e-008 Parameter Convergence has been set to: 1e-008

Default Initial (and Specified) P arameter Values Background = 0.0238095 Slope = 0.00474439 Power = 1.01848

Asymptotic Correlation Matrix of Paramet er Estimates

( *** The model parameter(s) -Backgroun d -Power have been estimated at a boundary point, or have been specified by the user, and do not appear in the correlati on 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 NAB-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 Devia nce Test d.f. P-value Full model -65.7706 3 Fitted model -65.8419 1 0.14 2715 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 = 16.9904B-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 200Fraction Affected

dose Gamma Multi-Hit Model with 0.95 Confidence Level

14:55 04/09 2009 BMDL BMD Gamma Multi-HitB-10 DRAFT – DO NOT CITE OR QUOTE

Multistage 1°

=================================================== ================= Multistage Model. $Revision: 2.1 $ $Date: 200 0/08/21 03:38:21 $ Input Data File: C:\BMDS\UNSAVED1.(d) Gnuplot Plotting File: C:\BMDS\UNSAVED1.plt Thu Sep 14 09:09:29 2006

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BMDS Model Run NCI 1978 Tubular Nephropathy Male R at - 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 positiv e

Dependent variable = PercentPositiveModerateMark edTubularNephropathy 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 Paramet er Estimates

( *** The model parameter(s) -Backgroun d have been estimated at a boundary point, or have been specified by the user, and do not appear in the correlati on 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 Devia nce Test d.f. P-value Full model -65.7706 3 Fitted model -65.8277 1 0.11 4158 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 % tw o-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 300Fraction Affected

dose Multistage Model with 0.95 Confidence Level

09:09 09/14 2006 BMD

BMDL MultistageB-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

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BMDS Model Run NCI 1978 Tubular Nephropathy Male R at - Weibull Model

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ ~~~~~~~~~~~~~~~~~~

The form of the probability function is:

P[response] = background + (1-background)*[1-EXP (-slope*dose^power)]

Dependent variable = PercentPositiveModerateMark edTubularNephropathy 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: 1 e-008 Parameter Convergence has been set to: 1e-008

Default Initial (and Specified) P arameter Values Background = 0.0238095 Slope = 0.00453277 Power = 1.00295

Asymptotic Correlation Matrix of Paramet er Estimates

( *** The model parameter(s) -Backgroun d -Power have been estimated at a boundary point, or have been specified by the user, and do not appear in the correlati on 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 Devia nce Test d.f. P-valueB-13 DRAFT – DO NOT CITE OR QUOTE

Full model -65.7706 3 Fitted model -65.8419 1 0.14 2715 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 = 16.9904B-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 300Fraction Affected

dose Weibull Model with 0.95 Confidence Level

09:13 09/14 2006

BMDL BMD WeibullB-15 DRAFT – DO NOT CITE OR QUOTE

NCI (1978) Tubular Nephropathy in Female Rats Multistage 2°

=================================================== ================= Multistage Model. $Revision: 2.1 $ $Date: 200 0/08/21 03:38:21 $ Input Data File: C:\BMDS\UNSAVED1.(d) Gnuplot Plotting File: C:\BMDS\UNSAVED1 .plt Thu Apr 09 16:18:29 2009

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BMDS Model Run - NCI 1978 Tubular Nephropathy Fema le 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 positiv e

Dependent variable = PercentPositiveModerateMark edTubularNephropathy 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 Paramet er Estimates

( *** The model parameter(s) -Backgroun d -Beta(1) have been estimated at a boundary point, or have been specified by the user, and do not appear in the correlati on 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 Devia nce Test d.f. P-value Full model -56.7357 3 Fitted model -57.0429 1 0.61 4339 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 % tw o-sided confidence interval for the BMDB-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 300Fraction Affected

dose Multistage Model with 0.95 Confidence Level

09:21 09/14 2006 BMD

BMDL MultistageB-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\tm p9D5.plt Wed Apr 08 11:27:18 2009

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BMDS Model Run NTP 1989 Tubular Nephropathy Male R at - 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 positiv e

Dependent variable = PercentPositiveModerateMark edTubularNephropathy 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 Paramet er Estimates

Background Beta(1)

Background 1 -0.72

Beta(1) -0.72 1

Parameter Estimate s

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 TableB-19 DRAFT – DO NOT CITE OR QUOTE

Model Log(likelihood) # Param's Devia nce Test d.f. P-value Full model -100.939 3 Fitted model -100.952 2 0.025 8029 1 0.8724 Reduced model -103.852 1 5.8 2641 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 % tw o-sided confidence interval for the BMDB-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 14Fraction Affected

dose Multistage Model with 0.95 Confidence Level

11:27 04/08 2009 BMD BMDL MultistageB-21 DRAFT – DO NOT CITE OR QUOTE

Quantal-linear Model

================================================== ================== Quantal Linear Model using Weibull Model (Ver sion: 2.12; Date: 05/16/2008) Input Data File: C:\USEPA\BMDS2\Temp\tmpA17.(d ) Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tm pA17.plt Wed Apr 08 13:34:31 2009

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BMDS Model Run NTP 1989 Tubular Nephropathy Male R at - Quantal-linear Model

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ ~~~~~~~~~~~~~~~~~~

The form of the probability function is:

P[response] = background + (1-background)*[1-EXP (-slope*dose)]

Dependent variable = PercentPositiveModerateMark edTubularNephropathy 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: 1 e-008 Parameter Convergence has been set to: 1e-008

Default Initial (and Specified) P arameter Values Background = 0.362745 Slope = 0.0329154 Power = 1 Sp ecified

Asymptotic Correlation Matrix of Paramet er 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 correlati on matrix )

Background Slope

Background 1 -0.66

Slope -0.66 1

Parameter Estimate s

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 Devia nce Test d.f. P-value Full model -100.939 3B-22 DRAFT – DO NOT CITE OR QUOTE

Fitted model -100.952 2 0.025 8029 1 0.8724 Reduced model -103.852 1 5.8 2641 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 = 1.8769B-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 14Fraction Affected

dose Quantal Linear Model with 0.95 Confidence Level

13:34 04/08 2009 BMDL BMD Quantal LinearB-24 DRAFT – DO NOT CITE OR QUOTE

NTP (1989) Female Rat Nephropathy

Gamma Model

================================================== ================== Gamma Model. (Version: 2.13; Date: 05/16/200 8) Input Data File: C:\USEPA\BMDS2\Temp\tmpD9.(d) Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tm pD9.plt Fri Apr 10 10:19:37 2009

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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 distr ibution function

Dependent variable = PercentPositiveModerateMark edTubularNephropathy 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: 1 e-008 Parameter Convergence has been set to: 1e-008

Default Initial (and Specified) P arameter Values Background = 0.245098 Slope = 0.0111213 Power = 1.3

Asymptotic Correlation Matrix of Paramet er 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 correlati on matrix )

Background Slope

Background 1 -0.55

Slope -0.55 1

Parameter Estimate s

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 boundB-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 Devia nce Test d.f. P-value Full model -93.9362 3 Fitted model -93.9519 2 0.031 2372 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 = 10.7248B-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 100Fraction Affected

dose Quantal Linear Model with 0.95 Confidence Level

12:21 12/16 2008 BMDL BMD Quantal Linear BMD Lower BoundB-27 DRAFT – DO NOT CITE OR QUOTE

Quantal-linear Model

================================================== ================== Quantal Linear Model using Weibull Model (Ver sion: 2.12; Date: 05/16/2008) Input Data File: C:\USEPA\BMDS2\Temp\tmpE4.(d) Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tm pE4.plt Fri Apr 10 10:36:29 2009

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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 = PercentPositiveModerateMark edTubularNephropathy 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: 1 e-008 Parameter Convergence has been set to: 1e-008

Default Initial (and Specified) P arameter Values Background = 0.245098 Slope = 0.00666772 Power = 1 Sp ecified

Asymptotic Correlation Matrix of Paramet er 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 correlati on matrix )

Background Slope

Background 1 -0.55

Slope -0.55 1

Parameter Estimate s

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 Devia nce Test d.f. P-value Full model -93.9362 3B-28 DRAFT – DO NOT CITE OR QUOTE

Fitted model -93.9519 2 0.031 2372 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 = 10.7248B-29 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 100Fraction Affected

dose Quantal Linear Model with 0.95 Confidence Level

10:36 04/10 2009 BMDL BMD Quantal LinearB-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\tm pE3.plt Fri Apr 10 10:34:27 2009

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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 = PercentPositiveModerateMark edTubularNephropathy 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: 1 e-008 Parameter Convergence has been set to: 1e-008

Default Initial (and Specified) P arameter Values Background = 0.245098 Slope = 0.00666772 Power = 1

Asymptotic Correlation Matrix of Paramet er 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 correlati on matrix )

Background Slope

Background 1 -0.55

Slope -0.55 1

Parameter Estimate s

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 thusB-31 DRAFT – DO NOT CITE OR QUOTE

has no standard error.

Analysis of Deviance Table

Model Log(likelihood) # Param's Devia nce Test d.f. P-value Full model -93.9362 3 Fitted model -93.9519 2 0.031 2372 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 = 10.7248B-32 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 100Fraction Affected

dose Weibull Model with 0.95 Confidence Level

10:34 04/10 2009 BMDL BMD WeibullB-33 DRAFT – DO NOT CITE OR QUOTE

NTP (1989) Linear Mineralization in Male Rats

Probit Model

================================================== ================== Probit Model. (Version: 3.1; Date: 05/16/200 8) Input Data File: C:\USEPA\BMDS2\Temp\tmpA33.(d ) Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tm pA33.plt Wed Apr 08 14:24:02 2009

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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 distri bution function

Dependent variable = PercentPositiveLinearMinera lization 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: 1 e-008 Parameter Convergence has been set to: 1e-008

Default Initial (and Specified) P arameter Values background = 0 Sp ecified intercept = -1.67551 slope = 0.149038

Asymptotic Correlation Matrix of Paramet er Estimates

( *** The model parameter(s) -backgroun d have been estimated at a boundary point, or have been specified by the user, and do not appear in the correlati on matrix )

intercept slope

intercept 1 -0.87

slope -0.87 1

Parameter Estimate s

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 0.191579B-34 DRAFT – DO NOT CITE OR QUOTE

Analysis of Deviance Table

Model Log(likelihood) # Param's Devia nce Test d.f. P-value Full model -71.6113 3 Fitted model -71.8283 2 0.43 3989 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 = 3.21773B-35 DRAFT – DO NOT CITE OR QUOTE 0 0.1 0.2 0.3 0.4 0.5 0.6 0.7 0.8

0 2 4 6 8 10 12 14Fraction Affected

dose Probit Model with 0.95 Confidence Level

14:24 04/08 2009 BMDL BMD ProbitB-36 DRAFT – DO NOT CITE OR QUOTE

NTP (1989) Male Rat Hyperplasia of Pelvic Transitio nal Epithelium LogLogistic Model

================================================== ================== Logistic Model. (Version: 2.12; Date: 05/16/2 008) Input Data File: C:\USEPA\BMDS2\Temp\tmp4D5.(d ) Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tm p4D5.plt Wed Aug 12 14:26:53 2009

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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: 1 e-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 Paramet er Estimates

( *** The model parameter(s) -backgroun d -slope have been estimated at a boundary point, or have been specified by the user, and do not appear in the correlati on matrix )

intercept

intercept 1

Parameter Estimate s

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 Devia nce Test d.f. P-value Full model -40.4963 3 Fitted model -41.2103 1 1.4 2796 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 = 4.48322B-38 DRAFT – DO NOT CITE OR QUOTE 0 0.05 0.1 0.15 0.2 0.25 0.3

0 2 4 6 8 10 12 14Fraction Affected

dose Log-Logistic Model with 0.95 Confidence Level

14:26 08/12 2009 BMDL BMD Log-LogisticB-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: 05/16/200 8) Input Data File: C:\USEPA\BMDS2\Temp\tmpF14.(d ) Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tm pF14.plt Thu Oct 08 08:59:00 2009

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ndDegenRenalTubulesDataNoSeverityMaleRat.dax

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ ~~~~~~~~~~~~~~~~~~

The form of the probability function is:

P[response]= background+(1-background)*CumGamma[ slope*dose,power], where CumGamma(.) is the cummulative Gamma distr ibution 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: 1 e-008 Parameter Convergence has been set to: 1e-008

Default Initial (and Specified) P arameter Values Background = 0.136364 Slope = 0.0871864 Power = 1.3

Asymptotic Correlation Matrix of Paramet er Estimates

Background Slope Power

Background 1 0.52 0.64

Slope 0.52 1 0.93

Power 0.64 0.93 1

Parameter Estimate s

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 3.0996B-40 DRAFT – DO NOT CITE OR QUOTE

Analysis of Deviance Table

Model Log(likelihood) # Param's Devia nce Test d.f. P-value Full model -14.3635 4 Fitted model -14.4712 3 0.21 5359 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 0.4 0.6 0.8 1

0 10 20 30 40 50 60Fraction Affected

dose Gamma Multi-Hit Model with 0.95 Confidence Level

08:59 10/08 2009

BMDL BMD Gamma Multi-HitB-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\tm pF17.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 positiv e

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 Paramet er Estimates

Background Beta(1)

Background 1 -0.4

Beta(1) -0.4 1

Parameter Estimate s

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 Devia nce Test d.f. P-valueB-42 DRAFT – DO NOT CITE OR QUOTE

Full model -14.3635 4 Fitted model -14.4712 2 0.21 5361 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 % t wo-sided confidence interval for the BMDB-43 DRAFT – DO NOT CITE OR QUOTE 0 0.2 0.4 0.6 0.8 1

0 10 20 30 40 50 60Fraction Affected

dose Multistage Model with 0.95 Confidence Level

09:00 10/08 2009 BMD

BMDL MultistageB-44 DRAFT – DO NOT CITE OR QUOTE

Quantal-linear Model

================================================== ================== Quantal Linear Model using Weibull Model (Ver sion: 2.12; Date: 05/16/2008) Input Data File: C:\USEPA\BMDS2\Temp\tmpF18.(d ) Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tm pF18.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: 1 e-008 Parameter Convergence has been set to: 1e-008

Default Initial (and Specified) P arameter Values Background = 0.136364 Slope = 0.047491 Power = 1 Sp ecified

Asymptotic Correlation Matrix of Paramet er 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 correlati on matrix )

Background Slope

Background 1 -0.29

Slope -0.29 1

Parameter Estimate s

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 Devia nce Test d.f. P-value Full model -14.3635 4 Fitted model -14.4712 2 0.21 5361 2 0.8979B-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 0.2 0.4 0.6 0.8 1

0 10 20 30 40 50 60Fraction Affected

dose Quantal Linear Model with 0.95 Confidence Level

09:02 10/08 2009

BMDL BMD Quantal LinearB-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\tm pF51.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 positiv e

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 Paramet er Estimates

Background Beta(1)

Background 1 -0.51

Beta(1) -0.51 1

Parameter Estimate s

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 Devia nce Test d.f. P-value Full model -18.2358 4 Fitted model -18.3071 2 0.14 2532 2 0.9312 Reduced model -22.4934 1 8.5 1521 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 % tw o-sided confidence interval for the BMDB-48 DRAFT – DO NOT CITE OR QUOTE 0 0.2 0.4 0.6 0.8

0 10 20 30 40 50 60Fraction Affected

dose Multistage Model with 0.95 Confidence Level

09:34 10/08 2009 BMD

BMDL MultistageB-49 DRAFT – DO NOT CITE OR QUOTE

Quantal-linear Model

=================================================== ================= Quantal Linear Model using Weibull Model (Ver sion: 2.12; Date: 05/16/2008) Input Data File: C:\USEPA\BMDS2\Temp\tmpF52.(d ) Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tm pF52.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: 1 e-008 Parameter Convergence has been set to: 1e-008

Default Initial (and Specified) P arameter Values Background = 0.136364 Slope = 0.0120518 Power = 1 Sp ecified

Asymptotic Correlation Matrix of Paramet er 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 correlati on matrix )

Background Slope

Background 1 -0.33

Slope -0.33 1

Parameter Estimate s

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 Devia nce Test d.f. P-value Full model -18.2358 4 Fitted model -18.3071 2 0.14 2532 2 0.9312B-50 DRAFT – DO NOT CITE OR QUOTE

Reduced model -22.4934 1 8.5 1521 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 60Fraction Affected

dose Quantal Linear Model with 0.95 Confidence Level

09:35 10/08 2009 BMDL BMD Quantal LinearB-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: 05/16/200 8) Input Data File: C:\USEPA\BMDS2\Temp\tmp4D6.(d ) Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tm p4D6.plt Wed Aug 12 14:31:38 2009

================================================== ==================

BMDS Model Run - Gorzinski et al (1985) - Male Rat – Hyp ertrophy/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 distr ibution 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: 1 e-008 Parameter Convergence has been set to: 1e-008

Default Initial (and Specified) P arameter Values Background = 0.0454545 Slope = 0.0907614 Power = 1.3

Asymptotic Correlation Matrix of Paramet er Estimates

( *** The model parameter(s) -Backgroun d -Power have been estimated at a boundary point, or have been specified by the user, and do not appear in the correlati on matrix )

Slope

Slope 1

Parameter Estimate s

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 Devia nce Test d.f. P-value Full model -9.35947 4 Fitted model -9.44226 1 0.16 5576 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 = 0.710032B-53 DRAFT – DO NOT CITE OR QUOTE 0 0.2 0.4 0.6 0.8 1

0 10 20 30 40 50 60Fraction Affected

dose Gamma Multi-Hit Model with 0.95 Confidence Level

14:31 08/12 2009

BMDL BMD Gamma Multi-HitB-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\tm p4D9.plt Wed Aug 12 14:35:51 2009

================================================== ==================

BMDS Model Run - Gorzinski et al (1985) - Male rat s - Hypertrophy/Dilation of Proximal Tubules - Weibull Model using Weibull Mode l (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: 1 e-008 Parameter Convergence has been set to: 1e-008

Default Initial (and Specified) P arameter Values Background = 0.0454545 Slope = 0.0491052 Power = 1

Asymptotic Correlation Matrix of Paramet er Estimates

( *** The model parameter(s) -Backgroun d -Power have been estimated at a boundary point, or have been specified by the user, and do not appear in the correlati on matrix )

Slope

Slope 1

Parameter Estimate s

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 Devia nce Test d.f. P-value Full model -9.35947 4 Fitted model -9.44226 1 0.16 5576 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 60Fraction Affected

dose Weibull Model with 0.95 Confidence Level

14:35 08/12 2009

BMDL BMD WeibullB-56 DRAFT – DO NOT CITE OR QUOTE

Quantal-linear Model

================================================== ================== antal-linear Model using Weibull Model (Versi on: 2.12; Date: 05/16/2008) Input Data File: C:\USEPA\BMDS2\Temp\tmp4DA.(d ) Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tm p4DA.plt Wed Aug 12 14:37:26 2009

================================================== ==================

BMDS Model Run - Gorzinski et al (1985) - Male rat s - 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: 1 e-008 Parameter Convergence has been set to: 1e-008

Default Initial (and Specified) P arameter Values Background = 0.0454545 Slope = 0.0491052 Power = 1 Sp ecified

Asymptotic Correlation Matrix of Paramet er Estimates

( *** The model parameter(s) -Backgroun d -Power have been estimated at a boundary point, or have been specified by the user, and do not appear in the correlati on matrix )

Slope

Slope 1

Parameter Estimate s

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 TableB-57 DRAFT – DO NOT CITE OR QUOTE

Model Log(likelihood) # Param's Devia nce Test d.f. P-value Full model -9.35947 4 Fitted model -9.44226 1 0.16 5576 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 60Fraction Affected

dose Weibull Model with 0.95 Confidence Level

14:37 08/12 2009

BMDL BMD WeibullB-58 DRAFT – DO NOT CITE OR QUOTE

NTP (1989) Female Rat Hepatocellular Necrosis Gamma Model

================================================== ================== Gamma Model. (Version: 2.13; Date: 05/16/200 8) Input Data File: C:\USEPA\BMDS2\Temp\tmpB62.(d ) Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tm pB62.plt Thu Apr 09 09:14:08 2009

================================================== ==================

BMDS Model Run NTP 1989 Hepatocellular Necrosis Fe 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 distr ibution function

Dependent variable = PercentPositiveHepatocellul arNecrosis 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: 1 e-008 Parameter Convergence has been set to: 1e-008

Default Initial (and Specified) P arameter Values Background = 0.0454545 Slope = 0.00743289 Power = 2.82109

Asymptotic Correlation Matrix of Paramet er Estimates

( *** The model parameter(s) -Backgroun d have been estimated at a boundary point, or have been specified by the user, and do not appear in the correlati on matrix )

Slope Power

Slope 1 0.95

Power 0.95 1

Parameter Estimate s

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 thusB-59 DRAFT – DO NOT CITE OR QUOTE

has no standard error.

Analysis of Deviance Table

Model Log(likelihood) # Param's Devia nce Test d.f. P-value Full model -16.7382 6 Fitted model -17.3091 2 1.1 4186 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 = 60.1812B-60 DRAFT – DO NOT CITE OR QUOTE 0 0.2 0.4 0.6 0.8 1

0 100 200 300 400 500Fraction Affected

dose Gamma Multi-Hit Model with 0.95 Confidence Level

09:14 04/09 2009 BMDL BMD Gamma Multi-HitB-61 DRAFT – DO NOT CITE OR QUOTE Modeling for Cancer Assessment

NTP (1989) BMD Modeling of Adenoma/Carcinoma in Mal e 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\tm p6E8.plt Mon Apr 13 14:38:06 2009

================================================== ==================

BMDS Model Run NTP 1989 Kidney Adenoma-Carcinoma M ale 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 positiv e

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: 2.2 2045e-016 Parameter Convergence has been set to: 1.49012e-00 8

**** We are sorry but Relative Function and Parame ter Convergence   ** ** are currently unavailable in this model. Ple ase keep checking   ** ** the web sight for model updates which will ev entually   ** ** incorporate these convergence criterion. Def ault values used.  

Default Initial Parameter Values Background = 0.014541 Beta(1) = 0 Beta(2) = 0.00799069

Asymptotic Correlation Matrix of Paramet er 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 correlati on matrix )

Background Beta(2)

Background 1 -0.67

Beta(2) -0.67 1B-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 Devia nce Test d.f. P-value Full model -33.5473 3 Fitted model -33.6008 2 0.10 6829 1 0.7438 Reduced model -36.7395 1 6.3 8433 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 % tw o-sided confidence interval for the BMD

Multistage Cancer Slope Factor = 0.0407692B-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 4Fraction Affected

dose Multistage Cancer Model with 0.95 Confidence Level

14:38 04/13 2009 BMD BMDL Multistage Cancer Linear extrapolationB-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\tm p7B8.plt Tue Apr 14 08:30:03 2009

================================================== ==================

BMDS Model Run NCI 1978 Hepatocellular Carcinoma Male Mi ce - 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 positiv e

Dependent variable = PercentHepatocellularCarcin oma 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: 2.2 2045e-016 Parameter Convergence has been set to: 1.49012e-00 8

**** We are sorry but Relative Function and Parame ter Convergence   ** ** are currently unavailable in this model. Ple ase keep checking   ** ** the web sight for model updates which will ev entually   ** ** incorporate these convergence criterion. Def ault values used.  

Default Initial Parameter Values Background = 0.141096 Beta(1) = 0 Beta(2) = 7.77012e-005

Asymptotic Correlation Matrix of Paramet er 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 correlati on matrix )

Background Beta(2)

Background 1 -0.73

Beta(2) -0.73 1

Parameter Estimate sB-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 Devia nce Test d.f. P-value Full model -71.2862 3 Fitted model -71.7199 2 0.86 7331 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 % tw o-sided confidence interval for the BMD

Multistage Cancer Slope Factor = 0.00724545B-66 DRAFT – DO NOT CITE OR QUOTE 0 0.1 0.2 0.3 0.4 0.5 0.6 0.7 0.8

0 20 40 60 80 100Fraction Affected

dose Multistage Cancer Model with 0.95 Confidence Level

08:30 04/14 2009 BMD BMDL Multistage Cancer Linear extrapolation