DRAFT - DO NOT CITE OR QUOTE EPA/635/R-10/003B www.epa.gov/iris

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# TOXICOLOGICAL REVIEW 2

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# OF 4

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# DICHLOROMETHANE 6 

# (METHYLENE CHLORIDE) 7

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### (CAS No. 75-09-2) 9

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## In Support of Summary Information on the 13 

## Integrated Risk Information System (IRIS) 14

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## December 2009 16 17

18 19 NOTICE 20 21 This document is an Interagency Science Consultation draft . This information is distributed 22 solely for the purpose of pre-dissemination peer re view under applicable information quality 23 guidelines. It has not been formally disseminated b y EPA. It does not represent and should not 24 be construed to represent any Agency determination or policy. It is being circulated for review 25 of its technical accuracy and science policy implic ations. 26 27 28 29 U.S. Environmental Protection Agency 30 Washington, DC 3132 DISCLAIMER 33

34 This document is a preliminary draft for review pur poses only. This information is 35 distributed solely for the purpose of pre-dissemina tion peer review under applicable information 36 quality guidelines. It has not been formally disse minated by EPA. It does not represent and 37 should not be construed to represent any Agency det ermination or policy. Mention of trade 38 names or commercial products does not constitute en dorsement or recommendation for use. 39

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41ii DRAFT - DO NOT CITE OR QUOTE CONTENTS—TOXICOLOGICAL REVIEW OF DICHLOROMETHANE 42 (CAS No. 75-09-2) 43

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CONTENTS—TOXICOLOGICAL REVIEW OF DICHLOROMETHANE ... ............................ ii 45 LIST OF TABLES.....................................

................................................... ................................ vii 46 LIST OF FIGURES ....................................

................................................... .............................. xiii 47 LIST OF ACRONYMS ...................................

................................................... ........................ xvii 48 FOREWORD ...........................................

................................................... ................................. xix 49 AUTHORS, CONTRIBUTORS, AND REVIEWERS ............... ................................................. xx 50 1. INTRODUCTION ...................................

................................................... ................................1 51 2. CHEMICAL AND PHYSICAL INFORMATION ..............

................................................... ...3 52 3. TOXICOKINETICS .................................

................................................... ...............................5 53 3.1. ABSORPTION...................................

................................................... ............................ 5 54 3.1.1. Oral — Gastrointestinal Tract Absorption ...

................................................... ....... 5 55 3.1.2. Inhalation—Respiratory Tract Absorption....

................................................... ...... 5 56 3.2. DISTRIBUTION .................................

................................................... ........................... 7 57 3.3. METABOLISM...................................

................................................... ........................... 9 58 3.3.1. The CYP2E1 Pathway.........................

................................................... .............. 11 59 3.3.2. The GST Pathway............................

................................................... .................. 14 60 3.4. ELIMINATION..................................

................................................... .......................... 19 61 3.5. PHYSIOLOGICALLY BASED TOXICOKINETIC MODELS ... ................................. 21 62 3.5.1. Probabilistic Mouse PBTK Dichloromethane Mo del (Marino et al., 2006) ........ 26 63 3.5.2. Probabilistic Human PBTK Dichloromethane Mo del (David et al., 2006) ......... 31 64 3.5.3. Evaluation of Rat PBTK Dichloromethane Mode ls............................................. 38 65 3.5.4. Comparison of Mouse, Rat and Human PBTK Mod els....................................... 39 66 HAZARD IDENTIFICATION..............................

................................................... .....................43 67 4.1. STUDIES IN HUMANS ............................

................................................... .................. 43 68 4.1.1. Introduction—Case Reports, Epidemiologic, a nd Clinical Studies ..................... 43 69 4.1.2. Noncancer Studies..........................

................................................... ................... 43 70 4.1.2.1. Case Reports of Acute, High-dose Exposures ........................................ 43 71 4.1.2.2. Controlled Experiments Examining Acute Effects .................................. 44 72 4.1.2.3. Observational Studies Focusing on Clinical Chemistr ies, Clinical 73 Examinations, and Symptoms

................................................... .............. 45 74 4.1.2.4. Observational Studies Using Workplace Medical Progr am Data ......... 51 75 4.1.2.5. Studies of Ischemic Heart Disease Mortality Risk ................................. 55 76 4.1.2.6. Studies of Suicide Risk

................................................... ......................... 56 77 4.1.2.7 Studies of Infectious Disease Risk

................................................... ....... 57 78 4.1.2.8. Studies of Reproductive Outcomes

................................................... ...... 57 79 4.1.2.9. Summary of Noncancer Studies

................................................... ........... 60 80 4.1.3. Cancer Studies.............................

................................................... ...................... 62 81 4.1.3.1. Identification and Selection of Studies for Evaluat ion of Cancer 82 Risk

................................................... ................................................... ... 62 83 4.1.3.2. Description of the Selected Studies

................................................... ...... 63 84 4.1.3.3. Cellulose Triacetate Film Base Production Cohorts .............................. 63 85 4.1.3.4. Cellulose Triacetate Fiber Production Cohorts ..................................... 70 86 4.1.3.5. Solvent-Exposed Workers—Hill Air Force Base, Utah .......................... 75 87 4.1.3.6. Case-Control Studies of Specific Cancers and Dichlo romethane .......... 76 88 4.1.3.7. Summary of Cancer Studies by Type of Cancer ..................................... 83 89iii DRAFT - DO NOT CITE OR QUOTE

4.2. SUBCHRONIC AND CHRONIC STUDIES AND CANCER BI OASSAYS IN 90 ANIMALS—ORAL AND INHALATION ........................ ............................................ 92 91 4.2.1. Oral Exposure: Overview of Noncancer and Ca ncer Effects............................... 92 92 4.2.1.1. Toxicity Studies of Subchronic Oral Exposures: Hepa tic Effects .......... 93 93 4.2.1.2. Toxicity Studies of Chronic Oral Exposures: Hepatic Effects and 94 Carcinogenicity

................................................... .................................... 96 95 4.2.2. Inhalation Exposure: Overview of Noncancer and Cancer Effects.................... 102 96 4.2.2.1. Toxicity Studies of Subchronic Inhalation Exposures : General, 97 Renal, and Hepatic Effects

................................................... ................ 103 98 4.2.2.2. Toxicity Studies from Chronic Inhalation Exposures ........................... 107 99 4.3. REPRODUCTIVE/DEVELOPMENTAL STUDIES—ORAL AND 100 INHALATION.........................................

................................................... .................. 124 101 4.3.1. Reproductive Toxicity Studies ..............

................................................... .......... 127 102 4.3.1.1. Oral (Gavage) Studies

................................................... ....................... 127 103 4.3.1.2. Inhalation Studies

................................................... .............................. 128 104 4.3.2. Developmental Toxicity Studies .............

................................................... ........ 129 105 4.3.2.1. Oral (Gavage) Studies and Culture Studies ......................................... 129 106 4.3.2.2. Inhalation Studies

................................................... .............................. 130 107 4.4. OTHER DURATION- OR ENDPOINT-SPECIFIC STUDIES . .................................. 132 108 4.4.1. Short-term (2-Week) Studies of General and Hepatic Effects in Animals ........ 132 109 4.4.2. Immunotoxicity Studies in Animals..........

................................................... ...... 133 110 4.4.3. Neurotoxicology Studies in Animals.........

................................................... ...... 135 111 4.4.3.1. Neurotoxicology Studies—Oral Exposures .......................................... 140 112 4.4.3.2. Neurotoxicology Studies—Inhalational Exposure ................................ 141 113 4.5. MECHANISTIC DATA AND OTHER STUDIES IN SUPPORT OF THE 114 MODE OF ACTION.....................................

................................................... ............. 148 115 4.5.1. Genotoxicity Studies .......................

................................................... ................ 148 116 4.5.1.1. In Vitro Genotoxicity Assays

................................................... ............. 148 117 Bacterial, Yeast, and Fungi mutagenicity assays ............................................... 148 118 4.5.1.2. In Vivo Genotoxicity Assays

................................................... .............. 158 119 4.5.2. Mechanistic Studies of Liver Effects .......

................................................... ....... 167 120 4.5.3. Mechanistic Studies of Lung Effects........

................................................... ....... 170 121 4.5.4. Mechanistic Studies of Neurological Effects

................................................... .. 175 122 4.6. SYNTHESIS OF MAJOR NONCANCER EFFECTS ......... ........................................ 177 123 4.6.1. Oral Exposures .............................

................................................... ................... 177 124 4.6.1.1. Summary of Human Data

................................................... .................. 177 125 4.6.1.2. Summary of Animal Data

................................................... ................... 178 126 4.6.2. Inhalation Exposures .......................

................................................... ................ 182 127 4.6.2.1. Summary of Human Data

................................................... .................. 182 128 4.6.2.2. Summary of Animal Studies

................................................... ............... 183 129 4.6.3. Mode of Action Information .................

................................................... .......... 190 130 4.6.3.1. Mode of Action for Nonneoplastic Liver Effects ................................... 190 131 4.6.3.2. Mode of Action for Nonneoplastic Lung Effects ................................... 191 132 4.6.3.3. Mode of Action for Neurological Effects .............................................. 191 133 4.6.3.4. Mode of Action for Reproductive and Developmental E ffects .............. 192 134 4.6.3.5. Mode of Action for Immunotoxicity

................................................... ... 193 135 4.7. EVALUATION OF CARCINOGENICITY ................ ................................................. 19 4 136 4.7.1. Summary of Overall Weight of Evidence ......

................................................... . 194 137 4.7.2. Synthesis of Human, Animal, and Other Suppo rting Evidence ......................... 195 138iv DRAFT - DO NOT CITE OR QUOTE

4.7.3. Mode of Action Information .................

................................................... .......... 204 139 4.7.3.1. Hypothesized Mode of Action

................................................... ............ 204 140 4.7.3.2. General Conclusions About the Mode of Action for Tu mors in 141 Rodents and Possible Relevance to Humans ........................................ 211 142 4.8. SUSCEPTIBLE POPULATIONS AND LIFE STAGES ...... ....................................... 213 143 4.8.1. Possible Childhood Susceptibility..........

................................................... ......... 213 144 4.8.2. Possible Gender Differences ................

................................................... ........... 215 145 4.8.3. Other Susceptible Populations..............

................................................... ........... 215 146 5. DOSE-RESPONSE ASSESSMENTS......................

................................................... ...........217 147 5.1. ORAL REFERENCE DOSE (RfD) ....................

................................................... ....... 217 148 5.1.1. Choice of Principal Study and Critical Effe ct—with Rationale and 149 Justification......................................

................................................... ............... 217 150 5.1.2. Derivation Process for Noncancer Reference Values ........................................ 220 151 5.1.3. Evaluation of Dose Metrics for Use in Nonca ncer Reference Value 152 Derivations........................................

................................................... .............. 223 153 5.1.4. Methods of Analysis—Including Models (PBTK, BMD, etc.).......................... 224 154 5.1.5. RfD Derivation—Including Application of Uncertainty Factors (UFs) ............. 229 155 5.1.6. Previous RfD Assessment

................................................... ................................ 230 156 5.1.7. RfD Comparison Information

................................................... .......................... 230 157 5.2. INHALATION REFERENCE CONCENTRATION (RfC)..... .................................... 234 158 5.2.1. Choice of Principal Study and Critical Effe ct—with Rationale and 159 Justification......................................

................................................... ............... 234 160 5.2.2. Derivation Process for Reference Concentrat ion Values ................................... 238 161 5.2.3. Methods of Analysis—Including Models (PBTK, BMD, etc.).......................... 238 162 5.2.4. RfC Derivation—Including Application of Unc ertainty Factors (UFs)............. 243 163 5.2.5. Previous RfC Assessment....................

................................................... ............ 245 164 5.2.6. RfC Comparison Information.................

................................................... ......... 245 165 5.3. UNCERTAINTIES IN THE ORAL REFERENCE DOSE AND INHALATION 166 REFERENCE CONCENTRATION............................ ................................................. 25 0 167 5.4. CANCER ASSESSMENT ............................

................................................... ............. 262 168 5.4.1. Cancer Oral Slope Factor ...................

................................................... ............. 263 169 5.4.1.1. Choice of Study/Data—with Rationale and Justificati on ..................... 263 170 5.4.1.2. Derivation Process for Cancer Oral Slope Factor .............................. 263 171 5.4.1.3. Dose-Response Data

................................................... .......................... 266 172 5.4.1.4. Dose Conversion and Extrapolation Methods: Cancer O ral Slope 173 Factor

................................................... ................................................ 266 174 5.4.1.5. Oral Cancer Slope Factor

................................................... ................. 273 175 5.4.1.6. Alternative Derivation Based on Route-to-Route Extr apolation .......... 273 176 5.4.1.8. Previous IRIS Assessment: Cancer Oral Slope Factor ........................ 276 177 5.4.1.9. Comparison of Cancer Oral Slope Factors Using Diffe rent 178 Methodologies

................................................... .................................... 276 179 5.4.2. Cancer Inhalation Unit Risk ................

................................................... ............ 278 180 5.4.2.1. Choice of Study/Data—with Rationale and Justificati on ..................... 278 181 5.4.2.2. Derivation Process for Cancer Inhalation Unit Risk ........................... 279 182 5.4.2.3. Dose-Response Data

................................................... .......................... 279 183 5.4.2.4. Dose Conversion and Extrapolation Methods: Cancer I nhalation 184 Unit Risk

................................................... ............................................ 280 185 5.4.2.5. Cancer Inhalation Unit Risk

................................................... .............. 287 186 5.4.2.8. Previous IRIS Assessment: Cancer Inhalation Unit Ri sk ..................... 291 187v DRAFT - DO NOT CITE OR QUOTE 5.4.2.8. Comparison of Cancer Inhalation Unit Risk Using Dif ferent 188 Methodologies

................................................... .................................... 292 189 5.4.3. Differences Between Current Assessment and Previous IRIS PBTK-based 190 Assessment.........................................

................................................... ............. 294 191 5.4.4. Application of ADAFs .......................

................................................... ............. 296 192 5.4.4.1. Application of ADAFs in Oral Exposure Scenarios ............................. 297 193 5.4.4.2. Application of ADAFs in Inhalation Exposure Scenar ios .................... 297 194 5.4.5. Uncertainties in Cancer Risk Values........

................................................... ....... 298 195 6. MAJOR CONCLUSIONS IN THE CHARACTERIZATION OF HA ZARD AND 196 DOSE RESPONSE ......................................

................................................... .......................310 197 6.1. HUMAN HAZARD POTENTIAL .......................

................................................... ..... 310 198 6.2. DOSE-RESPONSE ................................

................................................... .................... 313 199 6.2.1. Oral RfD ...................................

................................................... ....................... 313 200 6.2.2. Inhalation RfC .............................

................................................... .................... 314 201 6.2.3. Uncertainties in Reference Dose and Referen ce Concentration Values ............ 315 202 6.2.4. Oral Cancer Slope Factor ...................

................................................... ............. 316 203 6.2.5. Cancer Inhalation Unit Risk ................

................................................... ............ 320 204 6.2.6. Uncertainties in Cancer Toxicity Values....

................................................... ..... 323 205 7. REFERENCES .....................................

................................................... ...............................325 206 207 208 APPENDIX A: EXTERNAL REVIEW PANEL COMMENTS ........ ...................................... A-1 209 210 APPENDIX B: HUMAN PBTK DICHLOROMETHANE MODEL ...... ..................................B-1 211 B-1. Human Model Description.......................

................................................... .....................B-1 212 B-2. Revisions to Parameter Distributions of David et al. (2006) ...................................... .....B-3 213 B-3. CYP2E1 and GST-T1 .............................

................................................... ......................B-4 214 B-4. Analysis of Human Physiological Distributions for PBPK Modeling ............................B-8 215 B-4.1. Age.........................................

................................................... ....................................B-8 216 B-4.2. Gender......................................

................................................... ..................................B-9 217 B-4.3. Body Weight .................................

................................................... ...........................B-10 218 B-4.4. Alveolar Ventilation........................

................................................... .........................B-12 219 B-4.5. Cardiac Output..............................

................................................... ...........................B-13 220 B-4.6. Fat Fraction ................................

................................................... ..............................B-14 221 B-4.7. Liver Fraction..............................

................................................... .............................B-14 222 B-4.8. Tissue volume normalization.................

................................................... ..................B-16 223 B-5. Summary of Revised Human PBTK Model ..........

................................................... ....B-16 224 225 APPENDIX C: RAT DICHLOROMETHANE PBTK MODELS ...... ......................................C-1 226 C-1. PBTK Models Selected for Comparison Against Di chloromethane Kinetic Data 227 Sets ...............................................

................................................... ...............................C-1 228 C-2. Model Performance.............................

................................................... ..........................C-4 229 C-2.1. Evaluation of Model Structure for Descriptio n of Carboxyhemoglobin Levels...........C-4 230 C-2.2. Evaluation of Model Structure for Prediction of Upt ake, Blood and Liver 231 Concentrations, and Expiration of Dichloromethane ..

................................................... C-5 232 C-2.3. Evaluation of Model Structure on Relative Fl ux of CYP and GST Metabolism of 233 Dichloromethane ....................................

................................................... .....................C-9 234 C-2.4. Evaluation of Model Predictions of Oral Abso rption of Dichloromethane................C-11 235 C-3. Model Option Summary ..........................

................................................... ...................C-15 236vi DRAFT - DO NOT CITE OR QUOTE 237 APPENDIX D: SUMMARY OF BENCHMARK DOSE (BMD) MODELI NG OF 238 NONCANCER ENDPOINTS ................................ .......................................... D-1 239 D-1. Oral Reference Dose: BMD Modeling of Nonneopl astic Liver Lesion Incidence 240 Data For Rats Exposed to Dichloromethane In Drinkin g Water For 2 Years 241 (Serota et al., 1986a) .............................

................................................... ..................... D-1 242 D-2. Inhalation Reference Concentration: BMD Model ing of Nonneoplastic Liver 243 Lesion Incidence Data For Rats Exposed to Dichlorom ethane by Inhalation For 2 244 Years (Nitschke et al., 1988a) .....................

................................................... ............... D-5 245 246 APPENDIX E: SUMMARY OF BENCHMARK DOSE (BMD) MODELI NG OF CANCER 247 ENDPOINTS..........................................

................................................... ........E-1 248 E-1. Oral Cancer Slope Factors: BMD Modeling of Li ver Tumor Incidence Data for 249 Mice Exposed to Dichloromethane in Drinking Water f or 2 Years (Serota et al., 250 1986b; Hazelton Laboratories, 1983)................

................................................... ..........E-1 251 E-1.1. Modeling results for the internal liver meta bolism metric...................................... ......E-3 252 E-1.2. Modeling results for the whole body metaboli sm metric .......................................... ...E-5 253 E-2. Cancer Inhalation Unit Risk: BMD Modeling of Liver and Lung Tumor 254 Incidence Data for Male Mice Exposed to Dichloromet hane via Inhalation for 2 255 Years (Mennear et al., 1988; NTP, 1986). .........

................................................... .......E-7 256 E-2.1. Modeling results for the internal liver meta bolism metric, liver tumors ......................E- 9 257 E-2.2. Modeling results for the internal lung metab olism metric, lung tumors.....................E-11 258 E-2.3. Modeling results for the whole body metaboli sm metric, liver tumors......................E-13 259 E-2.4. Modeling results for the whole body metaboli sm metric, lung tumors ......................E-15 260 261 APPENDIX F. COMPARATIVE CANCER INHALATION UNIT RISK BASED ON FEMALE 262 MICE DATA ..........................................

................................................... ........F-1 263 264 APPENDIX G: COMPARATIVE CANCER INHALATION UNIT RIS K BASED ON 265 BENIGN MAMMARY GLAND TUMORS IN RATS................ ................... G-1 266 267 APPENDIX H: SOURCE CODE AND COMMAND FILES FOR DICHL OROMETHANE 268 PBTK MODELS ........................................

................................................... ... H-1 269vii DRAFT - DO NOT CITE OR QUOTE 270 LIST OF TABLES 271 272 Table 2-1. Physical properties and chemical identi ty of dichloromethane.............................. ...... 3 273 Table 3-1. Distribution of radioactivity in tissue s 48 hours after inhalation exposure of mature 274 male Sprague-Dawley rats (n = 3) for 6 hours .......

................................................... .. 7 275 Table 3-2. Brain and perirenal fat dichloromethane and blood CO concentrations in male Wistar 276 rats exposed by inhalation to dichloromethane at co nstant exposure concentrations 277 compared with intermittently high exposure concentr ations....................................... 8 278 Table 3-3. Mean prevalences of the GST-T1 null (–/ –) genotype in human ethnic groups ........ 15 279 Table 3-4. GST-T1 enzyme activities toward dichlor omethane in human, rat, mouse, and 280 hamster tissues (liver, kidney, and erythrocytes)..

................................................... .. 17 281 Table 3-5. Values for parameter distributions in a B6C3F 1 mouse probabilistic PBTK model for 282 dichloromethane compared with associated values for point parameters in earlier 283 deterministic B6C3F 1 mouse PBTK models for dichloromethane............. ............... 29 284 Table 3-6. Internal daily doses for B6C3F 1 mice exposed to dichloromethane for 2 years 285 (6 hours/day, 5 days/week) calculated with differen t PBTK models ........................ 30 286 Table 3-7. Results of calibrating metabolic parame ters in a human probabilistic PBTK model for 287 dichloromethane with individual kinetic data for 42 exposed volunteers and MCMC 288 analysis ...........................................

................................................... ........................ 32 289 Table 3-8. Parameter distributions used in human M onte Carlo analysis for dichloromethane by 290 David et al. (2006) ................................

................................................... .................. 34 291 Table 3-9. Parameter distributions for the human P BTK model for dichloromethane used by 292 EPA................................................

................................................... ......................... 36 293 Table 3-10. Parameter values for the rat PBTK mode l for dichloromethane used by EPA ........ 39 294 Table 3-11. Parameters in the mouse, rat, and huma n PBTK model for dichloromethane used by 295 the EPA............................................

................................................... ....................... 41 296 Table 4-1. Percentage of male General Electric pla stic polymer workers reporting neurologic 297 symptoms or displaying abnormal values in measures of neurological function, 298 hepatic function, and cardiac function.............

................................................... ....... 54 299 Table 4-2. Ischemic heart disease mortality risk i n four cohorts of dichloromethane-exposed 300 workers ............................................

................................................... ....................... 56 301 Table 4-3. Suicide risk in two cohorts of dichloro methane-exposed workers............................ . 57 302 Table 4-4. Mortality risk in Eastman Kodak cellulo se triacetate film base production workers, 303 Rochester, New York................................

................................................... .............. 66 304 Table 4-5. Mortality risk by cumulative exposure i n Eastman Kodak cellulose triacetate film 305 base production workers, Rochester, New York ....... ................................................ 67 306 Table 4-6. Mortality risk in Imperial Chemical Ind ustries cellulose triacetate film base 307 production workers, Brantham, United Kingdom: 1,473 men employed 1946–1988, 308 followed through 1994..............................

................................................... .............. 70 309 Table 4-7. Mortality risk in Hoechst Celanese Corp oration cellulose triacetate fiber production 310 workers, Rock Hill, South Carolina: 1,271 men and w omen employed 1954–1977, 311 followed through 1990..............................

................................................... .............. 72 312 Table 4-8. Cancer mortality risk in Hoechst Celane se Corporation cellulose triacetate fiber 313 production workers, Cumberland, Maryland: 2,909 men and women employed 314 1970–1981, followed through 1989...................

................................................... ..... 74 315 Table 4-9. Summary of cohort studies of cancer ris k and dichloromethane exposure................ 85 316 Table 4-10. Summary of case-control studies of can cer risk and dichloromethane exposure..... 87 317viii DRAFT - DO NOT CITE OR QUOTE

Table 4-11. Incidences of histopathologic changes in livers of male and female F344 rats 318 exposed to dichloromethane in drinking water for 90 days....................................... 94 319 Table 4-12. Incidences of histopathologic changes in livers of male and female B6C3F 1 mice 320 exposed to dichloromethane in drinking water for 90 days....................................... 96 321 Table 4-13. Studies of chronic oral dichloromethan e exposures (up to 2 years)........................ . 97 322 Table 4-14. Incidences of nonneoplastic liver chan ges and liver tumors in male and female F344 323 rats exposed to dichloromethane in drinking water f or 2 years................................. 98 324 Table 4-15. Incidences for focal hyperplasia and t umors in the liver of male B6C3F 1 mice 325 exposed to dichloromethane in drinking water for 2 years...................................... 100 326 Table 4-16. Studies of chronic inhalation dichloro methane exposures .................................. ... 108 327 Table 4-17. Incidences of nonneoplastic histologic changes in male and female F344/N rats 328 exposed to dichloromethane by inhalation (6 hours/d ay, 5 days/week) for 2 years 110 329 Table 4-18. Incidences of selected neoplastic lesi ons in male and female F344/N rats exposed to 330 dichloromethane by inhalation (6 hours/day, 5 days/ week) for 2 years .................. 112 331 Table 4-19. Incidences of nonneoplastic histologic changes in B6C3F 1 mice exposed to 332 dichloromethane by inhalation (6 hours/day, 5 days/ week) for 2 years .................. 114 333 Table 4-20. Incidences of neoplastic lesions in ma le and female B6C3F 1 mice exposed to 334 dichloromethane by inhalation (6 hours/day, 5 days/ week) for 2 years .................. 115 335 Table 4-21. Incidences of selected neoplastic hist ologic changes in male and female Sprague336 Dawley rats exposed to dichloromethane by inhalatio n (6 hours/day, 5 days/week) 337 for 2 years ........................................

................................................... ..................... 119 338 Table 4-22. Incidences of selected nonneoplastic h istologic changes in male and female 339 Sprague-Dawley rats exposed to dichloromethane by i nhalation (6 hours/day, 340 5 days/week) for 2 years ...........................

................................................... ............ 121 341 Table 4-23. Incidences of selected neoplastic hist ologic changes in male and female Sprague342 Dawley rats exposed to dichloromethane by inhalatio n (6 hours/day, 5 days/week) 343 for 2 years ........................................

................................................... ..................... 123 344 Table 4-24. Summary of studies of reproductive and developmental effects of dichloromethane 345 exposure in animals ................................

................................................... .............. 125 346 Table 4-25. Reproductive outcomes in F344 rats exp osed to dichloromethane by inhalation for 347 14 weeks prior to mating and from GDs 0–21 ......... ............................................... 128 348 Table 4-26. Studies of neurobehavioral changes fro m dichloromethane, by route of exposure and 349 type of effect .....................................

................................................... .................... 137 350 Table 4-27. Studies of neurophysiological changes as measured by evoked potentials resulting 351 from dichloromethane, by route of exposure......... .................................................. 1 38 352 Table 4-28. Studies of neurochemical changes from dichloromethane, by route of exposure. 139 353 Table 4-29. Results from in vitro genotoxicity ass ays of dichloromethane with bacteria, yeast, or 354 fungi..............................................

................................................... ........................ 149 355 Table 4-30. Results from in vitro genotoxicity ass ays of dichloromethane with mammalian 356 systems, by type of test...........................

................................................... .............. 155 357 Table 4-31. Results from in vivo genotoxicity assa ys of dichloromethane in insects............... 158 358 Table 4-32. Results from in vivo genotoxicity assa ys of dichloromethane in mice.................. 160 359 Table 4-33. Results from in vivo genotoxicity assa ys of dichloromethane in rats and hamsters 360

................................................... ................................................... ........................... 164 361 Table 4-34. Comparison of in vivo dichloromethane genotoxicity assays targeted to lung or liver 362 cells, by species ..................................

................................................... .................. 165 363 Table 4-35. NOAELs and LOAELs in selected animal s tudies involving oral exposure to 364 dichloromethane for short-term, subchronic, or chro nic durations ......................... 180 365ix DRAFT - DO NOT CITE OR QUOTE

Table 4-36. NOAELs and LOAELs in animal studies in volving inhalation exposure to 366 dichloromethane for subchronic or chronic durations , hepatic, pulmonary, and 367 neurologic effects.................................

................................................... ................. 185 368 Table 4-37. NOAELs and LOAELs in selected animal s tudies involving inhalation exposure to 369 dichloromethane, reproductive and developmental eff ects ..................................... 188 370 Table 4-38. Incidence of liver tumors in male B6C3 F 1 mice exposed to dichloromethane in a 2- 371 year oral exposure (drinking water) study a

................................................... ........... 197 372 Table 4-39. Incidences of liver tumors in male and female F344 rats exposed to dichloromethane 373 in drinking water for 2 years......................

................................................... ........... 198 374 Table 4-40. Incidences of selected neoplastic lesi ons in B6C3F 1 mice exposed to 375 dichloromethane by inhalation (6 hours/day, 5 days/ week) for 2 years .................. 200 376 Table 4-41. Incidences of selected neoplastic lesi ons in F344/N rats exposed to dichloromethane 377 by inhalation (6 hours/day, 5 days/week) for 2 year s.............................................. 201 378 Table 4-42. Incidences of mammary gland tumors in two studies of male and female Sprague379 Dawley rats exposed to dichloromethane by inhalatio n (6 hours/day, 5 days/week) 380 for 2 years ........................................

................................................... ..................... 202 381 Table 4-43. Comparison of internal dose metrics in inhalation and oral exposure scenarios, in 382 male mice and rats .................................

................................................... ............... 204 383 Table 5-1. Incidence data for nonneoplastic liver lesions and internal liver doses, based on 384 various metrics, in male and female F344 rats expos ed to dichloromethane in 385 drinking water for 2 years (Serota et al., 1986a) ..

................................................... 225 386 Table 5-2. BMD modeling results for incidence of noncancer liver l esions in male and female 387 F344 rats exposed to dichloromethane in drinking wa ter for 2 years, based liver388 specific CYP metabolism dose metric (mg dichloromet hane metabolism via CYP 389 pathway per liter liver tissue per day)............

................................................... ....... 227 390 Table 5-3. RfD for dichloromethane based on PBTK m odel-derived probability distributions of 391 human drinking water exposures extrapolated from no nneoplastic liver lesion 392 incidence data for male rats exposed via drinking w ater for 2 years, based on liver393 specific CYP metabolism dose metric (mg dichloromet hane metabolized via CYP 394 pathway per liter liver tissue per day)............

................................................... ....... 228 395 Table 5-4. Potential points of departure with appl ied UFs and resulting RfDs......................... 232 396 Table 5-5. Incidence data for nonneoplastic liver lesions (hepatic vacuolation) and internal liver 397 doses, based on various metrics, in female Sprague- Dawley rats exposed to 398 dichloromethane via inhalation for 2 years (Nitschk e et al., 1988a) ....................... 239 399 Table 5-6. BMD modeling results for incidence of n oncancer liver lesions in female Sprague400 Dawley rats exposed to dichloromethane by inhalatio n for 2 years, based on liver 401 specific CYP metabolism metric (mg dichloromethane metabolized via CYP 402 pathway per liter liver tissue per day)............

................................................... ....... 241 403 Table 5-7. Inhalation RfC for dichloromethane base d on PBTK model-derived probability 404 distributions of human inhalation exposure extrapol ated from nonneoplastic liver 405 lesion data for female rats exposed via inhalation for 2 years, based on liver-specific 406 CYP metabolism dose metric (mg dichloromethane meta bolized via CYP pathway 407 per liter liver tissue per day) ....................

................................................... ............. 242 408 Table 5-8. Potential points of departure with appl ied UFs and resulting RfCs ......................... 248 409 Table 5-9. Statistical characteristics of human eq uivalent applied doses in specific populations 410 of the GST-T1 -/- group............................................. ............................................... 260 411 Table 5-10. Statistical characteristics of HECs in specific populations of the GST-T1 -/- group 412

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Table 5-11. Incidence data for liver tumors and in ternal liver doses, based on GST metabolism 414 dose metrics, in male B6C3F 1 mice exposed to dichloromethane in drinking water 415 for 2 years ........................................

................................................... ..................... 267 416 Table 5-12. BMD modeling results and tumor risk fa ctors for internal dose metric associated 417 with 10% extra risk for liver tumors in male B6C3F 1 mice exposed to 418 dichloromethane in drinking water for 2 years, base d on liver-specific GST 419 metabolism and whole body GST metabolism dose metri cs................................... 269 420 Table 5-13. Cancer OSFs for dichloromethane based o n PBTK model-derived internal liver 421 doses in B2C3F1 mice exposed via drinking water for2 years, based on liver422 specific GST meabolism and whole body metabolism do se metrics, by population 423 genotype .........................................

................................................... ..................... 272 424 Table 5-14. Alternative route-to-route cancer OSFs for dichloromethane extrapolated from male 425 B6C3F 1 mouse inhalation liver tumor incidence data using a tissue-specific GST 426 metabolism dose metric, by population genotype ..... .............................................. 274 427 Table 5-15. Cancer OSF based on a human BMDL 10 using administered dose for liver tumors in 428 male B6C3F 1 mice exposed to dichloromethane in drinking water for 2 years ...... 275 429 Table 5-16. Comparison of OSFs derived using vario us assumptions and metrics, based on 430 tumors in male mice................................

................................................... .............. 277 431 Table 5-17. Incidence data for liver and lung tumo rs and internal doses, based on GST 432 metabolism dose metrics, in male and female B6C3F 1 mice exposed to 433 dichloromethane via inhalation for 2 years .........

................................................... . 280 434 Table 5-18. BMD modeling results and tumor risk fa ctors associated with 10% extra risk for 435 liver and lung tumors in male and female B6C3F 1 mice exposed by inhalation to 436 dichloromethane for 2 years, based on liver-specifi c GST metabolism and whole 437 body GST metabolism dose metrics...................

................................................... .. 283 438 Table 5-19. IURs for dichloromethane based on PBTK model-derived internal liver and lung 439 doses in B6C3F 1 male mice exposed via inhalation for 2 years, base d on liver440 specific GST metabolism and whole body metabolism d ose metrics, by population 441 genotype...........................................

................................................... ..................... 286 442 Table 5-20. Upper bound estimates of combined huma n IURs for liver and lung tumors 443 resulting from lifetime exposure to 1 μ g/m 3 dichloromethane, based on liver-specific 444 GST metabolism and whole body metabolism dose metri cs, by population genotype 445

................................................... ................................................... ........................... 289 446 Table 5-21. Inhalation units risks based on human BMDL 10 values using administered 447 concentration for liver and lung tumors in B6C3F 1 mice exposed by inhalation to 448 dichloromethane for 2 years ........................

................................................... ......... 291 449 Table 5-22. Comparison of IURs derived by using va rious assumptions and metrics.............. 293 450 Table 5-23. Comparison of key B6C3F 1 mouse parameters differing between prior and curre nt 451 PBTK model application .............................

................................................... ......... 295 452 Table 5-24. Application of ADAFs to dichloromethan e cancer risk following a lifetime (70- 453 year) oral exposure ................................

................................................... ............... 297 454 Table 5-25. Application of ADAFs to dichloromethan e cancer risk following a lifetime (70- 455 year) inhalation exposure..........................

................................................... ............ 298 456 Table 5-26. Summary of uncertainty in the derivati on of cancer risk values for dichloromethane 457

................................................... ................................................... ........................... 299 458 Table 5-27. Statistical characteristics of human i nternal doses for 1 mg/kg-day oral exposures in 459 specific populations ...............................

................................................... ............... 308 460 Table 5-28. Statistical characteristics of human i nternal doses for 1 mg/m 3 inhalation exposures 461 in specific subpopulations.........................

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Table 6-1. Comparison of OSFs derived by using var ious assumptions and metrics, based on 463 liver tumors in male mice ..........................

................................................... ........... 319 464 Table 6-2. Comparison of IURs derived by using var ious assumptions and metrics................ 322 465 466 Table B-1. Parameter distributions used in human M onte Carlo analysis for 467 dichloromethane by David et al. (2006) .............

................................................... ............ B-2 468 469 Table B-2. Parameters for BW distributions as func tions of age and gender ......................... B- 10 470 471 Table B-3. Parameter distributions for the human P BTK model used by the EPA................. B-18 472 473 Table C-1. Parameter values used in rat PBTK model s.................................................. ........... C-3 474 475 Table C-2. Effect of PBTK model configuration on p redicted dichloromethane metabolite 476 production in the liver of rats from inhalation of 200 or 1,000 ppm dichloromethane 477 for 4 hours (Andersen et al., 1991) or 2,000 or 4,0 00 ppm dichloromethane for 478 6 hours/day, 5 days/week, for 2 years (NTP, 1986)..

................................................... .... C-10 479 480 Table C-3. Observations and predictions of total e xpired dichloromethane resulting from 481 gavage doses in rats ...............................

................................................... ....................... C-15 482 483 Table D-1. Incidence data for nonneoplastic liver lesions and internal liver doses, based 484 on various metrics, in male and female F344 rats ex posed to dichloromethane in 485 drinking water for 2 years (Serota et al., 1986a) ..

................................................... .......... D-1 486 487 Table D-2. BMD modeling results for incidence of noncancer liver l esions in male and female 488 F344 rats exposed to dichloromethane in drinking wa ter for 2 years, based liver-specific 489 CYP metabolism dose metric (mg dichloromethane meta bolism via CYP pathway per liter 490 liver tissue per day)..............................

................................................... ........................... D-2 491 492 Table D-3. Incidence data for nonneoplastic liver lesions (hepatic vacuolation) and internal liver 493 doses, based on various metrics, in female Sprague- Dawley rats exposed to 494 dichloromethane via inhalation for 2 years (Nitschk e et al., 1988a) ................................. D- 5 495 496 Table D-4. BMD modeling results for incidence of n oncancer liver lesions in female 497 Sprague-Dawley rats exposed to dichloromethane by i nhalation for 2 years, based 498 on liver specific CYP metabolism metric (mg dichlor omethane metabolized via 499 CYP pathway per liter liver tissue per day)........

................................................... ............ D-6 500 501 Table E-1. Incidence data for liver tumors and int ernal liver doses, based on GST 502 metabolism dose metrics, in male B6C3F 1 mice exposed to dichloromethane in 503 drinking water for 2 years.........................

................................................... .......................E-1 504 505 Table E-2. BMD modeling results and tumor risk fac tors for internal dose metric 506 associated with 10% extra risk for liver tumors in male B6C3F 1 mice exposed to 507 dichloromethane in drinking water for 2 years, base d on liver-specific GST 508 metabolism and whole body GST metabolism dose metri cs..............................................E-2 509 510xii DRAFT - DO NOT CITE OR QUOTE

Table E-3. Incidence data for liver and lung tumor s and internal doses, based on GST 511 metabolism dose metrics, in male B6C3F1 mice expose d to dichloromethane via 512 inhalation for 2 years .............................

................................................... ..........................E-7 513 514 Table E-4. BMD modeling results and tumor risk fac tors associated with 10% extra risk 515 for liver and lung tumors in male B6C3F 1 mice exposed by inhalation to 516 dichloromethane for 2 years, based on liver-specifi c GST metabolism and whole 517 body GST metabolism dose metrics ...................

................................................... .............E-8 518 519 Table F-1. Incidence data for liver and lung tumor s and internal doses, based on GST 520 metabolism dose metrics, in female B6C3F 1 mice exposed to dichloromethane via 521 inhalation for 2 years .............................

................................................... ..........................F-1 522 523 Table F-2. BMD modeling results and tumor risk fact ors associated with 10% extra risk for liver 524 and lung tumors in female B6C3F 1 mice exposed by inhalation to dichloromethane for 2 525 years, based on liver-specific GST metabolism and w hole body GST metabolism dose 526 metrics............................................

................................................... ..................................F-3 527 528 Table F-3. IURs for dichloromethane based on PBTK model-derived internal liver and lung 529 doses in B6C3F 1 female mice exposed via inhalation for 2 years, based on l iver-specific 530 GST metabolism and whole body metabolism dose metri cs, by population genotype .....F-4 531 532 Table F-4. Upper bound estimates of combined human IURs for l iver and lung tumors resulting from 533 lifetime exposure to 1 μ g/m 3 dichloromethane, based on liver-specific GST metabo lism and whole 534 body metabolism dose metrics, by population genoty pe, using female mouse data for derivation of 535 risk factors

................................................... ................................................... ......................F-5 536 537 Table G-1. Incidence data for mammary gland tumors and internal doses, based on different dose 538 metrics, in male and female F344 rats exposed to di chloromethane via inhalation for 539 2 years ............................................

................................................... ................................. G-1 540 541 Table G-2. BMD modeling results associated with 10% extra risk for mammary gland tumors in 542 F344 rats exposed by inhalation to dichloromethane for 2 years, based on AUC for 543 dichloromethane in blood ...........................

................................................... .................... G-3 544 545 Table G-3. IURs for dichloromethane, based on benig n mammary tumors and PBTK model546 derived internal doses in F344N rats exposed via in halation for 2 years, based on AUC for 547 dichloromethane in blood dose metric...............

................................................... ............. G-4 548 549 550xiii DRAFT - DO NOT CITE OR QUOTE LIST OF FIGURES 551 552 Figure 3-1. Proposed pathways for dichloromethane metabolism......................................... ...... 10 553 Figure 3-2. Schematics of PBTK models (1986–2006) used in the development of estimates for 554 dichloromethane internal dosimetry. ................

................................................... ...... 22 555 Figure 3-3. Schematic of mouse PBTK model used by Marino et al. (2006).............................. 2 7 556 Figure 3-4. Schematic of human PBTK, used by David et al. (2006). .................................... .... 31 557 Figure 3-5. Schematic of rat PBTK model used in cu rrent assessment................................... .... 38 558 Figure 5-1. Exposure response array for oral expos ure to dichloromethane ............................. 219 559 Figure 5-2. Process for deriving noncancer oral Rf Ds and inhalation RfCs using rodent and 560 human PBTK models..................................

................................................... .......... 221 561 Figure 5-3. PBTK model-derived internal doses (mg dichloromethane metabolized via the CYP 562 pathway per liter liver per day) in rats and humans and their associated external 563 exposures (mg/kg-day), used for the derivation of R fDs. ....................................... 226 564 Figure 5-4. Comparison of RfDs derived from select ed point of departures for endpoints 565 presented in Table 5-4. ............................

................................................... ............. 233 566 Figure 5-5. Exposure response array for chronic (a nimal) or occupational (human) inhalation 567 exposure to dichloromethane (log Y axis)...........

................................................... . 235 568 Figure 5-6. Exposure response array for subacute t o subchronic inhalation exposure to 569 dichloromethane (log Y axis) .......................

................................................... ........ 236 570 Figure 5-7. PBTK model-derived internal doses (mg dichloromethane metabolized via the CYP 571 pathway/L liver/day) in rats and humans versus exte rnal exposures (ppm)............ 240 572 Figure 5-8. Comparison of RfCs derived from select ed point of departures for endpoints 573 presented in Table 5-8. ............................

................................................... ............. 249 574 Figure 5-9. Comparison dichloromethane oxidation r ate data with alternate kinetic models. .. 253 575 Figure 5-10. Sensitivity coefficients for long-ter m mass CYP- and GST-mediated metabolites 576 per liver volume from a daily drinking water concen tration of 10 mg/L in rats. .... 257 577 Figure 5-11. Sensitivity coefficients for long-ter m mass CYP- and GST-mediated metabolites 578 per liver volume from a long-term average daily inh alation concentration of 500 ppm 579 in rats. (KA is not included since it has no impac t on inhalation dosimetry.)........ 257 580 Figure 5-12. Frequency density of human equivalent applied doses in specific populations in 581 comparison to a general population (0.5- to 80-year -old males and females) estimate 582 for an internal dose of 15.1 mg dichloromethane met abolized by CYP per liter liver 583 per day; all groups were restricted to the GST-T1 -/- population)............................ 260 584 Figure 5-13. Frequency density of HECs in specific populations in comparison to a general 585 population (0.5- to 80-year-old males and females) estimate for an internal dose of 586 128.1 mg dichloromethane metabolized by CYP per lit er liver per day ; all groups 587 restricted to the GST-T1 -/- population). ...................................... ............................ 262 588 Figure 5-14. Process for deriving cancer OSFs and IURs by using rodent and human PBTK 589 models.............................................

................................................... ...................... 265 590 Figure 5-15. PBTK model-derived internal doses (mg dichloromethane metabolized via the GST 591 pathway per liter liver per day) in mice and humans and their associated external 592 exposures (mg/kg-day) used for the derivation of ca ncer OSFs based on liver tumors 593 in mice. ...........................................

................................................... ...................... 268 594 Figure 5-16. PBTK model-derived internal doses (mg dichloromethane metabolized via the GST 595 pathways per liter tissue per day) for liver (A) an d lung (B) in mice and humans, and 596 their associated external exposures (ppm), used for the derivation of cancer 597 inhalation unit risks..............................

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Figure 5-20. Histograms for a liver-specific dose of GST metabolism (mg GST metabolites per 599 liter liver per day) for the general population (0. 5- to 80-year-old males and females) 600 and specific age/gender groups, within the populati on of GST-T1 +/+ genotypes, given 601 a daily oral dose-rate of 1 mg/kg-day dichlorometha ne. ......................................... 308 602 Figure 5-21. Histograms for liver-specific dose of GST metabolism (mg GST metabolites per 603 liter liver per day) for the general population (0. 5- to 80-year-old males and females) 604 and specific age/gender groups, within the populati on of GST-T1 +/+ genotypes, given 605 a continuous inhalation exposure to 1 mg/m 3 dichloromethane.............................. 309 606 607 Figure B-1. Schematic of the David et al. (2006) P BTK model for dichloromethane in the 608 human ..............................................

................................................... ................................B-1 609 610 Figure B-2. Total CYP2E1 activity (Vmax) normalized to the average total activity in 14-18 611 year-old individuals (Vmax(14-18)) plotted against normalized body weight (BW) for 612 individuals ranging from six months to 18 years of age. Data source: Johnsrud et al. 613 (2003) .............................................

................................................... ................................B-6 614 615 Figure B-3. Body-weight scaled CYP2E1 activity (V max c) normalized to the average scaled 616 activity in 14-18 year-old individuals (V max c (14-18)) plotted against age individuals 617 ranging from six months to 18 years of age. Data s ource: Johnsrud et al. (2003)..........B-7 618 619 Figure B-4. U.S. age distribution, 6 months-80 yea rs (values from U.S. Census Bureau).........B-9 620 621 Figure B-5. U.S. age-specific gender distribution (values from U.S. Census Bureau) ..............B-9 622 623 Figure B-6. Function fits to age-dependent data fo r BW mean and standard deviations 624 for males and females in the United States (values from Portier et al., 2007).................B-11 625 626 Figure B-7. Example body weight (BW) histogram fro m Monte Carlo simulation for 0.5- 627 to 80-year-old males and females in the United Stat es (simulated n = 10,000)...............B-12 628 629 Figure B-8. Mean value respiration rates for males and females as a function of age (values from 630 Clewell et al., 2004) ..............................

................................................... ........................B-13 631 632 Figure B-9. Geometric standard deviations (GSDs) f or respiration rates for males and 633 females as a function of age (values from Arcus-Art h and Blaisdell, 2007) ...................B-13 634 635 Figure B-10. Fraction body fat (VFC) over various age ranges in males and females (data 636 from Clewell et al., 2004).........................

................................................... .....................B-15 637 638 Figure B-11. Fraction liver (VLC) as a function of age (data from Clewell et al., 2004) ........B-16 639 640 Figure C-1. Schematic of the Andersen et al. (1991 ) PBTK model for dichloromethane 641 in the rat .........................................

................................................... ..................................C-2 642 643 Figure C-2. Observations of exhaled dichloromethane (DCM) and carbon monoxide (CO) after a 644 bolus oral dose of 200 mg/kg in rats (data of Angel o et al., 1986b). ................................C-5 645 646xv DRAFT - DO NOT CITE OR QUOTE

Figure C-3. Observations and predictions (models A -D) of Gargas et al. (1986) data for 647 respiratory uptake by 3 rats of 100, 500, 1000, or 3000 ppm dichloromethane in a 9- 648 L closed chamber. ...............................

................................................... ..........................C-6 649 650 Figure C-4. Observations (points) and predictions (curves, Models A-D) of Angelo et al. 651 (1986b) data for dichloromethane in blood following 10 and 50 mg/kg iv injection 652 in rats. Model predictions with doses at 56% of th e nominal doses, i.e., 5.6 and 28 653 mg/kg, are shown for comparison as dashed lines for model D.........................................C- 7 654 655 Figure C-5. Observations and predictions (Models A -D) of Andersen et al. (1987) data for 656 dichloromethane in rat blood from inhalation of 200 and 1000 ppm dichloromethane for 4 657 hours..............................................

................................................... ...................................C-8 658 659 Figure C-6. Observations and predictions (Models A -D) of Angelo et al. (1986b) data for 660 percent of dichloromethane dose expired as dichloro methane following 10 and 661 50 mg/kg iv injection in rats. Model predictions w ith doses at 56% of the nominal 662 doses, i.e., 5.6 and 28 mg/kg, are shown for compar ison as dashed lines for model 663 D..................................................

................................................... .....................................C-9 664 665 Figure C-7. Observations and Model D predictions o f Angelo et al. (1986b) data for (A) percent 666 dose expired as dichloromethane, (B) blood dichloro methane, (C) percent expired as 667 carbon monoxide, and (D) liver dichloromethane in r ats given a single dichloromethane 668 gavage doses of 50 and 200 mg/kg, using a numerical ly fitted GI absorption rate (K A ) of 669 1.8. ............................................

................................................... ...................................C-12 670 671 Figure C-8. Model predictions with of blood carbox y-hemoglobin (COHb, percent of 672 total Hb) from a single gavage dose of 526 mg/kg DC M in rats, compared to the 673 data of Pankow et al. (1991). Model simulations pe rformed with Model D (heavy 674 read line, KA = 1.8/h) or with an alternate value o f KA = 0.47/h, fit to these data. ........C-13 675 676 Figure D-1. Predicted (logistic model) and observe d incidence of noncancer liver lesions in male 677 F344 rats exposed to dichloromethane in drinking wa ter for 2 years 678 (Serota et al., 1986a) .............................

................................................... ......................... D-3 679 680 Figure D-2. Predicted (log-probit model) and obser ved incidence of noncancer liver lesions in 681 female Sprague-Dawley rats inhaling dichloromethane for 2 years 682 (Nitschke 1988a) ...................................

................................................... ....................... D-7 683 684 Figure E-1. Predicted and observed incidence of an imals with hepatocellular carcinoma or 685 adenoma in male B6C3F 1 mice exposed to dichloromethane in drinking water 686 for 2 years, using liver-specific metabolism dose m etric (Serota et al., 1986b; Hazelton 687 Laboratories, 1983)................................

................................................... ..........................E-3 688 689 Figure E-2. Predicted and observed incidence of an imals with hepatocellular carcinoma or 690 adenoma in male B6C3F 1 mice exposed to dichloromethane in drinking water 691 for 2 years, using whole-body metabolism dose metri c (Serota et al., 1986b; Hazelton 692 Laboratories, 1983)................................

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Figure E-3. Predicted and observed incidence of ani mals with hepatocellular carcinoma or 695 adenoma in male B6C3F 1 mice exposed by inhalation to dichloromethane for 2 years, 696 using liver-specific metabolism dose metric (Mennea r et al., 1988; NTP, 1986)..............E-9 697 698 Figure E-4. Predicted and observed incidence of an imals with carcinoma or adenoma in 699 the lung of male B6C3F 1 mice exposed by inhalation to dichloromethane for 2 700 years, using liver-specific metabolism dose metric (Mennear et al., 1988; NTP, 701 1986)..............................................

................................................... ................................E-11 702 703 Figure E-5. Predicted and observed incidence of ani mals with hepatocellular carcinoma or 704 adenoma in male B6C3F 1 mice exposed by inhalation to dichloromethane for 2 years, 705 using whole-body metabolism dose metric (Mennear et al., 1988; NTP, 1986)..............E-13 706 707 Figure E-6. Predicted and observed incidence of an imals with carcinoma or adenoma in 708 the lung of male B6C3F 1 mice exposed by inhalation to dichloromethane for 2 709 years, using whole-body metabolism dose metric (Men near et al., 1988; NTP, 710 1986)..............................................

................................................... ................................E-15 711 712 Figure G-1. PBTK model-derived internal doses (dai ly average AUC for dichloromethane in 713 blood) in rats and humans, and their associated ext ernal exposures (ppm) used for the 714 derivation of cancer IURs, based on mammary tumors in rats. ........................................ G-2 715 716xvii DRAFT - DO NOT CITE OR QUOTE LIST OF ACRONYMS 717 718 719 ACGIH American Conference of Governmental Industrial Hyg ienists 720 ADAF age-dependent adjustment factor 721 AEGL acute exposure guideline level 722 AIC Akaike’s Information Criterion 723 ALT alanine aminotransferase 724 AP alkaline phosphatase 725 AST aspartate aminotransferase 726 ATSDR Agency for Toxic Substances and Disease Registry 727 AUC area under the curve of a concentration versus tim e plot 728 BAER brainstem-auditory evoked response 729 BMD benchmark dose 730 BMDL 10 95% lower bound on the BMD 731 BMR benchmark response level 732 BW body weight 733 CAEP cortical-auditory-evoked potential 734 CASRN Chemical Abstracts Service Registry Number 735 CHO Chinese hamster ovary 736 CI confidence interval 737 CMR Chemical Marketing Reporter 738 CNS central nervous system 739 COHb carboxyhemoglobin 740 CV coefficient of variation 741 CYP cytochrome P450 742 DNA deoxyribonucleic acid 743 EPA U.S. Environmental Protection Agency 744 FEP flash-evoked potential 745 FOB functional observational battery 746 GD gestation day 747 GSH reduced glutathione 748 GST glutathione S-transferase 749 HEC human equivalent concentration 750 HPRT hypoxanthine-guanine phosphoribosyl transferase 751 IARC International Agency for Research on Cancer 752 ICD-9 International Classification of Diseases 9 th ed. 753 IgM immunoglobulin M 754 IRIS Integrated Risk Information System 755 IUR inhalation unit risk 756 LOAEL lowest-observed-adverse-effect level 757 LOH loss of heterozygosity 758 MCHC mean corpuscular hemoglobin concentration 759 MCMC Markov Chain Monte Carlo 760 Mg milligrams 761 mRNA messenger ribonucleic acid 762 NADPH nicotinamide adenine dinucleotide phosphate 763 NIOSH National Institute of Occupational Safety and Heal th 764 NLM National Library of Medicine 765xviii DRAFT - DO NOT CITE OR QUOTE

NOAEL no-observed-adverse-effect level 766 NRC National Research Council 767 NTP National Toxicology Program 768 OR odds ratio 769 OSF oral slope factor 770 OSHA Occupational Safety and Health Administration 771 PBTK physiologically based toxicokinetic 772 PND postnatal day 773 QCC cardiac output 774 RfC reference concentration 775 RfD reference dose 776 SD standard deviation 777 SEM standard error of the mean 778 SEP somatosensory-evoked potential 779 SMR standardized mortality ratio 780 SRC Syracuse Research Corporation 781 SSB single-strand break 782 TWA time-weighted average 783 UF uncertainty factor 784 VPR ventilation:perfusion ratio 785xix DRAFT - DO NOT CITE OR QUOTE FOREWORD 786 787 The purpose of this Toxicological Review is to prov ide scientific support and rationale 788 for the hazard and dose-response assessment in IRIS pertaining to exposure to dichloromethane. 789 It is not intended to be a comprehensive treatise o n the chemical or toxicological nature of 790

dichloromethane. 791 The intent of Section 6, Major Conclusions in the Characterization of Hazard and Dose 792 Response, is to present the major conclusions reached in the derivation of the reference dose, 793 reference concentration and cancer assessment, wher e applicable, and to characterize the overall 794 confidence in the quantitative and qualitative aspe cts of hazard and dose response by addressing 795 the quality of data and related uncertainties. The discussion is intended to convey the limitations 796 of the assessment and to aid and guide the risk ass essor in the ensuing steps of the risk 797

assessment process. 798 For other general information about this assessment or other questions relating to IRIS, 799 the reader is referred to EPA’s IRIS Hotline at (20 2) 566-1676 (phone), (202) 566-1749 (fax), or 800 hotline.iris@epa.gov (email address). 801

802

803xx DRAFT - DO NOT CITE OR QUOTE AUTHORS, CONTRIBUTORS, AND REVIEWERS 804 805 806 CHEMICAL MANAGERS 807 808 Glinda S. Cooper, Ph.D. 809 Ambuja S. Bale, Ph.D., DABT 810 Office of Research and Development, IRIS Program 811 U.S. Environmental Protection Agency 812 Washington, DC 813 814 AUTHORS 815 816 Glinda S. Cooper, Ph.D. 817 Ambuja S. Bale, Ph.D., DABT 818 Andrew Rooney, Ph.D. 819 Paul Schlosser, Ph.D. 820 Allan Marcus, Ph.D. 821 Gene (Ching-Hung) Hsu, Ph.D., DABT 822 National Center for Environmental Assessment 823 Office of Research and Development 824 U.S. Environmental Protection Agency 825 Washington, DC 826 827 John C. Lipscomb, Ph.D., DABT 828 National Center for Environmental Assessment 829 U.S. Environmental Protection Agency 830 Cincinnati, OH 831 832 Peter McClure, Ph.D., DABT 833 Michael Lumpkin, Ph.D. 834 Fernando Llados, Ph.D. 835 Mark Osier, Ph.D., DABT 836 Daniel Plewak, B.S. 837 Syracuse Research Corporation 838 Syracuse, NY 839 840 Elizabeth Dupree Ellis, Ph.D. 841 Oak Ridge Institute for Science and Education 842 Center for Epidemiologic Research 843 Oak Ridge, TN 844 845

REVIEWERS 846 This document has been peer reviewed by EPA scienti sts. 847xxi DRAFT - DO NOT CITE OR QUOTE

INTERNAL EPA REVIEWERS 848 849 Ghazi Dannan, Ph.D 850 Karen Hogan, M.S. 851 Jennifer Jinot, Ph.D 852 Paul White, Ph.D 853 Samantha Jones, Ph.D. 854 Jamie Strong, Ph.D. 855 National Center for Environmental Assessment 856 Office of Research and Development 857 U.S. Environmental Protection Agency 858 859 David Herr, Ph.D 860 National Health and Environmental Effect Research L aboratory 861 Office of Research and Development 862 U.S. Environmental Protection Agency 863 864 865 8662/13/2008 1 DRAFT - DO NOT CITE OR QUOTE 1. INTRODUCTION 867 868

869 This document presents background information and j ustification for the Integrated Risk 870 Information System (IRIS) Summary of the hazard and dose-response assessment of 871 dichloromethane. IRIS Summaries may include oral r eference dose (RfD) and inhalation 872 reference concentration (RfC) values for chronic an d other exposure durations, and a 873 carcinogenicity assessment. 874 The RfD and RfC, if derived, provide quantitative i nformation for use in risk assessments 875 for health effects known or assumed to be produced through a nonlinear (presumed threshold) 876 mode of action. The RfD (expressed in units of mg/ kg-day) is defined as an estimate (with 877 uncertainty spanning perhaps an order of magnitude) of a daily exposure to the human 878 population (including sensitive subgroups) that is likely to be without an appreciable risk of 879 deleterious effects during a lifetime. The inhalat ion RfC (expressed in units of mg/m 3 ) is 880 analogous to the oral RfD, but provides a continuou s inhalation exposure estimate. The 881

inhalation RfC considers toxic effects for both the respiratory system (portal of entry) and for 882 effects peripheral to the respiratory system (extra respiratory or systemic effects). Reference 883 values are generally derived for chronic exposures (up to a lifetime), but may also be derived for 884 acute ( ≤ 24 hours), short-term (>24 hours up to 30 days), an d subchronic (>30 days up to 10% of 885 lifetime) exposure durations, all of which are deri ved based on an assumption of continuous 886 exposure throughout the duration specified. Unless specified otherwise, the RfD and RfC are 887 derived for chronic exposure duration. 888 The carcinogenicity assessment provides information on the carcinogenic hazard 889 potential of the substance in question and quantita tive estimates of risk from oral and inhalation 890 exposure may be derived. The information includes a weight-of-evidence judgment of the 891 likelihood that the agent is a human carcinogen and the conditions under which the carcinogenic 892 effects may be expressed. Quantitative risk estima tes may be derived from the application of a 893 low-dose extrapolation procedure. If derived, the oral slope factor is a plausible upper bound on 894 the estimate of risk per mg/kg-day of oral exposure . Similarly, an inhalation unit risk is a 895 plausible upper bound on the estimate of risk per μ g/m 3 air breathed. 896 Development of these hazard identification and dose -response assessments for 897 dichloromethane has followed the general guidelines for risk assessment as set forth by the 898 National Research Council (1983). EPA Guidelines a nd Risk Assessment Forum Technical 899 Panel Reports that may have been used in the develo pment of this assessment include the 900 following: Guidelines for the Health Risk Assessment of Chemic al Mixtures (U.S. EPA, 1986a), 901 Guidelines for Mutagenicity Risk Assessment (U.S. EPA, 1986b), Recommendations for and 902 Documentation of Biological Values for Use in Risk Assessment (U.S. EPA, 1988a), Guidelines 903 for Developmental Toxicity Risk Assessment (U.S. EPA, 1991), Interim Policy for Particle Size 9042 DRAFT - DO NOT CITE OR QUOTE

and Limit Concentration Issues in Inhalation Toxici ty Studies (U.S. EPA, 1994a), Methods for 905 Derivation of Inhalation Reference Concentrations a nd Application of Inhalation Dosimetry 906 (U.S. EPA, 1994b), Use of the Benchmark Dose Approach in Health Risk A ssessment (U.S. EPA, 907 1995), Guidelines for Reproductive Toxicity Risk Assessmen t (U.S. EPA, 1996), Guidelines for 908 Neurotoxicity Risk Assessment (U.S. EPA, 1998a), Science Policy Council Handbook : Risk 909

Characterization (U.S. EPA, 2000a), Benchmark Dose Technical Guidance Document (U.S. 910 EPA, 2000b), Supplementary Guidance for Conducting Health Risk A ssessment of Chemical 911

Mixtures (U.S. EPA, 2000c), A Review of the Reference Dose and Reference Concen tration 912

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

, 913 Supplemental Guidance for Assessing Susceptibility from Early-Life Exposure to Carcinogens 914 (U.S. EPA, 2005b), Science Policy Council Handbook: Peer Review (U.S. EPA, 2006a), and A 915 Framework for Assessing Health Risk of Environmenta l Exposures to Children (U.S. EPA, 916 2006b). 917 The literature search strategy employed for this co mpound was based on the Chemical 918 Abstracts Service Registry Number (CASRN) and at le ast one common name. Any pertinent 919 scientific information submitted by the public to t he IRIS Submission Desk was also considered 920 in the development of this document. The relevant literature was reviewed through April 2009. 9213 DRAFT - DO NOT CITE OR QUOTE 2. CHEMICAL AND PHYSICAL INFORMATION 922 923 924 Dichloromethane is a colorless liquid with a penet rating ether-like odor (Lewis, 925 1997). 1 Selected chemical and physical properties of dich loromethane are listed in Table 2-1. 926

927 Table 2-1. Physical properties and chemical identi ty of dichloromethane

Physical property/chemical identity Reference CAS number 75-09-2 Lide (2000) Synonyms methylene chloride, methylene dichloride, methyl bichloride O’Neil et al. (2001)

Molecular weight 84.93 O’Neil et al. (2001) Chemical formula CH 2 Cl 2 O’Neil et al. (2001) Boiling point 40°C Lide (2000) Melting point −95.1°C Lide (2000) Vapor pressure 1.15 × 10 2 mm Hg at 25°C Boublik et al. (1984) Density 1.3266 g/mL at 20°C Lide (2000) Vapor density 2.93 (air = 1.02) Holbrook (2003) Water solubility 1.30 × 10 4 mg/L at 25°C Horvath (1982) Other solubility Miscible in ethanol, ether, and dimethylformamide; soluble in carbon tetrachloride International Agency for Research on Cancer (IARC) (1999) Partition coefficient log K ow = 1.25 Hansch et al. (1995) Flash point Not flammable U.S. Coast Guard (1999) Auto ignition temperature 640°C Holbrook (2003) Latent heat of vaporization 3.30 × 10 5 J/kg U.S. Coast Guard (1999) Heat of fusion 16.89 cal/g U.S. Coast Guard (1999 ) Critical temperature 245.0°C Holbrook (2003) Critical pressure 6.171 × 10 6 Pa Holbrook (2003) Viscosity 0.430 cP at 20°C Lewis (1997) Henry’s constant 3.25 × 10 –3 atm m 3 /mol at 25°C Leighton and Calo (1981) OH reaction rate constant 1.42 × 10 –13 cm 3 /molecule sec at 25°C Atkinson (1989) Chemical structure



# C Cl

Cl H

H

928 929 Dichloromethane is produced by two methods of manu facturing (International Agency 930 for Research on Cancer [IARC], 1999). The older me thod involves the direct reaction of 931 methane with chlorine either at high temperatures o r at lower temperatures under catalytic or 932 photolytic conditions (Holbrook, 2003). The more c ommon method used today involves an 933 initial reaction of hydrochloric acid with methanol to yield methyl chloride. Excess methyl 934 chloride is then reacted in the gas phase thermally with chlorine to produce dichloromethane 935 (Holbrook, 2003). This process can also be carried out catalytically or photolytically. 9364 DRAFT - DO NOT CITE OR QUOTE

Dichloromethane became an important industrial che mical in the U.S. during 937 World War II (Hardie, 1964). Dichloromethane has b een used in paint strippers and removers, 938 as a propellant in aerosols, in the manufacture of drugs, pharmaceuticals, film coatings, 939 electronics, and polyurethane foam, and as a metal- cleaning solvent. Dichloromethane can also 940 be used in the decaffeination process of coffee and tea (ATSDR, 2000). The U.S. production 941 was 3.8 million pounds in 1941 and 8.3 million poun ds in 1944 (Searles and McPhail, 1949). 942 Dichloromethane production rose sharply in the deca des following the war due to the increased 943 demand for this substance for use mainly in paint s trippers (Hardie, 1964; Searles and McPhail, 944 1949). U.S. production in 1947, 1955, 1960, and 19 62 was approximately 19, 74, 113, and 945 144 million pounds, respectively (Hardie, 1964; Sea rles and McPhail, 1949). As other solvent 946 uses and its use in aerosol propellants became impo rtant, demand for this substance increased 947 further (Anthony, 1979). Dichloromethane productio n continued to rise dramatically through the 948 1970s; production capacities were 520 million pound s in 1973 and 830 million pounds in 1979 949 (Chemical Marketing Reporter [CMR], 1979, 1973). 950 After 1980, production of dichloromethane began to decline. Production capacities fell 951 from 722 million pounds in 1982 to 465 million poun ds in 1997 (CMR, 1997, 1982). The total 952 U.S. production capacity for dichloromethane in 200 0 was 535 million pounds (CMR, 2000). 953

The demand for dichloromethane decreased from 600 m illion pounds in 1979 to 200 million 954 pounds in 1999 (CMR, 2000, 1979). The decline in p roduction of and demand for 955 dichloromethane over the past 2 decades has been at tributed to increased regulation, the use of 956 alternative chemicals in aerosol spray cans, and co ncern over dichloromethane carcinogenicity 957 (Holbrook, 2003; ATSDR, 2000). 958 Dichloromethane in the environment will partition mainly to air (National Library of 959 Medicine [NLM], 2003). In air, dichloromethane exi sts as a vapor. Some of the 960 dichloromethane released to soil or water is expect ed to volatilize to air. In soil, 961 dichloromethane is expected to be highly mobile and may migrate to groundwater. The potential 962 for dichloromethane to bioconcentrate in aquatic or marine organisms is low. Dichloromethane 963 may biodegrade in soil or water under both aerobic and anaerobic conditions. 964

1 To avoid confusion, “dichloromethane” is used thro ughout this summary even if a specific paper used t he term “methylene chloride.”5 DRAFT - DO NOT CITE OR QUOTE 3. TOXICOKINETICS 965

966 967

3.1. ABSORPTION 968 3.1.1. Oral — Gastrointestinal Tract Absorption 969 There are currently no data available on absorption of dichloromethane following oral 970 intake in humans. However, after oral administrati on in animals, dichloromethane is rapidly and 971 nearly completely absorbed in the gastrointestinal tract (Angelo et al., 1986a, b; McKenna and 972 Zempel, 1981). Angelo et al. (1986b) reported that , following administration of single 973 radiolabeled oral doses (10, 50, or 200 mg/kg) to m ature male F344 rats, 97% of the label was 974 detected in the exhaled air within 24 hours, indica ting nearly complete absorption. At several 975 time points within 40 minutes of dose administratio n, less than 2% of the dose was found in the 976 lower part of the gastrointestinal tract, indicatin g that the majority of dichloromethane absorption 977 occurs in the upper gastrointestinal tract (Angelo et al., 1986b). Similar results were reported in 978

mature male B6C3F 1 mice exposed to up to 50 mg/kg (Angelo et al., 198 6a). In mature male 979 Sprague-Dawley rats administered a single dose (1 o r 50 mg/kg) of radiolabeled 980 dichloromethane, less than 1% of the label was foun d in feces collected for 48 hours after dose 981 administration (McKenna and Zempel, 1981). Absorpt ion of dichloromethane generally follows 982 first-order kinetics (Angelo et al., 1986a), and no evidence for a dichloromethane-specific carrier 983 has been presented. The vehicle appears to affect the rate, but not the extent, of gastrointestinal 984 absorption, with an aqueous vehicle resulting in a more rapid absorption of dichloromethane than 985 an oil-based vehicle (Angelo et al., 1986a). 986

987 3.1.2. Inhalation—Respiratory Tract Absorption 988

Several studies in humans have demonstrated the ab sorption of dichloromethane 989 following inhalation exposure. In a study by Astra nd et al. (1975), 14 male volunteers (ages 19– 990 29) were exposed to about 870 mg/m 3 (250 ppm) or 1,740 mg/m 3 (500 ppm) for 30 minutes 991 while resting or exercising on a bicycle ergometer. There was a pause of about 20 minutes 992 without exposure between rest and exercise periods. Uptake of dichloromethane was estimated 993 at about 55% while resting and about 40, 30, and 35 % at respective workloads of 50, 100, and 994

150 watts. Blood levels of dichloromethane correla ted directly with exposure concentrations, 995 and did not appear to increase when a workload was applied (Astrand et al., 1975). Similar 996 reports of rapid uptake and a direct correlation be tween dichloromethane exposure level and 997 blood levels in humans have been presented by other groups (DiVincenzo and Kaplan, 1981; 998 DiVincenzo et al., 1971). 999 With extended (1–2 hours or greater) exposure, upt ake tends to reach a steady-state level, 1000 at which point blood dichloromethane levels remain more or less constant (DiVincenzo and 1001 Kaplan, 1981; DiVincenzo et al., 1972; Riley et al. , 1966). DiVincenzo et al. (1972) reported 10026 DRAFT - DO NOT CITE OR QUOTE

that in humans exposed to 100 or 200 ppm of dichlor omethane for 2 hours (without physical 1003 exercise), dichloromethane was rapidly absorbed, re aching an approximate steady state, as 1004 assessed by levels of unchanged dichloromethane in the expired air, within the first 15– 1005 30 minutes of exposure. A later study by the same group (DiVincenzo and Kaplan, 1981) 1006 similarly reported a rapid absorption of dichlorome thane in volunteers exposed to 50–200 ppm 1007 for 7.5 hours on each of 5 consecutive days. A ste ady-state level, as assessed by levels of 1008 unchanged dichloromethane in the expired air, was r eached quickly (1–2 hours), with exhaled 1009 dichloromethane levels increasing with increasing e xposure level. A similar pattern was seen 1010 with blood dichloromethane levels. Estimated pulmo nary uptake was 69–75% and did not vary 1011 appreciably with exposure concentration. In anothe r experiment in which one of the 1012 investigators was seated during exposure to 100 ppm dichloromethane for 2 hours, 1013 concentrations of dichloromethane in expired air re ached an apparent plateau of about 70 ppm 1014 within the first hour of exposure (Riley et al., 19 66). 1015 Body fat may influence absorption of dichlorometha ne, as evidenced by data from an 1016 experiment involving 12 men ages 21–35, divided int o two groups (n = 6 per group) based on 1017 percent body fat (Engström and Bjurström, 1977). T he mean percent body fat in the leaner 1018 group was 7.8% (standard error of the mean [SEM] 1. 9), range 2.3–13.6%, compared with 1019 25.1% (SEM 2.8), range 18.3–36.2%, in the more over weight group. Total uptake of 1020 dichloromethane during a light exercise period (50 watts 2 ) for 1 hour with an exposure level of 1021 750 ppm was positively correlated with percent body fat (r = 0.81), and the estimated amount of 1022 dichloromethane in fat storage was also correlated with percent body fat (r = 0.84). 1023 A pattern of absorption similar to that seen in hu mans has been seen in animals. Initially, 1024 dichloromethane is readily absorbed following inhal ation exposure, as evidenced by rapid 1025 appearance of dichloromethane in blood, tissues, an d expired air (Withey and Karpinski, 1985; 1026 Stott and McKenna, 1984; Anders and Sunram, 1982; C arlsson and Hultengren, 1975; Roth et 1027 al., 1975). For example, absorption of inhaled 500 ppm dichloromethane in anesthetized, mature 1028 male F344 rats reached an apparent plateau within 1 0–20 minutes and was relatively constant for 1029 up to 2 hours (Stott and McKenna, 1984). In these experiments, absorption was calculated from 1030 measurements of exposure (nose only) and effluent c oncentrations and ventilation flow rate in 1031 intact animals; double tracheostomized rats were us ed to measure absorption in the isolated 1032 upper respiratory tract and the lower respiratory t ract. At a ventilation rate of 53 mL/minute, 1033 absorption expressed as mean percentage of dichloro methane available for absorption was 44% 1034 (standard deviation [SD] 10) in intact rats, 13.2% (SD 3.6) in the upper respiratory tract, and 1035 37% (SD 4.1) in the lower respiratory tract. 1036

1037

1038

2 A watt is the International System Unit of power and is equal to one joule of energy per second. It is a measure of the rate of energy use or production (i.e., the exe rcise effort that was exerted by the individuals in the study).7 DRAFT - DO NOT CITE OR QUOTE

3.2. DISTRIBUTION 1039 Results from studies of animals show that, followin g absorption, dichloromethane is 1040 rapidly distributed throughout the body and has bee n detected in all tissues that have been 1041 evaluated. Twenty minutes after a single intraveno us dose of 10 mg [ 14 C]-dichloromethane/kg to 1042

mature male B6C3F 1 mice (Angelo et al., 1986a), total label was great est in the liver 1043 (6.72 μ g-equivalents/g tissue), with lower levels reported in the lung (1.82 μ g-equivalents/g 1044 tissue), kidney (1.84 μ g-equivalents/g tissue), and the remainder of the c arcass 1045 (1.90 μ g-equivalents/g tissue). By 4 hours post administr ation, levels in the liver had fallen to 1046

3.08 μ g-equivalents/g tissue, lung levels were 0.64 μ g-equivalents/g tissue, and carcass levels 1047

were 0.23 μ g-equivalents/g tissue. The levels in the kidney r ose sharply in the first hour 1048 postexposure but then fell and remained steady at ~1 .60 μ g-equivalents/g tissue for the 1049 remaining 3 hours of the study (Angelo et al., 1986 a). McKenna et al. (1982) exposed groups of 1050 mature male Sprague-Dawley rats to 50, 500, or 1,50 0 ppm [ 14 C]-labeled dichloromethane for 1051 6 hours and examined tissues at 48 hours for presen ce of radiolabel; results are shown in 1052 Table 3-1. The greatest concentration of label was found in the liver, followed by the kidney and 1053 lung. 1054

1055 Table 3-1. Distribution of radioactivity in tissue s 48 hours after inhalation

exposure of mature male Sprague-Dawley rats (n = 3) for 6 hours

Mean ± SD, μg-equivalent dichloromethane/g tissue, by exposure level Tissue 50 ppm 500 ppm 1,500 ppm

Liver 8.4 ± 1.5 35.6 ± 7.5 44.2 ± 3.5 Kidney 3.3 ± 0.1 16.2 ± 2.4 30.5 ± 0.2 Lung 1.9 ± 0.2 11.0 ± 1.3 16.5 ± 1.6

Brain 0.8 ± 0.3 4.2 ± 1.3 6.7 ± 0.2 Epidydimal fat 0.5 ± 0.2 6.5 ± 0.5 4.1 ± 0.9

Skeletal muscle 1.1 ± 0.1 4.4 ± 1.9 7.7 ± 0.7

Testes 1.1 ± 0.2 5.5 ± 1.3 8.1 ± 0.5

Whole blood 1.1 ± 0.2 8.1 ± 1.9 8.9 ± 1.7 Remaining carcass 1.3 ± 0.2 5.9 ± 0.9 8.6 ± 1.4 1056 Source: McKenna et al. (1982). 1057 1058 1059 As noted in the preceding section, adipose tissue may affect the uptake of 1060 dichloromethane, and there is also evidence of a re lation between adiposity and dichloromethane 1061 storage. In the study by Engström and Bjurström (1 977) involving 12 men ages 21–35 exposed 1062 to 750 ppm dichloromethane during a 1 hour light ex ercise (50 watts) period, dichloromethane 1063 was measured in body fat biopsy specimens at 1, 2, 3, and 4 hours postexposure. All specimens 1064

were taken from the buttocks. The concentration of dichloromethane (per gram tissue) was 10658 DRAFT - DO NOT CITE OR QUOTE

negatively correlated with percent body fat, but th e total estimated amount of dichloromethane in 1066 fat tissue 4 hours postexposure was higher in subje cts with a higher amount of fat (r = 0.84). 1067 Carlsson and Hultengren (1975) exposed groups of 1 0 mature male Sprague-Dawley rats 1068 to [ 14 C]-dichloromethane for 1 hour at a mean concentrati on of 1,935 mg/m 3 (557 ppm) and SD 1069 of 90 mg/m 3 (26 ppm). The initial levels were highest in the white adipose tissue (approximately 1070

80 μ g dichloromethane per gram tissue) compared with ap proximately 35, 20, and 5 μg1071 equivalent dichloromethane/g tissue in the liver, k idney and adrenal glands, and brain, 1072 respectively. These initial levels in the adipose quickly fell to less than 10 μg-equivalent 1073 dichloromethane/g tissue; more moderate declines we re seen in the other tissues. 1074 With acute 6-hour exposure scenarios, peak exposur e concentrations may have a greater 1075

influence on dichloromethane levels in the brain an d perirenal fat than time-weighted average 1076 (TWA) concentrations during the exposure period (Sa volainen et al., 1981). In rats exposed over 1077 a 6-hour period for 5 days/week to a TWA of 1,000 p pm dichloromethane consisting of two 1078 1-hour peak concentrations (2,800 ppm) interspersed with exposure to 100 ppm, levels of 1079 dichloromethane in the brain and perirenal fat were significantly higher than corresponding 1080 levels in rats exposed to constant levels of 1,000 ppm. This difference was not seen with blood 1081 carbon monoxide (CO) levels (Table 3-2). With cons tant exposure concentrations of 500 or 1082 1,000 ppm, perirenal fat levels of dichloromethane approximately doubled following 2 weeks of 1083 exposure compared with 1 week of exposure, indicati ng that some storage of dichloromethane in 1084 fat tissue can occur with repeated exposure scenari os (Table 3-2). In contrast, brain levels of 1085 dichloromethane in rats exposed for 1 week were hig her than brain levels in rats exposed for 2 1086 weeks. One possible explanation of these observati ons is that there is an induction of enzymes 1087

involved in dichloromethane metabolism in liver and other tissues with repeated exposure and 1088 dichloromethane in fat is poorly metabolized. 1089

1090 Table 3-2. Brain and perirenal fat dichloromethane and blood CO

concentrations in male Wistar rats exposed by inhal ation to dichloromethane at constant exposure concentrations compared with intermittently high exposure concentrations

Exposure weeks 1 2 1 2 1 2 Exposure level a (TWA, ppm) Brain (nmol/g) Perirenal fat (nmol/g) Blood CO (nmo l/g) Control 0 0 0 0 40 ± 15 30 ± 10 500, constant 30 ± 7 9 ± 3 436 ± 47 918 ± 215 675 ± 195 781 ± 62 1,000, constant 33 ± 2 14 ± 3 1,316 ± 209 2, 171 ± 219 876 ± 80 825 ± 56 1,000, with two 1-hour peaks of 2,800 ppm 111 ± 18 50 ± 15 2,295 ± 147 2,431 ± 146 728 ± 84 873 ± 90

a Groups of 5 rats were exposed to 0, 50, or 1,000 pp m 6 hours/day or 100 ppm interspersed with two 1-ho ur peaks of 2,800 ppm for 5 days/week for 1 or 2 weeks . Tissue concentration values are mean ± SD.

Source: Savolainen et al. (1981). 10919 DRAFT - DO NOT CITE OR QUOTE

Placental transfer 1092 Dichloromethane is capable of crossing the placenta l barrier and entering the fetal 1093 circulation. Anders and Sunram (1982) reported tha t when pregnant Sprague-Dawley rats (n = 1094 3) were exposed to 500 ppm dichloromethane for 1 ho ur on gestational day (GD) 21, mean 1095 maternal blood levels were 176 nmol/mL (SEM 50), wh ile fetal levels were 115 nmol/mL (SEM 1096 40); interestingly, the levels of CO, a metabolite of dichloromethane, were similar in both the 1097 maternal blood (167 nmol/mL, SEM 12) and fetal bloo d (160 nmol/mL, SEM 31). Withey and 1098 Karpinski (1985) also reported higher maternal comp ared with fetal dichloromethane levels 1099 based on a study of five pregnant Sprague-Dawley ra ts exposed to 107–2,961 ppm of 1100

dichloromethane. Maternal blood levels of dichloro methane were 2–2.5-fold higher than those 1101

found in the fetal circulation. 1102

1103 Blood-brain barrier transfer 1104 Dichloromethane is thought to readily transfer acro ss the blood-brain barrier, as 1105 evidenced by the detection of radioactivity in brai n tissue 48 hours after exposures of rats to 1106

radiolabeled dichloromethane at concentrations of 5 0, 500, or 1,500 ppm for 6 hours (McKenna 1107 et al., 1982) (see Table 3-1), and the historical d emonstrations that dichloromethane has transient 1108 sedative and anesthetic properties in humans (for r eview of these reports, see Mattsson et al. 1109 [1990] and Winneke [1974]). Dichloromethane is no longer used as an anesthetic gas because 1110 the margin between anesthetic and lethal doses is n arrow (Winneke, 1974). 1111

1112 3.3. METABOLISM 1113 Metabolism of dichloromethane involves two primary pathways, outlined in Figure 3-1 1114 (Agency for Toxic Substances and Disease Registry [ ATSDR], 2000; Guengerich, 1997; Hashmi 1115 et al., 1994; Gargas et al., 1986). Dichloromethan e is metabolized to CO in a cytochrome 1116 P450 (CYP)-dependent oxidative pathway that is pred ominant at low exposure levels. The CYP1117 related pathway results in the addition of oxygen, followed by spontaneous rearrangement to 1118 formyl chloride, and then to CO; each spontaneous r earrangement releases H + and Cl

– ions. At 1119 higher exposure levels, the CYP pathway becomes sat urated and a second pathway begins to 1120 predominate. Glutathione S-transferase (GST)-catal yzed addition of glutathione (GSH) is the 1121 initial step in this pathway. The replacement of o ne of the chlorine atoms with the S-glutathione 1122 group results in formation of S-(chloromethyl)gluta thione and the release of H + and Cl

– ions. 1123 Hydration of S-(chloromethyl)glutathione results in an S-glutathionyl methanol molecule, which 1124 can spontaneously form formaldehyde or rearrange to form an S-glutathione formaldehyde 1125 molecule, and then further rearrange to formate. B oth formaldehyde and formate can then be 1126

further metabolized to CO 2

. 1127

112810 DRAFT - DO NOT CITE OR QUOTE GSTT1 

## C H C l

C l H Dichloromethane

CYP2E1

Formyl Chloride CO

Carbon Monoxide

COHb

Carboxyhemoglobin C H C l

GS H

S-(chloromethyl) glutathione

C H OH

GS H C H O

H

S-glutathionyl methanol Formaldehyde C H O

G - S | H

CO 2 CO 2 C H O

G - S

C H

O

OH Formic acid CO 2 (minor pathway)

1129 1130 Figure 3-1. Proposed pathways for dichloromethane metabolism. 1131 1132 Adapted from: ATSDR (2000); Guengerich (1997); Has hmi et al. (1994); Gargas et al. (1986). 1133 1134 1135 As described in the following discussion of the two pathways, a metabolic balance 1136 appears to exist between them, with the CYP pathway tending to be relatively more active at 1137 lower doses and the GST pathway metabolizing the ma jority of a dichloromethane dose at higher 1138 exposure levels once the CYP pathway has become sat urated. Exposure to other agents may 1139 shift this balance. For example, pretreatment with compounds that deplete GSH (e.g., buthionine 1140 sulfoximine, diethylmaleate, phorone) resulted in a n increase in blood carboxyhemoglobin 1141 (COHb) levels, following a single injection of dich loromethane, relative to animals that did not 1142 receive GSH depletion, indicating a shift to the CY P pathway (Oh et al., 2002). Similarly, co1143 exposure to agents that compete for CYP2E1 results in a shift toward the GST pathway and away 1144 from CO production (Lehnebach et al., 1995; Pankow and Jagielki, 1993; Pankow et al., 1145 1991a, b; Glatzel et al., 1987; Roth et al., 1975). 1146

114711 DRAFT - DO NOT CITE OR QUOTE

3.3.1. The CYP2E1 Pathway 1148 There is considerable evidence of the importance of the CYP2E1 metabolic pathway in 1149 studies in animals (Oh et al., 2002; Wirkner et al. , 1997; Kim and Kim, 1996; Lehnebach et al., 1150 1995; Pankow et al., 1991a, b; Pankow and Hoffmann, 1989; Pankow, 1988; Glatzel et al., 1987; 1151 Angelo et al., 1986a, b; Landry et al., 1983; Ander s and Sunram, 1982; McKenna et al., 1982; 1152 McKenna and Zempel, 1981; Rodkey and Collison, 1977 ; Carlsson and Hultengren, 1975; Roth 1153 et al., 1975; Fodor et al., 1973) and humans (Takes hita et al., 2000; DiVincenzo and Kaplan, 1154 1981; Astrand et al., 1975). These studies demonst rate that exposure to dichloromethane, 1155 regardless of exposure route, results in the format ion of CO, as assessed by direct measurements 1156 of elevated levels of CO in expired air and increas ed levels of COHb in the blood. 1157 The first step in the CYP2E1 pathway is the format ion of formyl chloride (Figure 1). 1158 Watanabe and Guengerich (2006) conducted a series o f studies to investigate the downstream 1159 metabolites of formyl chloride, and reported only m arginal (3% maximum at pH 9) formation of 1160

S -formyl GSH from formyl chloride in the presence of GSH. Therefore, most (>97%) of the 1161 formyl chloride is metabolized further to carbon mo noxide. Furthermore, CO formation from 1162 formyl chloride was independent of GSH presence in the assay. 1163

Results from numerous studies in rats in which CYP2 E1 metabolism was blocked or 1164 induced indicate that the generation of CO occurs a s a result of metabolism of dichloromethane 1165 by the CYP2E1 pathway (Figure 3-1). Co-exposure of rats to a high dose of ethanol 1166 (174 mmol/kg), which is metabolized by CYP2E1, and dichloromethane (1.6, 6.2, 15.6 mmol/kg) 1167 resulted in no increase in blood COHb, indicating t hat the metabolic pathway for CO formation 1168 had been either blocked or saturated (Glatzel et al ., 1987). Similar results have been seen with 1169 coadministration of other known CYP substrates, inc luding diethyldithiocarbamate (Lehnebach 1170 et al., 1995), methanol (Pankow and Jagielki, 1993) , benzene, toluene, and three xylene isomers 1171 (Pankow et al., 1991b). Pretreatment of animals wi th CYP inducers (e.g., benzene, toluene, 1172 xylenes, methanol, isoniazid), particularly those t hat induce CYP2E1, resulted in an increased 1173 level of CO formation, as assessed by COHb formatio n or measurement in expired air, following 1174 single exposures to dichloromethane (Kim and Kim, 1 996; Pankow and Jagielki, 1993; Pankow 1175 et al., 1991b; Pankow and Hoffmann, 1989; Pankow, 1 988). Pretreatment with disulfuram, a 1176 CYP2E1 blocker, resulted in a complete lack of form ation of COHb following dichloromethane 1177 exposure, indicating that CYP2E1 is the isozyme res ponsible for metabolism of dichloromethane 1178 (Kim and Kim, 1996). 1179 Evidence in hamster and rat studies suggests that the CYP2E1 pathway becomes 1180 saturated at high dichloromethane exposure levels; comparable data from studies in mice were 1181 not found. In hamsters, mean COHb percentages were elevated to a similar degree (about 28– 1182 30%, compared with <1% in controls) in three groups exposed by inhalation to 500, 1,500, or 1183 3,500 ppm dichloromethane for 6 hours (Burek et al. , 1984). After 21 months of exposure by 1184 this protocol, mean COHb percentages in the three e xposure groups remained similarly elevated, 118512 DRAFT - DO NOT CITE OR QUOTE

indicative of saturation of the CYP2E1 pathway in h amsters at exposure levels ≥ 500 ppm and a 1186

lack of accumulation of dichloromethane and CYP2E1 metabolites with chronic exposure. 1187 McKenna et al. (1982) found that blood COHb levels in rats increased when inhalation exposure 1188 concentration was increased from 50 to 500 ppm but that similar levels of COHb were reported 1189 following exposure to 1,500 ppm as following exposu re to 500 ppm; the peak blood COHb 1190 percentages were approximately 10%. In rats expose d to 0, 50, 200, or 500 ppm for 6 hours/day, 1191 5 days/week for 2 years, mean COHb percentages were 2.2, 6.5, 12.5, and 13.7%, respectively, 1192 suggesting that saturation of the CYP2E1 pathway is approached at 200 ppm (Nitschke et al., 1193 1988a). In male F344 rats exposed for 4 hours to d ichloromethane concentrations of about 150, 1194 300, 600, 1,000, and 2,000 ppm, mean COHb percentag es (estimated from a figure) were about 1195 4% at 150 ppm and about 8% at each of the four high er exposure concentrations (Gargas et al., 1196 1986). McKenna and Zempel (1981) reported that inc reasing the oral dose of labeled 1197 dichloromethane from 1 mg/kg to 50 mg/kg in rats re sulted in a lower fraction of the total dose 1198 being metabolized to CO. Single injections of 3 an d 6 mmol/kg of dichloromethane in rats 1199 resulted in nearly identical levels of blood COHb ( Oh et al., 2002). 1200 In human subjects exposed to dichloromethane in th e workplace, saturation of CYP 1201 metabolism appears to be approached in the 400–500 ppm range (Ott et al., 1983e). Blood 1202 samples were drawn during working hours from 136 fi ber production workers who were exposed 1203 to dichloromethane, acetone, and methanol. TWA exp osure concentrations for the workers were 1204 determined by personal monitoring techniques, and p ercent COHb levels in the blood samples 1205

were determined. Estimated TWA concentrations in t he exposed workers showed a bimodal 1206 distribution. The lower mode of exposure concentra tions showed the highest frequency in the 1207 150–200 ppm range, while the higher mode showed the highest frequency in the range of 450– 1208 500 ppm. Plots of percent COHb against TWA exposur e concentrations showed that saturation 1209 begins to be apparent in the 400–500 ppm range of e xposure concentrations. 1210

The liver is the tissue most enriched in CYP2E1 ca talytic activity, but CYP2E1 protein 1211 and messenger ribonucleic acid (mRNA) have been det ected in other human tissues, including 1212 the lung, brain, kidney, pancreas, bladder, small i ntestine, and blood lymphocytes (Nishimura et 1213 al., 2003). As such, the liver is expected to be t he main site of CYP metabolism of 1214 dichloromethane, but other tissues are also expecte d to metabolize dichloromethane via this 1215 pathway. Of particular relevance given the neurolo gic effects seen with dichloromethane are the 1216 distribution and inducibility of CYP2E1 in differen t areas of the brain (Miksys and Tyndale, 1217 2004). Individuals with decreased CYP2E1 activity may experience decreased generation of CO 1218

and an increased level of GST-related metabolites f ollowing exposure to dichloromethane. As a 1219 result, these individuals may be more susceptible t o the chronic effects of dichloromethane from 1220 GST-related metabolites than individuals with highe r levels of CYP2E1 activity. Conversely, 1221 individuals with higher CYP2E1 activity may experie nce relatively increased generation of CO 122213 DRAFT - DO NOT CITE OR QUOTE

at a given dichloromethane exposure level and, ther efore, may be more susceptible to the acute 1223 toxicity of dichloromethane (from CO). 1224 Results from studies examining human interindividu al variation in CYP2E1 activities 1225 (e.g., catalytic activities, protein levels, or mRN A levels) indicate that individuals may vary in 1226 their ability to metabolize dichloromethane through the CYP2E1 pathway. In a study of liver 1227 samples from 30 Japanese and 30 Caucasian individua ls, two- to threefold variation was found in 1228 the levels of CYP2E1 protein, whereas catalytic act ivity toward substrates associated with 1229 CYP2E1 (e.g., 7-ethoxycoumarin) displayed a wider r ange of values, approximately 25-fold; no 1230 clear gender-specific or ethnic differences were fo und in hepatic levels of CYP2E1 protein or 1231 enzymatic activities associated with CYP2E1 (Shimad a et al., 1994). In a study of 1232 interindividual variation in 70 healthy human subje cts (40 men and 30 women) given an oral 1233 dose of chlorzoxazone, a therapeutic agent whose me tabolism and blood clearance has been 1234 related to CYP2E1 levels, a three- to fourfold rang e in plasma half-life and clearance values was 1235 observed, with no clear or dramatic age- or gender- specific differences (Kim et al., 1995). A six1236 to sevenfold range in chlorzoxazone hydroxylation a ctivity was reported for a group of 1237 69 healthy, smoking and nonsmoking male and female volunteers with mixed ethnic 1238 backgrounds; the range was markedly increased when a group of 72 alcoholic inpatients was 1239 included (Lucas et al., 1999). In studies of human liver microsomes, four- to sixfold ranges in 1240 CYP2E1-dependent oxidation of trichloroethylene hav e been reported (Lipscomb et al., 2003, 1241 1997). CYP2E1 protein levels in 50 specimens of hu man lymphocytes from healthy individuals 1242 showed an approximate fivefold range (Bernauer et a l., 2000), and a 3.7-fold range in liver 1243 CYP2E1 mRNA levels was reported for a group of 24 p atients with chronic hepatitis (Haufroid 1244 et al., 2003). More recently, a threefold range wa s reported for maximal rates of hepatic 1245 CYP2E1-catalyzed metabolism of dichloromethane, whi ch were estimated with a modified 1246 physiologically based toxicokinetic (PBTK) model or iginally developed by Andersen et al. 1247 (1987) and kinetic data (e.g., dichloromethane brea th and blood concentrations) for 13 volunteers 1248 (10 males and 3 females) exposed to one or more con centrations of dichloromethane by 1249 inhalation for 7.5 hours (Sweeney et al., 2004). I n summary, most studies indicate a three- to 1250 sevenfold variability in CYP2E1 activity, as assess ed by various types of measurements, among 1251 “healthy” volunteers. However, various clinical fa ctors (i.e., obesity, alcoholism, use of specific 1252 medications) or co-exposures (i.e., to various solv ents) (Lucas et al., 1999) may result in greater 1253 variation, and thus the potential for saturation at lower exposures, within the general population. 1254 Several genetic polymorphisms for the human CYP2E1 gene have been described, but 1255

clear and consistent correlations with interindivid ual variation in CYP2E1 protein levels or 1256 associated enzyme activities have not been identifi ed (Ingelman-Sundberg, 2004; Lucas et al., 1257 2001; Kim et al., 1995; Shimada et al., 1994). The most frequently studied CYP2E1 1258 polymorphisms, RsaI/PstI , are located in the 5 ́-flanking region of the gene , and mutations are 1259 thought to lead to increased CYP2E1 protein express ion via transcription (Lucas et al., 2001). 126014 DRAFT - DO NOT CITE OR QUOTE

Available data indicate that the frequency of this polymorphism, as well as other CYP2E1 1261 polymorphisms, varies among ethnic groups. For exa mple, Stephens et al. (1994) examined 1262 blood samples from 126 African-Americans, 449 Europ ean Americans, and 120 Taiwanese 1263 subjects and found frequencies for a rare RsaI allele (C2) of 0.01 in African-Americans, 0.04 in 1264 European Americans, and 0.28 in Taiwanese subjects. In a study of 102 Mexicans, the reported 1265 mutation frequency at the RsaI C2 allele was 0.30 (Mendoza-Cantú et al., 2004). 1266

1267 3.3.2. The GST Pathway 1268 The other major pathway for dichloromethane metabol ism involves the conjugation of 1269 dichloromethane to GSH, catalyzed by GST. This res ults in the formation of a GSH conjugate 1270 that is eventually metabolized to CO 2 (Figure 3-1). The conjugation of dichloromethane to GSH 1271

results in formation of two reactive intermediates that have been proposed to be involved in 1272 dichloromethane toxicity, S-(chloromethyl)glutathio ne and formaldehyde. In studies with rat, 1273 mouse, and human liver cytosol preparations in the presence of GSH, examination of metabolites 1274

with 13 C-NMR indicated that S-(chloromethyl)glutathione wa s an intermediate in the pathway to 1275 formaldehyde (Hashmi et al., 1994). Formaldehyde f ormation from dichloromethane has been 1276 noted in human (Bruhn et al., 1998; Hallier et al., 1994; Hashmi et al., 1994), rat, and mouse 1277 (Casanova et al., 1997; Hashmi et al., 1994) cells in vitro. Formation of free hydrogen ion is also 1278 hypothesized, although no direct evidence supportin g this has been presented. 1279 The GST pathway has approximately a 10-fold lower a ffinity for dichloromethane than 1280 the CYP pathway (Reitz et al., 1989; Andersen et al ., 1987). At lower exposure concentrations, 1281 the CYP pathway is expected to predominate, but, as exposure concentrations increase, the GST 1282 pathway is expected to gain in relative importance as a dispositional pathway for absorbed 1283

dichloromethane. Based on in vitro studies with li ver preparations, the estimated Michaelis1284 Menten kinetic constants (K m s) in GST assays with dichloromethane were about 13 7 mM in a 1285

B6C3F 1 mouse preparation and about 44 mM in two human pre parations (Reitz et al., 1989). In 1286 contrast, estimated K m s in CYP assays were about 1.8, 1.4, and 2.0 mM in B6C3F 1 mouse, F344 1287 rat, and Syrian golden hamster preparations, respec tively. In four human liver preparations, 1288

estimated CYP K m s were about 2.6, 2.0, 0.9, and 2.8 mM (Reitz et al ., 1989). 1289 Early investigations indicated that in humans GSTs of the α -, μ -, and π -classes were not 1290 responsible for the metabolism of dichloromethane ( Bogaards et al., 1993). Tissue samples that 1291 metabolized substrates specific to those GST classe s did not conjugate dichloromethane to GSH. 1292 Later investigations identified the recently-charac terized GST theta class (Meyer et al., 1991), 1293 specifically GST-theta1-1 (GST-T1), as the GST isoe nzyme responsible for the metabolism of 1294 dichloromethane (Mainwaring et al., 1996; Blocki et al., 1994). In the absence of the GST-T1 1295 gene, no deoxyribonucleic acid (DNA)-protein cross- links were formed by human liver cells 1296 exposed to dichloromethane (Casanova et al., 1997), and formaldehyde production was not 1297 detected in human erythrocytes (Hallier et al., 199 4). In a mouse model with a disrupted GST- 129815 DRAFT - DO NOT CITE OR QUOTE

T1 gene, GST activity with dichloromethane in liver and kidney cytosol samples was 1299 substantially lower compared with wild-type GST mic e (Fujimoto et al., 2007). 1300 A polymorphism of the GST-T1 gene has been demonstr ated in humans. People with 1301 two functional copies of the gene (+/+) readily con jugate GSH to dichloromethane. Individuals 1302 having only one working copy of the gene (+/–) disp lay relatively decreased conjugation ability. 1303 Individuals with no functional copy of the gene (–/ –) do not express active GST-T1 protein and 1304

do not metabolize dichloromethane via a GST-related pathway (Thier et al., 1998). Results from 1305 studies of GST-T1 genotypes in human blood samples indicate that average prevalences of the 1306 GST-T1 null (–/–) genotype are higher in Asian ethn ic groups (47–64%) than in other groups, 1307 including Caucasians (19–20%), African-Americans (2 2%), and mixed groups (19%) (Raimondi 1308 et al., 2006; Garte et al., 2001; Nelson et al., 19 95) (see Table 3-3). Although information on the 1309 age distribution of study subjects was not generall y reported in these analyses, there is little 1310 reason to expect effect modification by age since t his is not a gene linked to early mortality. 1311 Based on data collected by Nelson et al. (1995) and U.S. 2000 census data (and assuming Hardy1312 Weinberg equilibrium), Haber et al. (2002) calculat ed U.S. average distributions of GST-T1 1313 genotypes as follows: 32% +/+; 48% +/–; and 20% –/–

. 1314

1315 Table 3-3. Mean prevalences of the GST-T1 null (–/ –) genotype in human ethnic groups

Reference Ethnic group Nelson et al. (1995) a Garte et al. (2001) b Raimondi et al. (2006) c Chinese 64.4% (n = 45) Not reported Not reported Korean 60.2% (n = 103) Not reported Not reported Caucasian 20.4% (n = 442) 19.7% (n = 5,577) 19.0% ( n = 6,875) Asian Not reported 47.0% (n = 575) 53.6% (n = 1,727 ) African-American 21.8% (n = 119) Not reported Not r eported Mexican American 9.7% (n = 73) Not reported Not rep orted Other Not reported Not reported 19.4 % (n = 1,485)

a Nelson et al. (1995) examined prevalence of the nul l GST-T1 genotype from analysis of blood samples fr om subjects of various ethnicities as noted above.

b Garte et al. (2001) collected GST-T1 genotype data in Caucasian (29 studies; 5,577 subjects) and Asian (3 studies, 575 subjects) ethnic groups; subjects were controls in case-control studies of cancer and various poly morphisms in genes for bioactivating enzymes.

c Raimondi et al. (2001) collected GST-T1 genotype da ta from 35 case-control studies of cancer and GST-T 1 genotype; data in this table are for control subjec ts. The “other” group in this study is defined as Latino, AfricanAmerican, and mixed ethnicities. 1316 1317 Results from a study of the distribution of activi ty levels for in vitro conjugation of 1318 dichloromethane with GSH in 22 human liver samples are roughly reflective of these estimates 1319 of the distribution of this polymorphism (Bogaards et al., 1993). No activity was found in 3/22 1320 of the liver samples. Eleven of the samples showed low activity levels (0.21–0.41 nmol 132116 DRAFT - DO NOT CITE OR QUOTE

product/minute/mg protein), and eight samples showe d high activity levels ranging from 0.82 to 1322 1.23 nmol/minute/mg protein. In another study of s even human subjects, lysates of erythrocytes 1323 showed high activities for producing formaldehyde from dichloromethane (presumably via GST1324 T1) in three subjects (15.4, 17.7, and 17.8 nmol pr oduct/minute/mg hemoglobin) and lower 1325 activity in the other four subjects (4.3, 6.0, 7.2, and 7.6 nmol product/minute/mg hemoglobin) 1326 (Hallier et al., 1994). 1327 Comparisons of mice, rats, humans, and hamsters for the ability to metabolize 1328 dichloromethane via the GST pathway in liver and lu ng tissues indicate that mice appear to be 1329 the most active at metabolizing dichloromethane (Sh erratt et al., 2002; Thier et al., 1998; 1330 Casanova et al., 1997, 1996; Hashmi et al., 1994; R eitz et al., 1989). Reitz et al. (1989) reported 1331 mean (± SD) GST enzymatic activity levels with dich loromethane as substrate (in units of nmol 1332 product formed/minute/mg protein) in liver cytosol preparations to be: 25.9 ± 4.2 units in 1333

B6C3F 1 mice (n = 15 determinations per preparation); 7.05 ± 1.7 units in F344 rats (n = 6); and 1334 1.27 ± 0.21 units (n = 6) in Syrian golden hamsters . Mean GST activity levels in liver 1335 preparations from four human subjects (accident vic tims screened for human immunodeficiency 1336 virus and hepatitis B and C and obtained through a transplant center) were 2.62 ± 0.44 units (n = 1337 10), –0.01 ± 0.04 units (n = 6), 2.71 ± 0.45 units (n = 6), and 3.03 ± 0.44 units (n = 6) (Reitz et 1338 al., 1989). The finding that one of the four indiv iduals was unable to conjugate dichloromethane 1339 with GST was reflective of the estimated frequency of the GST-T1 null genotype in the U.S. 1340 population (see Table 3-3). Mean GST activity leve ls in lung cytosol preparations showed a 1341 similar rank order among species: 7.3 ± 1.4 units i n mice (n = 4), 1.0 ± 0.1 units in rats (n = 4), 1342 0.0 ± 0.2 units in hamsters (n = 4), and 0.37 ± 0.2 5 units in a pooled lung preparation from the 1343 same four human subjects (n = 2). Reitz et al. (19 89) noted that relative abilities of these animal 1344 species to metabolize dichloromethane via the GST p athway correlated with their cancer 1345 sensitivities in long-term inhalation bioassays: (1 ) B6C3F 1 mice showed statistically significant 1346 increased incidence of liver and lung tumors in a 2 -year cancer bioassay (National Toxicology 1347 Program [NTP], 1986); (2) rats showed much less evi dence of increased incidence of liver 1348 tumors, and no increased risk of lung tumors at equ ivalent exposure concentrations but showed 1349 increased incidence of nonmalignant mammary tumors (NTP, 1986; Burek et al., 1984); and 1350 (3) Syrian golden hamsters did not show tumorigenic responses at any site (Burek et al., 1984). 1351 Thier et al. (1998) conducted a study evaluating th e activity of GST-T1 after treatment of 1352 dichloromethane in the cytosol of liver and kidney homogenates from hamsters (pooled male and 1353 females), rats (pooled male and female), male mice, and female mice and for humans classified 1354 as nonconjugators, low conjugators, or high conjuga tors of GST to dichloromethane. Little 1355 information is provided about the human samples oth er than that 13 kidney cancer patients were 1356 the source of the kidney samples; normal tissue ide ntified by pathological exam was used. Blood 1357 samples from 10 of these patients were collected an d enzyme activities measured in erythrocytes 1358 from 9 of these samples were reported. Results of conjugation of dichloromethane to GSH from 135917 DRAFT - DO NOT CITE OR QUOTE

these studies are presented in Table 3-4. As can b e seen from the table, activity levels (expressed 1360 as nmol/minute per mg of cytosolic protein) of huma ns varied considerably, with nonconjugators 1361 (presumed to be GST-T1 –/– ) having no detectable activity, low conjugators (p resumed to be 1362

GST-T1 +/– ) having moderate activity, and high conjugators (p resumed to be GST-T1 +/+ ) having 1363 approximately twice the activity seen in low conjug ators. In the liver, the activity of rat GST 1364 conjugation was over twofold that seen in human hig h conjugators, while levels in mice were 1365 >11-fold (males) or 18-fold (females) greater than those of human high conjugators. In the 1366 kidney, the activity of high-conjugator humans was approximately 1.8-fold that of rats and was 1367 comparable to the activity of both male and female mice. The data in Table 3-4 show the 1368 following order for GST-T1 activities with dichloro methane as substrate: in liver preparations, 1369 mouse >> rat > human high conjugators > human low c onjugators > hamster > human 1370 nonconjugators and, in kidney preparations, female mouse ≈ male mouse ≈ human high 1371 conjugators > rat ≈ human low > hamster > human nonconjugators. In ad dition, the data indicate 1372 that activity levels in liver, kidney, and erythroc ytes of human subjects are in correspondence 1373 with the nonconjugator, low conjugator, and high co njugator designations. 1374

1375 Table 3-4. GST-T1 enzyme activities toward dichlor omethane in human,

rat, mouse, and hamster tissues (liver, kidney, and erythrocytes)

Activity (nmol/min per mg protein) a Activity (nmol/min per mL) a Liver Kidney Erythrocytes Human, nonconjugators Not detectable (2) Not detectable (1) Not detectable (1) Human, low conjugators 0.62 ± 0.30 (11) 1.38 ± 0 .52 (8) 9.67 ± 2.49 (5) Human, high conjugators 1.60 ± 0.48 (12) 3.05 ± 0 .72 (4) 18.28 ± 0.46 (3) Rat 3.71 ± 0.28 (8) 1.71 ± 0.28 (8) Not measured Mouse, male 18.2 ± 2.22 (5) 3.19 ± 0.46 (5) Not mea sured Mouse, female 29.7 ± 6.31 (5) 3.88 ± 0.90 (5) Not m easured Hamster 0.27 ± 0.20 (6) 0.25 ± 0.21 (6) Not measure d

a Mean ± SD with number of samples noted in parenthes es.

Source: Adapted from Thier et al. (1998). 1376 1377 Sherratt et al. (2002) reported that, on a per mg b asis, native recombinant mouse GST-T1 1378 (purified after expression in Escherichia coli ) was approximately twofold more active toward 1379 dichloromethane than native recombinant human enzym e, as well as being approximately 1380 fivefold more efficient (as assessed by the ratio o f k cat /K m ). 1381

The distribution of GST-T1 in human tissues has bee n examined with antibodies raised 1382 against recombinant human GST-T1 (Sherratt et al., 2002, 1997). Immunoblotting of sodium 1383 dodecyl sulfate polyacrylamide gel electrophoresis gels loaded with tissue extracts from a 1384 73 year-old man who had died with brochopneumonia a nd atherosclerosis indicated the 138518 DRAFT - DO NOT CITE OR QUOTE

following order of expression of GST-T1: liver ≈ kidney > prostate ≈ small intestine > 1386

cerebrum ≈ pancreas ≈ skeletal muscle > lung ≈ spleen ≈ heart ≈ testis (Sherratt et al., 1997). 1387 It was estimated that the levels of cross-reacting materials in the cerebrum, pancreas, or skeletal 1388

muscle extracts were about 10% of those in the live r, whereas levels in the lung, spleen, heart, 1389

and testis were less than 5% of the levels in the l iver. Comparison of the amounts of cross1390 reacting material in soluble liver extracts from a B6C3F 1 mouse and five human subjects (i.e., 1391 normal liver tissue samples from biopsies of second ary liver tumors) found that levels of GST1392 T1 protein were higher in the mouse extracts than i n any of the human liver extracts (Sherratt et 1393 al., 2002). Densitometer analysis indicated that t he GST-T1 level in the mouse liver extract was 1394 about fivefold higher than those in human liver ext racts displaying the highest level. Cross1395 reacting material was not detectable in liver extra cts from one of the five human subjects, 1396 indicating that this individual may have been GST-T 1 null (Sherratt et al., 2002). 1397 Results from in situ hybridization with oligonucleo tide anti-sense probes for GST-T1 1398

mRNA levels and immunohistochemical studies with an tibodies to GST-T1 have indicated that 1399 there may be subtle differences between mice and hu mans in the intracellular localization of 1400 GST-T1 in the liver. Mainwaring et al. (1996) repo rted that staining for GST-T1 mRNA was 1401 higher in liver slices from B6C3F 1 mice than in liver slices from F344 rats and that staining in 1402 human liver samples was very low. Although the num ber of mouse and rat liver samples 1403 examined in this study was not indicated in the ava ilable report, it was reported that slices from 1404 five human liver samples were examined. No informa tion was provided regarding the clinical 1405 history of the sources of the human samples. In mo use liver, staining for GST-T1 mRNA was 1406 enhanced in the limiting plate hepatocytes, in nucl ei, in bile-duct epithelial cells, and in lesser 1407 amounts in the centrilobular cells in general. In rat liver, a similar pattern was observed, except 1408 no enhanced staining was observed in the limiting p late hepatocytes or in nuclei. Staining for 1409 GST-T1 mRNA in the human liver samples showed an ev en distribution throughout the liver 1410 lobule, and no mention of a specific nuclear locali zation was made (Mainwaring et al., 1996). 1411 Quondamatteo et al. (1998), using antibodies to GST -T1, subsequently reported a similar 1412 localization of GST-T1 protein in nuclei of cells i n mouse liver slices. In another study using 1413 antibodies raised against recombinant human GST-T1 or a peptide derived from the deduced 1414 mouse GST-T1 primary sequence, Sherratt et al. (200 2) reported that nuclear staining was 1415 observed in all cells in mouse liver slices (from f ive individual B6C3F 1 mice) showing the 1416 presence of mouse GST-T1; staining in the cytoplasm was only detected in cells with very high 1417

levels of GST-T1. In liver slices obtained from tw o human subjects (males, ages 60 and 1418 61 years, with a secondary liver tumor and what was described as a “cavernous hemangioma” 1419 without malignancy, respectively), the most intense nuclear staining was associated with bile 1420 duct epithelial cells, but there was heterogeneity of staining within hepatocytes; some cells 1421 showed nuclear staining, but others only exhibited cytoplasmic staining (Sherratt et al., 2002). 142219 DRAFT - DO NOT CITE OR QUOTE

In summary, the relative amount of dichloromethane metabolized via the GST pathway 1423 increases with increasing exposure concentrations. As the high affinity CYP pathway becomes 1424 saturated (either from high exposure levels of gene tic or other factors that decrease CYP2E1 1425 activity), the GST pathway increases in relative im portance as a dispositional pathway for 1426 dichloromethane. Two reactive metabolites (S-(chlo romethyl)glutathione and formaldehyde) 1427 resulting from this pathway have been identified. GST-T1 is the GST isozyme that catalyzes 1428 conjugation of dichloromethane with GST. Interindi vidual variation in the ability to metabolize 1429 dichloromethane via GST-T1 is associated with genet ic polymorphisms in humans. Estimated 1430 U.S. population prevalence of nonconjugators (–/– a t the GST-T1 locus) is about 20%, but higher 1431 prevalences (47–64%) have been reported for Asians (Raimondi et al., 2006; Haber et al., 2002; 1432 Garte et al., 2001; Nelson et al., 1995). The prev alences for low (+/– at the GST-T1 locus) and 1433 high (+/+) conjugators have been estimated at 48 an d 32%, respectively (Haber et al., 2002). 1434 The liver and kidney are the most enriched tissues in GST-T1, but evidence is available for the 1435 presence of GST-T1 in other tissues at lower levels , including the brain and lung. In humans, 1436 GST-T1 expression in the brain is lower than that s een in the liver or kidney but higher than in 1437 the lung. Comparisons of mice, rats, humans, and h amsters for the ability to metabolize 1438 dichloromethane via the GST pathway in liver (based on measurement of tissue-specific enzyme 1439 activity) indicate the following rank order: mice > rats > or ≈ humans > hamsters. This relative 1440 ranking corresponds to the rank order of the streng th of the association between inhalation 1441 exposure to dichloromethane and liver tumors in lon g-term cancer bioassays with mice, rats, and 1442 hamsters. In mouse liver tissue, GST-T1 appears to be localized in the nuclei of hepatocytes and 1443 bile-duct epithelium, but rat liver does not show p referential nuclear localization of GST-T1. In 1444 human liver tissue, some hepatocytes show nuclear l ocalization of GST-T1 and others show 1445 localization in cytoplasm, as well as in bile duct epithelial cells. The apparent species 1446

differences in intracellular localization of GST-T1 may play a role in species differences in 1447 susceptibility to dichloromethane carcinogenicity i f nuclear production of 1448 S-(chloromethyl)glutathione is more likely to lead to DNA alkylation than cytoplasmic 1449 production. 1450

1451

3.4. ELIMINATION 1452 Dichloromethane is eliminated mainly through exhal ation either of the parent compound 1453 or as the two primary metabolites CO 2 and CO (Angelo et al., 1986a, b; McKenna et al., 19 82; 1454 DiVincenzo and Kaplan, 1981; DiVincenzo et al., 197 2, 1971). In human studies, 1455 dichloromethane is rapidly eliminated from the body following the cessation of exposure, with 1456 much of the parent compound completely removed from the bloodstream and expired air by 1457 5 hours postexposure in experiments using exposure levels of 90, 100, or 210 ppm (DiVincenzo 1458 et al., 1972, 1971; Riley et al., 1966). Studies i n rats have similarly demonstrated that 1459 elimination from the blood is rapid, with eliminati on half-times in F344 rats on the order of 4– 146020 DRAFT - DO NOT CITE OR QUOTE

6 minutes following intravenous doses in the range of 10–50 mg/kg (Angelo et al., 1986a). In a 1461 study using Sprague-Dawley rats, Carlsson and Hulte ngren (1975) demonstrated variability in 1462 elimination rates between different types of tissue s, with the most rapid elimination seen in the 1463 adipose and brain tissue, while elimination from li ver, kidneys, and adrenals proceeded more 1464 slowly. 1465 In a study using human volunteers, DiVincenzo and K aplan (1981) reported a dose1466 related increase in CO in the expired breath after inhalation exposure to 50–200 ppm of 1467 dichloromethane, with a net elimination as CO on th e order of 25–35% of the absorbed dose. 1468 Similar results have been reported in animal studie s. Following gavage administration of 50 or 1469 200 mg/kg-day doses of [ 14 C]-labeled dichloromethane in water to groups of si x mature male 1470 F344 rats for up to 14 days, >90% of the label was recovered in the expired air within 24 hours 1471 of dose administration (Angelo et al., 1986b). Fol lowing administration of the first of 14 daily 1472 50 mg/kg-day doses, radioactivity in parent compoun d, CO 2 , and CO in the 24-hour expired 1473 breath accounted for 66, 17, and 16% of the adminis tered radioactivity, respectively; similar 1474 patterns were reported for 24-hour periods followin g administration of the seventh and 1475 fourteenth 50 mg/kg-day dose. Following administra tion of the first 200 mg/kg-day dose, 1476 radioactivity in parent compound, CO 2 , and CO in the 24-hour expired breath accounted fo r 77, 1477 9, and 6%, respectively, of the administered radioa ctivity (Angelo et al., 1986b). In mature, male 1478 Sprague-Dawley rats given a smaller dose (1 mg/kg) of [ 14 C]-labeled dichloromethane, 1479 radioactivity in parent compound, CO 2 , and CO in 48-hour expired breath accounted for 12 , 35, 1480 and 31%, respectively; these data indicate that, at lower dose levels, a greater percentage of the 1481 administered dose was metabolized by the CYP pathwa y and eliminated in the expired breath, 1482 compared with higher dose levels (McKenna and Zempe l, 1981). Similar patterns of 1483 radioactivity distribution in parent compound, CO 2 , and CO in expired breath were found in 1484

mature male B6C3F 1 mice following gavage administration of 50 mg/kg-d ay (in water), or 1485 500 or 1,000 mg/kg-day (in corn oil), [ 14 C]-labeled dichloromethane (Angelo et al., 1986a). For 1486 example, radioactivity in parent compound, CO 2 , and CO in 24-hour expired breath accounted 1487 for 61, 18, and 11% of the administered radioactivi ty, following administration of a single 1488 50 mg/kg dose to a group of six mice (Angelo et al. , 1986a). Exhalation rates were similarly 1489 high following inhalation exposure of mature male S prague-Dawley rats (>90%) (McKenna et 1490 al., 1982) or following intravenous administration of dichloromethane to mature male F344 rats 1491 (Angelo et al., 1986b). 1492 Elimination of dichloromethane in the urine of expo sed humans is generally small, with 1493 total urinary dichloromethane levels on the order o f 20–25 μ g or 65–100 μ g in 24 hours 1494 following a 2-hour inhalation exposure to 100 or 20 0 ppm, respectively (DiVincenzo et al., 1495 1972). However, a direct correlation between urina ry dichloromethane and dichloromethane 1496 exposure levels was found in volunteers, despite th e comparatively small urinary elimination 1497 (Sakai et al., 2002). Following administration of a labeled dose in animals, regardless of 149821 DRAFT - DO NOT CITE OR QUOTE

exposure route, generally <5–8% of the label is fou nd in the urine and <2% in the feces 1499 (McKenna et al., 1982; McKenna and Zempel, 1981; Di Vincenzo et al., 1972, 1971). 1500

1501

3.5. PHYSIOLOGICALLY BASED TOXICOKINETIC MODELS 1502

Several PBTK models for dichloromethane in animals and humans have been developed 1503

from 1986 to 2006. These models are mathematical r epresentations of the body and its 1504 absorption, distribution, metabolism, and eliminati on of dichloromethane and select metabolites, 1505 based on the structure of the Ramsey and Andersen ( 1984) model for styrene. The models’ 1506 equations are designed to mimic actual biological b ehavior of dichloromethane, incorporating in 1507 vitro and in vivo data to define physiological and metabolic equation parameters. As such, the 1508

models can simulate animal or human dichloromethane exposures and predict a variety of 1509

dichloromethane and metabolite internal dosimeters (i.e., instantaneous blood and tissue 1510 concentration, area under the curve [AUC] of concen tration versus time plots, rate of metabolite 1511 formation), allowing for the extrapolation of toxic ity data across species, route of exposure, and 1512 high to low exposure levels. The development of di chloromethane PBTK models has resulted in 1513 either increased biological detail and functionalit y or refinement of model parameters with newly 1514 available data. The former type of development pro vides more options for toxicity data 1515 extrapolation, while the latter serves to increase confidence in model predictions and decrease 1516 uncertainty in risk assessments for which the model s were, or will be, applied. This section of 1517 the document describes each of the models reported in the scientific literature and/or used by the 1518 regulatory community (i.e., Occupational Safety and Health Administration [OSHA], EPA) and 1519 their contribution to the advancement of predictive dosimetry and data extrapolation for 1520 dichloromethane. In some instances, model developm ent was accomplished by the addition of 1521 new biological compartments (e.g., tissue systems). Diagrams of the compartmental structure of 1522 the models are shown in Figure 3-2. Significant st atistical advances in parameter estimation also 1523 have been incorporated in model development. For t his reason, some animal and human PBTK 1524 models may be described as deterministic (Sweeney e t al., 2004; Casanova et al., 1996; Reitz et 1525 al., 1988a, b; U.S. EPA, 1988b, 1987a, b; Andersen et al., 1987; Gargas et al., 1986) in which 1526 point estimates for each model parameter are used, resulting in point estimates for dosimetry. 1527 Others may be described as probabilistic (Jonsson a nd Johanson, 2001; El-Masri et al., 1999; 1528 OSHA, 1997), in which probability distributions for each parameter were defined, resulting in 1529 probability distributions for dosimetry. The latte r approach, particularly utilizing a Bayesian 1530 hierarchical statistical model structure (described below) (David et al., 2006; Marino et al., 2006) 1531 to estimate parameter values, allows for the introd uction of intra- and interspecies variability into 1532 model predictions and quantitative assessment of mo del uncertainty. Both deterministic (U.S. 1533 EPA, 1988b, 1987a, b) and probabilistic (OSHA, 1997 ) applications have been used to develop 1534 regulatory values. As discussed below, subsequent applications of the developed models for 1535 cancer risk assessment have resulted in significant ly different estimates of human cancer risk. 153622 DRAFT - DO NOT CITE OR QUOTE 1537

1538 1539 1540 1541 1542 1543 1544 1545 1546 1547 1548 1549 1550 1551 1552 1553 1554 1555 1556 1557 1558 1559 1560 1561 1562 1563 1564 1565 1566 1567 1568 1569 1570 1571 1572 1573 1574 1575 1576 1577 1578 1579 Figure 3-2. Schematics of PBTK models (1986–2006) used in the 1580 development of estimates for dichloromethane intern al dosimetry. 1581 1582 Key references: Model A—Gargas et al. (1986); B—And ersen et al. (1987); C—Andersen et al. 1583 (1991); D—Casanova et al. (1996); E—Sweeney et al. (2004); F—OSHA (1997); G—Jonsson 1584 and Johanson (2001). Models C–G all build on the s tructure in model B. Models E and G have 1585 been applied in humans; all others have been applie d in humans and rodents (mice and/or rats). 1586 CYP = CYP pathway metabolites; GST = GST pathway me tabolites. 1587 Gas Exchange Lung

Fat

Richly Perfused

Slowly Perfused

Liver

CYP GST Blood Alveolar Air

Endogenous Production Gas Exchange

Fat

Richly Perfused

Slowly Perfused

Liver

CYP GST CO Sub Model CYP GST 3

- 2A 3

- 2B

Gas Exchange Lung

Fat

Richly Perfused

Slowly Perfused

Liver

CYP GST CYP

GST

Formaldehyde

DNA

- protein cross

- links Blood Alveolar Air

Endogenous Production Gas Exchange Lung

Fat

Richly Perfused

Slowly Perfused

Liver CYP GST

CYP GST CO Sub Model 3

- 2C 3

- 2D

Gas Exchange Lung

Richly Perfused

Perineral Fat

Subcutaneous Fat

Liver

CYP GST CYP

GST

Working Muscle

Resting Muscle Gas Exchange Lung

Fat

Richly Perfused

Slowly Perfused

Liver

CYP GST CYP

GST

Bone Marrow 3

- 2F

3- - 2G Blood Alveolar Air

Endogenous Production Gas Exchange Lung

Fat

Richly Perfused

Slowly Perfused

Liver CYP GST

CYP GST CO Sub Model

CYP 3

- 2E23 DRAFT - DO NOT CITE OR QUOTE

The deterministic rat model of Gargas et al. (1986 ), based on previous work by Ramsey 1588 and Andersen (1984) examining inhalation pharmacoki netics of styrene in rats, was the first 1589 PBTK model for dichloromethane. It was comprised o f four compartments (fat, liver, richly 1590 perfused tissues, and slowly perfused tissues [Figu re 3-2A]) and described flows and partitioning 1591 of parent material and metabolites through the comp artments with differential equations. 1592 Metabolism, which was restricted to the liver compa rtment, was described as two competing 1593 pathways: the GST pathway, described with a linear first-order kinetic model, and the CYP 1594 pathway, described with a saturable Michaelis-Mente n kinetic model. Rate constants for the 1595 CYP and GST pathways in rats were determined by opt imization of the model with in vivo gas 1596 uptake data. COHb production was modeled both endo genously and from CYP-mediated 1597

metabolism of dichloromethane. This model demonstr ated the dose-dependent flux through the 1598 competing CYP and GST metabolic pathways and the ef fect of CYP inhibition on COHb 1599 generation. 1600 Andersen et al. (1987) extended the rat model of G argas et al. (1986) to include a lung 1601 compartment, including CYP and GST metabolism pathw ays within the lung, in rats, mice, 1602 hamsters, and humans (Figure 3-2B). Physiological flow rates were allometrically scaled among 1603 species by 3⁄4 power of body weight (BW). Rate const ants for the CYP and GST pathways in 1604 rodents were determined by optimization of the mode l with in vivo gas uptake data. CYP rate 1605

constants for humans were derived from data on dich loromethane uptake in human subjects 1606 (number of subjects not reported). Human GST rate constants were derived by allometric 1607 scaling of the animal GST rate constants. Model pr edictions compared favorably with kinetic 1608 data for human subjects exposed by inhalation to di chloromethane (Andersen et al., 1987). 1609 Using the mouse cancer bioassay data from NTP (1986 ), Andersen et al. (1987) compared the 1610 linear body surface area-derived or the PBTK model- derived human liver and lung dose 1611 surrogates associated with tumor development (mg di chloromethane metabolized via GST 1612 pathway/volume tissue/day). They reported that PBT K model-extrapolated human liver and lung 1613 internal doses were for inhalation exposure 167- an d 144-fold lower and for drinking water 1614 exposure 45- and 213-fold lower, respectively, than body surface area scaled internal doses. The 1615 study authors suggested that the lower model-predic ted human internal dose surrogates were due 1616 to the need to saturate the CYP pathway before appr eciable tumorigenic metabolite levels could 1617 be attained, which is not captured by extrapolation based on body surface area. 1618 U.S. EPA (1988b, 1987a, b) slightly modified the A ndersen et al. (1987) model for mice 1619 by using different alveolar ventilation and cardiac flow rates and used the mouse and human 1620

models to derive human cancer risks from animal tum or incidence data. The flow rate 1621 parameters in the Andersen et al. (1987) model were based on a human breathing rate of 1622

12.5 m 3 /day (reflecting a resting rate), compared with the EPA value of 20 m 3 /day (reflecting 1623 average daily activity level) and a mouse breathing rate of 0.084 m 3 /day (based on allometric 1624 scaling of bioassay-specific BWs), compared with th e rate commonly used by EPA, 162524 DRAFT - DO NOT CITE OR QUOTE

0.043 m 3 /day (U.S. EPA, 1987a). The internal dose metric u sed in the applications of the model 1626

to cancer risk assessment was reflective of the amo unt of dichloromethane metabolized by the 1627 GST pathway. In addition to using the mouse and hu man PBTK models to account for species 1628 differences in dosimetry, a body surface area corre ction factor of 12.7 was applied to low-dose 1629 slopes of estimated dose-response relationships for liver and lung tumors in mice to account for 1630 presumed higher human responsiveness, relative to m ice, to dichloromethane-induced cancer 1631 (U.S. EPA, 1987a). The factor of 12.7 is the cube root of the ratio of human to mouse reference 1632 BWs; this BW scaling factor was applied to adjust f or interspecies toxicodynamic variability 1633 (i.e., presumed differences in the lifetime impact in mice and humans of a given daily amount of 1634 dichloromethane metabolically activated per liter o f tissue) (Rhomberg, 1995). A human cancer 1635 inhalation unit risk (IUR) of 4.7 × 10 –7 per (μg/m 3 ), based on this analysis, was placed on IRIS in 1636 September 1990. 1637 The Andersen et al. (1987) models were also modifi ed by addition of submodel structures 1638

for estimation of new dosimeters of interest. Ande rsen et al. (1991) added the capability to 1639 specifically describe the kinetics of dichlorometha ne, CO, and COHb in rats and humans with 1640 the addition of the Coburn-Forster-Kane equation to describe CO and COHb kinetics 1641 (Figure 3-2C). However, equations were not added f or metabolism of dichloromethane to CO in 1642 the lung. Casanova et al. (1996) extended the Ande rsen et al. (1987) mouse model to include a 1643 submodel that predicted the formation of formaldehy de and DNA-protein cross-links in the liver 1644 (Figure 3-2D). 1645 Further refinements of the Andersen et al. (1987) models allowed for incorporation of 1646

new data. New in vitro measurements of metabolic r ate constants in human and animal tissues 1647 were incorporated into the Andersen et al. (1987) m odels by Reitz and coworkers (Reitz, 1991; 1648 Reitz et al.1988a, b). Sweeney et al. (2004) modif ied the Andersen et al. (1987) human PBTK 1649 model, adding extrahepatic CYP metabolism in richly perfused tissues (Figure 3-2E) to obtain a 1650

better fit of the model to kinetics data for humans . Data for 13 volunteers (10 men and 1651 3 women) who were exposed to one or more concentrat ions of dichloromethane for 7.5 hours 1652

included dichloromethane concentrations in breath a nd blood, COHb concentrations in blood, 1653

and CO concentrations in exhaled breath. Individua l CYP V max c (maximal velocity) values were 1654 obtained by optimizing model predictions to match t ime-course data simultaneously for 1655

dichloromethane concentrations in blood and exhaled breath for each individual. Resultant 1656

individual values of CYP V max c ranged from 7.4 to 23.6 mg/hour/kg 0.7 , indicating an 1657 approximate threefold range in maximal CYP metaboli c activity. 1658 The significance of metabolic variability for the kinetics of dichloromethane in animals 1659 and humans was explored by several investigators us ing PBTK models. Dankovic and Bailer 1660 (1994) used the updated human model presented by Re itz et al. (1988a, b) to explore the 1661 consequences of interindividual variability in vitr o kinetic constants for the CYP and GST 1662 pathways (based on data for four human subjects) an d reported that predicted GST-metabolized 166325 DRAFT - DO NOT CITE OR QUOTE

doses to the lung and liver could range from about zero to up to fivefold greater than those 1664 predicted with the values of these rate constants u sed in the Reitz et al. (1988a, b) model. 1665 El-Masri et al. (1999) replaced parameter estimates in the mouse and human PBTK models 1666 presented by Casanova et al. (1996) with probabilit y distributions, including published 1667 information on the distribution of GST-T1 polymorph ism in human populations, and used Monte 1668

Carlo simulations to estimate distributions of canc er potency of dichloromethane in mice, 1669 distributions of the amount of DNA-protein cross-li nks formed in the liver of humans, and 1670 distributions of human cancer risks at given exposu re levels of dichloromethane. The analysis 1671 showed that, at exposure levels of 1, 10, 100, and 1,000 ppm dichloromethane, average and 1672 median cancer risk estimates were 23–30% higher whe n GST-T1 polymorphism was not 1673

included in the model. 1674 Given the demonstrated influence of population var iability in dichloromethane 1675

metabolism on PBTK model-derived cancer risk estima tes (El-Masri et al., 1999; Dankovic and 1676 Bailer, 1994), PBTK model development has included a more formal statistical treatment of data 1677 for physiological and metabolic variability. Bayes ian statistical approaches have been applied to 1678 develop probabilistic PBTK models for dichlorometha ne. Probabilistic models account for 1679 variability between individuals in model parameters by replacing point estimates for the model 1680 parameters with probability distributions. Calibra tion or fitting of probabilistic PBTK models to 1681 experimental toxicokinetic data is facilitated by a Bayesian technique called Markov Chain 1682 Monte Carlo (MCMC) simulation, which quantitatively addresses both variability and 1683 uncertainty in PBTK modeling (Jonsson and Johanson, 2003). 1684 OSHA (1997) used MCMC simulation to fit probabilis tic versions of the Reitz et al. 1685 (1988a, b) and Andersen et al. (1991, 1987) mouse a nd human models, which included 1686 probability distributions for all model parameters. GST- and CYP-mediated metabolism 1687 occurred in the liver and lung compartments (see Fi gure 3-2F). The model parameters were 1688 modified to focus on occupational exposure scenario s; that is, a parameter distribution for work 1689 intensity (using data from Astrand [1989]) was adde d, which adjusted physiological flow rates as 1690 a function of work intensity as measured in watts. In addition, updated measurements of 1691 blood:air and tissue:air partition coefficients (Cl ewell et al., 1993) were used to describe 1692 distributions for these parameters. The Clewell et al. (1993) blood:air partition coefficient of 23 1693 is higher than the value of 8.29 reported by Anders en et al. (1987) and used by EPA (U.S. EPA, 1694 1988b, 1987a, b). The newer Clewell et al. (1993) value for mice is the preferred value, since it 1695 is much closer to the values for rats (19.4) and ha msters (22.5) rather than humans (9.7), as 1696 reported by Andersen et al. (1987). Distributions of metabolic, physiological, and partitioning 1697 parameters in the mouse and human models were updat ed by using Bayesian methods with data 1698 for mice and humans in published studies of mouse a nd human physiology and dichloromethane 1699

kinetic behavior. 170026 DRAFT - DO NOT CITE OR QUOTE

Jonsson et al. (2001) used additional human kineti cs data to expand the PBTK model of 1701 Reitz et al. (1988a, b) and added new model compart ments (Figure 3-2G). These investigators 1702 used MCMC simulation to develop a probabilistic mod el from the Reitz et al. (1988a, b) human 1703 model by using published in vitro measurements of l iver V max for the CYP pathway (Reitz et al., 1704 1989) and kinetic data for five human subjects expo sed by inhalation to dichloromethane 1705 (Astrand et al., 1975). A working muscle compartme nt was added to the basic Andersen et al. 1706 (1987) and Reitz et al. (1988a, b) structure (see F igure 3-2G). Jonsson and Johanson (2001) 1707 refined and extended this probabilistic model by in cluding an additional fat compartment (to 1708 provide a better description of the experimental da ta for the time course of dichloromethane in 1709 subcutaneous fat), incorporating (with MCMC simulat ion) kinetic data for dichloromethane in an 1710 additional 21 human subjects and including three GS T-T1 genotypes/phenotypes 1711 (nonconjugators –/–, low conjugators +/–, high conj ugators +/+). Monte Carlo simulations were 1712 then used with the refined probabilistic model to p redict human liver cancer risk estimates at 1713 several dichloromethane exposure levels using an al gorithm similar to the one used by El-Masri 1714 et al. (1999), using DNA-protein cross-links as the internal dose metric. The mean, 50 th , 90 th

, 1715

and 95 th percentile human cancer risk values from Jonsson e t al. (2001) and El-Masri et al. 1716 (1999) were very similar, within onefold of one ano ther for simulated exposure levels up to 1717 100 ppm. 1718 The most statistically rigorous and data-intensive PBTK model development was 1719 performed by Marino et al. (2006) for mice and Davi d et al. (2006) for humans. Development of 1720 these models used multiple mouse and human data set s in a Bayesian hierarchical statistical 1721 structure to quantitatively capture population vari ability and reduce uncertainty in model 1722 dosimetry and the resulting risk values. EPA used these models in the derivation of reference 1723

values and cancer risk estimates in the current ass essment, and these models are described in 1724

more detail below. 1725

1726

3.5.1. Probabilistic Mouse PBTK Dichloromethane Mo del (Marino et al., 2006) 1727 Marino et al. (2006) used MCMC analysis to develop a probabilistic PBTK model for 1728 dichloromethane in mice, using the Andersen et al. (1987) model structure as a starting point 1729 (Figure 3-3). Metabolic kinetic parameters (V max c, K m , k f C, A1, and A2) (Table 3-5) were 1730 calibrated with this Bayesian methodology by using several experimental data sets. Distribution 1731 parameters (i.e., means and coefficients of variati on [CVs]) for other physiological parameters 1732 (i.e., BW, fractional flow rates, and fractional ti ssue volumes) and partition coefficients were 1733 taken from the general literature as noted by Clewe ll et al. (1993). Marino et al. (2006) noted 1734 that using distributions for these latter parameter s from the general literature (based on a large 1735 number of animals) was better than updating them ba sed on the relatively smaller number of 1736

animals in the available dichloromethane kinetic st udies. Clewell et al. (1993) determined 1737 blood:air and tissue:air partition coefficients (me ans and CVs) with tissues from groups of male 173827 DRAFT - DO NOT CITE OR QUOTE

and female B6C3F 1 mice. These partition coefficients were derived b y using a vial equilibration 1739 method similar to that used by prior investigators (Andersen et al., 1987; Gargas et al., 1986). 1740 Tissue:air partition coefficients were approximatel y two to three times lower than previously 1741 utilized values with the exception of the liver coe fficient, which was similar to previous values 1742 (Table 3-5). The blood:air partition coefficient ( 23) from Clewell et al. (1993) is higher than the 1743 previously reported value of 8.3 (Gargas et al., 19 86). The higher value is more in line with 1744 values measured in rats (19.4) and hamsters (22.5) and, thus, is more reasonable than the older 1745 value of 8.3. Table 3-5 shows mean and CVs for phy siological parameters and partition 1746 coefficients in the Marino et al. (2006) mouse mode l as well as values used in earlier 1747

deterministic PBTK mouse models for dichloromethane

. 1748

1749

1750

1751

1752

1753

1754

1755

1756

1757

1758

1759

1760

1761

1762

1763

1764

1765

1766 Figure 3-3. Schematic of mouse PBTK model used by Marino et al. (2006). 1767 1768 1769 The Bayesian calibration of the cardiac output con stant (QCC), ventilation:perfusion ratio 1770 (VPR), and metabolic parameters was divided into th ree sequential steps: using kinetic data from 1771

closed chamber studies with mice treated with an in hibitor of CYP2E1 (trans-1,2- 1772 dichloroethylene) in order to minimize the oxidativ e pathway and enable a more precise estimate 1773 of parameters for the GST pathway, followed by kine tic data for mice given intravenous 1774 injections of dichloromethane to estimate metabolis m parameters in the absence of pulmonary 1775 absorption processes and, finally, kinetic data for naïve mice exposed to dichloromethane in 1776 closed chambers (Marino et al., 2006). The initial prior distributions were based on mean values 1777 Blood Alveolar air

Endogenous production Gas exchange Lung

Fat

Richly perfused

Slowly perfused

Liver CYP GST

CYP GST CO sub model28 DRAFT - DO NOT CITE OR QUOTE

used by Andersen et al. (1987) for the metabolic pa rameters and by OSHA (1997) for the 1778 parameters for VPR, ratio of lung V max to liver V max (A1), and ratio of lung GST 1 st order kinetic 1779 constant (lung K F ) to liver K F (A2). Posterior distributions from the first Baye sian analysis were 1780 used as prior distributions for the second step, an d posterior distributions from the second step 1781 were used as prior distributions for the final upda ting. Final results from the Bayesian 1782 calibration of the mouse probabilistic model are sh own in Table 3-5. 178329Table 3-5. Values for parameter distributions in aB6C3F1mouse probabilistic PBTK model fordichloromethane compared with associated values forpoint parameters in earlier deterministic B6C3F1mousePBTK models for dichloromethaneMarino et al. (2006)aParameterPrior meanPrior CVFinal posteriormeanFinal posteriorCVU.S. EPA(1988b,1987a, b)Andersen et al.(1987)Fractional flow rates (fraction of QCC)bQFC FatQLC LiverQRC Rapidly perfused tissuesQSC Slowly perfused tissues0.050.240.520.190.600.960.500.400.050.240.520.190.050.240.520.19Fractional tissue volumes (fraction of BW)bVFC FatVLC LiverVLuC LungVRC Rapidly perfused tissuesVSC Slowly perfused tissues0.040.040.01150.050.780.300.060.270.300.300.040.040.01190.050.780.040.040.01190.050.78Partition coefficientscPB Blood:airPF Fat:bloodPL Liver:bloodPLu Lung:bloodPR Rapidly perfused:bloodPS Slowly perfused:blood235.11.60.460.520.440.150.300.200.270.200.20These parameters were taken from anextensive literature database derived froma large number of animals; therefore,further Bayesian updating does notinform on the true mean and variance forthese values.8.2914.51.711.711.710.968.2914.51.711.711.710.96Flow ratesQCC Cardiac output (L/hr/kg0.74)VPR ventilation:perfusion ratio28.01.520.580.7524.21.450.190.2014.3d1.028.0e1.0Metabolism parametersVmaxc Maximum CYP metabolic rate (mg/hr/kg0.7)KmCYP affinity (mg/L)kfC First-order GST metabolic rate constant (kg0.3/hr)A1 Ratio of lung Vmaxc to liver VmaxcA2 Ratio of lung kfC to liver kfC11.10.3961.460.4620.3222220.550.559.270.5741.410.2070.1960.210.420.280.360.3711.10.3961.460.4160.13711.10.3961.460.4160.1371784aMCMC analysis was used to update prior distributions (means and CVs) for flow rate and metabolic parameters in a sequential process with three sets of kinetic1785data from mouse studies, as explained further in the text. Final values for posterior distributions are given in this table.1786bSource: Andersen et al. (1987, 1991).1787cSource: Clewell et al. (1993).1788dBased on a mouse breathing rate of 0.043 m3/day.1789eBased on a mouse breathing rate of 0.084 m3/day.179030

Marino et al. (2006) used the Bayesian-calibrated mouse model to calculate internal dose 1791 metrics associated with exposure conditions in the NTP (1986) B6C3F 1 mouse cancer inhalation 1792 bioassay. The internal dose metric selected was mi lligrams (mg) dichloromethane metabolized 1793 by the GST pathway per liter tissue per day. This is the same dose metric used in earlier 1794 applications of PBTK models to derive human cancer IUR estimates based on cancer responses 1795 in mice (OSHA, 1997; Andersen et al., 1987; U.S. EP A, 1987a, b). Its use is consistent with 1796

evidence that dichloromethane metabolism via GST-T1 results in the formation of a reactive 1797 metabolite that damages DNA and results in the for mation of tumors (see section 4.7). The 1798

model was used to calculate values for this interna l dose metric in the lung and liver of mice in 1799 the NTP (1986) study, using the mean values of the final distributions for the parameters in the 1800

model. Resultant values were three- to four-fold h igher than values calculated with the Andersen 1801 et al. (1987) and U.S. EPA (1987a, b) versions of t he model (Table 3-6). Marino et al. (2006) 1802 noted that the difference could be primarily attrib uted to the changes in the partition coefficients 1803 based on Clewell et al. (1993) as well as to the Ba yesian updating of the metabolic parameters 1804 (see Table 3-5). 1805

1806 Table 3-6. Internal daily doses for B6C3F 1 mice exposed to

dichloromethane for 2 years (6 hours/day, 5 days/we ek) calculated with different PBTK models

PBTK model Target organ NTP (1986)

exposure level a Marino et al. (2006) U.S. EPA (1987a, b) Andersen et al. (1987)

Liver b Control 2,000 ppm 4,000 ppm 0 2,359.99 4,869.85 0

727.8 1,670 0

851 1,811 Lung b Control 2,000 ppm 4,000 ppm 0 474.991 973.343 0

111.4 243.7 0 123 256

a 2,000 ppm = 6,947 mg/m 3 ; 4,000 ppm = 13,894 mg/m 3

. b Internal dose expressed as mg dichloromethane metab olized by the GST pathway per liter tissue per day. 1807 1808 Marino et al. (2006) noted that inclusion of extra hepatic CYP metabolism in the slowly 1809 perfused tissue compartment in the mouse model had little impact on the formation of GST 1810 metabolites in the liver and lung, especially at ex posure levels used in the mouse NTP (1986) 1811 bioassay. To support this contention, the Andersen et al. (1987) model was modified to include 1812

10% of the liver rate of oxidative metabolism in th e slowly perfused tissue compartment (as 1813 suggested by Sweeney et al. [2004]), and the modifi ed model was used to calculate the formation 1814 of GST metabolites. If extrahepatic metabolism was included in the slowly perfused tissue 1815 compartment, there was a 5–6% reduction in the form ation of GST metabolites in the lung and 1816 liver at an exposure level of 50 ppm. At 2,000 or 4,000 ppm, however, there was only a 0.77 or 1817 0.37% reduction, respectively. Marino et al. (2006 ) did not discuss the impact of including 181831

extrahepatic metabolism in the rapidly perfused tis sue compartment; the same group of 1819 investigators developed a human PBTK model that inc luded CYP metabolism in the richly 1820 perfused compartment (David et al., 2006). 1821

1822

3.5.2. Probabilistic Human PBTK Dichloromethane Mo del (David et al., 2006) 1823 The basic model structure used by David et al. (200 6) was that of Andersen et al. (1987) 1824

with the addition of the CO submodel of Andersen et al. (1991), refinements from the Marino et 1825 al. (2006) mouse model, and an inclusion of CYP met abolism in richly perfused tissue (Figure 3- 1826 4). David et al. (2006) used Bayesian analysis to develop and calibrate metabolic parameters in a 1827 human probabilistic PBTK model for dichloromethane, using kinetic data from several studies of 1828 volunteers exposed to dichloromethane (n = 13 from DiVincenzo and Kaplan [1981]; n = 12 1829 from Engström and Bjurström [1977]; n = 14 from Ast rand et al. [1975]; n = 3 from Stewart et 1830 al. [1972a], and group means for metabolism paramet ers from Andersen et al. [1991]). Exhaled 1831

dichloromethane and CO and blood levels of dichloro methane and COHb were available in the 1832 studies by Andersen et al. (1991) and DiVincenzo an d Kaplan (1981). The other three studies 1833 included two or three of these measures. The only available data for levels of dichloromethane 1834 in fat came from the study of Engström and Bjurströ m (1977) (described in section 3.2 within 1835 adipose tissue). 1836

1837

1838

1839

1840

1841

1842

1843

1844

1845

1846

1847

1848 Figure 3-4. Schematic of human PBTK, used by David et al. (2006). 1849 1850 1851 Values (means and SDs or CVs) for the model parame ter distributions were selected from 1852 multiple sources considered to provide the most cur rent scientific evidence for each parameter 1853 (David et al., 2006). Mean values for cardiac outp ut (QCC), VPR, and all fractional tissue 1854

volumes and blood flow rates were based on mean val ues used by EPA (U.S. EPA, 2000d) in a 1855 PBTK model for vinyl chloride, as were values for C Vs for all physiological parameters, except 1856 CVs for VPR and fractional lung volume, which were set to those used by OSHA (1997). Means 1857 Blood Alveolar Air

Endogenous Production Gas Exchange Lung

Fat

Richly Perfused

Slowly Perfused

Liver CYP GST

CYP GST CO Sub Model

CYP32

for the CO submodel parameters were set equal to th ose in Andersen et al. (1991), except for 1858 those for the endogenous rate of CO production (REn COC) and the background amount of CO 1859 (ABCOC), which were based on data collected by DiVi ncenzo and Kaplan (1981). Means for 1860 partition coefficients, the ratio of lung V max to liver V max , and the ratio of lung K F to liver K F 1861 (A2) were those used by Andersen et al. (1987), whe reas prior means for V max c and K m were 1862 those used by Andersen et al. (1991). The prior me an for the metabolic parameter for CYP 1863 metabolism in the rapidly perfused tissue was set a t 0.03, slightly lower than the value suggested 1864 by Sweeney et al. (2004). Prior CVs for the metabo lic parameters were set at 200%. 1865 MCMC analysis was used to calibrate metabolic para meters in the human model in a 1866 two-step approach: (1) posterior distributions were estimated separately by using data from each 1867

of the five studies with kinetic data for humans ex posed to dichloromethane (with durations 1868 ranging from 1 to 8 hours and concentrations rangin g from 50 to 1,000 ppm); and (2) posterior 1869

distributions were estimated with combined data fro m the 42 individual subjects from the four 1870 studies with individual subject data (DiVincenzo an d Kaplan, 1981; Engström and Bjurström, 1871 1977; Astrand et al., 1975; Stewart et al., 1972a). Results from the Bayesian calibration with the 1872 combined kinetic data for individual subjects are s hown in Table 3-7. This analysis resulted in a 1873 narrowing of the distribution for the CYP2E1 metabo lism parameters V max and K m , from a fairly 1874 broad prior distribution with a CV of 200% for both parameters to 13.1 and 33.6%, respectively, 1875

for V max and K m . 1876

1877 Table 3-7. Results of calibrating metabolic parame ters in a human

probabilistic PBTK model for dichloromethane with i ndividual kinetic data for 42 exposed volunteers and MCMC analysis

Prior distributions Posterior distributions

Parameter Mean (arithmetic) CV Mean (arithmetic) CV

V max c —-maximal CYP metabolic rate (mg/hr/kg 0.7 ) 6.25 2 9.42 0.131

K m —CYP affinity (mg/L) 0.75 2 0.433 0.336

k f C—first-order GST metabolic rate (kg 0.3 /hr) 2 2 0.852 0.711 A1—ratio of lung V max c to liver V max c 0.00143 2 0.000993 0.399 A2—ratio of lung k f

C to liver k f

C 0.0473 2 0.0102 0.728

FracR—fraction of V max c in rapidly perfused tissues 0.03 2 0.0193 0.786

Source: David et al. (2006). 1878 1879 The parameter statistics shown in Table 3-7 (value s reported by David et al., 2006) are 1880 summary statistics of the converged parameter chain s obtained in that analysis, leaving out any 1881 evaluation of correlation or covariance among the u pdated parameters. As such, these statistics 1882 implicitly include both the inter-individual variab ility that would have been elucidated by the 188333

Bayesian analysis (variation between mean values fo r each individual for which data were 1884 available) and uncertainty in those values. 1885 David et al. (2006) further refined the human prob abilistic model to reflect 1886 polymorphisms in the GST pathway: homozygous posit ive (+/+) GST-T1 individuals, 1887 heterozygous (+/–) GST-T1, and homozygous negative (–/–) GST-T1 individuals with no GST 1888 activity. Distributions of GST activities for thes e genotypes in a group of 208 healthy male and 1889 female subjects from Sweden were scaled to obtain d istributions of k f C for each genotype that, 1890 when weighted by estimated frequencies of the genot ypes in the U.S. population, would result in 1891 an overall population mean equal to the k f C mean for the posterior distribution shown in 1892 Table 3-7 (0.852 kg 0.3 /hour). The resultant mean k f C values were 0.676 kg 0.3 /hour (SD 0.123) 1893 for heterozygous individuals and 1.31 kg 0.3 /hour (SD 0.167) for homozygous positive 1894 individuals. The final parameter distributions use d by David et al. (2006) are summarized in 1895

Table 3-8. 1896 As described in Appendix B, EPA undertook an evalua tion of the David et al. (2006) 1897 model and parameterization, focusing on the adequac y of the characterization of parameter 1898 distributions in the full human population. EPA’s c onclusion is that the reported distributions for 1899 physiological parameters in particular, but also ke y metabolic parameters, only represented a 1900 narrow set of adults (with the exception of BW). T he EPA therefore chose to use supplemental 1901 data sources to define these distributions in a way that should fully characterize the variability in 1902 the human population for individuals between six mo nths and eighty years of age. The EPA 1903 incorporated additional data concerning the variabi lity in CYP2E1 activity among humans, based 1904 on Lipscomb et al. (2003). The Lipscomb et al. (20 03) study was based on in vitro analysis of 1905 liver samples from 75 human tissue donors (activity towards trichloroethylene and measurements 1906 of protein content) to estimate a distribution of a ctivity in the population. These data support a 1907 wider distribution in CYP2E1 activity than had been used in the David et al. (2006) model, with 1908 approximately a sixfold range between the upper and lower bounds in Lipscomb et al. (2003) and 1909 a twofold range in David et al. (2006). Thus the E PA replaced the David et al. distribution 1910 parameters by using the same GM = 9.34, but GSD = 1 .73. Further, since even the data available 1911 to Lipscomb et al. (2003) were limited, and the log -normal distribution is naturally bounded to 1912 be greater than zero, the EPA chose to use a non-tr uncated distribution for this parameter. (Since 1913 the distribution form for CYP2E1 was set by David e t al. (2006) to be log-normal and the U.S. 1914 EPA chose to retain that form, even without specifi c bounds the distribution only includes values 1915 greater than zero.) Finally, the scaling of CYP2E1 for individuals under the age of 18 was 1916 adjusted based on the data of Johsrud et al. (2003) ; the EPA’s analysis of these data indicate 1917 CYP2E1 activity in children is better predicted whe n assumed to scale with body weight (BW) 1918 raised to the 0.88 power, as compared to the more g eneral power of 0.74, used by David et al. 1919 CYP2E1 activity for individuals over the age of 18 is still assumed to scale as BW 0.74

. 1920

192134

Table 3-8. Parameter distributions used in human Monte Carlo a nalysis for dichloromethane by David et al. (2006)

Distribution

Parameter Mean (arithmetic) SD Source BW Body weight (kg) 70.0 21.0 Humans a QCC Cardiac output (L/hr/kg 0.74 ) 16.5 1.49 Humans a VPR Ventilation:perfusion ratio 1.45 0.203 Humans a QFC Fat 0.05 0.0150 Humans a QLC Liver 0.26 0.0910 Humans a QRC Rapidly perfused tissues 0.50 0.10 Humans a QSC Slow perfused tissues 0.19 0.0285 Humans a Tissue volumes (fraction BW) VFC Fat 0.19 0.0570 Humans a

VLC Liver 0.026 0.00130 Humans a VLuC Lung 0.0115 0.00161 Humans a VRC Rapidly perfused tissues 0.064 0.00640 Humans a VSC Slowly perfused tissues (muscle) 0.63 0.189 Hum ans a

Partition coefficients PB Blood:air 9.7 0.970 Humans b

PF Fat:blood 12.4 3.72 Rats b

PL Liver:blood 1.46 0.292 Rats b PLu Lung:arterial blood 1.46 0.292 Rats b PR Rapidly perfused tissue:blood 1.46 0.292 Rats b PS Slowly perfused tissue (muscle:blood) 0.82 0.164 Rats b Metabolism parameters

V max c Maximum metabolism rate (mg/hr/kg 0.7 ) 9.42 1.23 Calibration c

K m Affinity (mg/L) 0.433 0.146 Calibration c A1 Ratio of lung V Max to liver V max 0.000993 0.000396 Calibration c A2 Ratio of lung KF to liver KF 0.0102 0.00739 Cali bration c FracR Fractional CYP2E1 capacity in rapidly perfuse d tissue 0.0193 0.0152 Calibration c First order metabolism rate (/hr/kg 0.3 )

Homozygous (–/–) 0 0 Calibration c Heterozygous (+/–) 0.676 0.123 Calibration c k f C Homozygous (+/+) 1.31 0.167 Calibration c

a US EPA, 2000d. Human PBTK model used for vinyl chlo ride.

b Andersen et al. (1987). Blood:air partition measur ed using human samples; other partition coefficient s based on estimates from tissue measures in rats.

c Bayesian calibration based on five data sets (see t ext for description); posterior distributions prese nted in this table.

Source: David et al. (2006). 1922 In addition, while the BW distribution in the David et al. (2006) PBTK model used 1923 ranges from 7 to 130 kg, thus covering 6-month-old children to obese adults, there are age1924 dependent changes and gender-dependent differences in ventilation rates and body fat that are 1925 not explicitly included. To more accurately reflec t the distribution of physiological parameters 1926 in the entire population, EPA replaced the unstruct ured distributions of David et al. (2006) with 1927

distributions based on available information that s pecifically account for population variability in 1928 age, gender, and age- and gender-specific distribut ions or functions for BW, QCC, alveolar 192935

ventilation, body fat (fraction), and liver fractio n (see Appendix B for more details of the 1930 evaluation of each of these parameters). 1931 The resulting set of parameter distribution charact eristics, including those used as defined 1932 by David et al. (2006) are described in Table 3-9. Using this revised set of distributions, 1933 including the (revised) CYP and (published) GST act ivity distributions, and other distributions 1934 used as defined by David et al. (2006), the model a s applied should reflect the full variability in 1935 the (U.S.) human population. 1936

1937361938Table 3-9. Parameter distributions for the human PBTK model for dichloromethane used by EPADistributionParameterShape(Geometric)meanaSD/GSDaLowerboundUpperboundSection or sourceBW Body weight (kg) Normalf (age, gender)1st%tile 99th%tileB-4.3; NHANES IVFlow ratesQAlvC Alveolar ventilation (L/hour/kg0.75) Normalf (age, gender)f (age)5th%tile95th%tileB-4.4; mean: Clewell et al. (2004);SD: Arcus-Arth and Blaisdell (2007)vprv Variability in ventilation:perfusion ratio Log-normal 1.00 0.203 0.69 1.42 VPR/VPRmeanof David et al. (2006)QCC Cardiac output (L/hour/kg0.75) QCCmean=f(QAlvC) QCC = QCCmean/vprv B-4.5; Clewell et al. (2004) (mean)Fractional flow rates (fraction of QCC)QFC Fat Normal 0.05 0.0150 0.0050 0.0950QLC Liver Normal 0.26 0.0910 0.010 0.533QRC Rapidly perfused tissues Normal 0.50 0.10 0.200.80QSC Slow perfused tissues Normal 0.19 0.0285 0.1050.276David et al. (2006);after sampling fromthese distributions, normalize:∑⋅=QjCQiCQCQiTissue volumes (fraction BW)VFC Fat Normalf (age, gender)0.3·mean 0.1·mean1.9·meanVLC Liver Normalf (age) 0.05·mean 0.85·mean1.15·meanVLuC Lung Normal 0.0115 0.00161 0.00667 0.0163VRC Rapidly perfused tissues Normal 0.064 0.00640 0.0448 0.0832VSC Slowly perfused tissues Normal 0.63 0.189 0.4310.829Fat mean: B-4.6 (Clewell et al., 2004); livermean: B-4.7 (Clewell et al., 2004);otherwise, David et al. (2006);aftersampling from these distributions,normalize:∑⋅⋅=VjCViCBWVi9215.0Partition coefficientsPB Blood:air Log-normal 9.7 1.1 7.16 13.0PF Fat:blood Log-normal 11.9 1.34 4.92 28.7PL, PLu,& PRLiver:blood, lung:arterial blood, andrapidly perfused tissue:bloodLog-normal 1.43 1.22 0.790 2.59PS Slowly perfused tissue (muscle):blood Log-normal0.80 1.22 0.444 1.46Geometric mean & GSD values listed here,converted from arithmetic mean and SDvalues of David et al. (2006)Metabolism parameters (based on Monte Carte calibration from five human data sets)VmaxcMaximum metabolism rate(mg/hr/kgXvmax)Lognormal9.341.73(none)(none)B-3 mean: David et al. (2006);GSD & bounds: Lipscomb et al. (2003);Xvmax = 0.88 for age <18;.Xvmax = 0.70 for age≥18.KmAffinity (mg/L) Log-normal 0.41 1.39 0.154 1.10 Geometric mean & GSD values listed here,37Table 3-9. Parameter distributions for the human PBTK model for dichloromethane used by EPADistributionParameterShape(Geometric)meanaSD/GSDaLowerboundUpperboundSection or sourceA1 Ratio of lung VMaxto liver VMaxLog-normal 0.00092 1.47 0.0002910.00292A2 Ratio of lung KF to liver KF Log-normal 0.0083 1.92 0.00116 0.0580FracRFractional MFO capacity in rapidlyperfused tissueLog-normal 0.0152 2.0 0.00190 0.122converted from arithmetic mean and SDvalues of David et al. (2006)First order metabolism rate ([hour/kg0.3]–1)Homozygous (–/–) Normal 0 0 0 0Heterozygous (+/–) Normal 0.676 0.123 0.001.291kfCHomozygous (+/+) Normal 1.31 0.167 0.002.145David et al. (2006)aArithmetic mean and SD listed for normal distributions; geometric mean and geometric SD (GSD) listed for log-normal distributions.193938

3.5.3. Evaluation of Rat PBTK Dichloromethane Mode ls 1940

Several deterministic PBTK rat models have been re ported in the scientific literature 1941 (Sweeney et al., 2004; Andersen et al., 1991, 1987; Reitz , 1991; Reitz et al.,1988a, b; U.S. EPA 1942 1988b, 1987a, b; Gargas et al., 1986). Unlike the mouse (Marino et al., 2006) and human (David 1943 et al., 2006), no hierarchical population model for dichloromethane in the rat exists in which 1944 parameter uncertainty is quantitatively integrated into model calibration. Rat data are not 1945 available that would allow for Bayesian calibration of individual metabolic parameters for the 1946 CYP or GST pathways. Thus, EPA assessed modified v ersions of deterministic rat PBTK 1947 models to select the most appropriate model for use in extrapolating internal dosimetry from rats 1948 to humans, for example in the determination of RfDs and RfCs based on effects seen in the rat. 1949 This work is described in detail in Appendix C and is based on evaluation of blood levels of 1950 dichloromethane and the percent saturation of hemog lobin as COHb (%COHb) and expired 1951 dichloromethane following intravenous injection (An gelo et al., 1986b), closed chamber gas 1952 uptake (Gargas et al., 1986), and dichloromethane a nd %COHb blood levels from a 4-hour 1953 inhalation exposure (Andersen et. al., 1991, 1987). Based on this work, the basic model 1954 structure of Andersen et al. (1991) was chosen, wit h the inclusion of lung dichloromethane 1955 metabolism via CYP (4% of liver metabolite producti on) and GST (14% of liver metabolite 1956 production) pathways (estimated from Reitz et al., 1989) (Figure 3-5) with metabolic parameters 1957 recalibrated against data of Andersen et al. (1991) , based on prediction agreement of the various 1958 parameters with the available rat data sets. Table 3-10 presents the parameter distribution data 1959

for this model. 1960

1961 Figure 3-5. Schematic of rat PBTK model used in cu rrent assessment. 1962 Blood Alveolar Air

Endogenous Production Gas Exchange Lung

Fat

Richly Perfused

Slowly Perfused

Liver CYP GST

CYP GST CO Sub Model39 Table 3-10. Parameter values for the rat PBTK mode l for dichloromethane used by EPA

Parameter Mean

Flow rates QCC (L/hour/kg 0.74 ) 15.9

VPR 0.94 Fractional flow rates (percent of QCC)

Fat 9

Liver 20 Rapidly perfused tissues 56 Slowly perfused tissues 15 Tissue volumes (percent BW)

Fat 7

Liver 4 Lung (scaled as BW 0.99 ) 1.15 Rapidly perfused tissues 5 Slowly perfused tissues 75

Partition coefficients

Blood:air 19.4

Fat:blood 6.19

Liver:blood 0.732 Lung:arterial blood 0.46 Rapidly perfused tissue:blood 0.732 Slowly perfused tissue (muscle):blood 0.408 Metabolism parameters Maximum metabolism rate (mg/hour/kg 0.7 ) 3.93 Affinity (mg/L) 0.524 Ratio of lung V Max to liver V Max 0.04 Ratio of lung KF to liver KF 0.14 1st order metabolism rate (liver KF) ([hour/kg 0.3 ] –1 ) 2.46 Oral absorption constant, k a (1/hr) 1.80 1963 3.5.4. Comparison of Mouse, Rat and Human PBTK Mod els 1964 The comparison of various parameters across specie s (Table 3-11) primarily shows the 1965 modest inter-species differences that are known to occur in physiological parameters, also 1966 including the approximately 2-fold differences in p artition coefficients which occur because of 1967 differences in rodent versus human blood lipid cont ent. The 2.5-fold lower V max c (CYP activity) 1968 in rats versus mice is also typical. The most stri king difference is the variation in A1 and A2. 1969 Those values, however, reflect the in vitro differe nces originally quantified by Lorenz et al. 1970 (1984) and used in the dichloromethane PBPK modelin g of Anderson et al. (1987). Thus these 1971 differences are based on independent measurements o f tissue-specific metabolic capacity, and 197240

while the specific values for mouse and human were refined through Bayesian analysis, the 1973 ultimate (posterior) values used are within a reaso nable range of the in vitro measurements and 1974 so do not appear to be artifactual. (Since in vivo kinetics often indicate some differences from 1975 what would be predicted without adjustment from in vitro, it is not surprising that such 1976 differences occur here.) These differences do expl ain why lung-specific metrics in particular 1977 lead to lower internal dose and hence risk predicti ons in humans compared to whole-body 1978

metrics. 197941Table 3-11. Parameters in the mouse, rat, and human PBTK model for dichloromethane used by the EPAMouseaRatbHumancParameterMean Value Mean CV/GSD(Shape, bounds)SourcesFractional flow rates (fraction of cardiac output)bQFC FatQLC LiverQRC Rapidly perfused tissuesQSC Slowly perfused tissues0.050.240.520.190.090.200.560.150.050.260.500.190.3 (N, 0.1-1.9)0.35 (N, 0.0385-2.05)0.2 (N, 0.4-1.6)0.15 (N, 0.553-1.453)David et al. (2006);thennormalized:∑⋅=QjCQiCQCQiFractional tissue volumes (fraction of body weight)bVFC FatVLC LiverVLuC LungVRC Rapidly perfused tissuesVSC Slowly perfused tissues0.040.040.01150.050.780.070.040.01150.050.75f(age, gender)f(age)0.01150.0640.630.3 (N, 0.1-1.9)0.05 (N, 0.85-1.15)0.14 (N, 0.58-1.42)0.1 (N, 0.7-1.3)0.3 (N, 0.684-1.32)Fat mean: §2.2.3.6;Liver mean: §2.2.3.7;otherwise David et al.(2006);then normalized:∑⋅⋅=VjCViCBWVi9215.0Partition coefficientscPB Blood/airPF Fat/bloodPL Liver/bloodPLu Lung/bloodPR Rapidly perfused/bloodPS Slowly perfused/blood23.05.11.60.460.520.4419.46.190.730.460.730.419.711.91.431.431.430.801.1(LN, 0.738-1.34)1.34(LN, 0.413-2.41)1.22(LN, 0.552-1.81)""1.22(LN, 0.555-1.83)Geometric mean (GM) &GSD/GM valuesconverted from arithmeticmean & SDs of David etal. (2006)Flow ratesQCC Cardiac output (L/hr/kg0.74)VPR ventilation/perfusion ratioQAlvC24.21.45QCC/VPR14.990.94QCC/VPRQCCmean=f(QAlvC)(variable)f (age, gender)QCC = QCCmean/vprv(varies)(LN, 0.69-1.42)f (age)(N, 5th-95th%)QCC: §2.2.3.5;vprv = VPR/VPRmean:David et al. (2006);QAlvC: §2.2.3.4;Metabolism parametersVmaxc Maximum CYP metabolic rate (mg/hr/kgXvmax)Xvmax CYP allometric scaling powerKmCYP affinity (mg/L)kfC First-order GST metabolic rate constant (kg0.3/hr)A1 Ratio of lung Vmaxc to liver VmaxcA2 Ratio of lung kfC to liver kfC9.270.70.5741.410.2070.1963.930.70.5242.460.040.149.340.88 for age <18;0.7 for age0.410 (-/-)e0.676 (+/-)e1.31 (+/+)e0.000920.00831.73(LN, [unbounded])1.39(LN, 0.376-2.68)-/-: NA+/-: 0.182 (N, 0-1.91)+/+: 0.128 (N, 0-1.64)1.47(LN, 0.316-3.17)1.92(LN, 0.140-6.99)Vmaxc:§2.2.2;others: David et al.(2006) (GM & GSD/GMvalues converted fromarithmetic mean & SDs )aBased on Marino et al. (2006) (source for all mouseparameters)bBased on Andersen et al. (1991), with the additionof lung metabolism of dichloromethane via the CYP (4% of liver metabolite production) and GST (14% ofliver42metabolite production) pathways. Physiological parameters and partition coefficients are from Andersen et al. (1991). The values for dichloromethane metabolismin the lung (as a fractional yield of liver metabolism for each pathway) were estimated from the in vitro ratios of enzyme activity (nmol/min/mg protein)in lung andliver cytosolic (GST) and microsomal (CYP) tissue fractions (Reitz et al., 1989). Metabolic parameterswere re-optimized against the inhalation data of Andersen etal. (1991) using a heteroscedasticity parameter value of 2, which uses relative error for the model fitting algorithm. See Appendix C for further details.cBased on David et al. (2006), with changes as noted. Additional sources include Clewell et al. (2004), Arcus-Arth and Blaisdell (2007), and Lipscomb et al. (2003).See identified sections for details. Distribution values (mean and a measure of dispersion) are provided with the CV (mean/SD) presented for normal (N) distributionsand the GSD (italicized) presented for log-normal (LN) distributions. Distributions were truncated, bounds are (upper-lower bound)/mean.eValues for the homozygous (–/–), heterozygous (+/–), and homozygous (+/+) GST-T1 genotypes, respectively.43 HAZARD IDENTIFICATION 1980 1981 1982

4.1. STUDIES IN HUMANS 1983 4.1.1. Introduction—Case Reports, Epidemiologic, a nd Clinical Studies 1984

There has been considerable interest in the influe nce of occupational exposure to 1985 dichloromethane in relation to a variety of conditi ons. The recognition that dichloromethane can 1986 be metabolized and bound to hemoglobin to form COHb , resulting in a reduction in the oxygen 1987 carrying capacity of the blood (Stewart et al., 197 2b), prompted investigations into risk of 1988

ischemic heart disease and other cardiovascular eff ects. Reports of neurological effects from 1989 acute, high-exposure situations contributed to conc ern about neurological effects of chronic 1990 exposure to lower levels of dichloromethane. A gen eral interest in potential cancer risk became 1991 more focused on lung and liver cancer because of th e observation of these specific tumors in the 1992 NTP (1986) experiments in mice. Details of the stu dies pertaining to the experimental and 1993 epidemiologic studies of noncancer outcomes (e.g., cardiac, neurologic, hepatic, reproductive) 1994 are presented in section 4.1.2, and studies of canc er risk are presented in section 4.1.3. 1995

1996

4.1.2. Noncancer Studies 1997

4.1.2.1. Case Reports of Acute, High-dose Exposures 1998 Numerous case reports have been published that des cribe health effects resulting from 1999 acute exposure to dichloromethane. Most of the rep orts describe health effects resulting from 2000 inhalation of dichloromethane or dermal contact, bu t a few involve ingestion. The COHb levels 2001 in some of these cases were relatively low (7.5–13% ), so the initial toxic effects of acute 2002 dichloromethane exposure appear to be due to its an esthetic properties as opposed to metabolic 2003

conversion of dichloromethane to CO. 2004 Bakinson and Jones (1985) reported on a series of 33 cases of acute inhalation exposures 2005 to dichloromethane that occurred in the workplace o ver the period 1961–1980. Thirteen had lost 2006 consciousness, and one of the workers died. Ninete en cases reported general neurological 2007 effects, 13 reported gastrointestinal symptoms, 4 r eported respiratory symptoms, and 1 reported 2008 hepatic symptoms. Of the 19 with general neurologi cal symptoms, all reported headache, and 2009 dizziness was reported by 11 workers. Five workers reported one of the following symptoms: 2010 drunkenness, confusion, lack of coordination, or pa resthesia. 2011 Rioux and Myers (1988) summarized the health effec ts reported for 26 cases of 2012 dichloromethane poisoning published in the literatu re between 1936 and 1986. Three cases 2013 resulted from abuse-related exposures, 2 from chron ic exposures, and 21 from acute exposures. 2014

The most common effects involved the central nervou s system (CNS) (unconsciousness, 2015 drowsiness, headache, and behavioral symptoms), pul monary edema and dyspnea, and 2016 dermatologic symptoms. Even severe symptoms could be reversed, but four deaths occurred. 201744

More than 10 other case reports of fatalities or p oisonings have been published since the 2018 summaries by Rioux and Myers (1988) and Bakinson an d Jones (1985), and many of these 2019 incidents involve inadequately ventilated occupatio nal settings (Jacubovich et al., 2005; Raphael 2020 et al., 2002; Fechner et al., 2001; Zarrabeitia et al., 2001; Goulle et al., 1999; Mahmud and 2021 Kales, 1999; Kim et al., 1996; Tay et al., 1995; Ma nno et al., 1992; Leikin et al., 1990; 2022 Shusterman et al., 1990). CNS depression and resul ting narcosis, respiratory failure, and heart 2023 failure are common features of these reports. In a survey of workers in furniture stripping shops, 2024 10 of the 21 workers stated that they sometimes exp erienced dizziness, nausea, or headache 2025 during furniture stripping operations (Hall and Rum ack, 1990). 2026 Chang et al. (1999) reported details of six patien ts who had ingested dichloromethane 2027 (four in a suicide attempt and two from accidental ingestion during a state of intoxication). The 2028 estimated amounts ingested were 350 mL or less. CO Hb levels, which were measured in only 2029 two of the cases, were 8.4 and 35% (with the latter being seen in a fatal case). As in exposures 2030 resulting from inhalation, the most common symptoms involved CNS depression, ranging from 2031 somnolence and weakness to deep coma. Tachypnea (n = 6) and corrosive gastrointestinal tract 2032 injury (n = 3) were also reported. Hepatic and ren al failure and pancreatitis were found in the 2033

two most severe cases. 2034

2035

4.1.2.2. Controlled Experiments Examining Acute Effects 2036 Several controlled experiments were conducted in t he 1970s, examining 2037 neurophysiological effects and levels of COHb resul ting from short-term (1–4 hours) exposures 2038 to dichloromethane at levels up to 1,000 ppm or lon ger-term exposures at levels up to 500 ppm. 2039

The 8-hour threshold limit value before 1975 was 50 0 ppm (National Institute of Occupational 2040 Safety and Health [NIOSH], 1986). These studies ar e described below. With the exception of 2041 Putz et al. (1979), there is no description in the published reports of the informed consent and 2042 other human subjects research ethics procedures und ertaken in these studies, but there is no 2043

evidence that the conduct of the research was funda mentally unethical or significantly deficient 2044 relative to the ethical standards prevailing at the time the research was conducted. 2045 In 1972, Stewart et al. (1972a, b) reported result s from four experiments that were 2046

initiated because of the chance observation of an e levation in COHb saturation levels in an 2047 individual (one of the investigators) the morning a fter he had spent 2 hours working with varnish 2048 remover. Participants were medical students and fa culty (including at least one of the 2049 coauthors). A total of 11 healthy nonsmoking volun teers were placed in an exposure chamber 2050 with mean concentrations of dichloromethane ranging from 213 to 986 ppm for 1 or 2 hours. 2051 These experiments indicated that dichloromethane ex posure at these levels resulted in COHb 2052 saturation levels that exceeded and were more prolo nged than those seen with threshold limit 2053 value exposures to CO. The exposures also resulted in symptoms of CNS depression indicated 2054 by visual evoked response changes and reports of li ght-headedness. Although return of COHb 205545

levels to background levels could take >24 hours, a ll of the other symptoms were reversible 2056 within a few hours after exposure ceased. 2057 Winneke (1974) measured auditory vigilance, visual flicker fusion frequency, and 2058 14 psychomotor tasks in a total of 38 women exposed to dichloromethane levels of 300–800 ppm 2059 for 4 hours in an exposure chamber. A comparison g roup (nine females, nine males) exposed to 2060 100 ppm CO for 5 hours was also included. Exposure to 800 ppm dichloromethane resulted in a 2061 statistically significant decrease in the performan ce of 10 of the 14 psychomotor tasks. In tests 2062 of auditory vigilance and visual flicker fusion, de pressed response was seen at 300 ppm and was 2063 further depressed at 800 ppm. These effects were n ot seen with CO exposure. 2064 Forster et al. (1974) exposed four healthy young m en to dichloromethane levels ranging 2065 from 0 to 500 ppm for 7.5 hours/day for a total of 26 days over a 6-week period to investigate 2066 alterations in hemoglobin affinity for oxygen and a ltered pulmonary function. While no changes 2067 were observed in pulmonary function, hemoglobin aff inity for oxygen was increased with no 2068 indication of adaptation to restore this affinity f or oxygen to normal. 2069 Putz et al. (1979) examined the behavioral effects seen after exposure to dichloromethane 2070 and to CO. Twelve healthy volunteers (six men and six women) each acted as his/her own 2071 control in separate 4-hour exposures to 70 ppm CO a nd 200 ppm dichloromethane. These levels 2072

were chosen so that the COHb level would reach 5% f rom both the CO and dichloromethane 2073 exposures. The experiments were conducted in a dou ble-blind manner so that neither the 2074 investigators nor the participant knew the exposure condition under study at any particular time. 2075 Informed consent was obtained, and the study was re viewed by the NIOSH Human Subject 2076 Review Board. The performance tests were dual task s (an eye-hand coordination task in 2077 conjunction with a tracking task), with five measur es of performance assessed at six time points 2078 over the 4-hour test period and an auditory vigilan ce task. Two levels of difficulty were assessed 2079

for each task to allow assessment of whether the ex posure effect was similar in low and high 2080 difficulty tasks. The tests of eye-hand coordinati on, tracking tasks, and auditory vigilance 2081 revealed significant impairment with both exposures under the more difficult task conditions. 2082 Effects were similar or stronger in magnitude for d ichloromethane compared with CO. 2083

2084

4.1.2.3. Observational Studies Focusing on Clinical Chemistr ies, Clinical Examinations, and 2085 Symptoms 2086 Studies in currently exposed workers 2087 Ott et al. (1983a, c, d) evaluated several paramete rs of hepatic, hematopoietic, and cardiac 2088 function in workers exposure to dichloromethane in a triacetate fiber production plant in Rock 2089 Hill, South Carolina. Two hundred sixty-six Rock H ill workers and a comparison group of 2090 251 workers in an acetate fiber production plant in Narrows, Virginia, were included in the 2091 examination of urinary and blood measures. These g roups included men and women, blacks and 2092 whites, and smokers and nonsmokers. The median 8-h our TWA exposure for dichloromethane 209346

ranged from 60 to 475 ppm in Rock Hill. Acetone at levels up to over 1,000 ppm was present in 2094 both plants, but dichloromethane and acetone exposu res were inversely related. 2095 There were differences in blood collection procedu res between the two plants and in the 2096 age, sex, race, and smoking history distribution of the study groups. The demographic and 2097 smoking differences were accounted for in the analy sis by stratification. Statistically significant 2098

differences were seen between the workers in the tw o plants for COHb, serum alanine 2099 aminotransferase (ALT), total bilirubin, and mean c orpuscular hemoglobin concentration 2100 (MCHC) (although the direction and magnitude of the se differences were not reported and the 2101

authors stated that the difference in serum ALT cou ld be due to the differences in blood 2102 collection procedures, which involved a sitting ver sus recumbent position of the subjects at the 2103 exposed and nonexposed plants, respectively) (Ott e t al., 1983c). Within the Rock Hill plant, 2104 analyses were also conducted to examine association s between dichloromethane exposure and 2105 the clinical parameters within specific race-sex gr oups by using multiple regression to control for 2106 smoking status, age, and time of venipuncture. Pos itive associations were seen with COHb in all 2107 race-sex groups (increases of 0.7–2.1% per 100 ppm increase in dichloromethane) and with total 2108 bilirubin (increases of 0.05–0.08 mg/dL per 100 ppm increase in dichloromethane) in all groups 2109 except nonwhite men (which was a much smaller group , n = 20, than the other groups). Red cell 2110 count, hematocrit, hemoglobin, and aspartate aminot ransferase (AST) were also positively 2111 associated with dichloromethane exposure in white f emales. The increase in total bilirubin level 2112 was not supported by parallel changes in other meas ures of liver function or red blood cell 2113 turnover, suggesting that this measure was not refl ecting liver damage or hemolysis. 2114 The increased red cell count, hemoglobin, and hema tocrit in women exposed to high 2115 levels of dichloromethane (up to 475 ppm, 8-hour TW A) may indicate a compensatory 2116 hematopoietic effect. The fact that these changes were not significant among men may be due to 2117 higher baseline hemoglobin, which was observed when comparisons were made between 2118 nonsmoking men and women. No such difference in th e baseline values was observed among 2119 the smoking men and women, suggesting that the comp ensatory advantage may be lost among 2120

smokers. 2121 Ott et al. (1983e) present results from a further investigation of changes in COHb, 2122 alveolar CO, and oxygen half-saturation pressure in relation to dichloromethane exposure. 2123 Blood samples were collected before and after shift from 136 Rock Hill and 132 Narrows 2124 workers. For the Rock Hill workers, personal monit oring for dichloromethane exposure was 2125 done during the shift. The TWA for dichloromethane ranged from 0–900 ppm, with a bimodal 2126 distribution (peaks around 150 and 500 ppm) resulti ng from the layout of the plant. The blood 2127 samples were used to determine blood COHb and alveo lar CO levels, and the partial oxygen 2128 pressure (P 50 ) (that is, the pressure required to keep 50% of th e blood oxygen-carrying capacity 2129 saturated with oxygen at pH 7.4 and 37°C). Separat e analyses were conducted for smokers and 2130 nonsmokers to account for the smoking-related effec ts on COHb. Linear relationships were seen 213147

between dichloromethane exposure and the before-shi ft COHb and alveolar CO levels, reflecting 2132 residual CO metabolism from the previous day’s expo sure. There were significant quadratic 2133 relationships between dichloromethane exposure and the postshift COHb and alveolar CO levels, 2134 indicating a partial saturation of the enzyme syste m metabolizing dichloromethane. The 2135

P 50 group means were lower among the exposed compared w ith the referents, among smokers 2136 compared with nonsmokers, and among men compared wi th women. Given the relationship 2137

between COHb and P 50 , an expected decrease in P 50 during the shift was observed among the 2138 exposed. 2139 Continuous 24-hour cardiac monitoring was also eva luated in a smaller sample of 2140 24 dichloromethane-exposed workers from the triacet ate fiber production plant in Rock Hill, 2141 South Carolina, and 26 workers from the comparison plant in Narrows, Virginia. This study (Ott 2142 et al., 1983d) was limited to white men ages 35 or more years. Special efforts were made to 2143 recruit men with a history of heart disease, becaus e this group was postulated to be most likely to 2144 demonstrate positive findings. The estimated TWA d ichloromethane exposure ranged from 2145 60 to 475 ppm in the exposed group. The evaluation examined ventricular and supraventricular 2146 ectopic activity and S-T segment depression in the exposed and nonexposed groups. 2147 Comparisons were also made between cardiac performa nce during work hours and nonwork 2148 hours to discern possible short-term effects of rec ent exposure. Comparing the findings for the 2149 24 exposed and 26 referent volunteers indicated no difference in ventricular or supraventricular 2150 ectopic activity or S-T-segment depression. There was no difference comparing work and 2151 nonwork hours among exposed. 2152 Soden et al. (1996) studied all active male worker s exposed to dichloromethane at a 2153 Hoechst Celanese triacetate film production plant i n Belgium. The production process was the 2154 same as the process at the Hoechst Celanese Rock Hi ll plant, except the Belgium plant was 2155 newer with better engineering controls to significa ntly reduce overall levels of the 2156 dichloromethane, acetone, and methanol used in the process. The objectives of the study were to 2157 determine the impact of varying levels of dichlorom ethane exposure on COHb levels, whether 2158 successive days of dichloromethane exposure affecte d the COHb levels, and what impact 2159 smoking had on COHb levels in conjunction with dich loromethane exposure. Workers were 2160 monitored semiannually for COHb at the end of the w ork shift and were personally monitored 2161 for exposure to the three solvents. Smoking status was defined based on a health assessment 2162 questionnaire, with smokers smoking at least one ci garette per day. Among nonsmokers, a dose 2163 response was found among COHb levels and average di chloromethane exposure levels in the 2164 range of 7–90 ppm. The maximum COHb was 4.00% at a n average exposure of 90 ppm 2165 (correlation coefficient = 0.58, p < 0.05). Smokers’ COHb levels were elevated when com pared 2166

with those of nonsmokers with similar dichlorometha ne air levels, but the dose-response 2167

correlation between dichloromethane air levels and COHb levels was weaker and not statistically 2168 significant (correlation coefficient 0.20). The ma ximum COHb level for smokers was 6.35% at 216948

an average dichloromethane air level of 99 ppm. Th e authors concluded that dichloromethane 2170 exposures up to the levels observed did not produce COHb levels that are likely to cause cardiac 2171

symptoms. 2172 Cherry et al. (1983, 1981) reported the results of health evaluations of two studies of 2173 triacetate film production workers. Cherry et al. (1981) recruited 46 of the 76 male workers at a 2174 triacetate film factory, where workers were exposed to dichloromethane and methanol in a ratio 2175 of 9:1 at air levels of dichloromethane ranging fro m 75 to 100 ppm. A small comparison group 2176 (n = 12) of workers at this factory who worked a si milar shift pattern (rapidly rotating shifts) but 2177 who were not exposed to dichloromethane was also in cluded. The men were asked whether they 2178 had ever experienced cardiac symptoms (pain in the arms, chest pain sitting or lying, or chest 2179 pain when walking or hurrying) and were asked about the presence, in the past 12 months, of 2180 neurological disorders (frequent headaches, dizzine ss, loss of balance, difficulty remembering 2181 things, numbness and tingling in the hands or feet) , affective symptoms (irritability, depression, 2182 tiredness), and stomachache (as an indicator of sym ptom overreporting). No difference in 2183 response was found in history of stomachache (repor ted by 15% of exposed workers compared 2184 with 17% nonexposed workers). Six of the exposed a nd none of the unexposed men responded 2185 positively to the cardiac symptoms. The exposed gr oup reported an excess of neurological 2186 symptoms; the number (and proportion) reporting zer o, one, two, and three or more symptoms 2187 were 26 (0.56), 8 (0.17), 9 (0.20), and 3 (0.07), r espectively, in exposed workers compared with 2188 11 (0.92), 1 (0.12), 0 (0.00), and 0 (0.00), respec tively, in controls ( p < 0.02 for chi-square test of 2189 linear trend). With respect to affective symptoms, the number (and proportion) reporting zero, 2190 one, two, and three symptoms were 28 (0.61), 6 (0.1 3), 7 (0.15), and 5 (0.11), respectively, 2191 among the exposed workers, and 9 (0.75), 2 (0.17), 1 (0.08), and 0 (0.0), respectively, among the 2192 unexposed workers. The authors concluded that ther e was no difference between exposed and 2193 nonexposed in reporting of affective symptoms based on a chi-square test of linear trend. There 2194 was no discussion of the statistical power of this test or of tests of the proportion reporting a 2195 specified number of symptoms (which may be a more a ppropriate test given the sample size), but 2196 it is clear that the statistical power of this test was very low. For example, taking the simple 2197 case of the comparison of the proportion reporting two or more symptoms and using the 2198 approximate estimates from this study (25 and 10% i n the exposed and unexposed, respectively), 2199 approximately 75 exposed and 300 unexposed workers would be needed for a power of 0.80 2200 (i.e., an 80% chance of rejecting the null hypothes is when the null hypothesis was false); the 2201 actual power with the sample size of 46 and 12 is l ess that 0.10. 2202 Based on these results, a follow-up study was cond ucted, which included a larger referent 2203 group. This study included the symptom list descri bed in the previous paragraph, a standardized 2204 clinical exam (including an electrocardiograph), an d neurological and psychological tests of 2205 nerve conduction, motor speed and accuracy, intelli gence, reading, and memory (Cherry et al., 2206 1981). Twenty-nine of the original 46 exposed work ers participated in the follow-up. The men 220749

who did not participate in the follow-up were simil ar in age and symptoms to the men who did. 2208 The new referent group was recruited from another p lant with the same owner and a very similar 2209 process but without dichloromethane exposure. One control, age-matched within 3 years, was 2210 selected for each exposed worker. No differences b etween the groups were found in the clinical 2211 exam, electrocardiogram, or nerve conduction tests. A statistically significant ( p < 0.05) deficit 2212 among the exposed workers was found for coarse moto r speed. On two tests of overall 2213 intelligence, the exposed group did significantly b etter than the referent, but, on a reading ability 2214 test designed to assess premorbid educational level , scores for the exposed group were slightly 2215 lower than for the referent group. (Only one of the se three differences, the trail making 2216 intelligence test, was statistically significant.) With respect to the report of neurological 2217 symptoms in the past year, the number (and proporti on) reporting zero, one, two, and three 2218 symptoms were 17 (0.59), 4 (0.14), 6 (0.21), and 2 (0.07), respectively, among the exposed 2219 workers, and 21 (0.72), 6 (0.21), 0 (0.0), and 2 (0 .07), respectively, among the unexposed 2220 workers, with a test of linear trend that was not s tatistically significant. The authors interpret th e 2221 results as indicating that the differences in neuro logical symptoms seen in the initial study were 2222 due to chance and that, taken as a whole, the expos ed workers had no detrimental effect 2223 attributable to dichloromethane exposure. Again, t he limitations of the statistical power of the 2224 analysis and alternative interpretations that might have resulted from approaches taken to 2225 improve the power were not discussed. These approa ches include combining the unexposed 2226 groups from the two analyses, using the full sample of the exposed group instead of the subset of 2227 29 who completed the clinical exam, or using a diff erent test (i.e., of a proportion rather than a 2228 linear trend), 2229 Cherry et al. (1983) compared dichloromethane-expo sed workers at an acetate film 2230 factory to nonexposed workers (from the same plant but from areas without solvent contact or 2231 from another film production factory in which solve nts were not used). The 56 exposed and 2232 36 unexposed workers were matched to within 3 years of age. Both factories were on rapid 2233 rotating shifts. Exposure to dichloromethane range d from 28–173 ppm, using individual air 2234 sampling pumps. Blood samples were taken to monito r dichloromethane levels at the beginning 2235 and end of the shift. Study participants were aske d to rate sleepiness, physical and mental 2236 tiredness, and general health on visual analog scal es with the extreme responses at either end. 2237 Participants were also given a digit symbol substit ution test and a test of simple reaction time. 2238 No differences were seen between exposed and unexpo sed groups at the beginning of the shift on 2239 the four visual analogue scales, but the exposed de teriorated more on each of the scales than did 2240

the controls. This difference in deterioration was statistically significant ( p ≤ 0.05) during the 2241 morning shift but was not statistically significant during the afternoon or night shifts. A 2242 significant correlation was shown between change in mood over the course of the shift and level 2243

of dichloromethane in the blood. No difference was seen between the exposed and referents on 2244 the tests of reaction time or digit substitution. However, among the exposed, deterioration in the 224550

digit substitution tests at the end of the shift wa s significantly related to blood dichloromethane 2246

levels. 2247 Anundi et al. (1993) studied 12 men who worked in a graffiti-removing company. Each 2248 worker filled out a questionnaire about previous oc cupational and non occupational exposure to 2249 solvents and use of protective equipment. Half-day breathing zone samples were taken for each 2250 of the 12 workers, and 15-minute samples were also taken for 10 workers. On the day the air 2251 sampling was done, a structured interview pertainin g to recent diseases or symptoms related to 2252 allergies, asthma, diseases of the skin, respirator y organs, gastrointestinal tract, urinary organs, 2253 neurological trauma and disease, and neuropsychiatr ic symptoms was conducted by a physician, 2254 and blood and urine samples were collected. The re sults were compared with those of 233 men 2255 from the area population. The 12 men (mean age 23 years) had worked between 3 months and 2256 4.5 years cleaning graffiti from underground statio ns. No respiratory protection was used, and 2257 the leather gloves were frequently soaked with solv ent. While mixed solvent was used to do the 2258 cleaning, dichloromethane was the predominant compo nent, as confirmed by the air samples. 2259 The geometric mean of the TWA calculated from the h alf-day samples was 127 mg/m 3 (range 2260 18–1,188 mg/m 3 ), with half of the samples exceeding the Swedish p ermissible exposure limit of 2261 120 mg/m 3 . The geometric mean of the 15-minute samples was 400 mg/m 3 (range 6– 2262 5,315 mg/m 3 ), with most samples exceeding the Swedish short-ti me exposure limit of 2263 300 mg/m 3 . Two workers had clinical laboratory data outside the normal range (urinary α 1 - or 2264 β 2 -microglobulin, serum ALT, γ -glutamyl transpeptidase), which could indicate pos sible kidney 2265 and liver damage. The authors stated that in both cases factors other than the solvent exposure 2266 (i.e., urinary tract medical condition preceding em ployment, history of renal stones) could have 2267 influenced these laboratory results. The prevalenc e of irritation of the eyes and upper respiratory 2268 tract (blocked nose and nasal catarrh) was much hig her in the graffiti-cleaning workers compared 2269 with the referent group (e.g., >70% of the workers compared with 18% of the comparison group 2270 reported a blocked nose; ~50% of workers and 15% of the comparison group reported eye 2271 irritation), but there were no or much smaller diff erences in abnormal tiredness, headache, 2272 nausea, or irritative cough. No acute effects on t he CNS were noted. 2273

2274

Studies in retired workers 2275 Lash et al. (1991) examined the hypothesis that lo ng-term exposure to dichloromethane 2276 produces lasting CNS effects as measured by long-te rm impairment on memory and attention 2277 centers. Retired aircraft maintenance workers empl oyed in at least 1 of 14 targeted jobs with 2278 dichloromethane exposure for 6 or more years betwee n 1970 and 1984 were compared to a like 2279 group of workers without dichloromethane exposure. The unexposed workers were also retired 2280

aircraft mechanics at the same base and held one of 10 jobs in the jet shop where little solvent 2281 was used. The exposed group made up of painters an d mechanics in the overhaul department 2282 was chosen to maximize the exposure contrast yet mi nimize differences in potential confounders 228351

between exposed and nonexposed. Exposures were typ ically within state and federal guidelines 2284 for dichloromethane exposure. From 1974 to 1986, w hen 155 measurements for 2285 dichloromethane exposure were made, mean breathing zone TWAs ranged from 82 to 236 ppm 2286 and averaged 225 ppm for painters and 100 ppm for m echanics. 2287 Data collection occurred in three phases, with an initial questionnaire to all retired 2288 members of the airline mechanics union to identify eligible workers, followed by a telephone 2289 survey to collect medical, demographic, and general employment criteria. Subjects who 2290 qualified were then recruited to participate in the medical evaluation. Sixty percent of the 2291 1,758 retirees responded to the questionnaire and 2 59 of these retirees met the eligibility criteria. 2292 Ninety-one men qualified for the medical evaluation based on the telephone survey; 25 retirees 2293 exposed to solvents, and 21 unexposed retirees part icipated in the evaluation. All were men 2294 between the ages of 55 and 75, without a history of alcoholism or any neurological disorder. The 2295 25 exposed participants worked an average of 11.6 y ears in dichloromethane-exposed jobs 2296 during the target period and 23.8 years in the indu stry. 2297 The medical evaluation included a questionnaire ab out the occurrence of 33 different 2298 symptoms in the past year, physiological measuremen t of odor and color vision senses, auditory 2299 response potential, hand grip strength, and measure s of reaction time (simple, choice, and 2300 complex), short-term visual memory and visual reten tion, attention, and spatial ability. The only 2301 large differences (i.e., effect size, or mean diffe rence between groups divided by the SD of the 2302 outcome measure, of 0.4 or greater) between the two groups were a higher score on verbal 2303 memory tasks (effect size approximately 0.45, p = 0.11) and lower score on attention tasks 2304 (effect size approximately –0.55, p = 0.08) and complex reaction time (effect size app roximately 2305 –0.40, p = 0.18) in the exposed compared with the control g roup. (Although not noted by the 2306 authors, the power to detect a statistically signif icant difference between the groups, given this 2307 sample size, was low [i.e., approximately 0.30 for an effect size of 0.40, using a two-tailed alpha 2308 of 0.05]) (Cohen, 1987). The authors investigated the possibility of response bias, given the low 2309 initial response to the mailed questionnaire recrui ting retirees and the small number of workers 2310 from the entire pool of eligible participants who a ctually participated in the medical evaluation. 2311 Attempts were made to contact 30% of the questionna ire nonrespondents, with 46% contacted 2312 and 31% completing the telephone interview. The on ly difference found between those who 2313 responded to the mailed questionnaire and those who did not was a higher percentage of 2314 diagnosed heart disease among the nonrespondents wh o were 2.5 years older and had been 2315 retired 1.7 more years than the respondents. Those who were eligible but did not participate in 2316 the medical evaluation were similar to the exam par ticipants on all characteristics included in the 2317 interview. The only difference was a higher preval ence of gout among the unexposed who did 2318 not participate compared to the unexposed who did p articipate. 2319

2320

4.1.2.4. Observational Studies Using Workplace Medical Progr am Data 232152

Kolodner et al. (1990) investigated the effect of occupational exposure to 2322

dichloromethane on six health outcomes identified i n the literature or based on biological 2323 plausibility. Participants in the study were male workers at least 19 years old at two 2324 General Electric plastic polymer plants where dichl oromethane was one of the chemicals used. 2325 Four dichloromethane exposure categories were estab lished based on full-shift personal air 2326 monitoring data (8-hour TWA) collected in 1979–1985 , job titles, and industrial hygienists’ 2327 knowledge of plant operations. The mean 8-hour TWA and number of workers in each of the 2328 four exposure groups were 49.0 ppm for the 19 worke rs in the highest, 10.9 ppm for the 2329 49 workers in the intermediate, 3.3 ppm for the 56 workers in the low, and <1.0 ppm for the 2330 772 workers in the minimal/no exposure group. 2331

Data from 1984 annual medical exams and 1985 absen ce data from payroll records were 2332 evaluated for possible health effects resulting fro m occupational exposure to dichloromethane. 2333 A high percentage of workers participated in the an nual medical exams, with only 5 of the 2334 896 eligible for inclusion in the study refusing th e exam completely in 1984. Six hypotheses 2335 were specifically tested regarding dichloromethane exposure in relation to different health 2336 outcomes: absence due to illness, hepatotoxicity (m anifested by nausea, weakness and fatigue, 2337 palpable liver, abdominal tenderness, jaundice, hep atomegaly, abnormal serum γ -glutamyl 2338 transferase, ALT, AST, or bilirubin), diabetes mell itus (manifested by weight loss, weakness and 2339 fatigue, polydypsia, polyuria, impaired vision, exc essive weight loss, elevated fasting blood 2340 sugar, and abnormal urinary glucose or urinary acet one), CNS toxicity (manifested by headache, 2341 lightheadedness, dizziness and vertigo, ataxia, wea kness and fatigue, and abnormalities detected 2342 in the central motor, central sensory, cranial nerv e, gait, neurocoordination, or Bibinski reflex 2343 examinations), cardiovascular abnormalities (manife sted by fatigue, dyspnea, chest pain with 2344 exertion, palpitations, or abnormalities detected i n the point maximum impulse exam, blood 2345 pressure measurements, or electrocardiogram), and n eoplastic breast changes (154 women were 2346 included in this portion of the study—manifested by painful breast, breast swelling, lump, nipple 2347 discharge, or abnormalities detected in the breast examination). 2348 Workers were placed in exposure categories based o n their current jobs. In addition, 2349 exposure to high noise levels occurred in both plan ts, and workers in each plant had exposure to 2350 another chemical, either phenol or phosgene. The a uthors noted that workers tended to move 2351 from entry-level jobs with high dichloromethane exp osure to supervisory jobs with lower 2352 dichloromethane exposure, based on the seniority sy stem in place at both plants. Thus, current 2353 exposure levels reported did not necessarily reflec t cumulative exposure. Because of the way the 2354 seniority system moved workers through jobs and the fact that workers were assigned to 2355 dichloromethane exposure categories based on their current job, age was inversely related to 2356 exposure and was controlled in the analysis of some of the continuous variables using analysis of 2357 covariance. Age adjustment was not employed in the analysis of dichotomous variables. The 235853

mean age was 35.3, 39.7, 37.1 and 29.5 years in the minimal/no, low, medium 3 , and high 2359 exposure groups, respectively. The small number of workers in the exposed groups limited the 2360 ability to evaluate the effects of dichloromethane exposure on health outcomes related to age, 2361 since age had to be adjusted in these analyses. Th e racial distribution did not differ among the 2362

exposure groups. 2363 The authors indicated that all the hypotheses were accepted with the exception of CNS 2364 symptoms. However, it should be noted that the sma ll size and younger age distribution in the 2365 high exposure group and the lack of adjustment for age in most of the analyses make it difficult 2366 to interpret the statistical testing that was perfo rmed. Data pertaining to neurological, hepatic, 2367 and cardiac function are shown in Table 4-1. Among the six neurological symptoms evaluated, a 2368 statistically significant positive exposure-effect relationship between dizziness/vertigo and 2369 dichloromethane exposure was identified. This tren d was driven most strongly by the low 2370 frequency of this reported symptom in the minimal/n o exposure group (1.2%), but there was no 2371 linear trend across the higher levels of exposure ( 7.5, 2.1, and 5.3% in the low, medium, and high 2372 exposure groups, respectively). 2373

2374

3 The “medium” exposure group is also referred to as the “intermediate” exposure group in Kolodner et a l. (1990).54 2375 Table 4-1. Percentage of male General Electric pla stic polymer workers

reporting neurologic symptoms or displaying abnorma l values in measures of neurological function, hepatic function, and car diac function

Exposure Group a Minimal/no (n = 772) Low

(n = 56) Medium

(n = 49) High

(n = 19) Neurological Headache 8.7 7.5 10.4 5.3 Lightheadedness 2.9 3.8 4.2 5.3 Dizziness/vertigo 1.2 7.5 2.1 5.3 Ataxia 0.0 1.9 0.0 0.0 Babinsky 0.0 0.0 0.0 0.0 Gait 0.0 0.0 0.0 0.0 Faintness/syncope b 0.1 0.0 2.1 0.0 Seizures b 0.4 0.0 2.1 0.0 Paresis/paralysis b 0.7 0.0 0.0 0.0 Parasthesis b 4.0 7.5 14.6 0.0 Head trauma/concussion b 0.8 1.9 0.0 0.0 Peripheral motor exam b,c 0.5 0.0 0.0 0.0 Peripheral sensory exam b,c 1.1 2.4 5.1 0.0 Rhomberg exam b,c 0.0 0.0 2.6 0.0 Hepatic Serum gamma glutamyl transferase 8.0 16.1 12.2 5. 3 Serum total bilirubin 3.0 1.8 2.0 10.0 Serum AST 1.8 3.6 4.1 0.0 Serum ALT 9.1 10.7 8.2 5.3 Cardiac d Palpitations: percent abnormal 1.2 9.1 2.1 0.0 Electrocardiogram borderline/abnormal 18.5 16.7 19.1 8.3 bradycardia/tachycardia abnormalities b 20.2 16.7 25.5 0.0 general rhythm abnormalities 12.0 11.1 17.0 8.3 atrial, atrioventricular, or sinus abnormalities 0. 8 0.0 0.0 0.0 bundle blocks or ventricular abnormalities 3.9 5.6 10.6 8.3 axis deviations 2.6 1.9 2.1 8.3 wave abnormalities 4.0 3.7 10.6 0.0 hypertrophy 3.8 3.7 6.4 0.0 evidence of infarction 2.3 5.6 2.1 0.0

a Mean 8-hour TWA exposure was <1.0, 3.3, 0.9, and 49 .0 ppm in the minimal/no, low, medium, and high gro ups, respectively; mean age 35.3, 39.7, 37.1, 29.5 years in the minimal/no, low, medium, and high groups, r espectively.

b The authors considered these to be screening variab les rather than hypothesis-testing variables.

c n = 629, 42, 39, and 14 in the minimal/no, low, med ium, and high groups, respectively

d For all cardiac outcomes except bradycardia/tachyca rdia, n = 728, 54, 47, and 12 in the minimal/no, lo w, medium, and high groups, respectively. For bradycardia/tac hycardia, n = 727 in the minimal/no group.

Source: Kolodner et al. (1990). 2376 2377 Soden (1993) compared health-monitoring data from dichloromethane-exposed workers 2378 in the Rock Hill triacetate fiber production plant to workers from another plant making polyester 2379 fibers owned by the same company in the same geogra phic area. Exposed and control workers 2380 were chosen from among workers who had worked at le ast 10 years in their respective areas and 2381 who participated in the company’s health-monitoring program between 1984 and 1986 and were 238255

still employed on December 31, 1986. Controls were matched by race, age, and gender to each 2383 Rock Hill worker for a sample size of 150 and 260 i n the exposed and control groups, 2384 respectively. (The aim of the study had been 1:2 ma tching.) The 8-hour TWAs among the Rock 2385 Hill workers were those reported by Lanes et al. (1 990), namely 475 ppm for dichloromethane, 2386 900 ppm for acetone, and 100 ppm for methanol. Non e of these exposures occurred at the 2387 polyester plant. There was a 90% participation rat e in the health-monitoring program. Six 2388 questions in the health history portion of the heal th-monitoring program concerned cardiac and 2389 neurological symptoms (chest discomfort with exerci se; racing, skipping, or irregular heartbeat; 2390 recurring severe headaches; numbness/tingling in ha nds or feet; loss of memory; dizziness). Part 2391 of this program included blood samples used for sta ndard clinical hepatic and hematologic 2392 parameters: serum ALT, AST, total bilirubin, and he matocrit. The clinical measures were 2393 available for 90 (60%) of the exposed and 120 (46%) of the control group because some 2394 participants declined this part of the health-monit oring program because similar tests had been 2395 part of recent personal medical care. 2396 There was little difference in the frequency of re ported symptoms between exposed 2397 workers and controls: chest discomfort reported by 2.0% of exposed and 4.0% of the controls, 2398 irregular heartbeat reported by 5.5% of exposed and 6.0% of the controls, recurring severe 2399 headaches reported by 3.5% of exposed and 5.5% of t he controls, numbness/tingling in hands 2400 and feet reported by 6.4% of exposed and 8.1% of th e controls, loss of memory reported by 1.3% 2401 of exposed and 0.4% of controls, and dizziness repo rted by 2.7% of exposed and 4.8% of 2402 controls (Soden, 1993). The levels of the blood va lues were similar in the exposed and control 2403 groups, except for a 3.1 IU/L decrease in serum AST activity ( p = 0.06). The authors concluded 2404 that this difference was not clinically significant but did not discuss the potential bias introduced 2405 by the selective participation in this part of the study. 2406

2407

4.1.2.5. Studies of Ischemic Heart Disease Mortality Risk 2408

Several studies have examined the relation between dichloromethane exposure and risk 2409 of cardiovascular-related mortality. The methodolo gical details of these studies are described in 2410 section 4.1.3.2.). No evidence of increased risk o f ischemic heart disease mortality was seen in 2411 two triacetate film production cohort studies (Hear ne and Pifer, 1999; Tomenson et al., 1997) or 2412 in two triacetate fiber production cohort studies ( Gibbs et al., 1996; Lanes et al., 1993). 2413

Information on this outcome was not included in the dichloromethane analysis of civilian Air 2414 Force base workers (Blair et al., 1998). The stand ardized mortality ratios (SMRs) for ischemic 2415 heart disease mortality were <1.0 in all of the coh orts and dose groups examined (Table 4-2). 2416 The “healthy worker effect” may have contributed to these observations. There are no case2417

control studies of ischemic heart disease and dichl oromethane exposure. 2418 241956

Table 4-2. Ischemic heart disease mortality risk i n four cohorts of dichloromethane-exposed workers

Obs a Exp b SMR 95% CI c Triacetate film production Hearne and Pifer (1999) Cohort 1 (men) 117 136.7 0. 86 0.71–1.03 Cohort 2 (men) 122 143.3 0.85 0.71–1.02 Tomenson et al. (1997) Men 114 123.9 0.92 0.76–1.10 Triacetate fiber production Lanes et al. (1993) Men and women 43 47.8 0.90 65–121

Men 50–100 ppm 96 100.1 0.96 0.78–1.2 350–700 ppm 98 106.8 0.92 0.75–1.1

Women 50–100 ppm 32 45.8 0.70 0.48–0.99 Gibbs et al. (1996)

350–700 ppm 0 3.4 – 0.0–1.1

a Obs = number of observed deaths.

b Exp = number of expected deaths.

c CI = confidence interval. 2420 2421 4.1.2.6. Studies of Suicide Risk 2422 Suicide risk is not an outcome that was a primary hypothesis or motivation of the cohort 2423 studies but may be relevant given the potential neu ropsychological effects of dichloromethane, 2424 as evidenced from studies of acute and chronic expo sure scenarios described previously. In a 2425 triacetate film production cohort in Rochester, New York, Hearne and Pifer (1999) reported 2426 14 observed deaths from suicide compared with 7.8 e xpected, for an SMR of 1.8 (95% 2427 confidence interval [CI] 0.98–3.0) (Table 4-3). Th is cohort (“Cohort 1”) consisted of 1,311 men 2428 who were first employed between 1946 and 1970 and w ere followed through 1994. Similar 2429 results were seen in a different, but somewhat over lapping, cohort in this study (“Cohort 2”) of 2430 1,013 men employed between 1964 and 1970 and follow ed through 1994 (see section 4.1.3.3.1). 2431 There was also evidence of increasing suicide risk with dichloromethane exposure, particularly 2432 in the highest exposure group, in the study of tria cetate fiber production workers in Maryland 2433 (Gibbs, 1992). The triacetate fiber production coh ort study in Rock Hill, South Carolina, has 2434 published what appears to be erroneous information about suicide risk. In the 1993 paper (Lanes 2435 et al., 1993), 4 observed and 5.21 expected cases w ere reported (SMR 0.77), but the SMR that 2436 was reported with these data was 1.19 (95% CI 0.39, 2.8). This ratio would correspond to 2437 6 observed and around 5.2 expected cases. Informat ion on suicide was not included in the other 2438 film and fiber cohort studies (Tomenson et al., 199 7) or in the analysis of civilian Air Force base 2439 workers (Blair et al., 1998). There are no case-co ntrol studies of suicide risk and 2440 dichloromethane exposure. 2441

244257

Table 4-3. Suicide risk in two cohorts of dichloro methane-exposed workers

Obs a Exp b SMR 95% CI Triacetate film production Hearne and Pifer (1999) Cohort 1 14 7.8 1.8 0.98–3. 0

Cohort 2 9 5.1 1.8 0.81–3.4 Triacetate fiber production c 50–100 ppm 8 6.4 1.3 0.54–2.5 Gibbs (1992) 350–700 ppm 8 4.4 1.8 0.78–3.6

a Obs = number of observed deaths.

b Exp = number of expected deaths.

c One additional study provided data on suicide risk, but some kind of error seems to be present: 4 obse rved and 5.21 expected cases were reported in Lanes et al. ( 1993), which would be an SMR of 0.77, but the SMR t hat was reported with these data was 1.19 (95% CI 0.39, 2.8 ). This ratio would correspond to 6 observed and a round 5.2 expected cases. 2443 2444 4.1.2.7 Studies of Infectious Disease Risk 2445 There is limited information pertaining to infectio us disease risk in relation to 2446 dichloromethane exposure. Only one of the cohort s tudies (Hearne and Pifer, 1999) reported 2447 data for the broad category of infectious and paras itic disease mortality. In Cohort 1 of this 2448 analysis, there were no observed deaths in this cat egory (5.6 expected), and in Cohort 2 there 2449 were 3 observed and 4.7 expected, for an SMR of 0.6 4. The detailed report by Gibbs (1992) of 2450 the cellulose triacetate fiber production cohorts i n Maryland (Gibbs et al., 1996) also contained 2451 information on the facility in South Carolina that was the site of the report by Lanes et al. (1993, 2452 1990). Slightly elevated risks of mortality due to influenza and pneumonia were seen among the 2453 male workers in the high exposure group in Maryland (7 observed, 5.62 expected, SMR 1.25) 2454 and in South Carolina (3 observed, 1.33 expected, S MR 2.26). Among females, there were few 2455 observed or expected cases (in Maryland, 1 observed , 0.23 expected, SMR 4.36; in South 2456 Carolina, 0 observed and 0.74 expected). 2457

2458

4.1.2.8. Studies of Reproductive Outcomes 2459 Pregnancy outcomes in women exposed to dichloromet hane have been investigated in 2460 two studies. Taskinen et al. (1986) studied sponta neous abortions among women employed in 2461 eight pharmaceutical factories between 1973 and 198 0. Data on pregnancy outcomes were 2462 collected from a national hospital and clinic disch arge registry in Finland from 1973 to 1981 by 2463 matching the worker rosters to the registry. Expos ure to dichloromethane was one of eight 2464

solvents or classes of solvents included in the stu dy. The study consisted of two parts. The first 2465 investigated the rate of spontaneous abortions (num ber of spontaneous abortions divided by the 2466 sum of spontaneous abortions and births) during, be fore, or after employment in the 2467 pharmaceutical industry. One hundred and forty-two spontaneous abortions and 1,179 births 2468 were identified among the female workers at the eig ht plants. Employment hire and termination 246958

dates were obtained from plant records. The sponta neous abortion rate was 10.9% during 2470 employment compared with 10.6% before and after emp loyment. These results compared to a 2471 rate of 8.5% in the general population in the geogr aphic area where the factories were located. 2472 The rate of spontaneous abortions among workers dec lined over the period of the study, with a 3- 2473 year moving average of 15% at the beginning declini ng to 9.5% at the end of the study. Over the 2474 same period, the industrial hygiene allegedly impro ved in the plants. Ten congenital 2475 malformations of different types were identified am ong the women (five among those who were 2476 employed in the pharmaceutical industry during the pregnancy and five among those whose 2477 pregnancies occurred before or after this employmen t). 2478 The second part of the study by Taskinen et al. (1 986) was a case-control study of the risk 2479 of spontaneous abortions in relation to workplace e xposures during pregnancy. The source 2480 population consisted of women who were employed in one of the eight Finnish pharmaceutical 2481 factories during at least 1 week of the first trime ster of pregnancy during the study period. Cases 2482 (n = 44) were selected from this population based o n hospital or clinic records indicating a 2483 spontaneous abortion, and 130 controls (women who h ad given birth) were age-matched 2484 (3:1 matching, age within 2.5 years) to each case. Occupational exposure data were obtained by 2485 questionnaires completed by the plant physician or the nursing staff, blinded to the case status of 2486 the study member, in consultation with labor protec tion chiefs and department foremen. The 2487 questionnaire requested information about job histo ry and job tasks, exposure to eight specific 2488 solvents or classes of solvents (aliphatic solvents , alicyclic solvents, toluene, xylene, benzene, 2489 chloroform, dichloromethane, and other solvents), a ntineoplastic agents, carcinogens, hormones, 2490 antibiotics, heavy lifting, known chronic diseases, acute diseases during pregnancy, smoking 2491 status, and previous pregnancies. Exposure frequen cy to each solvent was based on the 2492 cumulative weighted sum of the number of days/week the woman was exposed to the solvent. 2493 While overall response to the questionnaire was 93% , less than half the questionnaires contained 2494 information about smoking or previous pregnancies, precluding inclusion of these variables in 2495 the analysis. The distribution of broad categories of occupations (i.e., pharmaceutical workers 2496 and packers, laboratory assistants) was similar in both groups. However, exposure to each of the 2497 solvents was higher in the cases compared with cont rols, and the results for dichloromethane 2498 were relatively strong. For dichloromethane, the p revalence of exposure was 28.9 and 14.3% in 2499 cases and controls, respectively, resulting in an o dds ratio (OR) of 2.3 (95% CI, 1.0, 5.7). There 2500 was also evidence of an increasing risk with higher exposure frequency, with an odds ratio (OR) 2501 of 2.0 (95% CI 0.6, 6.6) with exposures of less tha n once a week and 2.8 (95% CI 0.8, 9.5) with 2502 exposures of once a week or more. An association w as also seen with exposure to four or more 2503 solvents (OR 3.4, 95% CI 1.0, 12.5), and weaker ass ociations were seen with other specific 2504 solvents (e.g., chloroform, toluene). 2505 Bell et al. (1991) investigated the relation betwe en birth weight and maternal exposure to 2506 airborne dichloromethane as a result of living arou nd the triacetate film facility in Rochester, 250759

New York. For this population-based cross-sectiona l study, birth certificates were obtained for 2508 all births in 1976–1987 in Monroe County, where the triacetate film facility is located. Multiple 2509 births and births of infants weighing <750 grams we re excluded. Data abstracted from the 2510 certificate included date of birth, census tract of residence, age, race, educational level of the 2511 mother and father, sex, gestational age, multiple b irths, month of the pregnancy that prenatal care 2512 began, total previous births, total previous live b irths, and conditions present during the 2513 pregnancy. An air dispersion modeling system for 2 50 air emissions, including 2514 dichloromethane, predicting average annual ground l evel concentrations in the surrounding 2515 community, was used to assign dichloromethane expos ure levels to each birth mother. One of 2516 four levels of exposure was assigned to each census tract based on the isopleth of exposure in 2517 which more than half of the census tract population resided. Because of the few births among 2518 nonwhites that occurred in areas of higher exposure , the study was restricted to whites 2519 (n = 91,302). The number of births that occurred i n each of the four exposure levels was 2520 n = 1,085 in the high-exposure group (50 μ g/m 3 [0.014 ppm]), n = 1,795 in the moderate2521 exposure group (25 μ g/m 3 [0.007 ppm]), n = 6,044 in the low-exposure group (10 μ g/m 3 2522 [0.003 ppm]), and n = 82,076 in the no-exposure gro up. At the levels of dichloromethane 2523 exposure in this population no significant adverse effect on birth weight was found. There was 2524 an 18.7 g decrease in birthweight (95% CI –51.6, 14 .2) in the high- compared with the no2525 exposure group, adjusting for maternal age, materna l education, parity, previous pregnancy loss, 2526 late start of prenatal care, sex of the child, and pregnancy complications. No significant 2527 association was found between any combination of ex posure levels and birth weight. There was 2528 no association between exposure group and risk of a low birthweight infant (i.e., <2,500 g, 2529 OR 1.0 [95% CI 0.81, 1.2] in the high- compared wit h the no-exposure group). The authors 2530 point out a number of problems with assignment of d ichloromethane exposure. It is possible that 2531 the dichloromethane exposure was overestimated usin g the model. Comparisons to ambient air 2532 sampling levels collected six times/year resulted i n the dichloromethane exposure derived from 2533 the model being twice as high as the ambient air sa mples. There was also inaccuracy in the 2534 assignment of dichloromethane exposure level to eac h birth because the exposure assignment 2535 was made using the predominant value of the isoplet h for a census tract. 2536 Two studies have investigated the occurrence of ol igospermia among men occupationally 2537 exposed to dichloromethane exposure. Kelly (1988) studied 34 men employed in an automotive 2538 plant as bonders, finishers, and press operators. These men were self-referred to a health center 2539 for a variety of complaints, including neurological symptoms, musculoskeletal symptoms, and 2540 shortness of breath. Twenty-six of the men were bo nders and eight were finishers or press 2541 operators. The job as bonder consisted of dipping hands into an open bucket of dichloromethane 2542 and splashing it onto plastic automobile parts. Th e dichloromethane exposure for bonders 2543 averaged 68 ppm with a range from 3.3 to 154.4 ppm. Eight men, all of whom were bonders, 2544 reported symptoms of testicular and epidydimal tend erness, with confirmation on medical exam. 254560

They ranged in age from 20 to 47 years old and had been bonders for up to 2.9 years. The COHb 2546 levels for the eight workers with genital symptoms ranged from 1.2 to 17.3%, with an average of 2547 6.9% anywhere from 4 to 90 hours postexposure. The COHb levels for the two men who 2548 smoked were among the highest, namely 7.3 and 17.3% . Four of the eight workers agreed to 2549 provide semen samples; their sperm counts were 2–26 × 10 6 /cm 3 . The authors stated that men 2550 with sperm counts as low as 25 × 10 6 /cm 3 may still be fertile, but none of these men had ha d any 2551 children since working with dichloromethane despite not using contraceptives. There was one 2552 miscarriage. All four men reported dipping their h ands into open buckets of dichloromethane 2553 without any protective equipment, and two men repor ted feeling dizzy, giddy, and high at work. 2554 Based on the results of the Kelly (1988) case repo rt, Wells et al. (1989) planned to do a 2555 study of oligospermia among 20 exposed and 20 unexp osed to dichloromethane. The exposed 2556

workers were unvasectomized men who had worked for the 3 months prior to recruitment in 2557 furniture stripping shops. Eleven men were recruit ed from among 14 eligible workers at six 2558 different shops where dichloromethane was utilized. Names of acquaintances of the exposed 2559 were solicited as potential referents. Only one ex posed man provided any names. Therefore, the 2560 study was redirected as a case report on the 11 exp osed men. The mean TWA dichloromethane 2561 exposure was 122 ppm (range 15–366 ppm) with a mean COHb of 5.8% (range 2.2–13.5%). 2562 The mean COHb for smokers, 10.2% (range 8.1–13.5), was higher than for nonsmokers, 3.9% 2563 (range 2.2–5.9), and the nonsmoker levels were high er than the 2% level considered to be the 2564 upper limit of normal in nonsmoking populations. T he mean sperm count was 54 × 10 6 /cm 3 2565 (range 23–128 × 10 6 /cm 3 ) compared to a population value of 47 × 10 6 /cm 3 for the same 2566 geographic based on samples analyzed at the same la boratory. Using the standard definition for 2567 oligospermia of 20 × 10 6 /cm 3 , none of the 11 workers had oligospermia. 2568

2569

4.1.2.9. Summary of Noncancer Studies 2570 The clinical and workplace studies of noncancer he alth effects of dichloromethane 2571 exposure have examined markers of disease and speci fic clinical endpoints relating to cardiac, 2572 neurological disease, hepatic function, and reprodu ctive health. 2573

2574 Cardiac effects 2575

The effect of dichloromethane on the formation of COHb (Stewart et al., 1972b) raised 2576 concerns about potential risk of cardiovascular dam age. To date, there is little evidence of 2577 cardiac damage related to dichloromethane exposure in the cohort studies of dichloromethane2578 exposed workers that examined ischemic heart diseas e mortality risk (Hearne and Pifer, 1999; 2579 Tomenson et al., 1997; Gibbs et al., 1996; Lanes et al., 1993) or in two small cardiac monitoring 2580 studies (Ott et al., 1983d; Cherry et al., 1981). However, limitations in these studies should be 2581 noted, including the healthy worker effect and the absence of data pertaining to workers who 2582 died before the establishment of the analytic cohor t (Gibbs et al., 1996; Gibbs, 1992). 258361

Neurological effects 2584 The acute effects of dichloromethane exposure on n eurological function seen in 2585 numerous case reports have also been established in experimental studies in humans (Putz et al., 2586 1979; Winneke, 1974; Stewart et al., 1972a, b). Re latively less is known about the long-term 2587 effects of chronic exposures in humans. Some data from studies of workers suggest that the 2588 effects of dichloromethane are relatively short-liv ed. For example, in the study by Cherry et al. 2589 (1983) of 56 exposed and 36 unexposed workers, alte rations in mood or in digit substitution test 2590 results were seen during the course of a work shift but were not seen at the beginning of a shift. 2591 No difference in four neurological symptoms was see n in an analysis of exposed workers 2592 (average exposure 475 ppm, ≥ 10-year duration) and an unexposed comparison group by Soden 2593 (1993). Other data suggest an increase in prevalen ce of neurological symptoms among workers 2594 (Cherry et al., 1981) and possible detriments in at tention and reaction time in complex tasks 2595 among retired workers (Lash et al., 1991). These l atter two studies are limited by the small 2596 sample size. Thus Cherry et al. (1981) and Lash et al. (1991) have low power for detecting 2597 statistically significant results and consequently should not be interpreted as definitive analyses 2598 showing no effects. Rather, these analyses provide some evidence of an increased prevalence of 2599 neurological symptoms among workers with average ex posures of 75–100 ppm (Cherry et al., 2600 1981) and long-term effects on specific neurologica l measures (i.e., attention and reaction time) 2601 in workers whose past exposures, at least for part of their work history, were in the 100–200 ppm 2602 range (Lash et al., 1991). The increased risk of s uicide (approximately a twofold increased risk) 2603 seen in two of the worker cohort studies (Hearne an d Pifer, 1999; Gibbs, 1992) is an additional 2604 indication of potential neurological consequences o f dichloromethane exposure. Adequate 2605 studies addressing these specific issues are not av ailable. Thus, given the suggestions from the 2606 currently available studies, the statement that the re are no long-term neurological effects of 2607 chronic exposures to dichloromethane cannot be made with confidence. 2608

2609 Hepatic effects 2610 Three studies provide data pertaining to markers o f hepatic damage (i.e., serum enzymes 2611 and bilirubin levels) (Soden, 1993; Kolodner et al. , 1990; Ott et al., 1983c). Two of these studies 2612 were based in the Rock Hill, South Carolina, cellul ose triacetate fiber plant (Soden, 1993; Ott et 2613 al., 1983c), with the most recent of the studies fo cusing on workers with more than 10 years 2614 duration in a high exposure area (average exposure estimated as 475 ppm). There is some 2615 evidence of increasing levels of serum bilirubin wi th increasing dichloromethane exposure in Ott 2616 et al. (1983c) and Kolodner et al. (1990), but ther e are no consistent patterns with respect to the 2617 other hepatic enzymes examined (serum γ -glutamyl transferase, serum AST, serum ALT). Thes e 2618 studies do not provide clear evidence of hepatic da mage in dichloromethane-exposed workers, to 2619 the extent that this damage could be detected by th ese serologic measures; however, these data 262062

are limited and thus the absence, presence, or exte nt of hepatic damage is not known with 2621 certainty. 2622

2623 Immune effects 2624 Only limited, and somewhat indirect, evidence perta ining to immune-related effects of 2625

dichloromethane in humans is available. No risk wa s seen in the broad category of infectious 2626 and parasite-related mortality reported by Hearne a nd Pifer (1999), but there was some evidence 2627 of an increased risk for influenza and pneumonia-re lated mortality at two cellulose triacetate 2628 fiber production work sites in Maryland and South C arolina (Gibbs, 1992). In the Maryland 2629 facility, an increased risk of cervical cancer was seen among the 938 female workers, with an 2630 SMR of 3.0 (95% CI, 0.96, 6.9) in the 50–100 ppm gr oup and 5.4 (95% CI 0.13, 30.1) in the 2631 350–700 ppm group (Gibbs et al., 1996). Cervical c ancer is viral mediated (human papilloma 2632 virus), and immunosuppression is a risk factor for development of this disease, as seen by the 2633 increased risk in immunocompromised patients and pe ople taking immunosuppressant 2634 medications (Leitao et al., 2008; Ognenovski et al. , 2004). 2635

2636 Reproductive effects 2637 Studies pertaining to various reproductive effects and dichloromethane exposure from 2638 workplace settings (Wells et al., 1989; Kelly, 1988 ; Taskinen et al., 1986) or environmental 2639 settings (Bell et al., 1991) have examined possible associations with spontaneous abortion 2640 (Taskinen et al., 1986), low birth weight (Bell et al., 1991), and oligospermia (Wells et al., 1989; 2641 Kelly, 1988). Of these, the data pertaining to spo ntaneous abortion provide the strongest 2642

evidence of an adverse effect of dichloromethane ex posure, particularly with respect to the case2643 control study in which the strongest association wa s seen specifically with the higher frequency 2644 category of dichloromethane exposure. It is a smal l study (44 cases, 130 controls), however, 2645 with limited quantitative exposure assessment and m ultiple exposures (although the association 2646 seen with dichloromethane was among the highest see n among the solvents) and so cannot be 2647 considered to firmly establish the role of dichloro methane in induction of miscarriage. However, 2648 the high exposure scenario, including the potential for substantial dermal exposure in the study 2649 of Kelly (1988), also suggests the potential for ad verse male reproductive effects. 2650

2651

4.1.3. Cancer Studies 2652

4.1.3.1. Identification and Selection of Studies for Evaluat ion of Cancer Risk 2653 Twelve epidemiologic studies of cancer risk were id entified and included in this 2654 evaluation: four cohorts for which the primary solv ent exposure was to dichloromethane (two in 2655 film production settings and two in cellulose triac etate fiber production), one large cohort of 2656 civilian employees at a military base with exposure s to a variety of solvents but that included an 2657 assessment specifically of dichloromethane exposure , and seven case-control studies of specific 265863

cancers with data on dichloromethane exposure. One additional study (Ott et al., 1985), a cohort 2659 of 1,919 men employed at Dow Chemical facilities, w as identified but was not included in the 2660 summary. The analysis was based on exposure to a c ombined group of chlorinated methanes 2661 (e.g., carbon tetrachloride, chloroform, methyl chl oride, and dichloromethane), and it was not 2662 possible from the data presented to assess the indi vidual effects of dichloromethane. 2663

2664

4.1.3.2. Description of the Selected Studies 2665 In this section, the study setting, methods (inclu ding exposure assessment techniques), 2666 results pertaining to incidence or mortality from s pecific cancers, and a brief summary of 2667 primary strengths and limitations are summarized fo r each of the 12 selected studies. When two 2668 papers of the same cohort were available, the resul ts from the longer period of follow-up are 2669 emphasized in the summary. Information from earlie r reports is used when these reports contain 2670 more details regarding working conditions, study de sign, and exposure assessment. The 2671 description of individual studies is followed by a summary of the evidence available from these 2672 studies relating to specific types of cancer. 2673

2674

4.1.3.3. Cellulose Triacetate Film Base Production Cohorts 2675

4.1.3.3.1. Cellulose triacetate film base production—Rochester , New York (Eastman Kodak). 2676 Friedlander et al. (1978) reported a cohort mortali ty study of workers in an Eastman Kodak 2677 facility in Rochester, New York. This study was ex panded and extended several times during 2678 the next 20 years (Hearne and Pifer, 1999; Hearne e t al., 1990, 1987). The latest analysis 2679 provided data on two overlapping cohorts. The firs t cohort (Cohort 1) consisted of 1,311 male 2680 workers employed in the roll coating division (n = 1,070) or the dope and distilling departments 2681 (n = 241) of the Eastman Kodak facility in Rocheste r, New York. Men who began working in 2682 these areas after 1945 and were employed in these a reas for at least 1 year (including seasonal or 2683 part-time work that equaled 1 full-time year equiva lent) from 1946 to 1970 were included. 2684 Follow-up time was calculated from the end of the f irst year of employment in the study area 2685 through December 31, 1994. The mean duration of wo rk in Cohort 1 was 17 years. The total 2686 number of person-years of follow-up was 46,112, and the mean duration of follow-up was 2687 35.2 years (range 25–49 years). The second cohort (Cohort 2) included 1,013 male workers in 2688 the roll coating division who were employed for at least 1 year in this division between 1964 and 2689 1970. Follow-up time was calculated from January 1 , 1964, or the date of first employment in 2690 the roll coating division for those who were employ ed there before 1964 and those who began in 2691 1964 or later, respectively. Follow-up continued t hrough December 31, 1994. The mean 2692 duration of work in Cohort 2 was 24 years. Total f ollow-up time was 26,251 person-years, and 2693 the mean duration of follow-up was 25.9 years (rang e 25–31 years). Cohort 2 was the focus of 2694 previous analyses by Friedlander et al. (1978) and Hearne et al. (1990, 1987). 269564

For both cohorts, causes of death were based on th e underlying causes of death recorded 2696 on the death certificates, which were routinely obt ained by the company for the processing of life 2697 insurance claims. The expected number of deaths wa s calculated using appropriate age-, sex-, 2698 calendar time-, and cause-specific death rates for men in New York State (excluding New York 2699 City). In addition, another referent group was als o used in the analysis of the second cohort. 2700 This other referent was based on the age-, sex-, ca lendar time-, and cause-specific death rates of 2701 other hourly male workers employed at the Eastman K odak plant in Rochester, New York. (An 2702 internal referent group was also described for Coho rt 1, but data for that analysis were not 2703 presented.) 2704 Dichloromethane was first used in the film product ion process at the Eastman Kodak 2705 facility around 1944 (Hearne et al., 1987). Cellul ose triacetate was dissolved in dichloromethane 2706 and then cast into a thin film onto revolving wheel s. The film was then cured by circulating hot 2707 air in the coating machines, and the solvent was re covered and redistilled. 1,2-Dichloropropane 2708 and 1,2-dichloroethane were also used as solvents f rom the 1930s to the 1960s, but 2709 dichloromethane was predominant (ratio 17:2:1 for d ichloromethane:1,2-dichloropropane:1,2- 2710 dichloroethane in general workplace air measurement s) (Hearne et al., 1987). 2711 The exposure assessment in the Rochester, New York, Eastman Kodak cohort studies 2712 was based on employment records (start and stop dat es for specific jobs in the relevant areas of 2713 the company) in combination with air monitoring dat a used to estimate the exposure level for a 2714 given job, location, and time period (Hearne et al. , 1987). Air monitoring began in the 1940s, 2715

but few data are available before 1959. In the mos t recent update (Hearne and Pifer, 1999), more 2716 than 1,500 area and 2,500 breathing zone air sample s were used in the exposure assessment 2717 process. Reductions in exposures in the dope depar tment and the distilling department began 2718 after 1965. The highest exposure jobs were operato r and maintenance workers (dope 2719 department) and filter washing and waste operator ( distilling department), with estimated 8-hour 2720 TWA exposures of 100–520 ppm between 1946 and 1985. There was little change in estimated 2721 exposures for jobs in the roll coating division fro m the 1940s through 1985, but some reduction 2722

was seen from 1986 to 1994. The mean 8-hour TWA ex posures were 39 ppm for Cohort 1 and 2723 26 ppm for Cohort 2. 2724 These data were used to estimate a cumulative expos ure index (i.e., the summation across 2725 all jobs held by an individual of the product of th e average dichloromethane concentration as 2726 ppm and the duration of employment in that job). T he authors refer to this as a “career exposure 2727 index.” Additional adjustment in these estimates w as made for respiratory protection, but the 2728 details of this adjustment were not described. For Cohort 1, the cumulative exposure categories 2729 used in exposure-effect analyses were <150, 150–349 , 350–799, and ≥ 800 ppm. For Cohort 2, 2730 the cumulative exposure categories were <400, 401–7 99, 800–1,199, and ≥ 1,200 ppm. The cut 2731 points were chosen to produce an approximately equa l number of expected total deaths in these 2732 categories. Two different methods to calculate exp ected number of deaths within each exposure 273365

category were used for each cohort analysis. For C ohort 1, an internal comparison was made 2734 based on the distribution of person-years within ea ch exposure category, and an external 2735 comparison was made applying New York State mortali ty rates. For Cohort 2, the internal 2736 comparison using the distribution of person-years w ithin each exposure category was also used, 2737 but the external comparison was based on mortality rates in other hourly workers at the 2738 Rochester, New York, Eastman Kodak work site. 2739 There was no increased risk of mortality for all s ites of cancer or for lung cancer in either 2740 cohort analysis (Table 4-4). Data pertaining to sm oking history, obtained from a survey of 2741 workers in the New York film production facility, i ndicate that smoking rates were similar in the 2742 exposed group, the internal comparison group, and t he general population; therefore, it is 2743 unlikely that differences in smoking could be maski ng an effect of dichloromethane. 2744 The only specific sites for which there were increa sed SMRs in both cohorts were brain 2745 and CNS cancer, Hodgkin’s lymphoma, and leukemia. Pancreatic cancer mortality risk was 2746

increased in Cohort 2 but not in Cohort 1. None of these associations were statistically 2747 significant, and the Hodgkin’s lymphoma observation s were based on a total of only four cases 2748 in both cohorts combined and so were very imprecise . Within Cohort 2, there was little 2749 difference in results for most sites, using the dif ferent referent groups, but the point estimates for 2750 the SMRs for brain and CNS cancer, Hodgkin’s lympho ma, and leukemia were somewhat higher 2751 using the New York State referent group compared wi th the internal Eastman Kodak referent 2752 group. An attenuation of the dichloromethane assoc iation seen in the analyses using the internal 2753 Kodak referent group would be expected if the risk of specific cancers was increased in this 2754 comparison group, possibly because of other workpla ce exposures. 275566Table 4-4. Mortality risk in Eastman Kodak cellulose triacetate film base production workers, Rochester,New YorkCohort 1:1,311 men employed 1946–1970, followedthrough 1994Cohort 2:1,013 men employed 1964–1970, followed through 1994New York referent groupNew York referent groupKodak referent groupCancer typeObsbExpbSMR95% CIObs Exp SMR 95% CIExp SMR 95% CICancer, all sites 93 105.8 0.88 0.71–1.0891 102.0 0.89 0.72–1.1094.7 0.96 0.77–1.18Livera1 2.4 0.42 0.01–2.361 2.4 0.42 0.01–2.331.8 0.55 0.01–3.07Pancreas 5 5.5 0.92 0.30–2.148 5.3 1.51 0.65–2.985.1 1.55 0.67–3.06Lunga27 36.0 0.75 0.49–1.0928 34.2 0.82 0.55–1.1931.3 0.89 0.59–1.29Braina6 2.8 2.16 0.79–4.694 2.1 1.88 0.51–4.812.7 1.46 0.39–3.75Lymphatic system 5 6.6 0.75 0.24–1.786 5.7 1.06 0.39–2.305.7 1.05 0.38–2.28Non-Hodgkin’s 2 4.1 0.49 0.06–1.763 3.5 0.85 0.17–2.503.6 0.84 0.17–2.46Hodgkin’s 2 1.1 1.82 0.20–6.572 0.6 3.13 0.35–11.300.9 2.23 0.25–8.05Multiple myeloma 1 1.5 0.68 0.01–3.791 1.5 0.65 0.01–3.621.3 0.79 0.01–4.39Leukemia 8 3.9 2.04 0.88–4.036 3.5 1.73 0.63–3.764.4 1.37 0.50–2.98aLiver includes liver and biliary duct; lung includes lung, trachea, and bronchus; brain includes brainand CNS.bObs = number observed deaths, Exp = number of expected deaths.Source: Hearne and Pifer (1999).275667

The authors presented the exposure-effect analysis, based on the estimated cumulative 2757 dichloromethane exposure groups, for all sites of c ancer, pancreatic cancer, lung cancer, brain 2758 cancer, and leukemia (Table 4-5). 2759

2760 Table 4-5. Mortality risk by cumulative exposure i n Eastman Kodak

cellulose triacetate film base production workers, Rochester, New York

Cohort, cancer, referent group SMRs (number of observed deaths) Cohort 1 a Cumulative exposure (ppm years) <150 150–349 350–799 ≥ 800 Cancer, all sites internal New York 0.81 (20) 0.67 1.02 (19) 0.93 1.10 (28)

0.95 1.07 (26) 1.00 Pancreas internal New York 0.74 (1) 0.68 0.00 (0) 0.00 0.77 (1)

0.65 2.34 (3) 2.18 Lung b

internal New York 0.78 (5) 0.52 1.07 (6) 0.90 1.25 (9)

0.86 0.90 (7) 0.77 Brain b

internal New York 0.58 (1) 1.10 0.78 (1) 1.77 1.65 (3)

3.99 0.85 (1) 1.78 Leukemia internal New York 0.83 (2) 1.61 0.00 (0) 0.00 0.48 (1)

0.98 2.73 (5) 5.79 Cohort 2 c Cumulative exposure (ppm years) <400 400–799 800–1,199 ≥ 1,200

Cancer, all sites Internal New York 0.89 (18) 0.76 0.96 (33) 0.93 1.11 (23)

1.13 1.08 (17)

1.12 Pancreas Internal New York 2.58 (4) 2.86 0.00 (0) 0.00 0.95 (2)

1.83 1.43 (2)

2.67 Lung b

Internal New York 0.95 (6) 0.80 1.15 (12) 1.00 0.94 (6)

0.89 0.82 (4)

0.79 Brain b

Internal New York 0.00 (0) 0.00 1.13 (2) 2.02 1.37 (1)

1.75 1.49 (1)

2.50 Leukemia Internal New York 0.00 (0) 0.00 0.84 (2) 1.26 0.75 (1)

1.10 2.70 (3)

4.84

a Cohort 1: 1,311 men employed 1946–1970 in the roll coating division, dope department, or distilling de partment, followed through 1994; mean exposure (cumulative ex posure years) 66, 244, 543, and 1,782 ppm-years in the four dose groups, respectively.

b Lung includes lung, trachea, and bronchus; brain in cludes brain and CNS.

c Cohort 2: 1,013 men employed 1964–1970 in the roll coating division, followed through 1994; mean expos ure (cumulative exposure years) 168, 581, 981, and 1,67 0 ppm-years in the four dose groups, respectively.

Source: Hearne and Pifer (1999). 2761 276268

There is no evidence of an exposure-effect for all site cancer mortality or lung cancer 2763 mortality risk. The relatively sparse number of de aths for the other specific cancer types makes 2764 it difficult to interpret the data. The patterns f or pancreatic cancer differ between the two 2765 cohorts, with increased risk at the higher dose in Cohort 1 and a U-shaped curve seen in 2766 Cohort 2. For brain cancer mortality, a higher SMR was seen in the groups with cumulative 2767 exposure levels of 800 ppm-years or greater compare d with lower exposure groups. For 2768 leukemia, in both cohorts, an increased mortality r isk is seen in the highest exposure group 2769 (mean approximately 1,700 ppm-years). 2770 A strength of the Eastman Kodak cohort studies was the sampling data for 2771

dichloromethane that allowed an assessment of each worker’s exposure, using the monitoring 2772 data and the worker’s job history, making exposure- effect analyses possible. Follow-up of the 2773 vital status of the cohort was >99% (Hearne and Pif er, 1999). There was also some information 2774 on smoking history, too, for workers in the plant, based on a survey conducted in 1986 (Hearne 2775 et al., 1987). A difficulty in interpreting the da ta, however, is that there was some overlap 2776

between the cohorts: 707 of the men were included i n both Cohort 1 and Cohort 2. Data are not 2777 presented in a way that would allow the reader to e liminate duplicate cases and person-years so 2778 that cases are only counted once when examining bot h cohorts. A strength of the Cohort 2779 1 sampling strategy, compared with that of Cohort 2 , is that Cohort 1 is limited to workers who 2780 began work at the plant after 1945. These workers would not have had workplace exposure to 2781 methanol and acetone, which were used at the plant in the film production process prior to that 2782 time. Also, follow-up began with the beginning of employment in the relevant area. In contrast, 2783 Cohort 2 was limited to workers who were employed f rom 1964 to 1970, so exposed workers 2784

who left or died before 1964 were not included. Th e relatively small number of cases with 2785 specific low incidence cancers (e.g., brain, leukem ia) is also a limitation of the analyses of both 2786 of the cohorts in this study. In addition, the exp osure levels in both cohorts (mean 8-hour TWA 2787 39 and 26 ppm in Cohorts 1 and 2, respectively) is relatively low compared with values seen in 2788 other workplaces, including the cellulose triacetat e fiber production cohorts described in Ott et 2789 al. (1983a) and Gibbs et al. (1996). Also, the out come assessment is based on mortality 2790 (underlying cause from death certificates) rather t han incidence data, and, because the Kodak 2791 studies were limited to men, there is no informatio n on risk of breast cancer or other female 2792 reproductive cancers. 2793

2794

4.1.3.3.2. Cellulose triacetate film base production—Brantham, United Kingdom (Imperial 2795 Chemical Industries). Tomenson et al. (1997) reported the results of a retrospective cohort 2796 mortality study of 1,473 men who worked at a film-b ase production facility in Brantham, 2797 England, anytime between 1946 and 1988 in jobs that were considered to have dichloromethane 2798 exposure. The start of the follow-up period was no t specified by the authors but is likely to have 2799 been 1946 or the date of first employment at the pl ant. Follow-up of the cohort continued 280069

through December 31, 1994, and vital status was bas ed on national records (United Kingdom 2801 National Health Service Central Register and the De partment of Social Security). Causes of 2802 death were based on the underlying causes of death recorded on the death certificates. The 2803 expected number of deaths was calculated using age- , sex-, calendar time-, and cause-specific 2804 death rates for England and Wales. In addition, a comparison using mortality rates for the local 2805 areas (Tendring and Samford) for 1968–1978 and anal yses limited to workers who had been 2806 employed for at least 3 months were also made, but the results of these analyses were not 2807 presented. The mean duration of work in the cohort was 9 years, the total number of person2808 years was 39,759, and the mean duration of follow-u p was 27.0 years (7–49 years). 2809 This facility produced cellulose diacetate film fro m 1950 to 1988, with other types of 2810 films also manufactured beginning in the 1960s. Di chloromethane was the solvent used in this 2811 process, and exposure occurred in the production of the triacetate film base and the casting of the 2812 film into rolls. The exposure assessment was based on more than 2,700 personal or air 2813 monitoring samples collected since 1975. An exposu re matrix was constructed, assigning jobs to 2814 1 of 20 work groups with similar exposure potential for each of four different time periods 2815 (before 1960, 1960–1969, 1970–1979, and 1980–1988). For the 1980–1988 period, exposure 2816 estimates for specific jobs were based on about 330 personal monitoring samples. For the earlier 2817 time periods, information about work tasks and loca tion was used in combination with the 2818 information about the number of, use of, speed of, and problems with casting machines at 2819

different times from their initial introduction in 1950. The highest exposures were estimated to 2820 be in the casting machine operators and cleaners. Lifetime cumulative exposure to 2821 dichloromethane was calculated as the product of th e mean level of exposure for the assigned 2822 work group and the duration of employment in that j ob summed across all jobs. Three categories 2823 of cumulative exposure were used for the analysis o f ever exposed workers: <400, 400–700, and 2824 800+ ppm-years. Approximately 30% of the workers i n the cohort were classified as 2825 “unassigned” for the calculation of exposure group because sufficient information needed to 2826 determine exposures (i.e., the location and tasks a ssigned to laborers and maintenance workers) 2827 was not available. The mean 8-hour TWA exposure wa s estimated as 19 ppm for the cohort. 2828 There was no increased risk of mortality for all si tes of cancer (Table 4-6), and the SMRs 2829 for most of the specific cancer sites examined (sto mach, colon, rectum, liver, pancreas, lung, and 2830 prostate) were less than 1.0. The only specific si tes for which there was an increased SMR (i.e., 2831 1.1 or higher) were brain and CNS cancer and leukem ia, and these estimates were based on few 2832 (less than five) observed cases (Table 4-6). Tomen son et al. (1997) present the exposure-effect 2833 analysis, based on the estimated cumulative dichlor omethane exposure groups, for all sites of 2834 cancer, pancreatic cancer, and lung cancer, and the re is no evidence of an increasing effect with 2835 increasing exposure level in these analyses. A for mal exposure-effect analysis for brain cancer 2836 or leukemia was not presented. However, the author s described two of the brain cancer cases as 2837 having “minimal” exposure to dichloromethane (and t hus presumably would have been in the 283870

<400 ppm-year cumulative exposure group). One case was estimated as having 572 ppm-years 2839 cumulative exposure, and the other case was an elec trician classified in the unassigned exposure 2840 group. He had worked for 21 years at an exposure l evel “that was unlikely to have exceeded 2841 15 ppm 8 hour TWA.” 2842

2843 Table 4-6. Mortality risk in Imperial Chemical Ind ustries cellulose

triacetate film base production workers, Brantham, United Kingdom: 1,473 men employed 1946–1988, followed through 1994

Cancer type Observed Expected a SMR 95% CI Cancer, all sites 68 104.6 0.65 0.51–0.82 Liver and biliary duct 0 1.5 – – Pancreas 3 4.4 0.68 0.14–1.99 Lung, trachea, bronchus 19 41.3 0.46 0.29–0.75 Brain and CNS system 4 2.8 1.45 0.40–3.72 Lymphatic and hematopoietic 6 7.1 0.85 0.31–1.84 Leukemia 3 2.7 1.11 0.23–1.84

a Expected, calculated from observed and SMR data rep orted by the authors by using the following formula : expected = 100 × observed ÷ SMR; SMRs and CIs were not calculated for categori es with zero observed cases.

Source: Tomenson et al. (1997). 2844 2845 A strength of this study was the monitoring data a vailable that allowed assignment of 2846 cumulative exposure categories for use in exposure- effect analyses. However, 30% (439) of 2847 exposed workers had insufficient work histories to determine lifetime cumulative exposure. Air 2848 measurements were not available until 1975, and per sonal measures were not available until 2849 1980. In addition, the duration of exposure was re latively low (mean, 9 years), the mean 2850 exposure level was relatively low (mean 8-hour TWA, 19 ppm), and there were very few deaths 2851 from specific types of cancer, which limit the stat istical power of the study to examine 2852 associations among dichloromethane and specific can cers. Other limitations, as were also noted 2853 in the Kodak cohort studies, include the use of mor tality rather than incidence to define risk, the 2854 reliance solely on underlying causes of death from death certificates to classify specific cancer 2855 types and the lack of information on breast cancer risk. 2856

2857

4.1.3.4. Cellulose Triacetate Fiber Production Cohorts 2858

4.1.3.4.1. Cellulose triacetate fiber production—Rock Hill, So uth Carolina (Hoechst Celanese 2859 Corporation). Two cohorts of cellulose triacetate fiber workers have been studied in Rock Hill, 2860 South Carolina (Lanes et al., 1993, 1990; Ott et al ., 1983a, b), and Cumberland, Maryland (Gibbs 2861 et al., 1996; Gibbs, 1992). Workers were exposed t o dichloromethane, methanol, and acetone in 2862

both facilities. 2863 Ott et al. (1983a, b) conducted a retrospective co hort mortality study of 1,271 acetate 2864 fiber production workers (551 men and 720 women) em ployed at least 3 months from 1954 to 286571

1977 at Dow Chemical Company, Rock Hill, South Caro lina. This analysis focused on ischemic 2866 heart disease mortality risk, and there was no pres entation of cancer risk. The Rock Hill cohort 2867 study was updated twice, through September 30, 1986 (Lanes et al., 1990), and December 31, 2868 1990 (Lanes et al., 1993), and analyses of cancer m ortality risks were included in these later 2869 reports. Causes of death information was obtained from death certificates, with coding based on 2870 the underlying and contributing causes (Ott et al., 1983a). The referent used in the updates was 2871 the general population of York County, South Caroli na, and analyses were adjusted for age, race, 2872 gender, and calendar period. Because the results o f the mortality risk analyses were similar for 2873 both updates, those from the 1993 paper are present ed here. The mean duration of work in the 2874 cohort was not reported, but 56% worked for fewer t han 5 years (calculated from Tables 3 and 2875 4 of Ott et al., 1983b). The mean duration of foll ow-up was 23.6 years in the analysis through 2876 1986 (Lanes et al., 1990) but was not reported in t he later paper (Lanes et al., 1993). The 2877 1993 report added approximately 4.25 years of follo w-up, which would result in an estimate of 2878 approximately 28 years follow-up for this report. 2879 The Rock Hill, South Carolina, plant began produci ng cellulose triacetate fiber in 1954. 2880

Dichloromethane was used as the solvent for the ini tial mixing with cellulose triacetate flakes. 2881

This mixture was then filtered and transferred to t he extrusion area for drying and winding. Air 2882

measurements taken in 1977–1978 were assumed to be representative of operations since 2883 dichloromethane use began in 1954, based on review of processing operations. The median 2884 8-hour TWA exposures were estimated as 140, 280, an d 475 ppm in the low, moderate, and high 2885 categories of exposure (Ott et al., 1983a). Employ ment records provided information on jobs 2886 held and dates employed, and this was used in conju nction with the exposure estimates for 2887 specific jobs and work areas to classify individual exposures. However, detailed work history 2888 information was only available for 475 (37%) of the workers (Lanes et al., 1990), and it is not 2889 clear how the exposure assessment was applied to wo rkers with missing job history data. 2890

Methanol was also used in the cellulose triacetate fiber production process, and methanol 2891 exposure was estimated as one-tenth that of dichlor omethane. Acetone exposure was used in the 2892 production of acetate (cellulose diacetate) fiber a t an adjacent part of the plant. The exposure to 2893 acetone was inversely related to that of dichlorome thane, with estimated median 8-hour TWAs 2894 of 1,080 ppm acetone in the low dichloromethane exp osure group and 110 ppm acetone in the 2895 moderate and high dichloromethane groups in the Roc k Hill plant (Ott et al., 1983a). 2896 In the latest follow-up (Lanes et al., 1993), ther e was no increase in mortality risk from 2897 cancer (all sites) or from cancer of the lung or pa ncreas (Table 4-7). The SMR for liver and bile 2898 duct cancer, based on four observed cases, was 2.98 (95% CI 0.81, 7.63). This was lower than 2899 the SMR of 5.75 (95% CI 1.82, 13.8) that was report ed in the 1990 analysis, based on these same 2900 four cases but on a shorter follow-up period (and t hus lower number of expected cases). Three 2901

of these cases were bile duct cancers. This was th e first cohort study that included women, and 2902 this study provided data on breast cancer risk. Th ere were 3 observed breast cancer deaths 290372

compared with 5.59 expected, yielding an SMR of 0.5 4 (95% CI 0.11, 1.57). No data were 2904 provided pertaining to reproductive risk factors (e .g., pregnancy history) for breast cancer among 2905 the women in this cohort, so it is difficult to ass ess whether these potential confounders are likely 2906 to have been distributed differently in the cohort compared with the referent group. Information 2907 about brain cancer, Hodgkin’s lymphoma, and leukemi a (Table 4-7) was not included in this 2908 report but was included in the report by Gibbs (199 2) (see Table 11 of that report). 2909

2910 Table 4-7. Mortality risk in Hoechst Celanese Corp oration cellulose

triacetate fiber production workers, Rock Hill, Sou th Carolina: 1,271 men and women employed 1954–1977, followed through 1990

Cancer type Observed Expected SMR b 95% CI b, c

Cancer, all sites 39 47.7 0.82 0.58–1.52 Liver and biliary duct 4 1.34 2.98 0.81–7.63

Pancreas 2 2.42 0.83 0.10–2.99 Lung, trachea, bronchus 13 16.21 0.80 0.43–1.37

Brain and CNS a 1 1.5 0.67 0.2–3.71 Hodgkin’s lymphoma a 0 0.24 – –

Leukemia a 1 1.11 0.90 0.02–5.0 Breast cancer (women) 3 5.59 0.54 0.11–1.57

a Data for brain and CNS cancer, Hodgkin’s lymphoma, and leukemia were reported in Gibbs (1992).

b SMRs and CIs were not calculated for categories wit h zero observed cases.

c CIs were calculated from Breslow and Day (1987, Tab le 2.10).

Source: Lanes et al. (1993). 2911 2912 There are a number of limitations of this study, i ncluding the small size of the cohort, 2913 small number of observed cancer deaths, availabilit y of detailed work history information for 2914 only 37% of the workers, and use of mortality rathe r than incidence data. The exposure levels at 2915 this plant were high, but the duration of exposure for most of the cohort was relatively short 2916 (<5 years). It is the first cohort study, however, that included women and provided information 2917

on breast cancer risk. 2918

2919

4.1.3.4.2. Cellulose triacetate fiber production—Cumberland, M aryland (Hoechst Celanese 2920 Corporation). Gibbs et al. (1996) studied a cohort of 2,909 cel lulose triacetate fiber production 2921 workers (1,931 men and 978 women) at a Hoechst Cela nese plant in Cumberland, Maryland. 2922 This retrospective cohort mortality study included all workers who were employed on or after 2923 January 1, 1970, and who worked at least 3 months. This study also included a very small 2924 comparison group (256 men, 46 women) that was descr ibed as a “0” or “no” exposure group of 2925 workers at the plant who worked in jobs that were n ot considered to have had dichloromethane 2926 exposure, for a total of 2,187 men and 1,024 women in the exposed and nonexposed groups. 292773

The plant closed in 1981, and mortality was followe d through 1989. Since 1955, 2928 employees of this plant were exposed to dichloromet hane, methanol, acetone, and finishing oils 2929 used as lubricants. Before 1955, acetone was the o nly exposure. Industrial hygiene monitoring 2930 focusing on dichloromethane, acetone, and methanol began in the late 1960s. Exposure 2931 groupings (low, 50–100 ppm, and high, 350–700 ppm) were assigned by area in which 2932 employees worked. The extrusion and spinning worke rs and jet wipers were among the high 2933 exposure group (300–1,250 ppm 8-hour TWA). The SMR analysis that was reported used 2934 Allegany County, Maryland, as the comparison group. Cause of death information was obtained 2935 from death certificates, but the authors did not st ate whether they used underlying or underlying 2936 and contributing cause of death information. The m ean duration of work in the cohort was not 2937 reported. The total follow-up period included 49,8 28 person-years (16,292 in the high exposure 2938 group and 33,536 in the low exposure group), and th e mean duration of follow-up was 17.2 years 2939 (range 8–20 years). These data were found in Hearn e and Pifer (1999, Table 7). 2940

There was little evidence of an increase in mortali ty risk from cancer (all sites) or from 2941 cancer of the liver and bile duct, pancreas, or bra in in men or in women (Table 4-8). An 2942 increasing risk with increasing exposure level was seen for prostate cancer mortality in men. 2943

The p -value for the trend was not given, but the authors describe it as a “nonstatistically 2944 significant dose-response relationship.” A statist ically significant SMR for prostate cancer death 2945 was seen in the 350–700 ppm group when latency (at least 20 years since first exposure) was 2946 included in the analysis (SMR = 2.08, p < 0.05). Cervical cancer mortality risk was increase d, 2947 but the small number of cases in the high exposure group did not allow a precise assessment of 2948 the pattern with respect to exposure level. There was no increased risk of breast cancer. 2949

295074 2951 Table 4-8. Cancer mortality risk in Hoechst Celane se Corporation cellulose

triacetate fiber production workers, Cumberland, Ma ryland: 2,909 men and women employed 1970–1981, followed through 1989

Men (n = 1,931) Women (n = 978) Cancer type, exposure level a Obs b Exp b SMR 95% CI c Obs b Exp b SMR 95% CI c Cancer, all sites 121 42 50–100 ppm 64 70.0 0.91 0.70–1.2 37 44.79 0.83 0.58–1.1 350–700 ppm 57 75.6 0.75 0.57–0.98 5 4.61 1.1 0.35–2.5 Liver 2 0 50–100 ppm 1 1.33 0.75 0.02–4.2 0 1.04 – 350–700 ppm 1 1.24 0.81 0.02–4.5 0 0.10 – Pancreas 3 1 50–100 ppm 2 2.24 0.89 0.1–3.2 1 1.73 0.58 0.01–3.2 350–700 ppm 1 2.90 0.35 0.01–1.9 0 0.18 – Lung 35 11 50–100 ppm 20 25.7 0.78 0.48–1.2 9 8.24 1.1 0.50–2.1 350–700 ppm 15 27.3 0.55 0.31–0.91 2 0.87 2.3 0.28–8.3 Brain a 2 2 50–100 ppm 1 1.88 0.53 0.01–2.96 2 0.66 3.1 0.37–10.9 350–700 ppm 1 1.94 0.52 0.01–2.87 0 0.07 Hodgkin’s a 50–100 ppm 1 0.4 2.5 0.06–13.9 0 0.23 350–700 ppm 0 0.41 0 0.02 Leukemia a 50–100 ppm 4 2.14 1.9 0.51–4.8 0 1.25 350–700 ppm 1 2.28 0.44 0.01–2.4 0 0.13 Prostate 22 Not applicable 50–100 ppm 9 6.41 1.4 0.64–2.7 350–700 ppm 13 7.26 1.8 0.95–3.1 Cervical Not applicable 6 50–100 ppm 5 1.69 3.0 0.96–6.9 350–700 ppm 1 0.19 5.4 0.13–30.1

Breast a 0 10 50–100 ppm 0 0.03 9 9.8 0.92 0.42–1.7 350–700 ppm 0 0.02 1 1.07 0.93 0.02–5.2

a Data for brain and CNS cancer, Hodgkin’s lymphoma, leukemia, and breast cancer reported in Gibbs (1992 ).

b Obs = number of observed deaths, Exp = number of ex pected deaths. Referent group = Allegany County, Maryland. SMRs and CIs were not calculated for cat egories with zero observed cases.

c CIs were calculated from Breslow and Day (1987, Tab le 2.10).

Sources: Gibbs et al. (1996); Gibbs (1992). 2952 2953 A primary limitation of this study is that workers who were exposed before 1970 but 2954 were not working at the plant in 1970 were not incl uded in the cohort. The authors had 2955 attempted to create a cohort of all workers who wer e employed on or after January 1, 1954, but 2956 problems with the completeness of the personnel fil e made it impossible to use this study design. 295775

From what the author (Gibbs, 1992) was able to dete rmine, the records of workers who had died, 2958 left the company, or retired before the mid to late 1960s (when a new personnel system was 2959 developed) were not available. Additional limitati ons include the small size of the cohort, small 2960 number of observed cancer deaths, and use of mortal ity (death certificate) data. This is 2961 particularly problematic for cancers with relativel y high survival rates (such as prostate cancer 2962 and cervical cancer), since incidence rates are not estimated well by mortality rates in this 2963

situation. 2964

2965

4.1.3.5. Solvent-Exposed Workers—Hill Air Force Base, Utah 2966 Spirtas et al. (1991) and Blair et al. (1998) eval uated exposure to dichloromethane in 2967 relation to mortality risk in successive retrospect ive cohort studies of 14,457 civilian workers 2968 employed at Hill Air Force Base in Utah for at leas t 1 year from 1952–1956. The analysis was 2969

limited to the workers that were white or who had m issing data on race, resulting in a sample 2970 size of 14,066 (10,461 men, 3,605 women). Spirtas et al. (1991) examined mortality through 2971 1982 (3,832 deaths), and Blair et al. (1998) update d mortality through 1990 (4,195 deaths). The 2972 underlying and contributing causes of death informa tion from death certificates was used to 2973 classify cause-specific mortality. SMRs were calcu lated by using mortality rates from the Utah 2974 population, and an internally standardized life tab le method was used to adjust for age at entry 2975 into the cohort and competing causes of death. In the Blair et al. (1998) analysis, adjusted 2976 relative risks (rate ratios) were estimated from a Poisson regression analysis with unexposed 2977

workers as the referent. The mean duration of work was not reported. In the analysis through 2978 1982 (Spirtas et al., 1991), there were 22,770 pers on-years of follow-up in men and 2979 3,091 person-years of follow-up in women who were c lassified as exposed to dichloromethane. 2980 The total number of workers classified as exposed t o dichloromethane was 1,222 (Stewart et al., 2981 1991), which would yield an estimated mean of appro ximately 21 years follow-up through 1982. 2982 The total number of person-years included in the la ter report (Blair et al., 1998), with the 2983 addition of 8 more years of follow-up, was not repo rted but would be expected to increase the 2984 mean follow-up time to approximately 29 years. 2985 Two industrial hygienists developed the exposure a ssessment based on walkthrough 2986 surveys, interviews with management and labor repre sentatives, review of historical records, job 2987 descriptions, monitoring data and other information pertaining to chemicals used, and 2988 organization of the work site (Blair et al., 1998; Spirtas et al., 1991). Each worker was assigned 2989 exposure by using information on the worker’s job h istory, which included job titles, department 2990 codes, and dates of employment. The most detailed exposure assessment was done for 2991 trichloroethylene, the primary focus of the study. Dichloromethane, one of 25 other exposures 2992 analyzed, was classified as a dichotomous exposure (ever exposed, never exposed). 2993 Blair et al. (1998) presented the mortality risk f or three specific cancers in relation to 2994

15 of the 25 chemicals classified as dichotomized e xposures. The rate ratios for non-Hodgkin’s 299576

lymphoma and multiple myeloma in relation to dichlo romethane in men were 3.0 (95% CI 0.9, 2996 10.0) and 3.4 (95% CI 0.9, 13.2), respectively. Th ese rate ratios, (particularly those for multiple 2997 myeloma), were considerably higher than the rate ra tios for any of the other chemicals examined, 2998 in which the next highest observed rate ratio was 1 .8 for Freon. No cases of either of these 2999

cancers were observed in women with dichloromethane exposure, but the rate ratio for breast 3000 cancer in these women was 3.0 (95% CI 1.0–8.8). As sociations of similar magnitude (rate ratios 3001 of 3.0–4.0) were also seen among breast cancer and some other exposures (Freon, solder flux, 3002 isopropyl alcohol, and trichloroethane). 3003 This is the largest of the cohort studies that wer e identified that included women and 3004 specifically reported data pertaining to breast can cer risk. The major limitation of this study is 3005 that the exposure assessment for dichloromethane wa s based on a dichotomized classification. In 3006 addition, exposure to many different types of solve nts was common; thus, it is difficult to 3007 completely separate the effects of individual expos ures. Some aspects of reproductive history, 3008 such as age at first pregnancy, are known risk fact ors for breast cancer. Reproductive history 3009 was not included in this analysis, but Blair et al. (1998) note that it is unlikely that these factors 3010 would confound the results of a few specific chemic als, since the referent group was an internal 3011 group within the cohort (and thus would be expected to be similar in terms of socioeconomic 3012 status) and there was no association overall betwee n solvent exposures and breast cancer 3013 mortality. 3014

3015

4.1.3.6. Case-Control Studies of Specific Cancers and Dichlo romethane 3016 Seven site-specific cancer case-control studies in cluded dichloromethane as an exposure 3017

of interest. These studies involve six cancer site s: brain and CNS (Cocco et al., 1999; Heineman 3018 et al., 1994), breast (Cantor et al., 1995), kidney (Dosemeci et al., 1999), pancreas (Kernan et al., 3019 1999), rectum (Dumas et al., 2000), and childhood l eukemia (Infante-Rivard et al., 2005). A 3020 synopsis of cohort studies in humans is provided in Table 4-9. 3021

3022

4.1.3.6.1. Case-control studies of brain cancer. Heineman et al. (1994) studied the association 3023 between astrocytic brain cancer (International Clas sification of Diseases 9 th ed. [ICD-9] codes 3024 191, 192, 225, and 239.7) and occupational exposure to chlorinated aliphatic hydrocarbons. 3025 Cases were identified by using death certificates f rom southern Louisiana, northern New Jersey, 3026 and the Philadelphia area. This analysis was limit ed to white males who died between 1978 and 3027 1981. Controls were randomly selected from the dea th certificates of white males who died of 3028 causes other than brain tumors, cerebrovascular dis ease, epilepsy, suicide, and homicide. The 3029 controls were frequency matched to cases by age, ye ar of death, and study area. 3030 Next of kin were successfully located for intervie w for 654 cases and 612 controls, which 3031 represents 88 and 83% of the identified cases and c ontrols, respectively. Interviews were 3032 completed for 483 cases (74%) and 386 controls (63% ). There were 300 cases of astrocytic 303377

brain cancer (including astrocytoma, glioblastoma, mixed glioma with astrocytic cells). The 3034 ascertainment of type of cancer was based on review of hospital records, which included 3035 pathology reports for 229 cases and computerized to mography reports for 71 cases. After the 3036 exclusion of 66 controls with a possible associatio n between cause of death and occupational 3037 exposure to chlorinated aliphatic hydrocarbons (som e types of cancer, cirrhosis of the liver), the 3038 final analytic sample consisted of 300 cases and 32 0 controls. 3039 In the next-of-kin interviews, the work history in cluded information about each job held 3040 since the case (or control) was 15 years old (job t itle, description of tasks, name and location of 3041 company, kinds of products, employment dates, and h ours worked per week). Occupation and 3042 industry were coded based on four-digit Standard In dustrial Classification and Standard 3043 Occupational Classification (Department of Commerce ) codes. The investigators developed 3044 matrices linked to jobs with likely exposure to dic hloromethane, five other chlorinated aliphatic 3045 hydrocarbons (carbon tetrachloride, chloroform, met hyl chloroform, tetrachloroethylene, and 3046 trichloroethylene), and organic solvents (Gomez et al., 1994). This assessment was done blinded 3047 to case-control status. Exposure was defined as th e probability of exposure to a substance (the 3048 highest probability score for that substance among all jobs), duration of employment in the 3049 exposed occupation and industry, specific exposure intensity categories, average intensity score 3050 (the three-level semiquantitative exposure concentr ation assigned to each job multiplied by 3051 duration of employment in the job, summed across al l jobs), and cumulative exposure score 3052 (weighted sum of years in all exposed jobs with wei ghts based on the square of exposure 3053 intensity [1, 2, 3] assigned to each job). Secular trends in the use of specific chemicals were 3054 considered in the assignment of exposure potential. Exposures were lagged 10 or 20 years to 3055 account for latency. Thus, this exposure assessmen t procedure was quite detailed. 3056 Adjusting for age and study area, the OR for the a ssociation between any exposure to 3057 dichloromethane and risk of astrocytic brain cancer was 1.3 (95% CI 0.9, 1.8). There was a 3058 statistically significant trend ( p < 0.05) with increasing probability of exposure to 3059 dichloromethane with an OR = 1.0 (95% CI, 0.7, 1.6) for low probability, OR = 1.6 (95% CI 0.8, 3060 3.0) for medium probability, and OR = 2.4 (95% CI 1 .0, 5.9) for high probability compared with 3061 the referent group of unexposed men. An increased risk with higher duration of exposure was 3062 also observed, with OR = 1.7 (95% CI 0.9, 3.6) for 21 or more years of work in exposed jobs for 3063 all exposed workers and OR = 6.1 (95% CI 1.1, 43.8) for the combination of 21 years or more of 3064 work in a high probability of exposure job. Simila r results were seen in additional analyses, 3065 controlling for age, study area, employment in elec tronics occupations and industries, and 3066 exposure to carbon tetrachloride, tetrachloroethyle ne, and trichloroethylene. There was also 3067 evidence of an association between astrocytic brain cancer risk and dichloromethane exposure, 3068 based on the average intensity score, with an OR = 1.1 (95% CI 0.7, 1.7) for the low-medium 3069 intensity group and an OR = 2.2 (95% CI 1.1, 4.1) f or the high intensity group, and trend p -value 3070 <0.05. The combination of high intensity and high duration (21 or more years) was strongly 307178

associated with risk (OR = 6.1 [95% CI 1.5, 28.3]), and a weaker association (OR = 1.4 [95% CI 3072 0.6, 3.2]) was seen for high intensity and shorter duration (2–20 years). The association between 3073 cumulative exposure score (low, medium, and high) a nd astrocytic brain cancer risk was 3074 nonlinear (ORs of 0.9, 1.9, and 1.2 in the low, med ium, and high exposure categories, 3075 respectively). 3076 The strengths of this case-control study include a large sample size, detailed work 3077 histories (including information not just about usu al or most recent industry and occupation but 3078 also about tasks and products for all jobs held sin ce age 15), and comprehensive exposure 3079 assessment and analysis along several different dim ensions of exposure. The major limitations 3080 were the lack of direct exposure information and po tential inaccuracy of the descriptions of work 3081

histories that were obtained from next-of-kin inter views. Heineman et al. (1994) acknowledge 3082 these limitations in the report, and, in response t o a letter by Norman and Boggs (1996) 3083 criticizing the methodology and interpretation of t he study, Heineman et al. (1996) noted that, 3084 while the lack of direct exposure information must be interpreted cautiously, it does not 3085

invalidate the results. Differential recall bias b etween cases and controls was unlikely because 3086 work histories came from next-of-kin for both group s, the industrial hygienists made their 3087 judgments blinded to disease status, and the strong associations that were seen with the exposure 3088

measures for dichloromethane were not seen with the other solvents included in the analysis. 3089 The relatively strong and statistically significant associations between dichloromethane and 3090 astrocytic brain tumors were seen along multiple me asures of exposure, suggesting that the 3091 results were unlikely to be spurious. Nondifferent ial misclassification would, on average, 3092 attenuate true associations and would be unlikely t o result in the types of exposure-response 3093 relationships that were observed in this study. 3094 Norman and Boggs (1996) described an apparent inco nsistency in the estimated trends in 3095 dichloromethane and carbon tetrachloride exposure b ased on the methodology used in this case3096 control study (described in more detail in Gomez et al. [1994]). In response, Gomez (1996) 3097 noted that the apparent inconsistency was actually due to an error in the labeling of the lines on 3098 one of the figures in the report rather than an inc onsistency with the estimated trends. Another 3099 point raised by Norman and Boggs (1996) was that th e Heineman et al. (1994) findings were 3100 surprising in light of the lack of brain carcinogen esis in animals. In response, Heineman et al. 3101 (1996) pointed out that carcinogens commonly cause different cancers in animals and humans. It 3102 can also be noted that brain tumors are exceedingly rare in animal bioassays (Sills et al., 1999). 3103 Norman and Boggs (1996) also suggested that the res ults of the Heineman et al. (1994) study be 3104 given no weight when compared with the results of t he cohort studies. The authors responded by 3105 pointing out that the cohort studies had low statis tical power and large CIs around their point 3106

estimates but were not inconsistent with an associa tion between dichloromethane and brain 3107 cancer (Heineman et al., 1996). This point is stre ngthened further by the more recent results 3108 from the Rochester, New York, Eastman Kodak cohort (Hearne and Pifer, 1999), described 310979

previously, since an increased SMR for brain and CN S cancers was seen in the longer follow-up 3110 period of this cohort. 3111 In another case-control study of brain cancer and dichloromethane exposure, Cocco et al. 3112 (1999) identified 12,980 female cases of cancer of the brain and CNS through the underlying 3113 cause of death listing (ICD codes 191 and 192) on d eath certificates from 24 states from 1984 to 3114 1992. (This collection of death certificates is a data set created by the National Center for Health 3115 Statistics, NIOSH, and the National Cancer Institut e to facilitate research on occupational 3116 exposures and mortality risk.) The cases included 161 women with meningioma (ICD-9 codes 3117 192.1, 192.3). Four women who died of nonmalignant diseases, excluding neurological 3118 disorders, were chosen as controls for each case. The controls were frequency matched to the 3119 cases by state, race, and 5-year age group. Occupa tion data were based on the occupation fields 3120 in the death certificates. This job was coded base d on the three-digit industry and three-digit 3121 occupation (Department of Census) codes. The inves tigators developed job exposure matrices 3122 that were applied to these industry/occupation code s. The job exposure matrices included 3123 probability and intensity scores for 11 occupationa l hazards, one of which was dichloromethane, 3124 but also included other solvents, electromagnetic f ields, chlorinated aliphatic hydrocarbons, 3125 benzene, lead, nitrosamines, insecticides, herbicid es, and public contact. The investigators used 3126 logistic regression models to estimate ORs, adjusti ng for each workplace exposure, marital 3127 status, three levels of socioeconomic status (based on occupation), and age at death. For each 3128 chemical, four levels of intensity and probability were defined (unexposed, low, medium, and 3129 high). 3130 A weak association between dichloromethane exposur e and brain/CNS cancer was seen 3131 (OR 1.2 [95% CI 1.1, 1.3]) (Cocco et al., 1999). T here was no exposure-related trend in the 3132 association between probability or intensity of exp osure and brain cancer. A similar but more 3133 imprecise association was seen with meningioma canc er (OR 1.2 [95% CI 0.7, 2.2]). There were 3134 too few cases of meningioma to stratify by exposure probability and intensity. 3135 The major limitations of this study are the use of mortality rather than incidence data and 3136 the reliance on occupation data from death certific ates. The death certificate occupation data are 3137 based on “usual” occupation, which may be more pron e to misclassification in studies of women 3138 because of gender-related differences in work patte rns (i.e., shorter duration jobs for women 3139 compared with men). A relatively broad job exposur e matrix was applied to the job information, 3140 and typically more generic job exposure matrices re sult in less sensitive assessment with limited 3141 ability to detect exposure-response trends (Teschke et al., 2002). Nondifferential 3142 misclassification of outcome and exposure would gen erally result in attenuated effect estimates. 3143

3144

4.1.3.6.2. Case-control studies of breast cancer. Cantor et al. (1995) conducted a case-control 3145 study of occupational exposures and breast cancer, using the 24 state (1984–1989) death 3146 certificate data described in the previous section. Cases were women with breast cancer coded as 314780

the underlying cause of death (ICD-9 code 174). Fo ur female controls per case were selected 3148 from all noncancer deaths, frequency matched by age (5-year age groups) and ethnicity (black, 3149 white). The occupation listed on the death certifi cate was coded based on the three-digit industry 3150 and three-digit occupation (Department of Census) c odes, and this was used with a job exposure 3151 matrix developed by the investigators to assess 31 workplace exposures, one of which was 3152 dichloromethane. Four exposure probability and thr ee exposure level scores were assigned. 3153 ORs for probability and level were calculated for e ach ethnic group, adjusting for age at death 3154 and a measure of socioeconomic status (based on occ upation). After excluding subjects whose 3155 death certificate occupations were listed as homema ker, there were 29,397 white cases and 3156 4,112 black cases (total 33,509) and 102,955 white controls and 14,839 black controls (total 3157 117,794). 3158

There was little evidence of an association betwee n exposure probability and breast 3159 cancer mortality using the probability exposure met ric. The ORs were 1.05 (95% CI 0.97, 1.1) 3160 and 0.76 (95% CI 0.3, 2.0) in probability level 3 a nd level 4, respectively, for white women and 3161 1.13 (95% CI 0.9, 1.4) in probability level 3 for b lack women. (There were too few black 3162 women in exposure probability level 4 for analysis. ) Weak associations were seen with exposure 3163 level. In white women, an OR of 1.17 (95% CI 1.1, 1.3) was seen with the highest exposure 3164 level, and in black women the OR in this exposure g roup was 1.46 (95% CI 1.2, 1.7). In the 3165 analysis that jointly considered exposure level and probability ratings but excluded the lowest 3166 probability of exposure, the OR for the highest cat egory of exposure level was 1.28 in whites 3167 ( p < 0.05) and 1.21 in blacks. 3168 As with the Cocco et al. (1999) case-control study that used a similar methodology, the 3169 limitations of this study include the use of an out come defined by mortality rather than incidence, 3170 use of usual occupation information as recorded in death certificates, and use of a very broad job 3171 exposure matrix to classify 31 different exposures. Although information on pregnancy and 3172 lactation history (known risk factors for breast ca ncer) was not available, the authors did adjust 3173 for socioeconomic status by using the occupation da ta, which may have corrected for some of the 3174 potential confounding due to reproductive history. 3175

3176

4.1.3.6.3. Case-control studies of pancreatic cancer. Kernan et al. (1999) conducted a case3177 control study of 63,097 pancreatic cancer cases, us ing the 24-state (1984–1993) death certificate 3178 data. The diagnosis of pancreatic cancer was based on underlying cause of death (ICD-9 code 3179 157). Four controls who had died during the same t ime period of causes other than cancer were 3180 selected for each case, frequency-matched by state, race, gender, and 5-year age group 3181 (n = 252,386). Usual occupation and industry, base d on the occupation data in the death 3182 certificate, were coded by using the three-digit (D epartment of Census) codes. A job-exposure 3183 matrix was used with the industry and occupation co des to evaluate exposure intensity and 3184 probability (each categorized as high, medium, or l ow) for formaldehyde, dichloromethane, 318581

10 other solvents, and a combined “organic solvents ” measure. Race- and gender- specific 3186 analyses were conducted by using logistic regressio n to estimate ORs and 95% CIs, adjusting for 3187 age, marital status (ever, never married), resident ial area (metropolitan, nonmetropolitan), and 3188 region (east, south central, south, and west). 3189 The point estimates for the ORs in the low, medium, and high intensity categories in the 3190 four race-gender groups ranged from 0.8 to 1.3, wit h no exposure-effect trend seen in any group. 3191 The only statistically significant OR was for high exposure intensity in white females (OR 1.3 3192 [95% CI 1.1–1.6]), with ORs of 1.0 (95% CI 0.9, 1.1 ) for medium intensity and 1.1 (1.0, 1.2) for 3193 low intensity in this group. An elevated OR was se en with high exposure probability in black 3194 males (OR 2.2 [95% CI 1.0, 4.8]) but not in white f emales (OR 1.0 [95% CI 0.8, 1.4]) or white 3195 males (OR 1.0 [05% CI 0.8, 1.3]), and the ORs were 0.9 for medium exposure probability in 3196 these three groups. There were relatively few blac k females in this study, resulting in imprecise 3197 estimates (OR 2.0 [95% CI 0.8, 5.4] for medium expo sure and OR 1.5 [95% CI 0.6, 3.6] for high 3198 exposure). 3199 The limitations of this study, as with the other c ase-control studies that used the 24-state 3200 death certificate data set, include the reliance on cause of death data from death certificates rather 3201

than medical-record validated incidence data and th e use of death certificate occupation data. 3202 The job exposure matrix used with the occupation da ta was more focused than those used in 3203 Cocco et al. (1999) and Cantor et al. (1995). Alth ough the analysis adjusted for some 3204 sociodemographic characteristics, it did not includ e measures of smoking history or diabetes, 3205 which are known risk factors for pancreatic cancer (Lowenfels and Maisonneuve, 2005). 3206

3207

4.1.3.6.4. Case-control studies of renal cancer. Dosemeci et al. (1999) reported data from a 3208 population-based case-control study of the associat ion between occupation exposures and renal 3209 cancer risk. The investigators identified newly di agnosed patients with histologically confirmed 3210

renal cell carcinoma from the Minnesota Cancer Surv eillance System from July 1, 1988, to 3211 December 31, 1990. The study was limited to white cases, and age and gender-stratified controls 3212 were ascertained by using random digit dialing (for subjects ages 20–64) and from Medicare 3213 records (for subjects 65–85 years). Of the 796 cas es and 796 controls initially identified, 3214 438 cases (273 men, 165 women) and 687 controls (46 2 men, 225 women) with complete 3215 personal interviews were included in the occupation al analysis. 3216 Data were obtained through in-person interviews th at included demographic variables, 3217 residential history, diet, smoking habits, medical history, and drug use. The occupational history 3218

included information about the most recent and usua l industry and occupation (coded using the 3219 standard industrial and occupation codes, Departmen t of Commerce), job activities, hire and 3220 termination dates, and full- and part-time status. A job exposure matrix developed by the 3221 National Cancer Institute was used with the coded j ob data to estimate exposure status to 3222 dichloromethane and eight other chlorinated aliphat ic hydrocarbons. 322382

ORs were adjusted for age, smoking, hypertension a nd use of drugs for hypertension, and 3224 body mass index. No association between renal cell carcinoma and exposure to dichloromethane 3225 was observed in men (OR 0.85 [95% CI 0.6, 1.2]), wo men (OR 0.95 [95% CI 0.4, 2.2]), or both 3226 sexes combined (OR 0.87 [95% CI 0.6, 1.2]). 3227 A strength of this study includes the use of incid ent cases of renal cancer from a defined 3228 population area, with confirmation of the diagnosis using histology reports. The occupation 3229 history was based on usual and most recent job, in combination with a relatively focused job 3230 exposure matrix. In contrast to the type of exposu re assessment that can be conducted in cohort 3231 studies within a specific workplace, however, expos ure measurements, based on personal or 3232 workplace measurements, were not used, and a full l ifetime job history was not obtained. 3233

3234

4.1.3.6.5. Case-control studies of rectal cancer. Dumas et al. (2000) reported data from a case3235 control study of occupational exposures and rectal cancer conducted in Montreal, Quebec, 3236 Canada. The investigators identified 304 newly dia gnosed cases of primary rectal cancer, 3237 confirmed on the basis of histology reports, betwee n 1979 and 1985; 257 of these participated in 3238 the study interview. One control group (n = 1,295) consisted of patients with other forms of 3239 cancer (excluding lung cancer and other intestinal cancers) recruited through the same study 3240 procedures and time period as the rectal cancer cas es. A population-based control group 3241 (n = 533), frequency matched by age strata, was dra wn by using electoral lists and random digit 3242 dialing. The occupational assessment consisted of a detailed description of each job held during 3243 the working lifetime, including the company, produc ts, nature of work at site, job activities, and 3244 any additional information from the interviews that could furnish clues about exposure. The 3245 percentage of proxy respondents was 15.2% for cases , 19.7% for other cancer controls, and 3246 12.6% for the population controls. 3247 A team of industrial hygienists and chemists blind ed to subjects’ disease status translated 3248 jobs into potential exposure to 294 substances with three dimensions (degree of confidence that 3249 exposure occurred, frequency of exposure, and conce ntration of exposure). Each of these 3250 exposure dimensions was categorized into none, any, or substantial exposure. Logistic 3251 regression models adjusted for age, education, prox y versus subject responder status, cigarette 3252 smoking, beer drinking, and body mass index. Using the cancer control group, the OR for any 3253 exposure to dichloromethane was 1.2 (95% CI 0.5, 2. 8) and the OR for substantial exposure 3254 (confident that exposure occurred with 5 or more ye ars of exposure at medium or high frequency 3255 and concentration) was 3.8 (95% CI 1.1, 12.2). The results using the population-based control 3256 group for this exposure were not presented. 3257 The strengths of this study were the large number of incident cases, specific information 3258 about job duties for all jobs held, and a definitiv e diagnosis of rectal cancer. However, the use of 3259 the general population (rather than a known cohort of exposed workers) reduced the likelihood 3260 that subjects were exposed to dichloromethane, resu lting in relatively low statistical power for 326183

the analysis. The job exposure matrix, applied to the job information, was very broad since it 3262

was used to evaluate 294 chemicals. 3263

4.1.3.6.6. Case-control studies of childhood leukemia. Infante-Rivard et al. (2005) examined 3264 the association between maternal occupational expos ures, before and during pregnancy, and risk 3265 of childhood acute lymphoblastic leukemia (ICD-9 co de 204.0) by using data from a population3266 based case-control study in Quebec, Canada. Incide nt cases diagnosed from 1980–2000 were 3267 identified from the cancer hospitals in the provinc e, and diagnosis was confirmed based on 3268 clinical records from an oncologist or hematologist . Between 1980 and 1993, cases ages 0– 3269 9 years at diagnosis were included, and from 1994 t o 2000 the age range was expanded to 3270 14 years. The number of eligible cases identified was 848, and, of these, 790 parents (93%) 3271 participated in the study. Population-based contro ls, individually matched to the sex and age at 3272 diagnosis of the cases, were identified from govern ment registries of all children in the province 3273 (1980–1993) and the universal health insurance file s (1994–2000). The parents of 790 (86%) of 3274 the 916 eligible controls who were identified parti cipated in the study. 3275 Data were collected by using a structured telephon e interview. Some information (i.e., 3276 job title, dates, type of industry, industry name a nd address) was obtained for all jobs held since 3277 age 18, and additional information (e.g., materials and machines used, typical activities) was 3278 obtained for jobs held by the mother from 2 years b efore the pregnancy through the birth of the 3279 child. Specialized exposure modules were also used to collect information about specific jobs 3280 (e.g., nurse, waitress, hair dresser, textile dry c leaner). All of this information was reviewed by 3281 chemists and industrial hygienists, blinded to case -control status, to classify exposure to over 3282 300 chemicals, although the primary focus of the st udy was on solvents (21 individual 3283 substances, including dichloromethane, and six mixt ures). The exposure assessment included 3284 ratings of confidence (possible, probable, and defi nite), frequency of exposure during a normal 3285 workweek (<5, 5–30, or >30% of the time), and level of concentration (low = slightly above 3286 background, high = highest possible exposure in the study population, and medium for in3287 between levels). 3288 A weak association was seen between any dichlorome thane exposure during the 2 years 3289 before pregnancy up to the birth and risk of leukem ia in the child (OR 1.34 [95% CI 0.54, 3.34]), 3290 and results were similar when limited to exposures during pregnancy. Stronger associations 3291 were seen with probable or definite exposure (OR 3. 22 [95% CI 0.88, 11.7]) compared with 3292 possible or no exposure. The estimates for categor ies based on concentration and frequency 3293

were similar but there was no evidence for an incre asing risk with increasing exposure level. 3294

3295

4.1.3.7. Summary of Cancer Studies by Type of Cancer 3296

The cohort and case-control studies with data rele vant to the issue of dichloromethane 3297 exposure and cancer risk are summarized in Tables 4 -9 and 4-10, respectively. The strongest of 3298 the cohort studies, in terms of design, are two of the triacetate film base production cohorts 329984

(Cohort 1 in New York and the United Kingdom cohort , reported in Hearne and Pifer [1999] and 3300 Tomenson et al. [1997], respectively). These are t he cohorts with the most extensive exposure 3301 assessment information. The start of eligibility f or cohort entrance corresponds with the 3302 beginning of the time when the exposure potential a t the work site began, and the follow-up 3303 period is relatively long (mean >25 years). Althou gh Cohort 2 of the New York film base 3304 production study has similar exposure data and foll ow-up, this cohort was limited to workers 3305 employed between 1964 and 1970 and therefore would have missed anyone leaving (possibly 3306 because of illness or death) before this time. In addition, because of the overlap between 3307 Cohort 1 and Cohort 2, including both cohorts in an evaluation would be double counting 3308 experiences of some individuals. Several limitatio ns of the triacetate film base production 3309 cohorts should be noted, however. One of these lim itations concerns the generalizability of the 3310 results, given the relatively low exposure level (m ean 8-hour TWA <40 ppm) compared to the 3311 other cellulose triacetate fiber production cohorts (Gibbs et al., 1996; Ott et al., 1983a). 3312 Exposures in small, poorly ventilated work areas ar e also often much higher than those seen in 3313 these film base production cohorts (Estill and Spen cer, 1996; Anundi et al., 1993). Other 3314 limitations include the limited power to detect a r isk of low-incidence cancers (including brain 3315 and leukemia), the lack of women and thus lack of d ata pertaining to breast cancer, and the use 3316 of mortality rather than incidence data. Although the exposure levels in the cohorts involved in 3317 cellulose triacetate fiber production were much hig her than those of the film production cohorts, 3318 the duration of exposure was relatively short in th e South Carolina cohort (Lanes et al., 1993), 3319 and the majority of workers were missing job histor y data. In the Maryland triacetate fiber 3320 production plant, duration of exposure was not repo rted and the length of follow-up was 3321 relatively short (mean, 17 years) (Gibbs et al., 19 96). Also, the cohort began in 1970, even 3322 though production began in 1955, and the missing pe rsonnel records made it impossible to 3323 recreate an inception cohort. The exposure assessm ent in the study of civilian Air Force base 3324 workers (Blair et al., 1998) allowed for only a dic hotomized classification of exposure, and there 3325 was considerable exposure to other solvents among t hese workers. This Air Force base study 3326 was the largest of the cohort studies that included women and presented data pertaining to breast 3327

cancer. 332885Table 4-9. Summary of cohort studies of cancer risk and dichloromethane exposureCohortTotal n, exposure levelaandduration, length of follow-up Inclusion criteriabExposure assessment;Outcome assessment ResultscHearne and Pifer (1999)Cellulose triacetate filmbase production;New YorkCohort 1n = 1,311 menMean, 39 ppmmean duration, 17 yrmean follow-up, 35 yrBegan working after1945; worked at least1 yrWork history (job records) andpersonal/air monitoring;death certificate (underlying causes)Elevated mortality risks seen forbrain cancer, Hodgkin’s disease,and leukemia (SMRs around 2.0);no risk for liver, lung, or pancreaticcancer (see Table 4-4)Cohort 2n = 1,013 menmean, 26 ppmmean duration, 24 yrmean follow-up, 26 yrEmployed at least 1 yrbetween 1964and 1970(potential exposurebegan 1946)Work history (job records) andpersonal/air monitoring;death certificate (underlying causes)Elevated mortality risks seen forbrain cancer, Hodgkin’s disease,leukemia, and pancreatic cancer(SMRs between 1.5 and 3); no riskfor liver or lung cancer (see Table 4-4)Tomenson et al. (1997)Cellulose triacetate filmbase production;United Kingdomn = 1,473 menmean, 19 ppmmean duration, 9 yrmean follow-up, 27 yrEmployed anytimebetween 1946 and 1988Work history (job records) andpersonal/air monitoring;death certificate (underlying causes)Elevated mortality risks seen forbrain cancer (SMR 1.45); weakelevation for leukemia; no risk forliver, lung, or pancreatic cancer (seeTable 4-5)Lanes et al. (1993)Cellulose triacetate fiberproduction;South Carolinan = 551 men and 720 women (totaln = 1,271); median 140, 280, and475 ppm in low, moderate, and high,respectively; 56% <5 yr workduration; mean follow-up, ~28 yrWorked at least3 months in thepreparation or extrusionareas from 1954 to1977Job history data and personal/airmonitoring of specific areas (but jobhistory data available for 37%);death certificate (underlying andcontributing causes)Elevated mortality risk for livercancer (SMR 2.98, lower than seenin earlier study of this cohort); norisk for lung, pancreatic, or braincancer (see Table 4-7)Gibbs et al. (1996)Cellulose triacetate fiberproduction;Marylandn = 1,931 men and 978 women (totaln = 2,909); 50–100 ppm in low and350–700 ppm in high exposure;duration not reported; mean follow-up17 yrEmployed on or afterJanuary 1, 1970, for atleast 3 months(potential exposurebegan 1955)Work history (job records) andpersonal/air monitoring;death certificate (fields used notstated)Elevated mortality risk for prostatecancer (men, SMRs 1.4 and 1.8)),cervical cancer (women, SMR≥3.0), and lung cancer in women(high exposure, SMR 2.3), but notin men; weak risk for liver cancer,no risk for pancreatic or braincancer (see Table 4-8)86Table 4-9. Summary of cohort studies of cancer risk and dichloromethane exposureCohortTotal n, exposure levelaandduration, length of follow-up Inclusion criteriabExposure assessment;Outcome assessment ResultscBlair et al. (1998)Air Force Base, Utahn = 10,461 men and 3,605 women(total n = 14,066)ddichotomized (yes, no)exposure duration not reportedmean follow-up ~29 yrEmployed at least 1 yrfrom 1952 to 1956(potential exposurebegan 1939)Work history (job records) andindustrial hygiene assessment basedon work site review (dichotomizedexposure);(underlying and contributing causes)Elevated mortality risk for nonHodgkin’s lymphoma (RR 3.0) andmultiple myeloma (RR 3.4) in men,and breast cancer in women (RR3.0) (see section 4.1.3.5)a8-hour TWA.bIf dichloromethane was used at the plant before thefirst date of entrance into the cohort, the year that potential exposure began is noted.cResults are described as elevated if SMR was around1.5 or higher. There is limited statistical powerfor these cause-specific analyses in these cohortstudies; thestatistical significance of individual estimates isnot presented in this table. RR = relative risk.dIncludes whites and unknown race.3329873330Table 4-10. Summary of case-control studies of cancer risk and dichloromethane exposureCancer type,referenceLocationn cases, n controls (source), demographicgroup Exposure assessment ResultsBrainHeineman et al. (1994)Louisiana, New Jersey, Philadelphia300 cases, 320 controls (death certificates);cancer confirmed by hospital records; whitemenJob exposure matrix applied to detailedinformation on all jobs held (at least1 year) since age 15, as obtained fromnext-of-kin interviews; probability,duration, intensity, and cumulativeexposure scores; six solvents evaluatedOR 1.3 for any exposure; increased riskwith increased probability (trendp-value<0.05, OR 2.4 for high probability),increased duration, increased intensity;strongest effects seen in high probabilityplus high duration (OR 6.1) or highintensity and high duration (OR 6.1)combinations; no association withcumulative exposure score (see section4.1.3.6.1)BrainCocco et al. (1999)24 states, U.S.12,980 cases, 51,920 controls (deathcertificates); womenJob exposure matrix applied to deathcertificate occupation; probability, andintensity scores; 11 exposures evaluatedWeak association overall (OR 1.2), notrend with probability or intensity scores(see section 4.1.3.6.1)BreastCantor et al. (1995)24 states, U.S.33,509 cases, 117,794 controls (deathcertificates); black and white womenJob exposure matrix applied to deathcertificate job data, probability, andexposure level; 31 substances evaluatedLittle evidence of association withexposure probability; weak associationwith exposure level in whites and inblacks (see section 4.1.3.6.2)PancreasKernan et al. (1999)24 states, U.S.63,037 cases, 252,386 controls (deathcertificates); black and white men andwomenJob exposure matrix applied to deathcertificate occupation, probability, andintensity scores; 11 chlorinated solventsand formaldehyde evaluatedLittle evidence of associations withintensity or probability (see section4.1.3.6.3)KidneyDosemeci et al. (1999)Minnesota438 incident cases (Minnesota cancerregistry), 687 controls (random digit dialingand Medicare records); cancer confirmed byhistology; men and womenJob exposure matrices applied to mostrecent and usual job, as ascertained frominterviews; nine solvents evaluatedNo evidence of increased risk associatedwith dichloromethane (OR 0.85 in men,0.95 in women) (see section 4.1.3.6.4)RectumDumas et al. (2000)Montreal, Canada257 incident cases, 1,295 other cancercontrols from 19 hospitals; 533 populationbased controls (electoral rolls and randomdigit dialing), cancer confirmed byhistology; menJob exposure matrix applied to detailedinformation on all jobs held, as ascertainedfrom interviews; 294 substances evaluatedLittle evidence of an association with anyexposure (OR 1.2), but increased risk in asmall, “substantial exposure” group (OR3.8) (using cancer controls; analysis ofpopulation controls not given for thisexposure) (see section 4.1.3.6.5)88Table 4-10. Summary of case-control studies of cancer risk and dichloromethane exposureCancer type,referenceLocationn cases, n controls (source), demographicgroup Exposure assessment ResultsChildhood leukemia (acutelymphoblastic leukemia)Infante-Rivard et al.(2005)Quebec, Canada790 incident cases (hospitals—allprovinces), 790 population-based controls(government population registries); cancerbased on oncologist or hematologistdiagnosisages 0–14,aboth sexesSystematic review of detailed informationon all jobs held by the mother from2 years before pregnancy through birth ofthe child; 21 individual substances and sixmixtures evaluated (mostly solvents);confidence, frequency, and concentrationof exposure ratedLittle evidence of association with anyexposure (OR 1.3), but strongerassociations (OR > 3.0, referent group =possible/no exposure) with probable ordefinite and with combinations offrequency and concentration (see section4.1.3.6.6)aFrom 1980 to 1993, study was limited to diagnoses of ages 0–9, but this was expanded between 1994 and2000 to ages 0–14.333189

Case-control studies offer the potential for incre ased statistical power for assessing 3332 associations with relatively rare cancers, such as brain cancer and leukemia. Case-control 3333 studies are often designed to examine incidence rat her than mortality, which is of particular 3334 importance in etiologic research for diseases with relatively high survival rates and diseases in 3335 which survival may be strongly related to factors t hat are difficult to adjust for without detailed 3336 data collection (e.g., access to health care). The re is a considerable range in the detail and 3337 quality of the exposure assessment used in case-con trol studies, however. Case-control studies 3338 rarely include specific measurements taken at speci fic work sites of individual study participants. 3339 Although it is more difficult to determine absolute exposure levels without these individual 3340 measurements, the exposure assessment methodology u sed in case-control studies can result in 3341 useful between-group comparisons of risk if the int ra-group variability is less than the inter3342 group variability in potential exposure levels. Am ong the case-control studies with data 3343 pertaining to cancer risk and dichloromethane expos ure, the two studies with the strongest 3344 designs are the study of brain cancer by Heineman e t al. (1994) and the study of childhood 3345 leukemia by Infante-Rivard et al. (2005). These ar e the studies that obtained detailed 3346 information about all jobs held (rather than just t he usual or most recent job), focused on a 3347 relatively small number of exposures, and used medi cal record data to confirm the diagnosis. 3348 Heineman et al. (1994) obtained the work history fr om interviews with next-of-kin, however, 3349 which is most likely to have resulted in nondiffere ntial misclassification of exposure, and thus 3350

attenuation in the observed associations. The use of death certificate data to classify disease and 3351 occupational exposures in the three studies using t he large 24 state death certificate database 3352 (brain cancer: Cocco et al. [1999]; breast cancer: Cantor et al. [1995]; pancreatic cancer: Kernan 3353 et al. [1999]) is also likely to have resulted in n ondifferential misclassification of both outcome 3354 and exposure (and thus attenuated associations). 3355 Considering the issues described above with respec t to the strengths and limitations of the 3356 available epidemiologic studies, a summary of the e pidemiologic evidence relating to 3357 dichloromethane exposure and specific types of canc er can be made, as described below. The 3358 available epidemiologic data suggest an association between dichloromethane and brain cancer 3359 and liver cancer, but not lung cancer. 3360

3361

4.1.3.7.1. Brain and CNS cancer. An increased risk of brain and CNS cancers was seen in the 3362 strongest cohort studies; SMRs were 2.16 in Cohort 1 in New York (Hearne and Pifer, 1999) and 3363 1.45 in the United Kingdom cohort (Tomenson et al., 1997). These estimates are based on a 3364 small number of observations (six cases in New York and four in the United Kingdom) and so 3365 are relatively imprecise. It is only in the latest follow-up of the New York film base production 3366 cohort that an elevated SMR was observed, further s uggesting that the statistical power of the 3367 other cohort studies for examining risk of this dis ease may be quite low. Two case-control 3368 studies of dichloromethane exposure and brain cance r have been conducted (Cocco et al., 1999; 336990

Heineman et al., 1994). The Heineman et al. (1994) study, which is the stronger study in terms 3370 of exposure assessment strategy and confirmation of diagnosis, reported relatively strong trends 3371 with increasing probability, duration, and intensit y measures of exposure, but a nonlinear trend 3372 was seen with the cumulative exposure metric. This difference could reflect a more valid 3373 measure of relevant exposures in the brain from the intensity measure, as suggested by the study 3374 in rats reported by Savolainen et al. (1981) in whi ch dichloromethane levels in the brain were 3375 much higher with a higher intensity exposure scenar io compared with a constant exposure period 3376 with an equivalent TWA (see section 3.2). The avai lable epidemiologic studies provide some 3377

evidence of an association between dichloromethane and brain cancer, and this area of research 3378 represents a data gap in the understanding of the c arcinogenic potential of dichloromethane. 3379

3380

4.1.3.7.2. Liver and biliary duct cancer. Liver and biliary duct cancer are relatively unco mmon 3381 (age-adjusted incidence 6.2 per 100,000 person-year s) (SEER website, seer.cancer.gov, accessed 3382 April 2006), so it is difficult to study in most oc cupational cohorts of limited size. The cohort 3383 study with the higher exposures, the Rock Hill, Sou th Carolina, triacetate fiber production plant, 3384 suggested an increased risk of liver cancer (Lanes et al., 1993, 1990). The SMR for liver and 3385 bile duct cancer was 2.98 (95% CI 0.81, 7.63) in th e latest update of this cohort. This 3386 observation was based on four cases; two of these c ases were biliary duct cancers. As the 3387 follow-up period has increased, the strength of thi s association has decreased, although it is 3388 relatively strong (albeit with wide CIs). The decr ease in the SMR with increasing follow-up 3389 reflects the increase in number of expected cases, because the four observed cases were seen 3390 earlier in the follow-up period. No other cohort s tudy has reported an increased risk of liver 3391 cancer mortality, although it should be noted that there is no other inception cohort study of a 3392 population with exposure levels similar to those ofthe Rock Hill plant, and no data from a case3393 control study of liver cancer are available pertain ing to dichloromethane exposure. The available 3394 epidemiologic studies, with biological plausibility inferred from the results from studies in mice 3395 and female rats (see section 4.2) (NTP, 1986; Serot a et al., 1986a, b; Nitschke 1988a), provide 3396

some evidence of an association between dichloromet hane and liver and biliary duct cancer, 3397 although it should be noted that this evidence is b ased on very limited epidemiologic data. 3398

3399

4.1.3.7.3. Lung cancer. In the stronger cohort studies (Cohort 1 in the N ew York Eastman 3400 Kodak Company triacetate film production study repo rted by Hearne and Pifer [1999] and the 3401 United Kingdom triacetate film production study rep orted by Tomenson et al. [1997]), the SMRs 3402 for lung cancer were well below 1.0. The New York study had also obtained data on smoking 3403 history that indicated it was unlikely that differe nces in smoking could be masking an effect of 3404 dichloromethane (Hearne et al., 1987). Lung cancer is a common cancer (age-adjusted incidence 3405 61 per 100,000 person-years) (SEER website, seer.ca ncer.gov, accessed April 2006), so the 3406 expected rates, even in small cohorts, are based on relatively robust estimates. The only group in 340791

any study that had an increased risk for lung cance r was the high-exposure women in the 3408 triacetate fiber production cohort in Maryland (Gib bs et al., 1996). However, this was based on 3409 only two cases and was a highly imprecise estimate (SMR 2.3 [95% CI 0.28, 8.3]). No case3410 control study of dichloromethane exposure and lung cancer risk is available. The available 3411 epidemiologic studies do not provide evidence for a n association between dichloromethane and 3412 lung cancer, although it should be noted that the s tudies with the best designs are limited to 3413 relatively low exposure levels. 3414

3415

4.1.3.7.4. Pancreatic cancer. An early study (Hearne et al., 1990) of Cohort 2 of the New York 3416 triacetate film production cohort had reported 8 ob served and 4.2 expected pancreatic cancer 3417 deaths, for a twofold increased SMR ( p = 0.13). This association was reduced in the subs equent 3418 follow-up (SMR 1.5 [95% CI 0.7, 3.0]) (Hearne and P ifer, 1999) but was not seen in the more 3419 methodologically sound Cohort 1 (SMR 0.92) or in an y of the other cohorts. A meta-analysis of 3420 the cohort studies (using the data of Hearne et al. [1990]) reported a summary association of 3421 1.42 (95% CI 0.80, 2.53) (Ojajärvi et al., 2001). This summary measure would be further 3422 reduced with the updated data for Cohort 2 and the addition of Cohort 1 from Hearne and Pifer 3423 (1999). The only case-control study of pancreatic cancer mortality risk and dichloromethane 3424 exposure (based on death certificate data) did not report consistent patterns with respect to 3425 intensity or exposure among the race-sex groups stu died. The available epidemiologic studies do 3426 not provide evidence for an association between dic hloromethane and pancreatic cancer. 3427

3428

4.1.3.7.5. Leukemia and lymphoma. Each of the individual hematopoietic cancers is r elatively 3429 uncommon, with age-adjusted incidence rates of 5 pe r 100,000 person-years or less (SEER 3430 website, seer.cancer.gov, accessed April 2006). Th e relatively inconsistent (point estimates 3431 ranging from 0.50 or less to 2.0 or higher) and imp recise measures of association between 3432 dichloromethane exposure and non-Hodgkin’s lymphoma , Hodgkin’s lymphoma, myeloma, and 3433 leukemia are thus expected, given the relatively sm all size of the available cohort studies. Only 3434 one case-control study of any of these diseases and dichloromethane is available, and this is a 3435 study of childhood leukemia (acute lymphoblastic le ukemia) in relation to maternal occupational 3436 history (Infante-Rivard et al., 2005). This is a l arge, population-based study of confirmed 3437 incident cases of leukemia, with a detailed exposur e assessment pertaining to the period before 3438 and during pregnancy. A threefold increased risk w as seen with probable or definite exposure 3439 (OR 3.22 [95% CI 0.88, 11.7]) compared with possibl e or no exposure. The available 3440 epidemiologic studies do not provide an adequate ba sis for the evaluation of the role of 3441 dichloromethane in any of the specific hematopoieti c cancers because of the small size of the 3442

cohort studies and the relative lack of case-contro l studies pertaining to these outcomes. 3443

344492

4.1.3.7.6. Breast cancer. Only one large cohort study included women and re ported data 3445 pertaining to breast cancer risk (Blair et al., 199 8), and this is a cohort with a limited exposure 3446 assessment (dichotomized) and multiple exposures. A relatively strong association was seen 3447 between dichloromethane exposure and breast cancer mortality in this study (rate ratio 3.0 [95% 3448 CI 1.0, 8.8]). Similar associations were seen with several other chemicals, and the potential 3449 effect of confounding and misclassification of thes e exposures may have biased the estimate in 3450 either direction. The only case-control study of b reast cancer risk and dichloromethane exposure 3451

used the 24-state death certificate data to classif y exposure and disease. The available 3452 epidemiologic studies do not provide an adequate ba sis for the evaluation of the role of 3453

dichloromethane in breast cancer because there are currently no cohort studies with adequate 3454 statistical power and no case-control studies with adequate exposure methodology to examine 3455 this relationship. 3456

3457

4.2. SUBCHRONIC AND CHRONIC STUDIES AND CANCER BI OASSAYS IN 3458

ANIMALS—ORAL AND INHALATION 3459 4.2.1. Oral Exposure: Overview of Noncancer and Ca ncer Effects 3460 Results from studies of animals exposed by the ora l route for short-term, subchronic, and 3461 chronic durations identify the liver and the nervou s system as the most sensitive targets for 3462 noncancer toxicity from repeated oral exposure to d ichloromethane. In a 90-day exposure study, 3463 nonneoplastic histopathologic changes in the liver were observed in F344 rats exposed to 3464 drinking water doses of ≥ 166 mg/kg-day (males) or ≥ 209 mg/kg-day (females) (Kirschman et al., 3465 1986). Similar changes were seen in F344 rats in a 2-year exposure of ≥ 50 mg/kg-day (Serota et 3466 al., 1986a). 3467 The 2-year oral exposure study in F344 rats did not produce evidence of increasing 3468 incidence of liver tumors across all of the dose gr oups in males or females (Serota et al., 1986a). 3469 In females, however, a jagged stepped pattern of in creasing incidence was observed. In a 3470 parallel study in B6C3F 1 mice (Serota et al., 1986b; Hazelton Laboratories, 1983), a clearer trend 3471 with respect to hepatic cancer was seen in males bu t not females. 3472 None of the chronic oral exposure studies included a systematic measurement of potential 3473 neurological effects. One 14-day study focusing on neurobehavioral changes is available, 3474 however. Changes in autonomic, neuromuscular, and sensorimotor functions were observed in 3475 F344 rats exposed for 14 days to gavage doses ≥ 337 mg/kg-day (Moser et al., 1995) (see section 3476 4.4.3 for more details). 3477 No effects on reproductive parameters were observe d in Charles River CD rats exposed 3478 for 90 days to gavage doses as high as 225 mg/kg-da y (General Electric Co., 1976) or in 3479 pregnant F344 rats exposed to gavage doses of up to 450 mg/kg-day on gestation days (GDs) 6– 3480 19 (Narotsky and Kavlock, 1995). However, no oral exposure studies examining developmental 3481 neurobehavioral effects have been conducted (see se ction 4.3 for more details). 348293

4.2.1.1. Toxicity Studies of Subchronic Oral Exposures: Hepa tic Effects 3483 Kirschman et al. (1986) examined the toxicity of di chloromethane in a 90-day drinking 3484 water study in F344 rats (20/sex/dose level). The nominal concentration of dichloromethane in 3485 the water was 0.15, 0.45, or 1.5%. Based on BW and water consumption data, average intakes 3486 were reported to be 0, 166, 420, or 1,200 mg/kg-day for males and 0, 209, 607, or 1,469 mg/kg3487 day for females. Clinical chemistry tests (hematol ogical and chemical variables in samples of 3488 blood and urine) and tissue histopathology were eva luated in groups of five rats/sex/dose level 3489 after 1 month of treatment. These endpoints were a lso evaluated in the remaining rats sacrificed 3490 after 90 days of exposure. 3491 Exposure to dichloromethane did not affect mortalit y or cause adverse clinical signs of 3492 toxicity. Gross necropsy was also unremarkable. R eported changes in mean values for clinical 3493 chemistry variables, compared with controls, includ ed elevated serum ALT activities for all 3494 treated males at 1 month and for the high-dose fema les at 3 months, elevated serum AST activity 3495 in high-dose females at 3 months, elevated serum la ctate dehydrogenase activities in mid- and 3496 high-dose females at 3 months, and decreases in ser um concentrations of fasting glucose, 3497 cholesterol, and triglycerides in all exposed group s of both sexes at 1 and 3 months. Actual 3498 values for clinical chemistry variables, however, w ere not presented in the report. 3499 No histopathologic alterations were seen in tissues after 1 month of treatment (a detailed 3500 description of tissues examined was not presented). In rats exposed for 3 months, exposure3501 related histopathologic changes were restricted to the liver. Elevated, statistically significant, 3502 incidences of hepatocytic vacuolation were observed in all exposed male groups and in the mid3503 and high-dose female groups (see Table 4-11). The most frequently observed vacuolation was 3504 described as generalized and occurring throughout t he lobule, and Oil Red-O-staining indicated 3505 most were lipid-containing vacuoles. The incidence s of generalized vacuolation scored as mild 3506 or moderate were higher in all of the female dose g roups compared with the controls. The 3507

authors stated that the no-observed-adverse-effect level (NOAEL) based on this study is less than 3508 200 mg/kg-day and the lowest-observed-adverse-effec t level (LOAEL) for males was 3509 166 mg/kg-day. The authors did not explicitly prov ide a LOAEL for females. The results 3510 indicate that 166 mg/kg-day and 209 mg/kg-day were the LOAELs for liver effects in male and 3511 female rats, respectively. 3512

351394 3514 Table 4-11. Incidences of histopathologic changes in livers of male and

female F344 rats exposed to dichloromethane in drin king water for 90 days

Lesion, by sex Controls Low dose Mid dose High dose Males—n per group a 15 15 15 15 Estimated mean intake (mg/kg-day) 0 166 420 1,200 Number (%) with Hepatocyte vacuolation (generalized, centrilobular, or periportal) 1 (7) 10 b (67) 9 b (60) 7 b (47)

Generalized vacuolation severity: minimal mild moderate 0 (0) 0 0 0 5 b (33)

4 0 1 8 b (53)

7 1 0 6 b (40)

6 0 0 Centrilobular severity: minimal mild moderate 0 (0) 0 0 0 1 (7)

1 0 0 0 (0)

0 0 0 2 (13)

0 2 0 Hepatocyte degeneration 0 (0) 0 (0) 0 ( 0) 2 (13) Focal granuloma 1 (7) 0 (0) 0 (0) 1 (7) Females—n per group a 15 15 15 15 Estimated mean intake (mg/kg-day) 0 209 607 1,469 Number (%) with Hepatocyte vacuolation (generalized, centrilobular, or periportal) 6 (40) 13 b (87) 15 b (100) 15 b (100)

Generalized vacuolation severity: minimal mild moderate 5 (33) 5 0 0 13 b (87)

8 4 1 15 b (100)

6 5 4 15 b (100)

8 6 1 Centrilobular severity: minimal mild moderate marked 0 (0) 0 0 0 0 0 (0)

0 0 0 0 1 (7)

0 1 0 0 11 b (28)

2 4 3 2 Hepatocyte degeneration 0 (0) 0 (0) 0 ( 0) 12 b (80) Focal granuloma 0 (0) 0 (0) 4 (27) c 6 b (40)

a 20 per group; 5 sacrificed at 1 month; these endpoi nts for the remaining 15 per group.

b Statistical significance testing not reported by au thors; Fisher’s exact test for comparison with cont rol p-- value <0.05 (two-sided).

c Statistical significance testing not reported by au thors; Fisher’s exact test for comparison with cont rol p - value <0.10 (two-sided). Authors stated LOAEL = 16 6 mg/kg-day in males but did not explicitly provide LOAEL for females; NOAEL is less than 200 mg/kg-day.

Source: Kirschman et al. (1986). 3515 Kirschman et al. (1986) conducted a similar 90-day study in B6C3F 1 mice (20/sex/dose 3516 level). The estimated average intakes were 0, 226, 587, or 1,911 mg/kg-day for males and 231, 3517 586, or 2,030 mg/kg-day for females. Six mice (two controls, two low dose, and two mid dose) 3518 died during the study from unknown causes. Adminis tration of dichloromethane did not cause 3519 adverse clinical signs of toxicity or affect food c onsumption, ophthalmology, or serum ALT 3520 activity. Gross necropsy examinations also were un remarkable. 3521 Histopathologic evaluation of tissues from mice ki lled after 1 month of treatment did not 3522 reveal any compound-related effects. Evaluation at 3 months showed subtle generalized or 352395

centrilobular changes in the liver (characterized a s increased vacuolation with fat deposition), 3524 which was evident in all exposed groups and most pr ominent in mid- and high-dose female 3525 groups (Table 4-12). The most frequently detected change was characterized as a generalized 3526

vacuolation. Some evidence was found for an increa se in severity of the generalized vacuolation 3527 with increasing exposure level, but the incidence o f this lesion in the control mice was 3528 substantial, especially in females (Table 4-12). I ncidences for centrilobular vacuolation were 3529 significantly increased only for the mid-dose femal e group. No other changes were found. 3530 Using the results from this study to select doses for a chronic study, Kirschman et al. 3531 (1986) expressed the opinion that the mid-dose leve l (587 mg/kg-day) was the LOAEL in this 3532 study. Although incidences for generalized vacuola tion were increased in the low- and mid-dose 3533 male groups, the incidences in the high-dose groups were not significantly increased compared 3534 with controls (Table 4-12). The study authors iden tified a LOAEL of 586 mg/kg-day for 3535

centrilobular vacuolation in male B6C3F 1 mice. The NOAEL for males was considered by the 3536 investigators to be between 226 and 587 mg/kg-day. 353796 3538 Table 4-12. Incidences of histopathologic changes in livers of male and

female B6C3F 1 mice exposed to dichloromethane in drinking water for 90 days

Lesion, by sex Controls Low dose Mid dose High dose Males—n per group a 14 14 14 15 Estimated mean intake (mg/kg-day) 0 226 587 1,911 Number (%) with Hepatocyte vacuolation (generalized, centrilobular, or periportal) 9 (64) 12 (86) 13 (93) 12 (80)

Generalized vacuolation, severity: minimal mild moderate marked 7 (50) 4 2 1 0 12 b (86)

3 7 2 0 13 b (93)

9 5 0 0 10 (67)

7 3 0 0 Centrilobular severity: minimal mild moderate 2 (14) 2 0 0 0 (0)

0 0 0 1 (7)

0 0 1 5 (33)

1 3 1 Females—n per group a 14 11 13 15 Estimated mean intake (mg/kg-day) 0 231 586 2,030 Number (%) with Hepatocyte vacuolation (generalized, centrilobular, or periportal) 13 (93) 11 (100) 13 (100) 13 (87)

Generalized vacuolation severity: minimal mild moderate marked 13 (93) 1 8 4 0 11 (100)

3 7 1 0 13 (100)

5 6 2 0 13 (87)

3 6 1 3 Centrilobular severity: minimal mild moderate marked 0 (0) 0 0 0 0 0 (0)

0 0 0 0 5 c (39)

0 2 3 0 1 (7)

0 1 0 0

a 20 per group; 5 sacrificed at 1 month.

b Statistical significance testing not reported by au thors; Fisher’s exact test for comparison with cont rol p - value = 0.10 for low dose group and p = 0.032 for mid-dose group (two-sided).

c Statistical significance testing not reported by au thors; Fisher’s exact test for comparison with cont rol p - value = 0.016 (two-sided). Authors say LOAEL = 587 mg/kg-day; NOAEL between 226 and 587 mg/kg-day for males; not explicitly stated for females.

Source: Kirschman et al. (1986). 3539 3540 4.2.1.2. Toxicity Studies of Chronic Oral Exposures: Hepatic Effects and Carcinogenicity 3541 Longer-term (up to 2-year) oral exposure studies in mice and rats are summarized in 3542

Table 4-13 and described in more detail below. The se studies provide additional information 3543 pertaining to hepatotoxicity and carcinogenicity. 3544

3545

3546

354797

Table 4-13. Studies of chronic oral dichloromethan e exposures (up to 2 years)

Reference, strain/species Number per group Exposure information Comments Serota et al. (1986a) F344 rats 85/sex/dose + 135

controls Drinking water, 2 years, target

dose 0, 5, 50, 125, 250 mg/kg-day Mean intake: males 0, 6, 52, 125,

235 mg/kg-day females 0, 6, 58, 136,

263 mg/kg-day Nonneoplastic liver effects

(foci/areas of alteration) in males and females (see Table 4-14)

Jagged stepped pattern of increasing incidence of neoplastic nodules or hepatcellular carcinoma in females (i.e., increased in the 50 and 250 mg/kg-day but not 125 mg/kg-day groups) (see Table 4- 14)

Serota et al. (1986b); Hazelton Laboratories (1983) B6C3F 1 mice Males 125, 200,

100, 100, 125 Females 100, 100, 50, 50, 50 Drinking water, 2 years, target

dose 0, 60, 125, 185, 250 mg/kgday Mean intake: males 0, 61, 124, 177,

234 mg/kg-day females 0, 59, 118, 172,

238 mg/kg-day Increasing trend of liver cancer

(hepatocellular adenoma or carcinoma) in males (see Table 4-15)

Maltoni et al. (1988) Sprague-Dawley

rats 50/sex/dose Gavage, up to 64 weeks 0, 100, 500 mg/kg-day, 4–5 days per week High mortality in high dose group led to termination of study at 64 weeks; non-statistically significant increase in malignant mammary tumors in female rats

Maltoni et al. (1988) Swiss mice 50/sex/dose + 60

controls Gavage, up to 64 weeks

0, 100, 500 mg/kg-day, 4–5 days per week High mortality in high dose group led to termination of study at 64 weeks

3548 3549 4.2.1.2.1. Chronic oral exposure in F344 rats (Serota et al., 1986a). Treatment with 3550 dichloromethane did not induce adverse clinical sig ns or affect survival in the F344 rats (Serota 3551 et al., 1986a). BWs of rats in the 125 and 250 mg/ kg-day groups were generally lower than in 3552 controls throughout the study. The authors stated that the differences, although small, were 3553 statistically significant, but the data were not sh own in the published report. Water consumption 3554 was lower throughout the study in both sexes of rat s from the 125 and 250 mg/kg-day groups 3555 relative to controls; food consumption was also low er in these groups during the first 13 weeks 3556 of treatment. Mean hematocrit, hemoglobin, and red blood cell count were increased in both 3557 sexes at dichloromethane levels of 50, 125, and 250 mg/kg-day for 52 and 78 weeks. Half of 3558 these increases were reported to be statistically s ignificant, but the report did not provide the 3559 numerical values or specify which parameters were s ignificant. Clinical chemistry results 3560 showed decreases in alkaline phosphatase (AP), crea tinine, blood urea nitrogen, total protein, and 3561 cholesterol in both sexes at 250 mg/kg-day, and mos t of these changes were statistically 356298

significant at one or both of the intervals evaluat ed. (Significant parameters not specified and the 3563 mean group values were not presented in the publish ed report.) No significant deviations in 3564 urinary parameters were observed. Organ weights we re not significantly affected by treatment 3565

with dichloromethane. 3566 No treatment-related histopathological effects wer e noted in the tissues examined except 3567 for the liver (Serota et al., 1986a). Examination of liver sections showed a dose-related positive 3568 trend (positive Cochran-Armitage trend test) in the incidences of foci/areas of cellular alteration 3569 in treated F344 rats (Table 4-14). Comparisons of incidences with control incidences indicated 3570 statistically significant elevations at all dose le vels except 5 mg/kg-day. These liver changes 3571

3572 Table 4-14. Incidences of nonneoplastic liver chan ges and liver tumors in

male and female F344 rats exposed to dichloromethan e in drinking water for 2 years

Target dose (mg/kg-day) Controls 0 a 5 50 125 250 Trend

p -value b 250 with

recovery c Males—n per group d 76 34 38 35 41 15 Estimated mean intake (mg/kg-day) 0 6 52 125 23 5 232 Number (%) with Liver foci/areas of alteration 52 (70) 22 (65) 35 (92) e 34 (97) e 40 (98) e <0.0001 15 (100) e Neoplastic nodules 9 (12) 1 (3) 0 (0) 2 (6) 1 (2) Not reported 2 (13)

Hepatocellular carcinoma 3 (4) 0 (0) 0 (0) 0 (0) 1 (2) Not reported 0 (0)

Neoplastic nodules and hepatocellular carcinoma 12 (16) 1 (3) 0 (0) 2 (6) 2 (5) Not reported 2 (13)

Females—n per group d 67 29 41 38 34 20 Estimated mean intake (mg/kg-day) 0 6 58 136 26 3 239 Number (%) with Liver foci/areas of alteration 34 (51) 12 (41) 30 (73) e 34 (89) e 31 (91) e <0.0001 17 (85) e Neoplastic nodules 0 (0) 1 (3) 2 (5) 1 ( 3) 3 (9) Not reported 2 (10)

Hepatocellular carcinoma 0 (0) 0 (0) 2 (5) 0 (0) 2 (6) Not reported 0 (0)

Neoplastic nodules and hepatocellular carcinoma 0 (0) 1 (3) 4 (10) f 1 (3) 5 (14) f p < 0.01 2 (10) f

3573

a Two control groups combined. 3574

b Cochran-Armitage trend test was used for trend test of liver foci/areas of alteration. For tumor mort ality-unadjusted 3575 analysis, a Cochran-Armitage trend test was used, a nd, for tumor mortality-adjusted analyses, tumor pr evalence 3576 analytic method by Dinse and Lagakos (1982) was use d. Similar results were seen in these two analyses

. 3577

c Recovery group was exposed for 78 weeks and then ha d a 26-week period without dichloromethane exposure ; 3578 n = 15 for nonneoplastic lesions and n = 17 for neo plastic lesions. 3579

d Number available at terminal sacrifice; starting w ith 135 controls (combining both control groups) an d 85 per sex 3580 per dose group except recovery group (n = 25); subt racted 5, 10, and 20 per group (except for recovery group) 3581 sacrificed at 25, 52, and 78 weeks, respectively, a nd subtracted unscheduled deaths, which ranged from 5 to 19 per 3582

group. 3583

e Significantly ( p < 0.05) different from control with Fisher’s exact test. 3584

f Significantly ( p < 0.05) different from controls with Fisher’s exac t test, mortality-unadjusted and mortality-adjusted 3585 analyses. 3586 3587 Source: Serota et al. (1986a). 358899

were first noted after treatment for 78 weeks and p rogressed until week 104. Livers of animals 3589

treated with dichloromethane also showed an increas ed incidence of fatty change, but incidence 3590 data for this lesion were not presented in the publ ished report. The recovery group also showed 3591

an increased incidence of areas of cellular alterat ions, but the fatty changes were less pronounced 3592 than in the 250 mg/kg-day group dosed for 104 weeks . The authors indicate that 5 mg/kg-day 3593 was a NOAEL and 50 mg/kg-day was a LOAEL for nonneo plastic liver changes in male and 3594 female F344 rats exposed to dichloromethane in drin king water for 2 years. 3595 Dichloromethane-exposed male rats showed no statist ically significant increased 3596 incidence of liver tumors. In females, there was a positive trend for increasing incidence of 3597 hepatocellular carcinoma or neoplastic nodules with increasing dose (Table 4-14) (Serota et al., 3598 1986a). Statistically significant increases in tum or incidences were observed in the 50 and 3599 250 mg/kg-day groups (incidence rates of 10% and 14 %, respectively) but not in the 125 mg/kg3600 day group (incidence rate of 3%). Incidence was al so increased (10%) in a group exposed for 3601 78 weeks followed by a 26-week period of no exposur e. The characterization of malignant 3602 potential of the nodules was not described, however , and no trend was seen in the data limited to 3603 hepatocellular carcinomas. The incidence of hepato cellular carcinoma or neoplastic nodules in 3604 this control group (0%) was lower than that seen in historical controls from the same laboratory 3605 (324 female F344 rats; 4 with carcinoma, 21 with ne oplastic nodules; 25/324 = 7.7%). 3606

3607 4.2.1.2.2. Chronic oral exposure in B6C3F 1 mice (Serota et al., 1986b; Hazelton Laboratories, 3608 1983). A 2-year drinking water study similar to the prev iously described study in F344 rats was 3609

also conducted in B6C3F 1 mice (Serota et al., 1986b; Hazelton Laboratories, 1983). The mice 3610 received target doses of 0, 60, 125, 185, or 250 mg /kg-day of dichloromethane in deionized 3611 drinking water for 24 months. Treatment groups con sisted of 100 female mice in the low-dose 3612 group and 50 in the remaining treatment groups. Th ere were 200, 100, 100, and 125 male mice 3613 (low- to high-dose groups) in the treated groups. One hundred females (in two groups of 50) and 3614 125 males (in two groups of 60 and 65 mice) served as controls. (The authors do not state why 3615 two groups of control mice were used, other than to say that the design was used due to the high 3616

and erratic incidence of liver tumors in historical control B6C3F 1 mice.) Based on water 3617 consumption and BW measurements, mean intakes were reported to be 61, 124, 177, and 3618 234 mg/kg-day for males and 59, 118, 172, and 238 m g/kg-day for females. Endpoints 3619 examined included clinical signs, BW and water cons umption, hematology at weeks 52 and 104, 3620 and gross and microscopic examinations of tissues a nd organs at termination. All tissues from 3621 the control and 250 mg/kg-day groups were examined microscopically, as well as the livers and 3622 neoplasms from all groups and the eyes of all males from all groups. 3623 Throughout the 2-year study, mice from both control and treatment groups exhibit ed a 3624 high incidence of convulsions (Serota et al., 1986b ; Hazelton Laboratories, 1983). The 3625 convulsions were noted only during handling for BW determinations, and efforts to establish a 3626100

basis for this response were unsuccessful. The inc idence of convulsions did not correlate with an 3627 increased mortality rate. Survival to 104 weeks wa s high (82% in males and 78% in females), 3628 and no evidence for exposure-related negative effec ts on survival were found. Exposure had no 3629 significant effect on BW or water consumption. Mea n leukocyte count was significantly 3630 elevated in males and females dosed with 250 mg/kg- day dichloromethane for 52 weeks, but the 3631

authors indicated that the mean values were within the normal historical range for the laboratory. 3632 Treatment-related nonproliferative histopathologic effects were restricted to the liver and 3633 consisted of a marginal increase in the amount of O il Red O-positive material in the liver of 3634 males and females dosed with 250 mg/kg-day (group i ncidences for this lesion, however, were 3635 not presented in the published report). The result s indicate that 185 mg/kg-day was a NOAEL 3636 and 250 mg/kg-day was a LOAEL for marginally increa sed amounts of fat in livers of male and 3637

female B6C3F 1 mice. 3638 Incidences for proliferative hepatocellular changes in female mice were not presented in 3639 the published reports (Serota et al., 1986b; Hazelt on Laboratories, 1983), but it was reported that 3640 exposed female mice did not show increased incidenc es of proliferative hepatocellular lesions. 3641

In the male B6C3F 1 mice, incidences for hepatic focal hyperplasia sho wed no evidence of an 3642 exposure-related effect (Table 4-15). The authors interpret the data regarding adenomas alone or 3643 carcinomas alone as showing no significantly elevat ed incidence compared with controls. The 3644

trend tests for each of these outcomes were 0.172 a nd 0.147 (Hazelton Laboratories, 1983), 3645 respectively, and none of the comparisons between i ndividual exposure groups and the controls 3646 was statistically significant at the chosen Bonferr oni-corrected level of <0.01. However, 3647 exposed male mice showed increased combined inciden ces of hepatocellular adenomas and 3648 carcinomas, with a linear trend p -value = 0.058 and individual p- values of <0.05. 3649 3650 Table 4-15. Incidences for focal hyperplasia and t umors in the liver of male

B6C3F 1 mice exposed to dichloromethane in drinking water for 2 years

Target dose (mg/kg-day) Controls 0 a 60 125 185 250 Trend

p- value b

n per group c 125 200 100 99 125 Estimated mean intake (mg/kg-day) 0 61 124 17 7 234 Number (%) with Focal hyperplasia d 10 (8) 14 (7) 4 (4) 10 (10) 13 (10) not reported Hepatocellular adenoma mortality-adjusted percent and p- value e 10 (8) (9) 20 (10)

(12)

p = 0.24 14 (14)

(17)

p = 0.064 14 (14)

(16)

p = 0.076 15 (12)

(12)

p = 0.13 0.172

Hepatocellular carcinoma mortality-adjusted percent and p- value e 14 (11) (13) 33 (17)

(19)

p = 0.082 18 (18)

(21)

p = 0.073 17 (17)

(19)

p = 0.11 23 (18)

(21)

p = 0.044 0.147

Hepatocellular adenoma or carcinoma mortality-adjusted percent and p- value e 24 (19) (21) 51 (26)

(29)

p = 0.071 30 (30)

(34)

p = 0.023 31 (31)

(34)

p = 0.019 35 (28)

(32) 0.036 0.058

3651101 a Two control groups combined. 3652

b Cochran-Armitage trend test (source: Hazelton Labo ratories [1983]) 3653

c Number at start of treatment. 3654

d Some mice with hyperplasia also had hepatocellular neoplasms, but the exact number was unspecified by Serota et 3655 al. (1986b). 3656

e Percent calculated based on number at risk, using K aplan-Meier estimation, taking into account mortali ty losses; p3657 value for comparison with control group, using asym ptotic normal test (source: Hazelton Laboratories [1983]). 3658 3659 Sources: Serota et al. (1986b); Hazelton Laborator ies (1983). 3660 3661 3662 Serota et al. (1986b) noted, in summary, that sligh t increases in proliferative 3663 hepatocellular lesions were found in exposed male b ut not female B6C3F 1 mice and that the 3664

increases were not dose related and were within the range of historical control incidences. The 3665 average incidence of hepatocellular adenomas and ca rcinomas in 354 control male B6C3F 1 mice 3666 in the laboratory in which the experiment was perfo rmed was 17.8% with a range of 5–40% 3667 (Serota et al., 1986b). Serota et al. (1986b) conc luded that dichloromethane “did not induce a 3668 treatment-related carcinogenic response in B6C3F 1 mice” under the conditions of this study. An 3669 alternative conclusion, as determined by EPA, is th at dichloromethane induced a carcinogenic 3670 response in male B6C3F 1 mice as evidenced by small, but statistically sign ificant, increases in 3671 hepatocellular adenomas and carcinomas at dose leve ls of 125, 185, and 250 mg/kg-day but not 3672 at 60 mg/kg-day and by a marginally increased trend test for combined hepatocellular adenomas 3673 and carcinomas. Results for the highest dose group (no effect on BW, histologic findings 3674 restricted to mild histologic changes in the liver [vacuolation], and a slight, but statistically 3675 significant, increase in incidence in liver tumors in males only) indicate that this mouse study 3676 may not have included the maximum tolerated dose. 3677

3678

4.2.1.2.3. Chronic oral exposure in Sprague-Dawley rats and Sw iss mice (Maltoni et al., 3679 1988). Maltoni et al. (1988) conducted gavage carcinogen icity studies in Sprague-Dawley rats 3680 and in Swiss mice. Groups of rats (50/sex/dose lev el) were gavaged with dichloromethane 3681 (99.9% pure) in olive oil at dose levels of 0 (oliv e oil), 100, or 500 mg/kg-day, 4–5 days/week 3682 for 64 weeks. This dosing regime was also used for groups of Swiss mice (50/sex/dose level 3683 plus 60/sex as controls). Endpoints monitored incl uded clinical signs, BW, and full necropsy at 3684 sacrifice (when spontaneous death occurred). For e ach animal sacrificed, histopathologic 3685 examinations were performed on the following organs : brain and cerebellum, zymbal glands, 3686 interscapular brown fat, salivary glands, tongue, t hymus and mediastinal lymph nodes, lungs, 3687 liver, kidneys, adrenals, spleen, pancreas, esophag us, stomach, intestine, bladder, uterus, gonads, 3688 and any other organs with gross lesions. High mort ality was observed in male and female high3689 dose rats (data not shown) and achieved significanc e ( p < 0.01) in males. The increased 3690 mortality became evident after 36 weeks of treatmen t and led to the termination of treatment at 3691 week 64. Explanation of the mortality was not prov ided by the study authors. As with the rats, 3692102

high mortality occurred in male and female mice fro m the high-dose group ( p < 0.01), and the 3693 exposure was terminated after 64 weeks. 3694 Little information is provided regarding nonneopla stic effects (Maltoni et al., 1988). 3695

Treatment with dichloromethane did not affect BW in the Sprague-Dawley rats. A reduction in 3696 BW was apparent in treated mice after 36–40 weeks o f treatment, but no data were shown to 3697 determine the magnitude of the effect. The lack of reporting of nonneoplastic findings from the 3698 histopathologic examinations precludes assigning NO AELs and LOAELs for possible 3699 nonneoplastic effects in these studies 3700 The Maltoni et al. (1988) studies of Sprague-Dawle y rats and Swiss mice did not find 3701 distinct exposure-related carcinogenic responses fo llowing gavage exposure to dichloromethane 3702 at dose levels up to 500 mg/kg-day, although the ea rly termination of the study (at 64 weeks) 3703 limits the interpretation of this finding. Dichlor omethane exposure was not related to the 3704 percentage of either study animal bearing benign an d malignant tumors or bearing malignant 3705 tumors or to the number of total malignant tumors p er 100 animals. High-dose female rats 3706 showed an increased incidence in malignant mammary tumors, mainly due to adenocarcinomas 3707 (8, 6, and 18% in the control, 100, and 500 mg/kg d ose groups, respectively; the number of 3708 animals examined was not provided), but the increas e was not statistically significant. A dose3709 related increase, although not statistically signif icant, in pulmonary adenomas was observed in 3710 male mice (5, 12, and 18% in control, 100, and 500 mg/kg-day groups, respectively). When 3711 mortality was taken into account, high-dose male mi ce that died in the period ranging from 52 to 3712 78 weeks were reported to show a statistically sign ificantly ( p < 0.05) elevated incidence for 3713 pulmonary tumors (1/14, 4/21, and 7/24 in control, 100, and 500 mg/kg-day groups, 3714 respectively). Details of this analysis were not p rovided. EPA applied a Fisher’s exact test to 3715

these incidences and determined a p- value of 0.11 for the comparison of the 500 mg/kg-d ay 3716 group (7/24) to the controls (1/14). 3717

3718 4.2.2. Inhalation Exposure: Overview of Noncancer and Cancer Effects 3719 Inhalation dichloromethane exposure studies in rats and mice, using subchronic and 3720 chronic durations, identify the CNS, liver, and lun gs as potential toxicity targets. Data from 3721

other studies indicate that hamsters are less susce ptible to the nonneoplastic and neoplastic 3722

effects of dichloromethane than are rats and mice. 3723 Increased incidences of nonneoplastic liver lesions were observed in Sprague-Dawley 3724 rats exposed to ≥ 500 ppm for 2 years (Nitschke et al., 1988a; Burek et al., 1984), F344 rats 3725 exposed to concentrations ≥ 1,000 ppm for 2 years (Mennear et al., 1988; NTP, 1 986), and 3726

B6C3F 1 mice exposed to ≥ 2,000 ppm for 2 years (Mennear et al., 1988; NTP, 1 986). 3727 Two-year inhalation exposure studies at concentrati ons of 2,000 or 4,000 ppm 3728 dichloromethane produced increased incidences of lu ng and liver tumors in B6C3F 1 mice 3729 (Mennear et al., 1988; NTP, 1986). Additional stud ies examining mechanistic issues regarding 3730103

this effect are described in sections 4.5.2 and 4.5 .3 (Maronpot et al., 1995; Foley et al., 1993; 3731 Kari et al., 1993). Significantly increased incide nces of benign mammary tumors (primarily 3732 fibroadenomas) were also observed in male and femal e F344/N rats exposed by inhalation to 3733 2,000 or 4,000 ppm for 2 years (Mennear et al., 198 8; NTP, 1986). In the male rats, the 3734 incidence of fibromas or sarcomas originating from the subcutaneous tissue around the 3735 mammary gland was also increased, but the differenc e was not statistically significant. In other 3736 studies in Sprague-Dawley rats with exposures of 50 –500 ppm (Nitschke et al., 1988a) and 500– 3737 3,500 ppm (Burek et al., 1984), the incidence of be nign mammary tumors was not increased, but 3738 in females the number of tumors per tumor-bearing r at increased at the higher dose levels. 3739 No obvious clinical signs of neurological impairmen t were observed in the 2-year 3740 bioassays involving exposure concentrations up to 2 ,000 ppm in F344 rats (Mennear et al., 1988; 3741 NTP, 1986) or 3,500 ppm in Sprague-Dawley rats (Nit schke et al., 1988a; Burek et al., 1984). In 3742

B6C3F 1 mice exposed to 4,000 ppm in B6C3F 1, there was some evidence of hyperactivity during 3743 the first year of the study and lethargy during the second year, with female mice appearing to be 3744 more sensitive (Mennear et al., 1988; NTP, 1986). Studies that evaluated batteries of 3745 neurobehavioral endpoints following subchronic or c hronic inhalation exposure are restricted to 3746

one in which no effects were observed more than 64 hours postexposure in an observational 3747 battery, a test of hind-limb grip strength, a batte ry of evoked potentials, or histologic 3748 examinations of brain, spinal cord, or peripheral n erves in F344 rats exposed to concentrations 3749 up to 2,000 ppm for 13 weeks (Mattsson et al., 1990 ) (see section 4.2.3) 3750 No effects on reproductive performance were found i n a two-generation reproductive 3751 toxicity study with F344 rats exposed to concentrat ions up to 1,500 ppm for 14 and 17 weeks 3752 before mating of the F0 and F1 generations, respect ively (Nitschke et al., 1988b) (described 3753 more completely in section 4.3). Developmental eff ects following exposure of Long-Evans rats 3754 to 4,500 ppm for 14 days prior to mating and during gestation (or during gestation alone) 3755 included decreased offspring weight at birth and ch anged behavioral habituation of the offspring 3756 to novel environments (Bornschein et al., 1980; Har din and Manson, 1980) (see section 4.3 for 3757 more details). In standard developmental toxicity studies involving exposure to 1,250 ppm on 3758 GDs 6–15, no adverse effects on fetal development w ere found in Swiss-Webster mice or 3759 Sprague-Dawley rats (Schwetz et al., 1975) (see sec tion 4.3). 3760

3761

4.2.2.1. Toxicity Studies of Subchronic Inhalation Exposures : General, Renal, and Hepatic 3762 Effects 3763 Data pertaining to general (e.g., BW, mortality), h epatic, and renal effects from several 3764 inhalation exposure studies in various species, wit h exposure periods of 3–6 months, are 3765 described below. (Studies providing detailed neuro logical data are described separately in 3766 section 4.4.3.) The earliest study involved severa l different species, with exposures of 3767 5,000 ppm for up to 6 months (Heppel et al., 1944). Two 14-week studies in dogs, monkeys, 3768104

rats, and mice were conducted with exposures at 0, 1,000, and 5,000 ppm (Haun et al., 1972, 3769 1971; Weinstein et al., 1972) and at 0, 25, and 100 ppm (Haun et al., 1972). Neurological effects 3770 and hepatic degeneration were seen at the 1,000 ppm dose. In the lower-dose portion of the 3771 Haun et al. (1972) study in mice, decreased cytochr ome P-450 levels in liver microsomes and 3772 some histopathologic liver changes (fat stains and cytoplasmic vacuolation) were seen at 3773 100 ppm but more obvious adverse effects were not o bserved. Leuschner et al. (1984) reported 3774 data from a high exposure (10,000 ppm) 90-day study of rats; beagle dogs were also included in 3775 this study, at an exposure level of 5,000 ppm. No evidence of toxicity was reported by the 3776 authors of this study. In a 13-week exposure study conducted by NTP (1986), decreased BWs 3777 and increased incidence of foreign body pneumonia w ere seen at 8,400 ppm in F344 rats, and 3778 histologic changes in the liver in B6C3F1 mice were seen at 4,200 ppm. 3779 The first experimental study of dichloromethane ex posure included dogs, rabbits, guinea 3780 pigs, and rats, with an exposure of approximately 5 ,000 ppm for 7 hours/day, 5 days per week 3781 for up to 6 months (Heppel et al., 1944). The stra ins of the animals, the comparability between 3782 exposed and unexposed group (in terms of sex distri bution and other attributes), and process by 3783 which animals were chosen for histologic examinatio n are not clearly described in the report. 3784 Exposed animals included adult dogs (1 male and 5 f emales), juvenile dogs (1 male and 1 female 3785 born in the exposure chamber and exposed daily from birth), adult rabbits (2 males and 3786 2 females), guinea pigs (14 males), and rats (15 ma les and 6 females). The nonexposed control 3787 group included 14 guinea pigs, 28 rats, 4 rabbits, and an unspecified number of dogs. Exposure 3788 produced no significant effects on BWs except in th e guinea pigs; after 131 exposures, average 3789 BWs were 0.820 and 1.025 kg for exposed and control guinea pigs, respectively. Three exposed 3790 guinea pigs died after 35, 90, and 96 exposures. N o other deaths occurred, except for one 3791 exposed female rat that died after 22 exposures and giving birth to a litter. Autopsy showed 3792

thrombi in the renal vessels associated with marked cortical infarction. No adverse clinical signs 3793 of toxicity (such as decreased activity or incoordi nation) were observed in exposed animals 3794 during the study. Urinalysis, hematology tests, an d tests of liver function performed on dogs 3795 during the study showed no treatment-related effect s. At termination, gross and microscopic 3796 examination of the major organs showed no pathologi cal changes after exposure to 5,000 ppm 3797 dichloromethane, with the exception that two of the exposed guinea pigs that died showed 3798 extensive pneumonia associated with moderate centri lobular fatty degeneration of the liver. The 3799 results indicate that 5,000 ppm was a NOAEL for non neoplastic systemic effects in dogs, rabbits, 3800 and rats exposed 7 hours/day, 5 days/week for up to 6 months. The findings of three deaths (two 3801 with pulmonary congestion and centrilobular fatty d egeneration) and 20% decreased average 3802 BW among the 14 exposed guinea pigs indicates that 5,000 ppm was a LOAEL in this species. 3803 Haun et al. (1972, 1971) and Weinstein et al. (1972 ) reported results from studies in 3804 which groups of 8 female beagle dogs, 4 female rhes us monkeys, 20 male Sprague-Dawley rats, 3805 and 380 female ICR mice were continuously exposed t o 0, 1,000, or 5,000 ppm dichloromethane 3806105

for up to 14 weeks in whole-body exposure chambers. Gross and histopathologic examinations 3807

were scheduled to be made on animals that died or w ere sacrificed during or at termination of the 3808 study. At 5,000 ppm, obvious nervous system effect s (e.g., incoordination, lethargy) were 3809 observed in dogs, monkeys, and mice. At 1,000 ppm, these effects were most apparent in dogs 3810 and monkeys (Haun et al., 1971). Food consumption was reduced in all species at 5,000 ppm 3811 and in dogs and monkeys at 1,000 ppm. All exposed animals either lost weight or showed 3812 markedly decreased BW gains compared with controls. For example, rats exposed to 1,000 or 3813 5,000 ppm for 14 weeks showed average BWs that were roughly 10 and 20% lower than control 3814 values. Significant numbers of dogs (4) and mice ( 123), as well as 1 monkey, died within the 3815 first 3 weeks of exposure to 5,000 ppm. Because of this high mortality, all surviving 5,000 ppm 3816 animals were sacrificed at 4 weeks of exposure, exc ept for one half (10) of the rats that went on 3817 to survive the 14-week exposure period. At 1,000 p pm, six of eight dogs died by 7 weeks, at 3818 which time the remaining two were sacrificed. Monk eys, rats, and all but a few mice survived 3819 exposure to 1,000 ppm for 14 weeks. 3820 Gross examination of tissues showed yellow, fatty l ivers in dogs that died during 3821 exposure to 1,000 or 5,000 ppm, “borderline” liver changes in 3 monkeys exposed to 5,000 ppm, 3822 and mottled liver changes in 4/10 rats exposed to 5 ,000 ppm for 14 weeks (Haun et al., 1971). 3823 Comprehensive reporting of the histologic findings from this study were not available, but Haun 3824 et al. (1972) reported that the primary target orga n was the liver and that in some exposed 3825 animals the kidney was also affected. Light and el ectron microscopy of liver sections from 3826 groups of 4–10 mice sacrificed after 1, 4, 8, and 1 2 hours and 1, 2, 3, 4, 6, and 7 days of 3827 exposure to 5,000 ppm showed hepatocytes with ballo on degeneration (dissociation of 3828 polyribosomes and swelling of rough endoplasmic ret iculum) as early as 12 hours of exposure 3829 (Weinstein et al., 1972). The degeneration peaked in severity after 2 days of exposure and, 3830 subsequently, partially reversed in severity. Info rmation on possible histopathologic changes in 3831 mice exposed to 1,000 ppm was not provided. 3832 The results from this study demonstrate that dogs a nd mice were more sensitive than 3833 were rats and monkeys to lethal effects, nervous sy stem depression, and possibly liver effects 3834 from continuous exposure to 1,000 or 5,000 ppm. Th e results indicate that continuous exposure 3835 to 1,000 ppm was an adverse effect level for mortal ity and effects on the nervous system and 3836 liver in dogs (exposed for up to 4 weeks) and for B W changes in rats (exposed for 14 weeks). 3837 The 5,000 ppm level induced mortality in beagle dog s, ICR mice, and rhesus monkeys (but not 3838 in Sprague-Dawley rats); obvious nervous system eff ects in dogs, mice, monkeys, and rats; and 3839 gross liver changes in dogs, mice, monkeys, and rat s. 3840 Haun et al. (1972) also conducted studies with grou ps of 20 mice, 20 rats, 16 dogs, and 3841 4 monkeys exposed continuously to 0, 25, or 100 ppm dichloromethane for 100 days (14 weeks). 3842 The animals presumably were of the same strains and sexes as those used in the studies involving 3843 exposure to 1,000 or 5,000 ppm dichloromethane (Hau n et al., 1972, 1971; Weinstein et al., 3844106

1972). All animals underwent necropsy and histopat hologic evaluation at termination of the 3845 exposure, but a list of the tissues examined and in cidence or severity data were not presented in 3846 the report. Hematology and clinical chemistry vari ables (including COHb levels) were measured 3847 in blood samples collected from dogs and monkeys at biweekly or monthly intervals during 3848 exposure. COHb levels were elevated in a dose-rela ted manner in monkeys and peaked at about 3849 5% (approximately 0.8% pre-exposure) after 6 weeks of exposure. COHb levels in dogs were 3850 unaffected by the 25 ppm exposure level and rose to about 2% (from about 0.6%) from 3851 week 4 on in high-dose dogs. Additional groups of mice were included for assessment of 3852 hexobarbital sleep times at monthly intervals; leve ls of cytochromes P-450, P-420, and b 5 in liver 3853 microsomes at monthly intervals; and spontaneous ph ysical activity at several intervals during 3854 the study. 3855 No clinical signs of toxicity or alterations in wei ght gain were seen in any of the species 3856 examined. In dogs and monkeys, hematology and clin ical chemistry results throughout the study 3857 and at termination were unremarkable, as were the r esults of the gross and histopathologic 3858 examinations. In mice exposed to 100 ppm, CYP leve ls in liver microsomes were significantly 3859 decreased (compared with control values) after 30, 60, and 90 days of exposure to 100 ppm, 3860 whereas levels of cytochrome b 5 and P-420 decreased after 30 days and increased aft er 90 days 3861 of exposure. At 25 ppm, no significant differences from control were seen in mouse liver levels 3862 of cytochromes. Mice exposed to 25 ppm showed no h istopathologic changes, while histologic 3863 changes in mice at 100 ppm were restricted to posit ive fat stains and some cytoplasmic 3864 vacuolation in the liver. In rats at both exposure levels, the livers showed positive staining for 3865 increased fat, and the kidneys showed evidence of n onspecific tubular degenerative and 3866 regenerative changes. Haun et al. (1972) indicate that no distinctively adverse effects were 3867 found in monkeys, dogs, rats, or mice continuously exposed to 25 or 100 ppm for up to 3868

14 weeks. Decreased CYP levels in liver microsomes and some histopathologic liver changes 3869 (fat stains and cytoplasmic vacuolation) were seen at the 100 ppm dose. 3870 Leuschner et al. (1984) exposed Sprague-Dawley rats (20/sex/dose level) to 0 or 3871 10,000 ppm and beagle dogs (3/sex/dose level) to 0 or 5,000 ppm dichloromethane in whole3872 body exposure chambers. Exposure periods were 6 ho urs/day for 90 consecutive days. 3873 Endpoints evaluated in both species included clinic al signs, food and water consumption, BW, 3874 hematology, clinical chemistry, urinalysis, and gro ss and microscopic evaluation of 27 organs at 3875 termination. Electrocardiography and blood pressur e measurements were also done in dogs. 3876 The only significant effect observed in rats was a slight redness of the conjunctiva 3877 1–10 hours after each exposure. In dogs, compound- related effects were restricted to slight 3878 sedation throughout the exposure period and slight erythema lasting up to 10 hours after 3879 exposure. In this 90-day study involving daily 6-h our exposures, 10,000 and 5,000 ppm were 3880 NOAELs for behavioral, clinical chemistry, hematolo gic, and histologic signs of toxicity in 3881 Sprague-Dawley rats and beagle dogs, respectively. 3882107

NTP (1986) exposed groups of F344 rats and B6C3F 1 mice (10/sex/dose level) to target 3883 concentrations of 0, 525, 1,050, 2,100, 4,200, or 8 ,400 ppm dichloromethane, 6 hours/day, 3884 5 days/week for 13 weeks in whole-body exposure cha mbers. Endpoints monitored included 3885 clinical signs, BW, and necropsy at termination. C omprehensive sets of tissues and organs in 3886 control and high-dose animals were histologically e xamined; tissues from the lower dose groups 3887

were examined to determine the no-observed-effect l evel. One male and one female rat from the 3888 8,400 ppm exposure group died before the end of the study, but the cause of death was not 3889 discussed. The final mean BWs of 8,400 ppm male an d female rats were reduced by 23 and 3890 11%, respectively, relative to controls. Foreign-b ody pneumonia was present in 4/10 male and 3891 6/10 female rats exposed to 8,400 ppm and in 1/10 f emale rat from the 4,200 ppm exposure 3892 group. The liver lipid/liver weight ratios for 8,4 00 ppm rats of both sexes and 4,200 ppm female 3893 rats were significantly lower than in controls. In mice, 4/10 males and 2/10 females exposed to 3894 8,400 ppm died before the end of the study, and the se deaths were considered treatment related. 3895 Histologic changes in exposed mice consisted of hep atic centrilobular hydropic degeneration (of 3896 minimal to mild severity) in 3/10 males and 8/10 fe males at 8,400 ppm and in 9/10 females from 3897 the 4,200 ppm exposure group. Histologic changes i n the 2,100 ppm mouse group were not 3898 mentioned. The liver lipid/liver weight ratio for the high-dose female mice was significantly 3899 lower than in controls. In this 13-week study invo lving 6-hour exposure periods for 3900 5 days/week, 4,200 ppm was a NOAEL and 8,400 ppm wa s a LOAEL for decreased BWs and 3901 increased incidence of foreign-body pneumonia in F3 44 rats. In B6C3F 1 mice, 2,100 ppm was a 3902 NOAEL and 4,200 ppm was a LOAEL for histologic chan ges in the liver. 3903

3904 4.2.2.2. Toxicity Studies from Chronic Inhalation Exposures 3905 Chronic inhalation exposure studies are summarized in Table 4-16. Details of each study 3906 are described below, with the results pertaining to nonneoplastic and neoplastic effects 3907 summarized in the following sections. 3908

3909108 3910 Table 4-16. Studies of chronic inhalation dichloro methane exposures

Reference, strain/species Number per group Exposure information Comments Mennear et al. (1988); NTP (1986) F344 rats 50/sex/dose 2 years, 6 hours/day, 5 days/week 0, 1,000, 2,000, 4,000 ppm Nonneoplastic liver effects and hemosiderosis in males and females (see Table 4-17) Weak trend for neoplastic nodule or hepatocellular carcinoma in females, benign mammary tumors in males and females (see Table 4-18) Mennear et al. (1988); NTP (1986) B6C3F 1 mice 50/sex/dose 2 years, 6 hours/day, 5 days/week 0, 2,000, 4,000 ppm Varied nonneoplastic effects (see Table 4-19) Liver and lung tumors (adenomas or carcinomas) in males and females (see Table 4-20) Burek et al. (1984) Syrian hamsters 95/sex/dose 2 years, 6 hours/day, 5 days/week 0, 500, 1,500, 3,500 ppm Decreased mortality Increased CoHb at 500 ppm (see section 4.2.2.2.3) Burek et al. (1984) Sprague-Dawley rats 92–97/sex/dose 2 years, 6 hours/day, 5 days/week 0, 500, 1,500, 3,500 ppm Nonneoplastic liver effects in males and females (see Table 4-21) Increased CoHb at 500 ppm Increased number of benign mammary tumors per tumor bearing rat (females) (see Table 4-21) Nitschke et al. (1988a) Sprague-Dawley rats 90/dose/sex 2 years, 6 hours/day, 5 days/week 0, 50, 200, 500 ppm Nonneoplastic liver effects in males and females (statistically significant in females) (seeTable 4-22) Increased CoHb at 50 ppm Increased number of benign mammary tumors per animal in females (see Table 4-23) Maltoni et al. (1988)

Sprague-Dawley rats, female 54–60/dose 2 years, 4 hours/day, 5 days/week for 7 weeks; 7 hours/day, 5 days/week for 97 weeks 0, 100 ppm No effects seen on total number of benign or malignant cancers

3911 3912 4.2.2.2.1. Chronic inhalation exposure in F344/N rats (Mennear et al., 1988; NTP, 1986). 3913 NTP conducted a 2-year inhalation exposure study in F344/N rats (Mennear et al., 1988; NTP, 3914 1986). The rats (50/sex/exposure level) were expos ed to dichloromethane (>99% pure) by 3915 inhalation in exposure chambers, 6 hours/day, 5 day s/week for 2 years. Exposure concentrations 3916 were 0, 1,000, 2,000, or 4,000 ppm. Mean daily con centrations never exceeded 110% of target 3917 and were <90% of target in only 23 of 1,476 analyse s. Endpoints monitored included clinical 3918 signs, mortality, and gross and microscopic examina tions of 32 tissues at study termination. 3919 Clinical examinations were conducted weekly for 3.5 months and biweekly until month 8. After 3920 8 months, the animals were clinically examined and palpated for tumors and masses monthly 3921 until the end of the study. 3922 Dichloromethane exposure did not significantly alt er BW gain or terminal BWs 3923 (Mennear et al., 1988; NTP, 1986). Survival of mal e rats was low in all exposed groups and the 3924109

control group, and no significant exposure-related differences were apparent. Most deaths 3925 occurred during the last 16 weeks of the study. Su rvival at week 86 was 36/50, 39/50, 37/50, and 3926 33/50 for the control, 1,000, 2,000, and 4,000 ppm groups, respectively. In female rats, there 3927 was a trend towards decreased survival, and the sur vival of high-dose female rats was 3928 significantly reduced, possibly due to leukemia. S urvival in the females at 86 weeks was 30/50, 3929 22/50, 22/50, and 15/50 for the control, 1,000, 2,0 00, and 4,000 ppm groups, respectively. 3930 Nonneoplastic lesions with statistically significan tly elevated incidences, compared with 3931 controls, included hepatocyte cytoplasmic vacuolati on and necrosis and liver hemosiderosis in 3932 males and females; renal tubular cell degeneration in males and females; splenic fibrosis in 3933 males; and nasal cavity squamous metaplasia in fema les (Table 4-17). The results indicate that 3934 1,000 ppm (6 hours/day, 5 days/week) was a LOAEL fo r nonneoplastic liver changes 3935 (hepatocyte cytoplasmic vacuolation and necrosis, h epatic hemosiderosis) in male and female 3936

F344/N rats. A NOAEL was not established because e ffects were observed at the lowest dose. 3937110 3938 Table 4-17. Incidences of nonneoplastic histologic changes in male and

female F344/N rats exposed to dichloromethane by in halation (6 hours/day, 5 days/week) for 2 years

Exposure (ppm) a

Lesion, by sex Controls 0 1,000 2,000 4,000 Males

n per group b 50 50 50 50 Number (%) c with Liver changes Hepatocyte cytoplasmic vacuolation 8 (16) 26 (53) d 22 (44) d 25 (50) d Hepatocyte focal necrosis 7 (14) 23 (47) d 6 (12) 16 (32) d Hepatocytomegaly 2 (4) 10 (20) 6 (12) 5 (10) Hemosiderosis 8 (16) 29 (59) d 37 (74) d 42 (84) d Bile duct fibrosis 8 (16) 10 (20) 17 (3 4) 23 (46) d Renal tubular cell degeneration 11 (22) 13 ( 26) 23 (46) d 10 (20) d Splenic fibrosis 2 (4) 6 (12) 11 (22 ) d 8 (16) d

Females

n per group e 50 50 50 50 Number (%) c with Liver changes Hepatocyte cytoplasmic vacuolation 10 (20) 43 (86) d 44 (88) d 43 (86) d Hepatocyte focal necrosis 2 (4) 32 (64) d 19 (38) d 9 (18) d Hepatocytomegaly 3 (6) 10 (20) d 18 (36) d 5 (10) Hemosiderosis 19 (38) 29 (58) d 38 (76) d 45 (90) d Bile duct fibrosis 4 (8) 3 (6) 10 (20) d 3 (6) Renal tubular cell degeneration 14 (28) 20 (4 0) 22 (44) 25 (51) d Splenic fibrosis 0 (0) 2 (4) 4 ( 8) 4 (8) Nasal cavity squamous metaplasia 1 (2) 2 (4) 3 (6) 9 (18) d

a 1,000 ppm = 3,474 mg/m 3 , 2,000 ppm = 6,947 mg/m 3 , 4,000 ppm = 13,894 mg/m 3

. b Number of male rats necropsied per group; only 49 1 ,000 ppm livers were examined microscopically.

c Percentages were based on the number of tissues exa mined microscopically per group.

d Statistical significance not reported in publicatio ns but significantly ( p ≤ 0.05) different from control as calculated by Fisher’s exact test.

e Number of females necropsied per group; only 49 4,0 00 ppm kidneys and spleens were examined microscopi cally.

Sources: Mennear et al. (1988); NTP (1986, Appendi x B, Tables C1 and C2). 3939 Incidences of mammary fibroadenomas were significa ntly increased in 4,000 ppm males 3940 and 2,000 and 4,000 ppm females, compared with cont rols (Table 4-18). Similar patterns were 3941

seen with the combination of fibroadenomas and aden omas (not shown in Table 4-18). In males, 3942

subcutaneous tissue fibroma or sarcoma was seen in 1/50, 1/50, 2/50, and 5/50 rats in the 0, 3943 1,000, 2,000, and 4,000 ppm groups, respectively, b ut these lesions were not seen in females. 3944 Incidences of female rats with liver neoplastic nod ules or carcinomas (combined) showed a 3945 significant trend test after survival adjustment on ly, but the incidences at the two highest dose 3946 levels were not significantly increased relative to the control (Table 4-18). 3947

Incidences for mononuclear cell leukemias in mid- and high-dose female rats were 3948 statistically significant after a survival-adjustme nt analysis. However, Mennear et al. (1988) 3949 considered the relationship between exposure to dic hloromethane and mononuclear cell leukemia 3950111

to be equivocal, based on the fact that most male r ats had leukemia (34/50, 26/50, 32/50, and 3951 35/50 in controls, 1,000, 2,000, and 4,000 ppm rats , respectively). Other neoplasms that had 3952 increased incidences included mesotheliomas (predom inantly in the tunica vaginalis) in males 3953 (0/50, 2/50, 5/50, and 4/50 in controls, 1,000, 2,0 00, and 4,000 ppm rats, respectively). This 3954

lesion was not considered to be related to dichloro methane exposure, because the concurrent 3955 control incidence (0/50) for this neoplasm was low relative to earlier inhalation studies 3956 conducted at this laboratory (4/100, 4%) and in oth er NTP studies with male F344/N rats 3957 (44/1,727) (mean historical percentage across NTP s tudies = 3 ± 2%). 3958 NTP (1986) concluded that there was “some evidence of carcinogenicity of 3959 dichloromethane” in male F344/N rats as shown by in creased incidence of benign mammary 3960 gland tumors and “clear evidence of carcinogenicity ” of dichloromethane in female F344/N rats 3961 as shown by increased incidence of benign mammary g land tumors. The summary of the hepatic 3962 effects in rats in the NTP (1986) report also notes the positive trend in the incidence of 3963 hepatocellular neoplastic nodules or carcinomas in females, which “may have been due to 3964 dichloromethane exposure.” 3965

3966

3967 3968112Table 4-18. Incidences of selected neoplastic lesions in male and female F344/N rats exposed todichloromethane by inhalation (6 hours/day, 5 days/week) for 2 yearsExposure (ppm)a0 (Controls) 1,000 2,000 4,000Neoplastic lesion, by sex n(%)b(%)cn(%)b(%)cn(%)b(%)cn(%)b(%)cTrendp-valuedMalesn per group50505050Liver—Neoplastic nodule or hepatocellular carcinoma2(4)(10)3(6)(13)4(8)(19)1(2)(6)0.55Liver—hepatocellular carcinoma 2(4)(10)1(2)(4)2(4)(10)1(2)(6)nrLung—Bronchoalveolar adenoma or carcinoma 11 (2)2 (4)1(2)Mammary glandAdenoma, adenocarcinoma, or carcinomaSubcutaneous tissue fibroma or sarcomaFibroadenomaMammary gland or subcutaneous tissue adenoma,fibroadenoma, fibroma, or sarcoma0101(0)(2)(0)(2)(6)(0)(6)0101(0)(2)(0)(2)(6)(0)(6)0224(0)(4)(4)(8)(9)(12)(21)1519e(2)(10)(2)(18)(23)(8)(49)0.008< 0.001<0.001Brain (carcinoma, not otherwise specified, invasive) 0(0)1(2)0 (0)0(0)Femalesn per group50505050Liver—Neoplastic nodule or hepatocellular carcinoma2(4)(7)1(2)(2)4(8)(14)5(10)(20)0.08Liver—hepatocellular carcinoma 0 (0)(0)0 (0)(0)1(2)(4)0 (0)(0)nrLung—Bronchoalveolar adenoma or carcinoma 1(2)1 (2)0(0)0(0)Mammary glandAdenocarcinoma or carcinomaAdenoma, adenocarcinoma, or carcinomaFibroadenomaMammary gland adenoma, fibroadenoma, oradenocarcinoma1156(2)(2)(10)(12)(16)(18)2211e13(4)(4)(22)(26)(41)(44)2213e14e(4)(4)(26)(28)(44)(45)0122e23e(0)(2)(44)(46)(79)(86)<0.001<0.001Brain (carcinoma, not otherwise specified, invasive, andoligodendroglioma)f1(2)0(0)2 (4)0(0)a1,000 ppm = 3,474 mg/m3, 2,000 ppm = 6,947 mg/m3, 4,000 ppm = 13,894 mg/m3.bPercentages based on the number of tissues examinedmicroscopically per group; for males, 49 livers and lungs were examined microscopically in the1,000 ppm groups, and only 49 brains were examinedmicroscopically in the 4,000 ppm group. For comparison, incidence in historical controls reported inNTP (1986) were 1% for female liver tumors and 16%for female mammary fibroadenomas.cMortality-adjusted percentage.dLife-table trend test, as reported by NTP (1986).nr = not reported.eLife-table test comparison dose group with control< 0.05, as reported by NTP (1986).fThe oligodendroglioma occurred in the 2,000 ppm group.Sources: Mennear et al. (1988); NTP (1986, Appendix A and Appendix E, Tables E1 and E2)113

4.2.2.2.2. Chronic inhalation exposure in B6C3F 1 mice (Mennear et al., 1988; NTP, 1986). A 3969 2-year inhalation exposure study in B6C3F 1 mice, similar to that in F344/N rats, was also 3970 conducted by NTP (Mennear et al., 1988; NTP, 1986). The mice (50/sex/exposure level) were 3971 exposed to dichloromethane (>99% pure) by inhalatio n at concentrations of 0, 2,000, or 3972 4,000 ppm in exposure chambers 6 hours/day, 5 days/ week for 2 years. As with the study in rats, 3973 mean daily concentrations in the mice never exceede d 110% of target and were <90% of target in 3974 only 23 of 1,476 analyses. Endpoints monitored inc luded clinical signs, mortality, and gross and 3975 microscopic examinations of 32 tissues at study ter mination. Clinical examinations were 3976 conducted weekly for 3.5 months and biweekly until month 8. After 8 months, the animals were 3977 clinically examined and palpated monthly for tumors and masses until the end of the study. 3978 The BW of 4,000 ppm males was comparable to contro ls until week 90 and 8–11% below 3979 controls thereafter. The BW of 4,000 ppm females w as 0–8% lower than that of controls from 3980 week 51 to 95 and 17% lower at study termination. No information was provided regarding food 3981 consumption during the study. Male and female mice from the high-dose groups (4,000 ppm) 3982 were hyperactive during the first year of the study ; during the second year, high-dose females 3983 appeared lethargic. Exposure was associated with d ecreased survivability of both male and 3984 female mice (males: 39/50, 24/50, and 11/50 and fem ales: 25/50, 25/50, and 8/50 in controls, 3985 2,000 ppm, and 4,000 ppm at 104 weeks, respectively ). In 4,000 ppm mice, statistically 3986 significant incidences of nonneoplastic lesions wer e found in the liver (cytologic degeneration), 3987 testes (atrophy), ovary and uterus (atrophy), kidne ys (tubule casts in males only), stomach 3988 (dilatation), and spleen (splenic follicles in male s only) (Table 4-19). In 2,000 ppm mice, the 3989 only nonneoplastic lesions showing statistically si gnificantly elevated incidences were ovarian 3990 atrophy, renal tubule casts, and hepatocyte degener ation in female mice (Table 4-19). The 3991 results indicate that 2,000 ppm, the lowest exposur e level, was a LOAEL for nonneoplastic 3992 changes in the ovaries, kidneys, and livers of fema le B6C3F 1 mice. A NOAEL was not 3993

established because effects occurred at the lowest exposure level. 3994

3995114 3996 Table 4-19. Incidences of nonneoplastic histologic changes in B6C3F 1

mice exposed to dichloromethane by inhalation (6 ho urs/day, 5 days/week) for 2 years

Exposure (ppm) a

Lesion, by sex Controls 0 2,000 4,000 Males : n per group b 50 50 50 Number (%) c with Liver changes Hepatocyte cytoplasmic vacuolation Hepatocyte focal necrosis Cytologic degeneration Not reported 0 (0) 0 (0) Not reported

0 (0) 0 (0) Not reported

2 (4) 22 (45) d Testicular atrophy 0 (0) 4 (8) 31 (62) d Renal tubule casts 6 (12) 11 (22) 20 (40) d Stomach dilatation 3 (6) 7 (15) 9 (18) d Splenic follicular atrophy 0 (0) 3 (6) 7 (15) d

Females : n per group e 50 50 50 Number (%) c with Liver changes Hepatocyte cytoplasmic vacuolation Hepatocyte focal necrosis Cytologic degeneration Not reported Not reported 0 (0) Not reported

Not reported 23 (48) d Not reported

Not reported 21 (44) d Ovarian atrophy 6 (12) 28 (60) d 32 (74) d Uterus atrophy 0 (0) 1 (2) 8 (17) d Renal tubule casts 8 (16) 23 (48) d 23 (49) d Glandular stomach dilatation 1 (2) 2 (4) 10 (20) d Splenic follicular atrophy 0 (0) 0 (0) 1 (2)

a 2,000 ppm = 6,947 mg/m 3 , 4,000 ppm = 13,894 mg/m 3

. b Number of male mice necropsied per group. The numb er biopsied in the 0, 2,000, and 4,000 ppm dose groups was 50, 49, and 49 for liver; 50, 49, a nd 50 for renal tubules; 49, 47, and 49 for stomach; and 49, 49, and 48 for spleen.

c Percentages were based on the number of tissues exa mined microscopically per group.

d Statistical significance not reported in publicatio ns but significantly different ( p ≤ 0.05) from control as calculated by EPA using Fisher’s exact t est.

e Number of females necropsied per group. The number biopsied in the 0, 2,000, and 4,000 ppm dose groups was 50, 48, and 48 for liver; 50, 47, a nd 43 for ovaries; 50, 48, and 47 for uterus; 49, 48, and 47 for renal tubule; 49, 47, and 48 for sto mach; and 49, 48, and 47 for spleen.

Sources: Mennear et al. (1988); NTP (1986, Appendi x C, Tables D1 and D2). 3997 3998 At both exposure levels, statistically significant ly elevated incidences were found for 3999 hepatocellular adenomas (males only); hepatocellula r carcinomas; hepatocellular adenomas and 4000 carcinomas, combined; bronchoalveolar adenomas; bro nchoalveolar carcinomas; and 4001 bronchoalveolar adenomas and carcinomas (Table 4-20 ). Statistically significant positive trend 4002 tests were found for each of these tumor types in f emale mice. The trend tests were significant 4003

for the liver tumors in male mice after life-table adjustment for reduced survival. The only other 4004 statistically significant carcinogenic response was for increased incidence of hemangiosarcomas 4005 or combined hemangiomas and hemangiosarcomas in mal e mice exposed to 4,000 ppm. NTP 4006115

(1986) concluded that the elevated incidences of li ver and lung tumors provided clear evidence 4007 of carcinogenicity in male and female B6C3F 1 mice. 4008

4009 Table 4-20. Incidences of neoplastic lesions in ma le and female B6C3F 1 mice

exposed to dichloromethane by inhalation (6 hours/d ay, 5 days/week) for 2 years

Exposure (ppm) a 0 (Controls) 2,000 4,000

Neoplastic lesion, by sex n (%) b (%) c n (%) b (%) c n (%) b (%) c Trend

p- valu e d

Males Liver Hepatocellular adenoma 10 (20) (23) 14 (29) (47) 14 (29) (68) 0.19 Hepatocellular 13 (26) (30) 15 (30) (44) 26 e (53) (76) 0.004 Hepatocellular adenoma or carcinoma 22 (44) (48) 24 (49) (67) 33 e (67) (93) 0.013 Lung Bronchoalveolar adenoma 3 (6) (8) 19 e (38) (56) 24 e (48) (79) <0.001 Bronchoalveolar carcinoma 2 (4) (5) 10 e (20) (34) 28 e (56) (93) <0.001 Bronchoalveolar adenoma or carcinoma 5 (10) (12) 27 e (54) (74) 40 e (80) (100) <0.001 Mammary adenocarcinoma f

– – – Hemangioma or hemangiosarcoma, combined 2 (4) (5) 2 (4) (8) 6 (12) (26) 0.08

Females Liver Hepatocellular adenoma 2 (4) (7) 6 (13) (21) 22 e (46) (83) <0.001 Hepatocellular carcinoma 1 (1) (4) 11 (23) (34) 32 e (67) (97) <0.001 Hepatocellular adenoma or carcinoma 3 (6) (10) 16 e (33) (48) 40 e (83) (100) <0.001 Lung Bronchoalveolar adenoma 2 (4) (7) 23 e (48) (58) 28 e (58) (91) <0.001 Bronchoalveolar carcinoma 1 (1) (4) 13 e (27) (46) 29 e (60) (92) <0.001 Bronchoalveolar adenoma or carcinoma 3 (6) (11) 30 e (63) (83) 41 e (85) (100) <0.001 Mammary adenocarcinoma 2 (4) (8) 3 (6) (10) 0 (0) (0) 0.21 Hemangioma or hemangiosarcoma, combined f – – –

a 2,000 ppm = 6,947 mg/m 3 , 4,000 ppm = 13,894 mg/m 3

. b Percentages based on the number of tissues examined microscopically per group; for males, 49 livers we re examined in the 2,000 and 4,000 ppm groups; for females, only 4 8 livers and lungs and 49 mammary glands were micro scopically examined in the 2,000 and 4,000 ppm groups. For co mparison, incidence in historical controls reported in NTP (1986) were 28% for male liver tumors, 31% for male lung t umors, 5% for female liver tumors, and 10% for fema le lung

tumors.

c Mortality-adjusted percentage.

d Life-table trend test, as reported by NTP (1986).

e Life-table test comparison dose group with control < 0.05, as reported by NTP (1986).

f Data not reported.

Sources: Mennear et al. (1988); NTP, (1986, Append ix E, Tables E3 and E4). 4010 4011 4.2.2.2.3. Chronic inhalation exposure in Syrian hamsters (Bur ek et al., 1984). Burek et al. 4012 (1984) conducted a chronic toxicity and carcinogeni city study in rats and hamsters. In the 4013 hamster study, groups of 95 Syrian golden hamsters of each sex were exposed to 0 (filtered air), 4014 500, 1,500, or 3,500 ppm dichloromethane (>99% pure ) under dynamic airflow conditions in 4015 whole-body exposure chambers 6 hours/day, 5 days/we ek for 2 years. Exposure started when the 4016116

animals were approximately 8 weeks of age. Interim sacrifices were conducted at 6, 12, and 4017 18 months. The hamsters were observed daily during exposure days and were palpated monthly 4018 for palpable masses starting the third month of the study. BWs were monitored weekly for the 4019 first 8 weeks of the study and monthly thereafter. Hematologic determinations included packed 4020 cell volume, total erythrocyte counts, total red bl ood cells, differential leukocyte counts, and 4021 hemoglobin concentration. The mean corpuscular vol ume, mean corpuscular hemoglobin, and 4022 MCHC were also determined. A reticulocyte count wa s also performed on all animals at the 18- 4023

month kill and on 10 animals/sex/dose at 24 months. Clinical chemistry determinations included 4024 serum AP and ALT activities, blood urea nitrogen le vels, and total protein and albumin. Urinary 4025 parameters measured were specific gravity, pH, gluc ose, ketones, bilirubin, occult blood, protein, 4026 and urobilinogen. Hematology, clinical chemistries , and urinalysis were performed at interim 4027

sacrifices and at termination. COHb was measured a fter a single 6-hour exposure and following 4028 22 months of exposure. Gross and microscopic exami nations were conducted on all tissues. In 4029 addition, the weights of the brain, heart, liver, k idneys, and testes were recorded. 4030 In the study using Syrian hamsters (Burek et al., 1984), hamsters were exposed to 4031 analytical concentrations of dichloromethane of 510 ± 27, 1,510 ± 62, and 3,472 ± 144 ppm for 4032 the target concentrations of 500, 1,500, and 3,500 ppm, respectively. No exposure-related 4033 clinical signs were observed in the hamsters throug hout the study. Significantly decreased 4034 mortality was observed in females exposed to 3,500 ppm from the 13th through the 24th month 4035 and from the 20th to the 24th month in females expo sed to 1,500 ppm. Exposure to 4036 dichloromethane had no significant effect on BW or on mean organ weights. Regarding 4037 hematology parameters (actual data were not shown), Burek et al. (1984) stated that a few 4038 statistically significant changes occurred, but no obvious pattern could be discerned and most 4039 values were within the expected range for the anima ls. There were no exposure-related 4040 alterations in clinical chemistry or urinalysis val ues. Male and female hamsters in all dose 4041 groups had significantly elevated COHb values after a single 6-hour exposure and after 4042 22 months of exposure, but at both time points ther e was no dose-response relationship above the 4043 first dose level and no apparent significant differ ences in the magnitude of the changes between 4044 the two time points. For example, mean values (± S D) for percentage COHb in male hamsters 4045 after 22 months of exposure were 3.3 (± 3.5), 28.4 (± 5.9), 27.8 (± 2.9), and 30.2 (± 4.9), for the 4046 control through 3,500 ppm groups, respectively. Si milar values were obtained for females at 4047

22 months and for males and females after the first day of exposure. Pathological evaluation of 4048 hamsters showed a lack of evidence of definite targ et organ toxicity. Specific observations 4049 mentioned by the authors included a trend of increa sing hemosiderin in the liver of male 4050 hamsters at 6 and 12 months; decreased amyloid depo sit in organs, such as the liver, kidneys, 4051 adrenal, and thyroid glands in exposed animals; and fewer biliary cysts in the liver. Increased 4052 hepatic hemosiderin at the 12 month sacrifice was o bserved in 1/5, 1/5, 3/5, and 5/5 male 4053 hamsters in the control through 3,500 ppm groups, r espectively. No exposure-related increased 4054117

incidences of hepatic hemosiderin, or other liver e ffects, were reported for the terminal sacrifice. 4055 The exposure-related decreases in geriatric changes (i.e., amyloid deposits and biliary cysts) 4056 were more prominent in females and were associated with the increased survivability in the 4057 exposed female hamsters compared with controls. Th e results indicate that 3,500 ppm was a 4058 NOAEL for adverse changes in clinical chemistry and hematological variables, as well as for 4059 nonneoplastic histologic changes in tissues, in mal e and female Syrian golden hamsters. A 4060 LOAEL was not established, based on the lack of adv erse changes in clinical chemistry and 4061 hematological variables as well as the absence of n onneoplastic histologic changes in tissues, in 4062 male and female Syrian golden hamsters. 4063

Evaluation of the total number of hamsters with a tumor, the number with a benign 4064 tumor, or the number with a malignant tumor reveale d no exposure-related differences in male 4065 hamsters. In the high-dose female group, there was a statistically significant increase in the total 4066 number of benign tumors at any tissue site (the rep ort did not specify which sites), but this was 4067 considered to be secondary to the increased surviva l of this group. Incidences of male or female 4068 hamsters with tumors in specific tissues were not s tatistically significantly elevated in exposed 4069 groups compared with control incidences. The resul ts indicate that no statistically significant, 4070 exposure-related carcinogenic responses occurred in male or female Syrian golden hamsters 4071 exposed (6 hours/day, 5 days/week) to up to 3,500 p pm dichloromethane for 2 years. 4072

4073

4.2.2.2.4. Chronic inhalation exposure in Sprague-Dawley rats (Burek et al., 1984). In the rat 4074 study, groups of 92–97 Sprague-Dawley rats of each sex were exposed (similar to the hamster 4075 study described in the previous section) to 0, 500, 1,500, or 3,500 ppm dichloromethane 4076 6 hours/day, 5 days/week for 2 years (Burek et al., 1984). Rats were approximately 8 weeks old 4077 when exposure started. Interim sacrifices were con ducted at 6, 12, 15, and 18 months. 4078 Endpoints monitored in rats were the same as in ham sters except that total protein and albumin in 4079

blood were not determined in rats. In addition to measurement at scheduled sacrifices, serum 4080 ALT activity was also measured after 30 days of exp osure. COHb was measured after 6, 11, 18, 4081 and 21 months of exposure. Bone marrow cells were collected for cytogenetic studies from five 4082 rats/sex/dose after 6 months of exposure. The scop e of the pathological examinations of the rats 4083 was the same as in the hamster study. 4084 No significant exposure-related signs of toxicity were observed in the rats during the 4085 study. A significant increase in mortality was see n in high-dose female rats from the 18th to the 4086 24th month of exposure, and this appeared to be exp osure-related. Exposure to dichloromethane 4087 had no significant effect on BW gain in either male s or females. The only exposure-related 4088 alterations in organ weights was a significant incr ease in both absolute and relative liver weight 4089 in high-dose males at the 18-month interim kill and a significant increase in relative liver weight 4090 in high-dose females also at 18 months. Statistica lly significant changes in hematologic 4091 parameters were restricted to increased mean corpus cular volume and mean corpuscular 4092118

hemoglobin values at 15 months in males. The clini cal chemistry tests revealed no significant 4093 exposure-related effects. Male and female rats in all exposed groups had significantly elevated 4094 COHb values at all time points, but no dose-respons e relationship was apparent. For example, 4095 mean (± SD) values for percentage COHb after 21 mon ths of exposure were 0.4 (± 0.7), 4096 12.8 (± 2.6), 14.8 (± 4.4), and 12.2 (± 5.7) for th e control through 3,500 ppm female rat groups, 4097 respectively. Exposure-related statistically signi ficant increases in incidences of nonneoplastic 4098 lesions were restricted to the liver (Table 4-21). The incidences of males or females with 4099 hepatocellular vacuolation consistent with fatty ch ange increased as the exposure concentration 4100 increased. Hepatocellular necrosis occurred at ele vated incidences in male rats exposed to 4101 1,500 or 3,500 ppm, compared with controls, but thi s endpoint was not reported in the female 4102 data. Liver lesions were initially observed after 12 months of treatment. There was some 4103 evidence that exposure at the two highest levels pr ovided some inhibition of the age-related 4104 glomerulonephropathy observed in the control rats a t termination. The results indicate that the 4105 lowest exposure level, 500 ppm, was a LOAEL for fat ty changes in the liver of male and female 4106 Sprague-Dawley rats and that exposure to ≥ 1,500 ppm induced hepatocellular necrosis in males. 4107119 4108 Table 4-21. Incidences of selected neoplastic hist ologic changes in male and

female Sprague-Dawley rats exposed to dichlorometha ne by inhalation (6 hours/day, 5 days/week) for 2 years

Exposure (ppm) a

Lesion, by sex Controls 0 500 1,500 3,500 Males—n per group 92 95 95 97 Number (%) with Liver changes Hepatocellular necrosis 2 b (2) 8 (8) 10 (10) c 11 (11) c Coagulation necrosis -- d

-- -- -- Hepatic vacuolation (fatty change) 16 b (17) 36 (38) c 43 (45) c 52 (54) c Foci of altered hepatocytes -- -- -- -- Foci of altered hepatocytes, basophilic -- -- -- -- Area of altered hepatocytes -- -- -- -- Multinucleated hepatocytes -- -- -- -- Glomerulonephropathy Severe Any degree 70 b (76) 92 b,e (100) 62 (65)

91 (96) 53 (56) c

93 (98) 39 (40) c

90 (93) Mammary changes Rats with benign mammary tumors 7 b (8) 3 (3) 7 (7) 14 (14) Total number of benign mammary tumors 8 6 11 17 Number of tumors per tumor-bearing rat f 1.1 2.0 1.6 1.2 Females—n per group 96 95 96 97 Number (%) with Liver changes Hepatocellular necrosis – – – – Coagulation necrosis 1 b (1) 0 (0) 2 (2) 7 (7) Hepatic vacuolation (fatty change) 33 b (34) 49 (52) c 56 (58) c 63 (65) c Foci of altered hepatocytes 35 b (37) 36 (38) 27 (28) 50 (52) c Foci of altered hepatocytes, basophilic 3 b (3) 0 (0) 4 (4) 10 (10) Area of altered hepatocytes 19 b (20) 24 (25) 28 (29) 35 (36) c Multinucleated hepatocytes 7 b (7) 36 (38) c 34 (35) c 29 (30) c Glomerulonephropathy Severe Any degree 5 (5) 62 b (65) 3 (3)

64 (67) 4 (4)

59 (62) 5 (5)

48 (50) c Mammary changes Rats with benign mammary tumors 79 (82) 81 (85) 80 (83) 83 (86) Total number of benign mammary tumors 165 218 245 287 Number of tumors per tumor-bearing rat f 2.1 2.7 3.1 3.5

a 500 ppm = 1,737 mg/m 3 , 1,500 ppm = 5,210 mg/m 3 , 3,500 ppm = 12,158 mg/m 3

. b Significant dose-related trend—Cochran-Armitage tre nd test p < 0.05.

c Significantly higher than control incidence by Fish er’s exact test.

d – = Reported as “no exposure effect” by Burek et al . (1984); data not given.

e Burek et al. (1984) reported that 93/92 male mice h ad glomerulonephropathy in the kidney in the contro l group; the incidence was corrected to 92/92.

f Calculated by EPA.

Source: Burek et al. (1984). 4109 4110 In females, an increasing trend was seen in the in cidence of foci or areas of altered 4111 hepatocytes. Female rats in all exposed groups sho wed increased incidence of multinucleated 4112 hepatocytes in the centrilobular region, compared w ith controls, but there was no evidence of 4113120

increasing incidence or severity with increasing ex posure level (Table 4-21). The foci and areas 4114 were apparent after 12 months and their number and size increased thereafter, but incidences for 4115 neoplastic nodules in the liver or hepatocellular c arcinomas were not increased in any exposure 4116 group. A statistically significant increased incid ence of salivary gland sarcomas was reported for 4117 male rats exposed to 3,500 ppm. Burek et al. (1984 ) considered this finding unusual and 4118 inconsistent with other existing data because the p rimary target organ for dichloromethane seems 4119 to be the liver. Incidences of rats with benign ma mmary gland tumors were not statistically 4120 significantly higher in exposed male or female grou ps compared with controls, and exposed male 4121 and female groups showed no significantly increased incidences for malignant mammary gland 4122 tumors. The average number of benign mammary tumor s per tumor-bearing rat increased with 4123 increasing exposure level. In females, the values were 2.1, 2.7, 3.1, and 3.5 in the control 4124 through 3,500 ppm groups, respectively; males showe d a similar response with increasing 4125 exposure level, albeit to a lesser extent (Table 4- 21). Burek et al. (1984) concluded that the 4126 significance of this benign mammary tumor response (i.e., increase in number of tumors per 4127 tumor-bearing rat) was unknown but speculated that the predisposition of this strain of rats 4128 (historical control incidences of female with benig n mammary tumors normally exceeded 80%) 4129 plus the high exposure to dichloromethane may have resulted in the response. 4130

4131

4.2.2.2.5. Chronic inhalation exposure in Sprague-Dawley rats (Nitschke et al., 1988a). 4132 Nitschke et al. (1988a) examined the toxicity and c arcinogenicity of lower concentrations of 4133 dichloromethane in Sprague-Dawley rats. Groups of 90 male and 90 female rats were exposed to 4134 0, 50, 200, or 500 ppm dichloromethane (>99.5% pure ) 6 hours/day, 5 days/week for 2 years. 4135 Interim sacrifices were conducted at 6, 12, 15, and 18 months (five rats/sex/interval). An 4136 additional group of 30 female rats was exposed to 5 00 ppm for 12 months and then exposed to 4137 room air for up to an additional 12 months, and ano ther group of 30 female rats was exposed to 4138 room air for the first 12 months, followed by expos ure to 500 ppm for the last 12 months of the 4139 study. These latter groups were included to examin e temporal relationships between exposure 4140 and potential carcinogenic response. All groups of rats were examined daily for signs of toxicity 4141 and all rats were examined for palpable masses prio r to the initial exposure and at monthly 4142

intervals after the first 12 months. BW was checke d twice a month for the first 3 months and 4143 monthly thereafter. Blood samples were collected a t interim sacrifices and analyzed for total 4144 bilirubin, cholesterol, triglycerides, potassium, e stradiol, follicle-stimulating hormone, and 4145 luteinizing hormone levels. In addition, COHb was determined at multiple times in blood 4146 collected from the tail vein. DNA synthesis (incor poration of 3 H-thymidine as a measure of 4147 cellular proliferation) was measured in the liver o f separate groups of female rats after exposure 4148 to the various concentrations for 6 and 12 months ( four females/exposure group per interval). 4149 All rats were subjected to a complete necropsy, and sections from most tissues were processed 4150 for microscopic examination. 4151121

Exposure to dichloromethane at any of the exposure levels did not significantly alter 4152 mortality rates, BWs, organ weights, clinical chemi stry values, or plasma hormone levels 4153 (Nitschke et al., 1988a). Blood COHb was elevated in a dose-related manner but not in an 4154 exposure duration-related fashion, suggesting lack of accumulation with repeated exposures. For 4155 example, mean (± SD) values for percentage COHb wer e 2.2 (± 1.3), 6.5 (± 1.1), 12.5 (± 0.8), 4156 and 13.7 (± 0.6) for male rats in the control throu gh 500 ppm groups, respectively, at the 4157 terminal sacrifice and were similarly affected at t he 6-month and 12-month intervals (e.g., 4158 respective values for males were 0.3(± 0.7), 2.8 (± 0.3), 9.6 (± 1.2), and 12.7 (± 1.6) at the 4159 12-month sacrifice). 4160 The results of the thymidine incorporation experime nt revealed no detectable alteration in 4161 the rate of liver DNA synthesis in the exposed grou ps compared with controls. Statistically 4162 significantly increased incidences of nonneoplastic liver lesions (hepatic vacuolation and 4163 multinucleated hepatocytes) occurred only in female s in the 500 ppm group (Table 4-22). Male 4164 rat incidence for hepatocyte vacuolation was elevat ed at 500 ppm but not to a statistically 4165 significant degree. In the group of female rats ex posed for only 12 months to 500 ppm, 4166 significantly increased incidences of nonneoplastic lesions, compared with controls, were 4167 restricted to liver cytoplasmic vacuolization (16/2 5 = 64%) and multinucleated hepatocytes 4168 (9/25= 36%) in rats exposed during the first 12 mon ths of the study; rats exposed only during the 4169 last 12 months of the study showed no elevated inci dences of the liver lesions. 4170

4171 Table 4-22. Incidences of selected nonneoplastic h istologic changes in male

and female Sprague-Dawley rats exposed to dichlorom ethane by inhalation (6 hours/day, 5 days/week) for 2 years

Exposure (ppm) a

Lesion, by sex Controls 0 50 200 500 Trend

p- value b Late

500 c Early

500 c Males—n per group 70 70 70 70 NA d NA Number (%) with Hepatic vacuolation (fatty change) 22 (31) – e – 28 (40) Multinucleated hepatocytes – – – – Females—n per group 70 70 70 70 25 25 Number (%) with Hepatic vacuolation (fatty change) 41 (59) 42 (60) 41 (59) 53 (76) f 0.01 15 (60) 16 (64) f Multinucleated hepatocytes 8 (11) 6 (9) 12 (1 7) 27 (39) f <0.0001 3 (12) 9 (36) f

a 50 ppm = 174 mg/m 3 , 200 ppm = 695 mg/m 3 , 500 ppm = 1,737 mg/m 3

. b Cochran-Armitage trend test.

c Late 500 = no exposure for first 12 months followed by 500 ppm for last 12 months; early 500 = 500 ppm for first 12 months followed by no exposure for last 12 month s. d NA= there were no male rats in these exposure group s.

e – = Incidences not reported.

f Significantly (

p ≤ 0.05) higher than control incidence by Fisher’s ex act test (Nitschke et al., 1988a).

Source: Nitschke et al. (1988a). 4172 4173122

A few fibrosarcomas or undifferentiated sarcomas i n the mammary gland were seen in 4174 the exposed rats, but these incidences were not sta tistically significant (Table 4-23). 4175 Significantly increased incidences of rats with neo plastic lesions were restricted to benign 4176 mammary tumors in female rats exposed for 2 years t o 200 ppm compared with controls 4177 (61/69 = 88%) (Table 4-23). However, significantly elevated incidences of this tumor type were 4178 not observed in 500 ppm females, and the 200 ppm in cidence was within the range for historical 4179 control values for benign mammary tumors in female Sprague-Dawley rats (79–82%) from two 4180 other chronic toxicity/carcinogenicity studies from the same laboratory. A slight, but statistically 4181 significant, increase in the number of palpable mas ses in subcutaneous or mammary regions (at 4182 23 months) per tumor-bearing rat was observed only in the 500 ppm female group. The numbers 4183 of benign mammary tumors per tumor-bearing rat were slightly elevated in the exposed groups 4184 compared with control groups, but no statistical an alysis of this variable was performed. In 4185 female rats exposed to 500 ppm (during the first or second 12 months of the study), slight but 4186 statistically significant elevations were found in the number of palpable masses in subcutaneous 4187 or mammary regions per tumor-bearing rat; the numbe rs of benign mammary tumors per tumor4188 bearing rat were slightly elevated compared with th ose of controls, but statistical analysis of this 4189 variable was not performed. 4190 A statistically significant increased incidence of brain or CNS tumors was not observed, 4191 but six astrocytoma or glioma (mixed glial cell) tu mors were seen in the exposed groups (4 in 4192 males, 2 in females). The authors concluded that t here was no distinct exposure-related 4193 malignant carcinogenic response in male or female S prague-Dawley rats exposed (6 hours/day, 4194 5 days/week) to up to 500 ppm dichloromethane for 2 years (Nitschke et al., 1988a). 4195

4196123

Table 4-23. Incidences of selected neoplastic hist ologic changes in male and female Sprague-Dawley rats exposed to dichlorometha ne by inhalation (6 hours/day, 5 days/week) for 2 years

Exposure (ppm) a

Lesion, by sex Controls 0 50 200 500 Late 500 b Early

500 b Males—n per group 70 70 70 70 0 0 Number (%) c with Liver tumors 0 (0) 0 (0) 0 (0) 0 (0) Lung tumors 0 (0) 0 (0) 0 (0) 0 (0) Mammary gland tumors Adenocarcinoma or carcinoma Fibroadenoma Fibroma Fibrosarcoma Undifferentiated sarcoma Fibroma, fibrosarcoma, or undifferentiated

sarcoma d 0 (0) 2 (4) 6 (11) 0 (0) 0 (0) 6 (11) 0 (0) 0 (0) 1 (6) 1 (6) 2 (4) 4 (6) 0 (0) 2 (3) 6 (11) 1 (6) 0 (0) 7 (12) 0 (0)

2 (3) 10 (16) 0 (0) 0 (0) 10 (16)

Brain tumors Astrocytoma or glial cell Granular cell 0 (0) 0 (0) 1 (1)

0 (0) 2 (3) 0 (0) 1 (1) 1 (1)

Females—n per group 70 70 70 70 25 25 Number (%) c with Liver tumors Neoplastic nodule(s) Hepatocellular carcinoma 4 (6) 1 (1) 4 (6)

0 (0) 3 (4) 2 (3) 4 (6) 1 (1) 0 (0)

0 (0) 1 (4)

0 (0) Lung tumors 0 (0) 0 (0) 0 (0) 0 (0) 0 (0) 0 (0) Mammary gland tumors Adenocarcinoma or carcinoma Adenoma Fibroadenoma Fibroma Fibrosarcoma 6 (9) 1 (1) 51 (74) 0 (0) 1 (1) 5 (7) 1 (1) 57 (83) 1 (1) 0 (0) 4 (6) 2 (3) 60 (87) 0 (0) 0 (0) 4 (6) 1 (1) 55 (80) 1 (1) 0 (0) 3 (12)

2 (8) 22 (88) 1 (4) 0 (0) 2 (8)

0 (0) 23 (92) 1 (1) 0 (0) Number with palpable masses in subcutaneous or mammary region 55 (78) 56 (81) 60 (87) 59 (86) 22 (88) 23 (92)

Number of palpable masses in subcutaneous or mammary region per tumor-bearing rat 1.8 2.1 2.0 2.2 e 2.3 e 2.7 e

Number with benign tumors 52 (75) 58 (84) 61 f (88) 55 (80) 23 (92) 23 (92) Number of benign tumors per tumor-bearing rat 2.0 2.3 2.2 2.7 2.2 2.6

Brain tumors Astrocytoma or glial cell Granular cell 0 (0) 1 (1) 0 (0) 0 (0) 0 (0) 0 (0) 2 (3) 1 (1) 0 (0)

0 (0) 0 (0)

0 (0)

a 50 ppm = 174 mg/m 3 , 200 ppm = 695 mg/m 3 , 500 ppm = 1,737 mg/m 3

b L ate 500 = no exposure for first 12 months followed by 500 ppm for last 12 months; early 500 = 500 ppm for first 12 months followed by no exposure for last 12 month s. No males were included in these exposure groups

. c Percentages were based on the number of tissues exa mined microscopically per group. In males, 69 lung s were examined microscopically in the 50 ppm groups, and only 57, 65, 59, and 64 mammary glands were examine d in the control, 50, 200, and 500 ppm groups, respectiv ely. In females, 69 mammary glands were examined microscopically in the control, 50, 200, and 500 pp m groups. d EPA summed across these three tumors, assuming no o verlap.

e Significantly (

p ≤ 0.05) higher than control by Haseman’s test (Nitsc hke et al., 1988a).

f Significantly (

p ≤ 0.05) higher than control incidence by Fisher’s ex act test (Nitschke et al., 1988a).

Source: Nitschke et al. (1988a). 4197124

4.2.2.2.6. Chronic inhalation exposure in Sprague-Dawley rats (Maltoni et al., 1988). Maltoni 4198 et al. (1988) conducted an inhalation exposure stud y in Sprague-Dawley rats. Two groups of 4199 female rats (54–60/dose) were exposed to 0 or 100 p pm dichloromethane for 104 weeks. The 4200 exposure period was 4 hours/day, 4 days/week for 7 weeks and then 7 hours/day, 5 days/week 4201 for 97 weeks. Endpoints monitored included clinica l signs, BW, and full necropsy at sacrifice 4202 (when spontaneous death occurred). For each animal sacrificed, histopathologic examinations 4203 were performed on the following organs: brain and c erebellum, zymbal glands, interscapular 4204 brown fat, salivary glands, tongue, thymus and medi astinal lymph nodes, lungs, liver, kidneys, 4205 adrenals, spleen, pancreas, esophagus, stomach, int estine, bladder, uterus, gonads, and any other 4206 organs with gross lesions. 4207 There was no evidence of increased mortality in the exposed group, and there was no 4208 effect on BW (Maltoni et al., 1988). Little inform ation was provided regarding nonneoplastic 4209 effects, precluding assignment of NOAELs and LOAELs for possible nonneoplastic effects in 4210 this study. Dichloromethane exposure was not relat ed to the percentage of rats with benign 4211 tumors and malignant tumors, malignant tumors, or t he number of total malignant tumors per 4212 100 animals. The percentage of rats with benign ma mmary tumors was 40.0% in controls and 4213 64.8% in the exposed group, and the percentage of m alignant mammary tumors was 3.3 and 4214 5.5% in controls and exposed, respectively. Neithe r of these differences was statistically 4215 significant. 4216

4217

4.3. REPRODUCTIVE/DEVELOPMENTAL STUDIES—ORAL AND I NHALATION 4218 Reproductive and development studies of dichloromet hane exposure are summarized in 4219

Table 4-24 and described in detail below. No effec ts on reproductive performance were 4220 observed in a 90-day gavage study in Charles River CD rats with doses up to 225 mg/kg-day 4221 (General Electric Co., 1976) or in a two-generation reproductive toxicity study with F344 rats 4222 exposed to concentrations up to 1,500 ppm for 14 or 17 weeks before mating of the F0 and 4223 F1 generations, respectively, as well as during the F1 gestational period (GDs 0–21) (Nitschke et 4224 al., 1988b). Reproductive parameters (e.g., number of litters, implants/litter, live fetuses/litter, 4225 percent dead/litter, percent resorbed/litter, or fe rtility index 4 ) were also examined in a study in 4226

male Swiss-Webster mice administered dichloromethan e (250 or 500 mg/kg) by subcutaneous 4227 injection three times/week for 4 weeks, and in a si milar study involving inhalation exposure to 0, 4228 100, 150, or 200 ppm dichloromethane; no statistica lly significant effects were seen in either 4229 protocol, although some evidence of a decrease in f ertility index was seen in the 150 and 4230 200 ppm groups (Raje et al., 1988). 4231

4232

4 Fertility index defined as number of females impreg nated divided by total number of females mated time s 100.1254233Table 4-24. Summary of studies of reproductive anddevelopmental effects of dichloromethane exposureinanimalsSpecies and n Exposure dose Exposure period ResultsReferenceOral and GavageCharles River rats(males and females),10 per sex per dosegroup0, 25, 75, 225 mg/kg(gavage)90 days before mating(10 days between lastexposure and matingperiod)No effects on fertility index, number of pups perlitter, pup survival, or F1 BW, hematology, andclinical chemistry tests (up to 90 days of age)GeneralElectric Co.(1976)Swiss-Webster mice(males), 20 per group0, 250, 500 mg/kg(subcutaneousinjection), 3×perweek4 weeks prior to mating(1 week between lastexposure and matingperiod)No effects on fertility index, number of litters,implants per litter, live fetuses per litter, resorptionrate; no testicular effectsRaje et al.(1988)F344 rats (females),17–21 per dose group0, 337.5,450 mg/kg-day(gavage)GDs 6–19 Decreased maternal weight gain; no effectonresorption rate, number of live litters, implants,live pups, or pup weightNarotsky andKavlock(1995)InhalationF344 rats (males andfemales, twogeneration), 30 persex per dose group(F0 and F1)0, 100, 500,1,500 ppm,6 hours/day14 weeks prior tomating (F0), GDs 0–21,and 17 weeks prior tomating, beginningPND 4, (F1)No effect on fertility index, litter size, neonatalsurvival, growth rates, or histopathologic lesionsNitschke et al.(1988b)Swiss-Webster mice(males), 20 per group0, 100, 150,200 ppm,2 hours/day6 weeks, prior tomating (2 days betweenlast exposure andmating period)Fertility index decreased in 150 and 200 ppmgroup (statistical significance depends on testused); no effects on number of litters, implants perlitter, live fetuses per litter, resorption rate; notesticular effects.Raje et al.(1988)Long-Evans rats(female), 16–21 perdose group0, 4,500 ppm 12–14 days beforemating and/or GDs 1–17Gestational exposure resulted in increased absoluteand relative maternal liver weight, decreased fetalBWHardin andManson(1980)Long-Evans rats(female), 16–21 perdose group0, 4,500 ppm 12–14 days beforemating and/or GDs 1–17Altered rate of behavioral habituation to novelenvironment (at 4 days of age). No effect oncrawling (at 10 days), movement in photocell cage(15 days), use of running wheel (45–108 days), andshock avoidance (4 months).Bornschein etal. (1980)126Table 4-24. Summary of studies of reproductive anddevelopmental effects of dichloromethane exposureinanimalsSpecies and n Exposure dose Exposure period ResultsReferenceSwiss-Webster mice(females), 30–40 pergroup0, 1,250 ppm,7 hours/dayGDs 6–15 Increased incidence of extra center of ossificationin sternum, increased (~10%) maternal bloodCOHb, increased maternal weight, increasedmaternal absolute liver weightSchwetz et al.(1975)Sprague-Dawley rats(females), 20–35 pergroup0, 1,250 ppm,7 hours/dayGDs 6–15 Decreased incidence of lumbar ribs or spurs,increased incidence of delayed ossification ofsternebrae, increased (~10%) maternal bloodCOHb, increased maternal absolute liver weightSchwetz et al.(1975)4234423542364237127

Following exposure of pregnant F344 rats to gavage doses of up to 450 mg/kg-day on 4238 GDs 6–19, maternal weight gain was decreased, but n o effects were found on the number of 4239 resorption sites, pup survivability, or pup weights at postnatal days (PNDs) 1 or 6 (Narotsky and 4240 Kavlock, 1995). The developmental effects followin g exposure of Long-Evans rats to 4241 4,500 ppm for 14 days prior to mating and during ge station (or during gestation alone) were 4242 decreased offspring weight at birth and changed beh avioral habituation of the offspring to novel 4243 environments (Bornschein et al., 1980; Hardin and M anson, 1980) (see section 4.3.2 for more 4244 details). In standard developmental toxicity studi es involving exposure to 1,250 ppm on GDs 6– 4245 15, no adverse effects on fetal development were fo und in Swiss-Webster mice or Sprague4246 Dawley rats, but the incidence of minor skeletal va riants (e.g., delayed ossification of sternebrae) 4247 was increased. (Schwetz et al., 1975) (see section 4.3.2). 4248

4249 4.3.1. Reproductive Toxicity Studies 4250

4.3.1.1. Oral (Gavage) Studies 4251 In a study sponsored by the General Electric Co. ( 1976), Charles River CD rats 4252 (10/sex/dose level) were administered 0, 25, 75, or 225 mg/kg-day dichloromethane by gavage in 4253 water for 90 days. The test material was dichlorom ethane (of unspecified purity) purchased from 4254 Dow Chemical Company. At approximately 100 days of age, the rats were mated 1 to 1 to 4255 produce the F1 generation. F1 rats (15/sex/dose le vel) received the same treatment as F0 for 4256 90 days, at which time they were sacrificed and nec ropsied. Comprehensive sets of 24 tissues 4257

from 10 male and 10 female F1 rats from the control and 225 mg/kg-day groups were examined 4258 microscopically after embedding, sectioning, and st aining. F1 rats were monitored for clinical 4259 signs, BW effects, and food consumption. Reproduct ive parameters examined were fertility 4260 index, number of pups per litter, and pup survival. F1 rats also underwent hematology and 4261 clinical chemistry tests and urinalysis at 1, 2, an d 3 months of the study and ophthalmoscopic 4262 examination at 3 months. There were no significant compound-related alterations in any of the 4263 endpoints monitored. 4264 Raje et al. (1988) administered dichloromethane (2 50 or 500 mg/kg) by subcutaneous 4265 injection three times per week for 4 weeks to male Swiss-Webster mice (20/group). Mating with 4266 unexposed females started 1 week after the last exp osure and continued for 2 weeks. After the 4267 mating period, the males were sacrificed, and the t estes were examined microscopically. On 4268 GD 17, the females were sacrificed and the uterine horns examined for live, dead, or resorbed 4269 fetuses. The authors reported that exposure to dic hloromethane had no statistically significant 4270 effects on number of litters, implants/litter, live fetuses/litter, percent dead/litter, percent 4271 resorbed/litter, or fertility index. Examination o f the testes showed no significant alterations 4272 compared with controls. 4273

4274

4275128

4.3.1.2. Inhalation Studies 4276 Nitschke et al. (1988b) conducted a two-generation reproductive toxicity study in rats. 4277 Groups of F344 rats (30/sex/dose level) were expose d by inhalation in whole-body chambers to 4278 0, 100, 500, or 1,500 ppm dichloromethane (99.86% p ure) 6 hours/day, 5 days/week for 4279 14 weeks and then mated to produce the F1 generatio n. Exposure of dams continued after 4280 mating on GDs 0–21 but was interrupted until PND 4. After weaning 30 randomly selected F1 4281 pups/sex/dose level were exposed as the parental ge neration for 17 weeks and subsequently 4282 mated to produce the F2 generation. The results sh owed no statistically significant exposure4283 related changes in reproductive performance indices (fertility, litter size), neonatal survival, 4284 growth rates, or histopathologic lesions in F1 (Tab le 4-25) or F2 weanlings sacrificed at time of 4285 weaning. According to the authors none of the valu es in Table 4-25 was significantly different 4286 from control values ( α = 0.05). 4287

4288 Table 4-25. Reproductive outcomes in F344 rats exp osed to

dichloromethane by inhalation for 14 weeks prior to mating and from GDs 0–21

Exposure (ppm) a 0 100 500 1,500 Fertility index b 77% 77% 63% 87%

Gestation index c 100% 100% 100% 100%

Gestation survival index d 99.6% 100% 100% 96.6% 4-day survival index e 91.0% 95.2% 98.5% 98.6% 28-day survival index f 99.4% 99.4% 100% 99.5%

Sex ratio on day 1 (M:F) 48:52 50:50 50:50 52:48

Litter size Day 0 11 ± 2 10 ± 2 10 ± 3 11 ± 2 Day 28 7 ± 2 7 ± 2 7 ± 2 8 ± 2

Pup BWs, g Day 1 5.2 ± 0.4 5.3 ± 0.5 5.3 ± 0.4 5.2 ± 0.4 Day 4 7.4 ± 0.7 7.5 ± 1.1 7.7 ± 0.7 7.3 ± 0.7 Day 28, male 44.6 ± 5.8 45.9 ± 5.0 47.0 ± 5. 4 45.0 ± 5.9 Day 28, female 43.2 ± 4.3 43.8 ± 4.5 44.4 ± 5.7 43.0 ± 4.8

a 100 ppm = 347 mg/m 3 , 500 ppm = 1,737 mg/m 3 , 1,500 ppm = 5,210 mg/m 3

, b Number of females delivering a litter expressed as a percentage of females placed with a male.

c Number of females delivering a live litter expresse d as a percentage of the number of females deliveri ng a litter.

d Percentage of newborn pups that were alive at birth

. e Percentage of pups surviving to day 4.

f Percentage of pups alive on day 4 and surviving to day 28.

Source: Nitschke et al. (1988b). 4289 4290129

Raje et al. (1988) exposed groups of male Swiss-We bster mice (20/group) to 0, 100, 150, 4291 or 200 ppm dichloromethane (HPLC grade, JT Baker Ch emical Co.) in inhalation chambers for 4292 2 hours/day, 5 days/week for 6 weeks. Mating with unexposed females started 2 days after the 4293 last exposure. As in the subcutaneous injection pr otocol described in the previous section, after 4294 the 2-week mating period, the males were sacrificed and the females were sacrificed on GD 17. 4295 Exposure of the male mice to dichloromethane had no statistically significant effects on number 4296 of litters, implants/litter, live fetuses/litter, p ercent dead/litter, or percent resorbed/litter, and no 4297 significant alterations in the testes were noted. The fertility index was 95%, 95%, 80%, and 80% 4298 in the control, 100, 150 and 200 ppm groups, respec tively. This decrease was not statistically 4299 significant as reported by the authors. Details of the statistical analyses were not provided. The 4300 overall Chi-square p -value was 0.27. Using a Cochran-Armitage exact tr end test on these data, 4301

EPA calculated a one-sided p- value of 0.059. Individual p -values for the comparison of each 4302 group with the control group were 0.97, 0.17, and 0 .17 for the 100, 150 and 200 ppm groups, 4303 respectively. The results for the combined 150 and 200 ppm groups were statistically different 4304 from the combined controls and 100 ppm group (Fishe r’s exact test, one-sided p- value = 0.048), 4305 suggesting a NOAEL of 100 ppm and LOAEL of 150 ppm. 4306

4307 4.3.2. Developmental Toxicity Studies 4308 The metabolism of dichloromethane into CO by CYP2E 1 raises concerns pertaining to 4309 developmental neurotoxicity. Gestational exposure to CO results in developmental toxicity and 4310 there are reports indicating that exposures as low as 75 ppm CO can result in significant 4311 neurological effects in offspring (Giustino et al., 1999). Neurobehavioral deficits in offspring 4312 include impaired avoidance behavior (De Salvia et a l., 1995) and memory (Giustino et al., 1999). 4313 Neurochemical changes , such as abnormal dopaminergic function (Cagiano et al., 1998) and 4314 disruption of neuronal proliferation (Fechter, 1987 ) , have also been observed. Oral and 4315 inhalation dichloromethane exposure studies have de monstrated increased blood CO levels (see 4316 section 3.3). In addition, increased blood CO leve ls were seen in rat fetuses exposed through 4317 maternal inhalation to 500 ppm dichloromethane on G D 21 (Anders and Sunram, 1982), and 4318 placental transfer of dichloromethane also occurs ( Withey and Karpinski, 1985; Anders and 4319 Sunram, 1982) 4320

4321

4.3.2.1. Oral (Gavage) Studies and Culture Studies 4322 Narotsky and Kavlock (1995) evaluated developmental effects of dichloromethane 4323 (99.9% pure) in F344 rats (17–21/dose group) treate d with 0, 337.5, or 450 mg/kg-day 4324 dichloromethane by gavage in corn oil on GDs 6–19. Dams were weighed on GDs 6, 8, 10, 13, 4325 16, and 20 and allowed to deliver naturally. They were sacrificed on PND 6 to count uterine 4326 implantation sites. Pups were grossly examined for developmental abnormalities and weighed 4327 on PNDs 1, 3, and 6. Dead pups or pups with no gro ss abnormalities were sacrificed and 4328130

examined for soft tissue abnormalities. Maternal w eight gain during pregnancy was significantly 4329 reduced in high-dose dams (by 33%, as estimated fro m Figure 5 of the paper); this group also 4330 exhibited rales and nasal congestion. Treatment wi th dichloromethane did not induce resorptions 4331 or alter the number of live litters on PND 1 or 6, the number of implants, the number of live pups 4332 on PND 1 or 6, or pup weight per litter. No gross or soft tissue abnormalities were observed. 4333 Rat embryos in culture medium were exposed to 0, 3 .46, 6.54, 9.79, or 11.88 μ mol/mL 4334

dichloromethane for 40 hours. At the end of the ex posure, embryos were observed for 4335 development of yolk sac vasculature, crown-rump len gth, total embryonic protein content, and 4336 number of somite pairs. A concentration of dichlor omethane of 6.54 μ mol/mL of culture 4337 medium resulted in decreased crown-rump length, dec reased somite number, and decreased 4338 amount of protein per embryo, whereas no effects we re seen at 3.46 μ mol/mL (Brown-Woodman 4339 et al., 1998). A time-course experiment conducted with a concentration of dichloromethane of 4340

9.22 μ mol/mL showed that marked differences in growth and development from controls were 4341 not significant until about 8 hours of culture. Br own-Woodman et al. (1998) noted that the 4342 concentrations that caused embryotoxicity in this s tudy were much higher than those found in 4343 individuals studied under controlled exposure condi tions and comparable to those found in 4344 postmortem blood after fatal inhalation. 4345

4346

4.3.2.2. Inhalation Studies 4347 Schwetz et al. (1975) exposed pregnant Swiss-Webst er mice (30–40/group) and Sprague4348 Dawley rats (20–35/group) by inhalation in whole-bo dy chambers to 0 or 1,250 ppm 4349 dichloromethane (97.86% pure) 7 hours/day on GDs 6– 15. Maternal BWs were recorded on 4350 GDs 6, 10, and 16 and on the day of sacrifice (GD 1 8 for mice, GD 21 for rats). At sacrifice, 4351 uterine horns were excised and examined for fetal p osition and number of live, dead, or absorbed 4352 fetuses. Fetuses were observed for gross, soft tis sue, and skeletal abnormalities. The only 4353 effects seen on developing fetuses were changes in the incidence of minor skeletal variants. In 4354 rats, the incidence of lumbar ribs or spurs was sig nificantly decreased compared with controls, 4355 whereas the incidence of delayed ossification of st ernebrae was significantly greater than in 4356 controls. In mice, a significant number of litters contained pups with a single extra center of 4357 ossification in the sternum. Exposure to dichlorom ethane produced significantly elevated blood 4358 COHb content in dams of both species (approximately 9–10% after 10 exposures versus 1–2% in 4359 controls). BWs in exposed mouse dams were signific antly increased (11–15%) compared with 4360 those in controls but were not affected in exposed rat dams. Mean absolute liver weights of 4361 exposed dams of both species were significantly ele vated compared with controls, but mean 4362 relative liver weights were not affected. The resu lts indicate that 1,250 ppm was a LOAEL for 4363 minimal maternal effects (increased COHb and increa sed absolute liver weight) and a LOAEL 4364

for adverse effects on the fetuses. 4365131

Hardin and Manson (1980) conducted a study in fema le Long-Evans rats to determine 4366 whether exposure before and during gestation is mor e detrimental to reproductive outcome than 4367 exposure either before or during gestation alone. Four groups of 16–21 rats were formed in 4368 which the rats were exposed by inhalation in whole- body chambers to 4,500 ppm 4369 dichloromethane (technical grade, >97% pure) 6 hour s/day for 12–14 days before breeding 4370 and/or on GDs 1–17 or were exposed to filtered air. Maternal BWs were measured every 4 days. 4371

Dams were euthanized on GD 21 and livers and uteri removed. Livers were weighed, and 4372 uterine horns were examined for fetal position and number of live, dead, or absorbed fetuses. 4373 Fetuses were observed for gross, soft-tissue, and s keletal abnormalities. Exposure during 4374 gestation (with or without pre-gestation exposure) significantly increased maternal liver weight 4375 (absolute and relative) by about 10–12 and 9–12%, r espectively, and decreased fetal BW by 4376 about 9–10% relative to those exposed to filtered a ir during gestation. None of the groups 4377 showed significant alterations in the incidence of gross, external, skeletal, or soft-tissue 4378 anomalies. Using the same study design and exposur e level, Bornschein et al. (1980) observed 4379 behavioral activities at various ages. Assessed ac tivities included head movement/pivoting when 4380 placed in a novel environment (4 days of age), limi ted crawling (10 days), movement in a 4381 photocell cage (15 days), use of running wheel (45– 108 days), and shock avoidance (4 months). 4382 Exposure during gestation (with or without pre-gest ation exposure) caused altered rates of 4383

behavioral habituation to novel environments in the pups tested as early as 10 days of age that 4384 were still present at 150 days of age. Growth, foo d and water consumption, wheel running 4385 activity, and avoidance learning were not significa ntly affected by exposure to dichloromethane. 4386 The results indicate that 4,500 ppm was a LOAEL for maternal effects (10% increased absolute 4387 and relative liver weight) and for effects on the f etuses (10% decreased fetal BW and altered 4388 behavioral habituation to novel environments). 4389 In a study of early-life (including gestational) e xposures, Maltoni et al. (1988) exposed 4390 54 pregnant Sprague-Dawley rats to 100 ppm dichloro methane via inhalation 4 hours/day, 4391 5 days/week for 7 weeks, followed by 7 hours/day, 5 days/week for 97 weeks. Exposure 4392 apparently started on GD 12. Groups of 60 male and 69 female newborns continued to be 4393 exposed after birth to 60 ppm dichloromethane 4 hou rs/day, 5 days/week for 7 weeks, followed 4394 by exposure 7 hours/day, 5 days/week for 97 weeks. Additional groups of 60 male and 4395 70 female newborn were exposed after birth to 60 pp m dichloromethane 4 hours/day, 4396 5 days/week for 7 weeks and then for 7 hours/day, 5 days/week for 8 weeks. BWs were 4397 measured every 2 weeks during exposure and every 8 weeks thereafter. At the end of exposure, 4398 animals were sacrificed and histologic examinations were performed on 20 tissue types. 4399 Early life exposures of Sprague-Dawley rats to dic hloromethane (Maltoni et al., 1988) 4400 did not affect mortality or BW in any group. Also, there was no significant effect of exposure to 4401 dichloromethane on the percentage of animals with b enign and malignant tumors and malignant 4402 tumors, the number of malignant tumors per 100 anim als, or the percentage of animals with 4403132

benign mammary tumors, malignant mammary tumors, le ukemias, pheochromocytomas, and 4404 pheochromoblastomas. The results provide no eviden ce that gestational exposure to 100 ppm 4405 dichloromethane during early life stages of develop ment increases the susceptibility of Sprague4406 Dawley rats to the potential carcinogenicity of dic hloromethane, but further conclusions from 4407 these results are precluded because the study inclu ded only one exposure level that was below 4408 the maximum tolerated dose for adult Sprague-Dawley rats. Experiments comparing cancer 4409 responses from early-life exposures with adult expo sures are not available for F344 rats or 4410

B6C3F 1 mice, the strains of animals in which carcinogenic responses to dichloromethane have 4411

been observed. 4412 In summary, the potential for gestational exposure to CO, resulting from maternal 4413 dichloromethane exposure via oral and inhalation ro utes, raises concerns regarding 4414 neurodevelopmental effects. In addition, dichlorom ethane transfer across the placenta has also 4415 been seen in inhalation exposure studies in rats (W ithey and Karpinski, 1985; Anders and 4416 Sunram, 1982). Although few developmental effects were observed at high exposures of 4417 dichloromethane (Bornschein et al., 1980; Schwetz e t al., 1975), there are no studies that have 4418 thoroughly evaluated neurobehavioral and neurochemi cal changes resulting from gestational 4419 dichloromethane exposure. The available data ident ify changes of behavior habituation at 4420 4,500 ppm (Bornschein et al., 1980) and increases i n COHb at 1,250 ppm (Schwetz et al., 1975). 4421 The behavioral changes observed at 4,500 ppm indica te developmental neurotoxic effects. No 4422 other neurological endpoints have been evaluated in the available developmental studies of 4423 dichloromethane, but increases in blood COHb strong ly suggest that dichloromethane is being 4424 metabolized to CO. Gestational exposure to CO can result in significant neurological effects in 4425 offspring, including neurobehavioral deficits (De S alvia et al., 1995), memory effects (Giustino 4426 et al., 1999), and neurochemical changes (Cagiano e t al., 1998; Fechter, 1987). As a result, it is 4427 unknown if developmental neurotoxicity could occur at lower exposures to dichloromethane. 4428

4429

4.4. OTHER DURATION- OR ENDPOINT-SPECIFIC STUDIES 4430 4.4.1. Short-term (2-Week) Studies of General and Hepatic Effects in Animals 4431 Two short-term (2-week) studies examined hepatic an d renal effects of dichloromethane 4432 exposure in F344 rats (Berman et al., 1995) and CD- 1 mice (Condie et al., 1983). Berman et al. 4433 (1995) administered dichloromethane by gavage in co rn oil for up to 14 days to groups of eight 4434 female F344 rats at dose levels of 0, 34, 101, 337, or 1,012 mg/kg-day. Starting at day 4, deaths 4435 occurred in the 1,012 mg/kg-day exposure group, wit h seven of eight rats dying before the end of 4436 the 14-day exposure period. In the dose groups tha t did not experience this high mortality, 4437 incidences of increased necrotic hepatocytes were 0 /8, 0/8, 0/8, and 3/8 for the 0, 34, 101, and 4438 337 mg/kg-day groups, respectively. The increase i n liver lesions was not accompanied by 4439

increases in serum activities of ALT or AST. Kidne ys, spleen, and thymus were also 4440 histopathologically examined in this study, but non e showed exposure-related lesions. The 4441133

results indicate that 101 mg/kg-day was a NOAEL and 337 mg/kg-day was a LOAEL for 4442 increased incidence of degenerative lesions in fema le rats exposed for 14 days. In a companion 4443 study with groups of eight female F344 rats that we re given single doses of 0, 101, 337, 1,012, or 4444 1,889 mg/kg-day, incidences of rats with increased necrotic hepatocytes were 1/8, 0/8, 8/8, 7/8, 4445 and 8/8, respectively (Berman et al., 1995). 4446 Condie et al. (1983) detected exposure-related live r lesions in a 14-day gavage study in 4447

which dichloromethane in corn oil was administered to male CD-1 mice at dose levels of 0, 133, 4448 333, or 665 mg/kg-day. Incidences of mice with min imal or slight cytoplasmic vacuolation were 4449 1/16, 0/5, 3/5, and 4/5 for the control through hig h-dose groups, respectively. The kidneys were 4450 also examined histopathologically in this study but showed no exposure-related lesions. No 4451 other tissues were prepared for histologic examinat ion. Blood urea nitrogen, serum creatinine, 4452 and serum ALT activities were not significantly alt ered by exposure. All dose levels 4453 significantly reduced to the same extent the active transport of p-aminohippurate into renal 4454 cortical slices in vitro, a measure of proximal tub ule function. The results most clearly identify 4455 133 mg/kg-day as a NOAEL and 333 mg/kg-day as a LOA EL for increased incidence of 4456 hepatocyte vacuolation in male mice. 4457

4458 4.4.2. Immunotoxicity Studies in Animals 4459 Aranyi et al. (1986) studied the effects of acute i nhalation exposures to 50 or 100 ppm 4460 dichloromethane on two measures of immune response (susceptibility to respiratory infection 4461 and mortality due to Streptoccocus zooepidemicus exposure and ability of pulmonary 4462 macrophages to clear infection with Klebsiella pneumoniae ). Female CD1 mice that were 5– 4463 7 weeks of age at the start of the exposure portion of the experiment were used for both assays. 4464 Up to five replicate groups of about 30 mice were c hallenged with viable S. zooepidemicus 4465 during simultaneous exposure to dichloromethane or to filtered air. Deaths were recorded over a 4466 14-day observation period. Clearance of 35 S-labeled K. pneumoniae by pulmonary macrophages 4467 was determined by measuring the ratio of the viable bacterial counts to the radioactive counts in 4468 each animal’s lungs 3 hours after infection; 18 ani mals were used per dose group. A single 3- 4469 hour exposure to 100 ppm dichloromethane significan tly increased the susceptibility to 4470 respiratory infection and greater mortality followi ng exposure to S. zooepidemicus ( p ≤ 0.01). 4471 Twenty-six deaths occurred in 140 (18.6%) mice chal lenged during a 3-hour exposure to 100 4472 ppm dichloromethane; in contrast, nine deaths occur red in 140 mice (6.4%) exposed to filtered 4473 air. The 3-hour exposure to 100 ppm dichloromethan e was associated with a statistically 4474 significant (

p ≤ 0.001) 12% decrease in pulmonary bactericidal acti vity (91.6 and 79.6% of 4475 bacteria killed in controls and 100 ppm group, resp ectively). No difference was seen in either 4476 mortality rate or bactericidal activity in experime nts using a single 3-hour exposure to 50 ppm or 4477 3-hour exposures to 40 ppm dichloromethane repeated daily for 5 days compared with control 4478 animals exposed to filtered air. These results sug gest that 3-hour exposure to 50 ppm 4479134

dichloromethane was a NOAEL and 100 ppm was a LOAEL for decreased immunological 4480 competence (immunosuppression) in CD-1 mice. 4481 Aranyi et al. (1986) also conducted a similar set o f experiments with 13 other chemicals 4482 (acetaldehyde, acrolein, propylene oxide, chlorofor m, methyl chloroform, carbon tetrachloride, 4483 allyl chloride, benzene, phenol, monochlorobenzene, benzyl chloride, perchloroethylene, and 4484 ethylene trichloride). Perchloroethylene and ethyl ene trichloride were the only chemicals in this 4485 group for which an increased mortality risk from st reptococcal pneumonia was seen (mortality 4486 risk 15.0 and 31.4% in controls and 50 ppm exposure groups for perchloroethylene and 13.4 and 4487 58.1% in controls and 50 ppm exposure groups for et hylene trichloride). Decreased bactericidal 4488 activity was also seen with acetaldehyde, acrolein, methyl chloroform, allyl chloride, benzene, 4489 benzyl chloride, perchloroethylene, and ethylene tr ichloride at one or more exposures. Results 4490 from several chemicals suggest that 5 days of expos ure results in greater decrease in bactericidal 4491 activity (i.e., acetaldehyde, acrolein, and benzene ), and others (e.g., perchloroethylene) suggest 4492 that 5 days of exposure does not result in greater suppression than a single exposure period. 4493

There was considerable variation in both measures o f immune response among the 4494 controls in the experiments (Aranyi et al., 1986). Among the controls in the experiments with 4495 the 13 chemicals other than dichloromethane, mortal ity in the streptococcal infectivity model 4496 ranged from 5.7–22.1%, with a mean of 12.7%. 5 Bactericidal activity in the klebsiella model 4497 among controls ranged from 67.9–94.7%, with a mean of 81.8%. The number of bacteria 4498 deposited in the lung in an inhalation bacterial in fectivity model can show considerable variation, 4499 (i.e., between 750 to 1,500 viable streptococcus or klebsiella organisms, [Ehrlich, 1980]). 4500 Therefore, concurrent controls are particularly imp ort due to the variation in preparation and 4501

aerosol administration of the bacteria in these ass ays. 4502 Warbrick et al. (2003) evaluated immunocompetence i n male and female Sprague4503 Dawley rats by measuring the immunoglobulin M (IgM) antibody responses following 4504 immunization with sheep red blood cells in addition to hematological parameters and 4505 histopathology of the spleen, thymus, lungs, and li ver. Groups of rats (8/sex/dose level) were 4506 exposed to 0 or 5,000 ppm dichloromethane 6 hours/d ay, 5 days/week for 28 days. Rats injected 4507 with cyclophosphamide served as positive controls. Five days before sacrifice (day 23 of 4508 exposure) all rats were injected with sheep red blo od cells. IgM levels in response to the sheep 4509 red blood cells were comparable between dichloromet hane-exposed and air-exposed rats, 4510 indicating that dichloromethane did not produce imm unosuppression in the animals under these 4511 exposure conditions. Cyclophosphamide-treated anim als had significantly lower levels of IgM 4512 in the blood serum, indicating immunosuppression. Rats exposed to dichloromethane showed 4513 reduced response to sound, piloerection, and hunche d posture during exposures. Neither BW 4514 gain nor the hematological parameters monitored wer e significantly affected by exposure to 4515

5 EPA did not include the duplicate assay of perchlor oethylene in calculating this summary statistic. I f this

additional assay is included, the mortality risk ra nges from 5.7–45.7%, with a mean of 15.0%.135

dichloromethane. Relative and absolute liver weigh ts were significantly increased in females, 4516 but not in males. Relative spleen weight was reduc ed in females, and no significant changes 4517 were seen in the weight of the thymus and lungs. H istopathology of the tissues examined was 4518 unremarkable. Exposure to 5,000 ppm dichloromethan e did not affect antibody production to the 4519 challenge with sheep red blood cells. 4520 In the 2-year drinking water study (Serota et al., 1986a, b) and 2-year inhalation study 4521 (Nitschke et al., 1988a), histopathologic analyses were conducted on the lymph nodes, thymus, 4522 and spleen among several other organs, and no signi ficant changes were noted. 4523 In summary, one study (Aranyi et al., 1986) demonst rated evidence of 4524 immunosuppression, including increased risk of stre ptococcal-pneumonia-related mortality and 4525

decreased clearance of klebsiella bacteria followin g a single dichloromethane exposure at 4526 100 ppm for 3 hours in CD-1 mice. The streptococca l and klebsiella bacterial inhalation assays 4527 are models of respiratory infection that test for l ocal immune effects associated with inhalation 4528 exposure rather than systemic immunosuppression. T he NOAEL identified in this study was 4529 50 ppm. In contrast, in a functional immune assay of systemic immunosuppression conducted in 4530 rats, Warbrick et al. (2003) did not observe change s in the antibody response to sheep red blood 4531 cells in a 28-day inhalation exposure to 5,000 ppm dichloromethane. Histopathologic analyses 4532 of immune system organs in chronic exposure studies for B6C3F 1 mice and F344 rats (Nitschke 4533 et al., 1988a; Serota et al., 1986a, b) revealed no changes from controls. However, no assays of 4534 functional immunity were included in these chronic studies. The limited database for 4535 dichloromethane does not suggest systemic immunosup pression, but the Aranyi et al. (1986) 4536 study provides evidence of route-specific local imm unosuppression from acute exposure studies 4537 in CD1 mice. Due to the acute exposure duration us ed in Aryani et al. (1986), the immune 4538 effects of short-term or chronic exposure to dichlo romethane are unclear. 4539

4540 4.4.3. Neurotoxicology Studies in Animals 4541 Neurological evaluations in animals during and afte r exposure to dichloromethane have 4542 resulted in CNS depressant effects similar to other chlorinated solvents (e.g., trichloroethylene, 4543 perchloroethylene) and ethanol. Overall, there are decreased motor activity, impaired memory, 4544 and changes in responses to sensory stimuli. Neuro behavioral, neurophysiological, and 4545 neurochemical/ neuropathological studies have been used to characterize the effects of 4546

dichloromethane on the CNS. A brief overview of th ese types of studies is provided below, 4547 followed by a detailed description of individual st udies. 4548 Neurobehavioral studies with dichloromethane used p rotocols to measure changes in 4549 spontaneous motor activity, a functional observatio nal battery (FOB) test (to evaluate gross 4550 neurobehavioral deficits), and a task developed to assess learning and memory. The FOB 4551 protocol includes various autonomic parameters, neu romuscular parameters, sensorimotor 4552 parameters, excitability measures, and activity. L earning and memory changes with 4553136

dichloromethane were studied by using a passive avo idance task. The oral and inhalation studies 4554 that examined neurobehavioral endpoints are summari zed in Table 4-26. 4555 Neurophysiological studies with dichloromethane exp osure consisted of measuring 4556 evoked responses in response to sensory stimuli. I n these studies, animals were implanted with 4557 electrodes over the brain region that responds to t he particular stimuli. For example, an electrode 4558 would be implanted over the visual cortex in an ani mal presented with a visual stimulus. Once 4559 the stimulus is presented to the animal, an evoked response is elicited from the brain region and 4560 transmitted to the implanted electrode. During adm inistration of a chemical, if there is a 4561 significant change in the magnitude, shape, and lat ency (among other measures) in the evoked 4562 response, then the chemical is considered to produc e neurological effects. A summary of studies 4563 examining dichloromethane exposure and neurophysiol ogical changes is shown in Table 4-27. 4564 In neurochemical/neuropathological studies with dic hloromethane, animals were first 4565 exposed to dichloromethane (via oral, inhalation, o r injection), and then the brains were 4566 removed. Changes in excitatory neurotransmitters, such as glutamate and acetylcholine, and the 4567 inhibitory neurotransmitter, GABA, were measured. Additionally, dopamine and serotonin 4568 levels, which are associated with addiction and moo d, were also measured. Other parameters 4569 that were measured included DNA/protein content and regional brain changes in the cerebellum 4570 and hippocampus. Measurement of neurochemical chan ges provides mechanistic information, 4571 and neurobehavioral and neurophysiological effects can be correlated to these results. Table 4572 4-28 summarizes studies of neurochemical changes an d dichloromethane. 4573137Inhalation exposureSpontaneous activityNMRI mouse, male 400–2,500 ppm 1 hour Initial increase in activity followed by apronounced decrease at exposures600 ppm and higherKjellstrand et al. (1985)Rat, male 5,000 ppm 1 hour, every other day for 10daysDecreased spontaneous locomotoractivityHeppel and Neal (1944)Wistar rat, male 500 ppm 6 hours/day, 6 days Increased preening frequency Savolainen et al. (1977)ICR mouse, female 5,000 ppm Continuous, 7 days Increased spontaneous activity in firstfew hours and then decreased activityWeinstein et al. (1972)Sprague-Dawley rat, male 1,000, 5,000 ppm Continuous, 14 weeks No neurobehavioral changes Haun et al.(1971)ICR mouse, female 1,000, 5,000 ppm Continuous, 14weeks Incoordination, lethargy Haun et al. (1971)Beagle dog, female 1,000, 5,000 ppm Continuous, 14weeks Incoordination, lethargy Haun et al. (1971)Rhesus monkey, female 1,000, 5,000 ppm Continuous,14 weeks Incoordination, lethargy Haun et al. (1971)ICR mouse, female 25, 100 ppm Continuous, 14 weeksIncreased spontaneous activity at 25 ppmThomas et al. (1972)Functional observational batteryF344 rat, male and female 50, 200, 2,000 ppm 6 hours/day, 5 days/week, 13weeks + 65 hours exposure freeNo effects observed on FOB, gripstrengthMattsson et al. (1990)Learning and memorySwiss-Webster mouse,male47,000 ppm Approximately 20 seconds +1 hour exposure free beforetraining; retested at days 1, 2,and 4Significant decrease in learning andrecall abilityAlexeef and Kilgore (1983)4574Table 4-26. Studies of neurobehavioral changes from dichloromethane, by route of exposure and type ofeffectSpecies Exposure(s) Duration Neurobehavioral effectReferenceOral and gavage exposureFunctional observational batteryF344 rat, female 101, 337, 1,012,1,889 mg/kg, gavageAcute—evaluated 4 and 24hours after dosingFOB neuromuscular and sensorimotorparameters significantly different fromcontrols at 1,012 and 1,889 mg/kg(337 mg/kg = NOAEL)Moser et al. (1995)F344 rat, female 34, 101, 337, 1,012mg/kg-day, gavage14 day—evaluated on days 4, 9,and 15.All FOB parameters (except activity)significantly affected from day 4at dosesof 337 and 1,012 mg/kg-dayMoser et al. (1995)13845754576aSEP = somatosensory-evoked potential; FEP = flash-evoked potential; BAER = brainstem-auditory-evoked response; CAEP = cortical-auditory-evoked4577potential; i.p. = intraperitoneal.45784579Table 4-27. Studies of neurophysiological changesas measured by evoked potentials resulting fromdichloromethane, by route of exposureSpecies Exposure(s) Duration SEPsameasured Effect ReferenceIntraperitonealLong-Evans rat, male 57.5, 115, 230,460 mg/kg, i.p.aAcute; tested at15 minutes, 1 hour, and5 hours after dosingFEPaSignificant changes in FEPs were notedin animals dosed 115 mg/kg and higher;FEP changes time and dose dependentHerr and Boyes(1997)Inhalation ExposureF344 rat, male 5,000, 10,000,15,000 ppmAcute, 1 hour; testedduring exposureElectroencephalogram,BAERa, CAEPa, FEP,SEPSignificant changes in SEP, FEP,BAER, and CAEP responses at allexposures; slight recovery noted at1 hour after exposureRebert et al. (1989)F344 rat, male and female 50, 200,2,000 ppmSubchronic, 6 hour/day,5 days/week, 13 weeks;tested 65 hours afterlast exposureFEP, CAEP, BAER,SEPNo significant changes noted in anyevoked potential measurementsMattsson et al.(1990)1394580a↑= increase;↓= decrease.4581Table 4-28. Studies of neurochemical changes fromdichloromethane, by route of exposureSpecies and sex Exposure Duration Regions EffectaReferenceOral exposureSprague-Dawley rat, male 534 mg/kg Acute, single dose; evaluated2 hours after dosedHippocampus,medulla, midbrain,hypothalamus↑acetylcholine in hippocampus↑dopamine and serotonin in medulla↓norepinephrine in midbrain↓norepinephrine and serotonin inhypothalamusKanada et al.(1994)Inhalation exposureWistar rat, male 1,000 ppm TWA (basalexposure of 100 ppm +2,800 ppm, 1 hour peakexposures at hours 1 and 4)6 hours/day, 5 days/week,2 weeksCerebrum,cerebellum↑NADPH diaphorase, succinatedehydrogenase in cerebrum↑cerebral RNA↓succinate dehydrogenase incerebellumSavolainen et al.(1981)Wistar rat, male 1,000 ppm TWA 6 hours/day, 5 days/week,2 weeks + 7 days exposurefreeCerebrum,cerebellum↓succinate dehydrogenase in bothregionsSavolainen et al.(1981)Wistar rat, male 1,000 ppm 6 hours/day, 5 days/week,2 weeksCerebrum,cerebellum↑acid proteinase↓succinate dehydrogenase incerebellumSavolainen et al.(1981)Wistar rat, male 1,000 ppm 6 hours/day, 5 days/week,2 weeks + 7 days exposurefreeCerebrum↓cerebral RNA Savolainen et al.(1981)Sprague-Dawley rat, male 70, 300, 1,000 ppm 6 hours/day, 3 days Caudate nucleus—medial↑catecholamine levels (70 ppm)↓catecholamine levels (300 and 1,000ppm)No effect on luteinizing hormonereleaseFuxe et al. (1984)Mongolian gerbil, male andfemale210, 350 ppm Continuous (24 hours/day),3 months + 4 monthsexposure freeHippocampus,cerebellumcerebral cortex↓DNA concentration per wet weight inhippocampus (210, 350 ppm) andcerebellar hemispheres (350 ppm)↑astroglial proteins in frontal andsensory motor cerebral cortexRosengren et al.(1986)Mongolian gerbil, male andfemale210 ppm Continuous (24 hours/day),3 monthsFrontal cortex,cerebellum↓glutamate, GABA,phosphoethanolamine in frontal cortex↑glutamate, GABA in posteriorcerebellar vermisBriving et al.(1986)Mongolian gerbil, male andfemale210 ppm Continuous (24 hours/day),3 months + 4 monthsexposure freeHippocampus,olfactory bulbs,cerebral cortex↓DNA concentration per wet weight inhippocampus onlyKarlsson et al.(1987)140

4.4.3.1. Neurotoxicology Studies—Oral Exposures 4582 Three studies evaluated the neurotoxic potential o f dichloromethane by either 4583 administering the solvent orally or by injection; t wo of these studies (Herr and Boyes, 1997; 4584 Kanada et al., 1994) only evaluated acute effects ( 2–5 hours) from single-dose exposures. 4585 Observed neurological effects included decreased sp ontaneous activity (Moser et al., 1995), 4586 changes in flash-evoked potential (FEP) measurement s (Herr and Boyes, 1997), and changes in 4587 catecholamine levels in the brain (Kanada et al., 1 994). 4588 Moser et al. (1995) conducted neurobehavioral eval uations in female F344 rats following 4589 an acute or 14-day oral administration of dichlorom ethane. A FOB protocol was utilized to 4590 determine changes in autonomic parameters (lacrimat ion, salivation, pupil response, urination, 4591 defecation), neuromuscular parameters (gait, righti ng reflex, forelimb and hind-limb grip 4592 strength, landing foot splay), sensorimotor paramet ers (tail pinch, click response, touch 4593 response), excitability measures (handling reactivi ty, arousal, clonic, and/or tonic movements), 4594 and activity (rearing, motor activity). A baseline FOB was performed on all rats prior to initial 4595

dichloromethane administration. After dichlorometh ane administration, a FOB was conducted at 4596 selected time points followed by a motor activity t est in a maze. In the acute study, rats were 4597 dosed with 0, 101, 337, 1,012, or 1,889 mg/kg dichl oromethane. At 4 and 24 hours after the 4598 administered dose, rats were tested for the neurolo gical parameters. Significant changes in the 4599 neuromuscular and sensorimotor parameters were obse rved and occurred mostly in rats 4600 administered with the highest dose. These signific ant changes were only observed at the 4-hour 4601 time point and not when measured at 24 hours. The NOAEL identified by the authors for this 4602 study was 337 mg/kg, based on no observable changes in the FOB. In the 14-day study, rats 4603 were administered 0, 34, 101, 337, or 1,012 mg/kg-d ay. FOB testing was conducted on days 4 4604 and 9 (before the daily dose) and approximately 24 hours after the last (14th) dose. With the 4605 exception of the activity measurements, all other n eurobehavioral parameters (neuromuscular, 4606 sensorimotor, autonomic, excitability) were signifi cantly affected from the 4th day through the 4607 entire 14-day exposure cycle. The NOAEL identified for the 14-day study was 101 mg/kg-day, 4608 based on FOB changes associated with the dichlorome thane exposure. 4609 A single dose acute neurophysiology study by Herr a nd Boyes (1997) evaluated the effect 4610 of dichloromethane on FEPs in adult male Long-Evans rats. Rats were implanted with epidural 4611 electrodes over the visual cortex area. After plac ement in an enclosed rectangular mirror 4612 chamber, PEPs were stimulated with a 10 μ sec flash. Baseline FEPs were collected and rats 4613 were injected intraperitoneally with 0 (corn oil, n = 16), 57.5 (n = 15), 115 (n = 15), 230 (n = 14), 4614 or 460 (n = 15) mg/kg dichloromethane. Animals wer e retested at 15 minutes, 1 hour, and 4615 5 hours after injection. Amplitude decreases in th e early FEP components were observed. The 4616 FEP amplitude changes were time and dose dependent with maximal effects at 15 minutes after 4617 dichloromethane dosage. All of the waveform amplit udes returned to control levels when 4618 measured at the 1-hour time point for all doses tes ted. Response latencies were still different 4619141

from controls when measured 5 hours after dosing, b ut the effect was less pronounced than at the 4620 15-minute and 1-hour time points. In this study, 5 7.5 mg/kg did not produce any significant 4621 changes in the FEP measures as compared to control and was considered this study’s NOAEL. 4622 The LOAEL was 115 mg/kg based on changes in the FEP amplitudes. 4623 Kanada et al. (1994) examined the effect of dichlor omethane on acetylcholine and 4624 catecholamines (dopamine, norepinephrine, serotonin ) and their metabolites in the midbrain, 4625 hypothalamus, hippocampus, and medulla from male Sp rague-Dawley rats (4–5/group) in a 4626 neurochemical/neuropathology study. The rats were sacrificed 2 hours after a single gavage dose 4627 of 0 or 534 mg/kg of undiluted dichloromethane. Ad ministration of dichloromethane 4628 significantly increased the concentration of acetyl choline in the hippocampus by approximately 4629 10% and increased dopamine and serotonin levels in the medulla by approximately 75%. 4630 Dichloromethane decreased norepinephrine levels in the midbrain and hypothalamus by 12– 4631 15%, and serotonin levels were decreased in the hyp othalamus by approximately 30%. There 4632 was a trend toward decreased dopamine in the hypoth alamus, but the variability between the 4633 animals was so high that the effect was not signifi cant. (These values for the percent changes 4634 were estimated by EPA from the figures presented in the paper.) The authors speculated that 4635 increased acetylcholine release associated with exp osure to dichloromethane and other solvents 4636 may originate from the nerve terminals. 4637

4638

4.4.3.2. Neurotoxicology Studies—Inhalational Exposure 4639 The database pertaining to neurotoxic effects from inhalation exposure to 4640 dichloromethane is considerably larger than the ora l exposure database. Acute (less than 1 day) 4641 and short-term (1–14 days) exposures resulted in an initial increase in spontaneous activity 4642 followed by a decrease for exposures between 500 an d 2,500 ppm (Kjellstrand et al., 1985; 4643 Savolainen et al., 1977). Higher (5,000 ppm) acute and short-term exposures resulted in 4644 decreased spontaneous activity and lethargy (Weinst ein et al., 1972; Heppel and Neal, 1944). 4645 Longer-term exposures (up to 14 weeks) produced dec reased motor activity and lethargy in 4646 several animals at 1,000 and 5,000 ppm (Haun et al. , 1971), and exposures at 25 ppm for 4647 14 weeks produced significant increases in activity in mice, starting at week 9. CNS depression 4648 was evidenced by decreased responses in the auditor y, visual, and somatosensory regions of the 4649 brain in a study of sensory-evoked potential effect s in 12 adult male F344 rats exposed to 0, 4650 5,000, 10,000, and 15,000 ppm for 1 hour periods (R ebert et al., 1989). Altered learning and 4651 memory abilities were demonstrated in young (3-, 5- , and 8-week-old) male Swiss-Webster mice 4652 exposed to 168 mg/L (~47,000 ppm) dichloromethane fo r approximately 20 seconds (until there 4653 was a loss of the righting reflex) (Alexeef and Kil gore, 1983). 4654

4655142

4.4.3.2.1. Inhalational exposure—neurobehavioral studies. 4656 Spontaneous motor activity—acute and short-term stu dies 4657 Heppel and Neal (1944) evaluated the neurological effects of 5,000 ppm dichloromethane 4658 in five male rats by measuring changes in spontaneo us activity during and after exposure. The 4659 five rats were not randomly selected, since the inv estigators chose to pick out the most active 4660 animals in the litter. During the 1-hour testing r uns, rats were placed in a rotating drum. 4661 Spontaneous activity was reported as the number of drum revolutions/hour. Twenty control test 4662 runs (1 run/day) were conducted prior to dichlorome thane exposure runs. After the pre-exposure 4663 period, rats were exposed to 5,000 ppm dichlorometh ane every other day for 1 hour, and activity 4664

was measured in the same manner as in the control r uns. Once dichloromethane exposure was 4665 stopped, the animals were allowed to recover for 30 minutes and a second 1-hour test run was 4666 performed to evaluate spontaneous activity during r ecovery. On nonexposure days, spontaneous 4667 activity was also measured in1-hour intervals to co mpare to the pre-exposure period. A total of 4668 five dichloromethane exposures, five postexposure, and five nonexposure trials were conducted 4669 over 10 days. Spontaneous activity significantly d eclined ( p < 0.01, Fisher’s t-test) during 4670 exposure to 5,000 ppm dichloromethane in comparison to nonexposure days. The average 4671

number of revolutions for all five rats over the te st runs was 576 on nonexposure days and 59 4672 revolutions during dichloromethane exposure. 4673 Weinstein et al. (1972) continuously exposed female ICR mice to 5,000 ppm 4674 dichloromethane for up to 7 days. Clinical behavio ral observations of the mice were made 4675 during dichloromethane exposure. Within the first few hours of exposure, spontaneous activity 4676 increased in comparison to control animals. After 24 hours of continuous exposure, there was a 4677 considerable decrease in spontaneous activity as no ted by observation only. The mice also 4678 appeared to be very lethargic and had a hunched pos ture and a rough hair coat, which are all 4679 signs of CNS depressive effects in rodents. These effects became progressively worse until after 4680 96 hours of exposure, where many mice resumed norma l activity. After the 7-day exposure, 4681 mice were nearly as active as the control animals b ut had a rougher coat and were judged to be 4682 emaciated and dehydrated. 4683 Male Wistar rats exposed to 500 ppm dichloromethan e 6 hours/day for 6 days exhibited 4684 an increase in preening frequency and time 1 hour a fter the last exposure relative to controls 4685 (Savolainen et al., 1977). However, there were no significant changes in other types of 4686 spontaneous activity. 4687 In the study by Kjellstrand et al. (1985), male NMR I mice were exposed to 4688 dichloromethane concentrations ranging from 400 to 2,500 ppm. At concentrations of 600 ppm 4689 and higher, exposures for 1 hour produced a biphasi c pattern of activity characterized by an 4690 initial increase in activity (as high as 200% of pr eexposure motor activity at 2,200 ppm, as 4691 estimated from Figure 6 in Kjellstrand et al. [1985 ]) during exposure followed by a decreased 4692 that reached the lowest point 1–2 hours after the e nd of exposure (as low as 40% motor activity 4693143

at 2,200 ppm, in comparison to preexposure, as esti mated from Figure 6 in Kjellstrand et al. 4694 [1985]). Motor activity returned to normal levels after the decreased activity observed 1–2 hours 4695 after exposure was stopped and indicated that the e ffect was reversible in this study design. 4696

4697 Spontaneous motor activity—subchronic (14 week) stu dies 4698 Haun et al. (1971) reported results from studies i n which female beagle dogs, female 4699 rhesus monkeys, male Sprague-Dawley rats, and femal e ICR mice were continuously exposed to 4700 0, 1,000, or 5,000 ppm dichloromethane for up to 14 weeks in whole-body exposure chambers. 4701 Gross and histopathologic examinations were made on animals that died or were sacrificed 4702 during or at termination of the study. At 5,000 pp m, obvious nervous system effects (e.g., 4703 incoordination, lethargy) were most apparent in dog s and also observed in monkeys and mice. 4704 Rats did not demonstrate any of these sedative effe cts. At 1,000 ppm, these effects were 4705 observed to a lesser extent in monkeys and mice, bu t dogs still displayed prominent CNS 4706 depressive behavior. Histopathologic analysis reve aled edema of the brain in three dogs that 4707 died during exposure to 5,000 ppm dichloromethane. No other gross brain-related changes were 4708 reported. The results indicate that continuous exp osure to 1,000 ppm was an adverse effect level 4709 for mortality and effects on the nervous system and liver in dogs (exposed for up to 4 weeks) and 4710 for BW changes in rats (exposed for 14 weeks). The 5,000 ppm level induced mortality in 4711 beagle dogs, ICR mice, and rhesus monkeys (but not Sprague-Dawley rats); obvious nervous 4712 system effects in dogs, mice, monkeys, and rats; an d gross liver changes in dogs, mice, monkeys, 4713

and rats. 4714 In the study by Thomas et al. (1972), female ICR mi ce were exposed continuously to 0, 4715 25, or 100 ppm dichloromethane for 14 weeks. Spont aneous activity of mice was evaluated by 4716 using closed circuit television for monitoring. Mi ce were evaluated in daily 2-hour testing 4717 sessions. The 25 and 100 ppm exposure groups were tested for 2 weeks prior to the onset of 4718 dichloromethane exposure. Starting at week 9, mice exposed to 25 ppm dichloromethane 4719 exhibited increases in spontaneous activity, but no quantitative measurements or statistical 4720 analysis were reported. The authors stated that no significant effect was observed in the group 4721 exposed to 100 ppm. 4722

4723 FOB—subchronic (13 week) study 4724 Only one study, a 13-week inhalation study in F344 rats (Mattsson et al., 1990) has 4725 conducted an FOB testing paradigm following a subch ronic exposure to dichloromethane. 4726 Groups of rats (12/sex/exposure level) were exposed to 0, 50, 200, or 2,000 ppm 4727 dichloromethane 6 hours/day, 5 days/week for 13 wee ks. An additional group of rats was 4728 exposed to 135 ppm CO to induce approximately 10% C OHb, approximately the level produced 4729 by saturation of oxidative metabolism of dichlorome thane. After the 13 weeks of exposure 4730 (beginning 65 hours after the last exposure), rats were subject to an FOB to evaluate any 4731144

neurobehavioral changes from the dichloromethane ex posure. Autonomic parameters were first 4732 characterized and then the rat was placed in a clea r plastic box to evaluate locomotor activity and 4733 then responsiveness to touch, sharp noise, and tail pinch. Hind-limb grip strength was also 4734 measured by using a strain gauge. All animals were examined clinically at weekly intervals and 4735 were tested at the end of the exposure period by FO B, grip strength, BW, temperature, and 4736 sensory-evoked potentials. No exposure-related eff ects were observed on the FOB, grip 4737 strength, or sensory-evoked potentials. No histopa thologic changes were noted in brains, spinal 4738 cords, or peripheral nerves from the high-dose dich loromethane group compared with control 4739 animals. In the absence of changes, lower concentr ations were not examined. 4740

4741 Learning and memory—acute study 4742 In a study by Alexeef and Kilgore (1983), a learni ng and memory evaluation was 4743 conducted following acute exposure to dichlorometha ne. Mice were exposed to 168 mg/L 4744 (~47,000 ppm) dichloromethane and were tested for le arning ability by using a passive4745 avoidance conditioning task. Male Swiss-Webster mi ce (3, 5, and 8 weeks old) were used in this 4746 study. In the passive avoidance task, mice were pl aced on a metal platform that extended into a 4747 hole. If the mouse went into the hole (a darkened area, which would be the preferred area for the 4748 mouse), it received a foot shock. Prior to the tra ining session, mice were exposed to either air or 4749 ~47,000 ppm dichloromethane. Animals were exposed t o dichloromethane until there was a loss 4750 of the righting reflex, which would take about 20 s econds on average, and then placed back in 4751 their home cage. One hour after exposure, animals were trained to learn the passive avoidance 4752

task. A mouse was considered to have learned the t ask once it remained on the platform for at 4753 least 30 seconds without entering the hole. Mice w ere then tested for recollection of the task at 4754 either 1, 2, or 4 days after the initial training s ession. In the learning phase of the task, 74% of 4755 the control mice retained the task in comparison to 59% of the dichloromethane-exposed group, 4756 indicating the significant effect of dichloromethan e on learning. There was also an age-related 4757 effect since exposed 3-week-old mice were less like ly to recall the task than five- or eight-week4758

old mice. There was no difference in task recall b etween the 5- and 8-week-old mice. 4759 Dichloromethane, at the exposure used in the study, was demonstrated to be non-analgesic, since 4760 pain-response times were comparable to those in air -exposed animals in the hot-plate pain test, 4761 and therefore the results of the passive avoidance test were not confounded by potential analgesic 4762 effects. As a result, it is demonstrated that expo sure to an acute and high concentration of 4763 dichloromethane alters learning ability in mice. 4764

4765

4.4.3.2.2. Inhalational exposure—neurophysiological studies. The effect of dichloromethane 4766 on sensory stimuli was evaluated by measuring senso ry-evoked responses during an acute 4767 exposure (Rebert et al., 1989) and following a subc hronic (13 week) exposure (Mattsson et al., 4768 1990). Rebert et al. (1989) evaluated the effects of dichloromethane on sensory-evoked 4769145

potentials (auditory, visual, and somatosensory) in F344 rats exposed to 0, 5,000, 10,000, and 4770 15,000 ppm dichloromethane for 1 hour in a head-onl y exposure chamber. Twelve adult male 4771 rats were implanted with chronic epidural electrode s placed over the visual and somatosensory 4772 cortices. Each rat served as its own control, with a 1-week recovery period between testing 4773 sessions. During each testing session, spontaneous electroencephalograms were recorded. 4774 Additionally, brainstem-auditory-evoked responses ( BAERs) (tone stimulus), cortical-auditory4775 evoked potentials (CAEPs) (click stimulus), FEPs (f lash stimulus), and somatosensory-evoked 4776 potentials (SEPs) (tail current stimulus) were meas ured in response to the stimuli. 4777 Dichloromethane decreased the SEP response to the t ail current stimulus, and earlier components 4778 of the FEP response were attenuated and eventually eliminated with increasing exposures. The 4779 BAER response profile was also significantly altere d. Dichloromethane completely abolished 4780 the CAEP at all concentrations tested. Slight reco very of this response was noted approximately 4781 1 hour after exposure. The collective results stro ngly suggest a CNS depressive profile for 4782

dichloromethane and indicate that this chemical aff ects the auditory, visual, and somatosensory 4783 regions of the brain. 4784 In a subchronic exposure study, male and female F34 4 rats were exposed to 4785 dichloromethane for 6 hours/day, 5 days/week for 13 weeks (Mattsson et al., 1990). Twelve 4786 animals of each sex were selected for exposure to 0 , 50, 200, or 2,000 ppm dichloromethane or 4787 135 ppm CO. For electrophysiological measures, rat s were surgically implanted with epidural 4788 electrodes 10 weeks after the onset of exposure. E lectrodes were placed over the somatosensory, 4789 visual, and cerebellar region. Electrophysiologica l measures that were recorded included FEP 4790 measurements, cortical flick fusion responses, CAEP s, BAERs, and SEPS recorded from the 4791 sensory (SEP-S) and cerebellar (SEP-C) regions. No ne of these measures were significantly 4792 altered by any dichloromethane or CO treatment in t his study. However, it should be noted that 4793 all of the electrophysiological measures were condu cted at least 65 hours after the last 4794 dichloromethane exposure. As a result, it can be c oncluded that a subchronic exposure to 4795 dichloromethane did not result in persistent change s in any of the neurophysiological measures 4796 that were evaluated in this study. It is not known if any neurological compensation occurred 4797 since SEP measurements were not taken during actual dichloromethane exposure in this 4798 subchronic study. 4799 Based on these two studies, the significant change s noted in several SEP measures during 4800 dichloromethane exposure were not observed after a subchronic exposure where animals were 4801 tested at least 65 hours after the last exposure. As a result, it is difficult to ascertain if tolera nce 4802 is developed to the dichloromethane-mediated change s in sensory potentials during an acute 4803 exposure or if these effects are still maintained d uring repeated exposure, since measurements 4804 were not taken during the subchronic exposure. 4805146

4.4.3.2.3. Inhalational exposure—neurochemistry and neuropatho logy studies. The studies 4806 evaluating specific neurochemical changes in relati on to dichloromethane exposure include 4807 studies of effects of short-term (3-day to 2-week) exposures (Fuxe et al., 1984; Savolainen et al., 4808 1981) and subchronic (3-month) exposures (Karlsson et al., 1987; Briving et al., 1986; 4809 Rosengren et al., 1986). 4810 Savolainen et al. (1981) examined three different e xposure schemes in male Wistar rats. 4811 The rats were exposed to 500, 1,000, or 1,000 ppm T WA dichloromethane for 6 hours/day, 4812 5 days/week for 2 weeks. (Note: The abstract of thi s paper describes the exposures as 500, 1,000, 4813 and 100 ppm TWA, but, based on information in the b ody of the paper, the abstract appears to be 4814 incorrect.) The 1,000 ppm TWA exposure consisted o f a basal 100 ppm exposure with two 4815 2,800 ppm 1-hour peak concentrations (at 1 and 4 ho urs) resulting in a time-weighted exposure 4816 of 1,000 ppm. Brains were removed from rats at the end of study and analyzed. The 1,000 ppm 4817 TWA group displayed increases in cerebral RNA. Oth er changes noted for this group in the 4818 cerebrum included significant increases in NADPH di aphorase and succinate dehydrogenase 4819 activity. In the 1,000 ppm constant exposure group , acid proteinase activity was below the levels 4820

observed in control animals in the first week but i ncreased to levels above control animals in the 4821 second week. In the cerebellum, there were no chan ges in RNA concentration, and there was a 4822 decrease in succinate dehydrogenase activity in bot h the 1,000 and 1,000 ppm TWA groups. 4823 After a 7-day withdrawal, RNA levels in the cerebru m were significantly lower in the 1,000 ppm 4824 group. Succinate dehydrogenase levels remained low ered in the 1,000 ppm TWA group after the 4825 7-day exposure-free period. No significant effects were seen at 500 ppm. 4826 Fuxe et al. (1984) evaluated changes in brain cate cholamine levels after a 3-day exposure 4827 to dichloromethane, using male Sprague-Dawley rats. Rats were exposed to 70, 300, and 4828 1,000 ppm dichloromethane 6 hours/day for 3 consecu tive days. Additional groups of rats were 4829 exposed to the same levels of dichloromethane and g iven intraperitoneal injections of the 4830 tyrosine hydroxylase inhibitor, α -methyl-dl- p -tyrosine methyl ester (H44/68), 2 hours prior to 4831 sacrifice. Brains were removed, stained, and evalu ated for catecholamine changes 16–18 hours 4832 after the last exposure. Catecholamine levels were measured in the hypothalamus, frontal cortex, 4833 and caudate nucleus among other brain regions. At all exposures, there was a significant 4834 decrease by approximately 10–15% of catecholamine c oncentrations in the posterior 4835 periventricular region of the hypothalamus. In the medial part of the caudate nucleus, which is 4836 involved in memory processes, catecholamine levels were significantly higher (12%) in the 4837 70 ppm group but significantly lower in the 300 ppm (1%) and 1,000 ppm (8%) groups 4838 compared with controls. The impact of dichlorometh ane was also evaluated on the 4839 hypothalamic-pituitary gonadal axis. The hypothala mus regulates secretion of reproductive 4840 hormones, such as follicle-stimulating hormone and luteinizing hormone. The levels of the 4841 hormone release were not significantly changed with dichloromethane exposure. However, 4842 when rats were dosed concurrently with H44/68 and d ichloromethane, statistically significant 4843147

inversely dose-related increases in luteinizing hor mone levels were observed (330, 233, and 4844 172% higher than controls in the 70, 300, and 1,000 ppm groups, respectively). The study 4845 overall demonstrates significant changes in catecho lamine levels in the hypothalamus and 4846 caudate nucleus. No significant changes in catecho lamine levels in the frontal cortex were 4847 reported. Catecholamine level changes in the hypot halamus did not appear to significantly affect 4848 hormone release; however, decreased catecholamine l evels in the caudate nucleus at higher 4849 exposures may lead to memory and learning impairmen t. 4850

A series of studies were conducted in male and fema le Mongolian gerbils exposed 4851 continuously to 210 ppm (Karlsson et al., 1987; Bri ving et al., 1986), 350 ppm, or 700 ppm 4852 (Rosengren et al., 1986) dichloromethane for 3 mont hs, followed by a 4-month exposure-free 4853 period. High mortality rates occurred at 350 ppm ( 6/10 males and 3/10 females by 71 days) and 4854 700 ppm (10/10 males and 9/10 females by 52 days). Rosengren et al. (1986) monitored two 4855 astroglial proteins, S-100 and GFA, as well as DNA concentrations in the brain. Decreased 4856 DNA concentrations were noted in the hippocampus at both the 210 and 350 ppm exposures. At 4857 350 ppm, there was also decreased DNA concentration in the cerebellar hemispheres, indicating 4858 a decreased cell density in these regions, probably due to cell loss. Increased astroglial proteins 4859 were found in the frontal and sensory motor cerebra l cortex, which directly correlated to the 4860 astrogliosis that was observed in those areas. Up- regulation of these astroglial proteins is a good 4861 indicator of neuronal injury (Rosengren et al., 198 6). 4862 Karlsson et al. (1987) measured DNA concentrations in different regions of the gerbil 4863 brain. After the solvent-free exposure period, bra ins were removed and the olfactory bulbs and 4864 cerebral cortices were dissected. Brain weights an d weights of the dissected brain regions were 4865 the same between control and dichloromethane-expose d animals. The total protein concentration 4866 per wet weight was not significantly different betw een dichloromethane-exposed and control 4867 animals. However, DNA concentrations per wet weigh t were significantly decreased in the 4868 hippocampus after dichloromethane exposure. No oth er examined regions demonstrated 4869 significant changes in DNA concentrations after dic hloromethane exposure. This selective DNA 4870 concentration decrease observed in the hippocampus is a sign of neurotoxicity and may possibly 4871 explain why some studies have noted memory and lear ning deficits with dichloromethane 4872 exposure. In a companion paper, in which only the 210 ppm level was tested, it was found that 4873 exposure to dichloromethane decreased the levels of glutamate, γ -aminobutyric acid, and 4874 phosphoethanolamine in the frontal cortex, while gl utamine and γ -aminobutyric acid were 4875 increased in the posterior cerebellar vermis (Brivi ng et al., 1986). Increased levels of glutamate 4876 in the posterior cerebellar vermis could reflect an activation of astrocytic glia, since glutamine 4877 synthetase is localized exclusively in astrocytes. The gerbils did not have a solvent-free 4878 exposure period as in the other two studies (Karlss on et al., 1987; Rosengren et al., 1986). The 4879 exposure regime in these studies did not affect BW or brain weight. Furthermore, the 4880 neurochemical changes observed in these studies wer e not attributed to formation of CO. 4881148

Neurological changes have been investigated by measuring changes in neurotransmitter 4882 levels and changes in neurotransmitter localization . Changes in catecholamine levels in the 4883 caudate nucleus after an acute exposure (Fuxe et al ., 1984) as well as decreased DNA content in 4884 the hippocampus after a subchronic dichloromethane exposure (Rosengren et al., 1986) suggest 4885 that memory functions are altered since both brain regions are associated with learning and 4886 memory. The results from Fuxe et al. (1984) direct ly correlated with the finding that learning 4887 and memory were impaired in mice after an acute (si ngle) and very high exposure (47,000 ppm) 4888 to dichloromethane (Alexeef and Kilgore, 1983). Ad ditionally, changes in the hippocampus also 4889 suggest memory effects after a long-term, continual exposure to dichloromethane, although no 4890 conclusive evidence has been presented to date. In another subchronic, continuous exposure to 4891 350 ppm dichloromethane for 3 months, decreased DNA concentration was observed in the 4892 cerebellar hemispheres of Mongolian gerbils and is suggestive of cell loss (Rosengren et al., 4893 1986). However, in a 2-week exposure study in male Wistar rats, RNA changes were not noted 4894 in the cerebellum, although enzyme activity was sig nificantly decreased in this region (but was 4895 increased in the cerebrum) (Savolainen et al., 1981 ). These results suggest that the cerebellum is 4896 a target for dichloromethane. Noted neurobehaviora l effects that may be linked to impaired 4897 cerebellar function include changes in motor activi ty and impaired neuromuscular function 4898 (Moser et al., 1995). 4899

4900

4.5. MECHANISTIC DATA AND OTHER STUDIES IN SUPPORT OF THE MODE OF 4901

ACTION 4902 4.5.1. Genotoxicity Studies 4903

4.5.1.1. In Vitro Genotoxicity Assays 4904 Bacterial, Yeast, and Fungi mutagenicity assays 4905

Numerous in vitro studies have demonstrated dichlo romethane as being mutagenic in 4906 bacterial assays, yeast, and fungi, and several stu dies provide evidence that the genotoxic action 4907 of dichloromethane in bacterial systems is enhanced in the presence of GSH (e.g., Dillon et al., 4908 1992; Their et al., 1993; Oda et al., 1996; DeMarin i et al., 1997; Pegram et al., 1997) (Table 4- 4909 29). Considering the results are primarily depende nt on the presence of GSH, activation likely 4910 involves the GST-T1 metabolic pathway, which produc es two proposed DNA-reactive 4911 metabolites, S-(chloromethyl)glutathione and formal dehyde. 4912

Dichloromethane induced mutations in Salmonella typhimurium strains containing GSH 4913 (e.g. TA100, TA98). These effects were not markedl y influenced by the addition of exogenous 4914 mammalian liver fractions, suggesting that endogeno us metabolism in these strains was 4915 sufficient to activate dichloromethane (Green, 1983 ; Jongen et al., 1982; 1978; Gocke et al., 4916 1981). In support of this hypothesis, dichlorometh ane exposure of NG-11, a glutathione4917

deficient variant of S. typhimurium strain TA100, produced twofold fewer base-pair mut ations 4918149Table 4-29. Results from in vitro genotoxicity assays of dichloromethane with bacteria, yeast, or fungiResultsAssay Test system Concentration(s)Without metabolicactivation–S9With metabolicactivation+S9 ReferenceBacteriaReverse mutationS. typhimuriumTA98a, TA100a6-hour exposure to 0, 7,000,and 14,000 ppm+ + Jongen et al. (1978)Reverse mutationS. typhimuriumTA98, TA100Up to 3,600μg/plate + ++ Gocke et al. (1981)Reverse mutationS. typhimuriumTA1535b, TA1537b,TA1538bUp to 3,600μg/plate – – Gocke et al. (1981)Reverse mutationS. typhimuriumTA1006-hour exposure to 0, 7,000,and 14,000 ppm+ ++ Jongen et al. (1982)Reverse mutationS. typhimuriumTA100Up to 84,000 ppm, 3-dayexposure+ + Green (1983)Reverse mutationS. typhimuriumTA100, TA1535,TA1950a,E. coliWU361089a10μL/plate + for TA100, TA1950,WU361089– for TA1535Not determined Osterman-Golkar et al.(1983)Reverse mutationS. typhimuriumTA100+ NotdeterminedZeiger (1990)Reverse mutationS. typhimuriumTA100, NG54c2- and 6-hour exposures to 0,2,500, 5,000, 7,500, and10,000 ppm+ + Dillon et al. (1992)Reverse mutationS. typhimuriumTA100, TA1535 andTA1538 (+GSTA 1-1and GSTP 1-1)0, 50, 100, and 200μL/plate + for TA100– for TA1535, TA1538Not determined Simula et al. (1993)Reverse mutationS. typhimuriumTA1535 (+GST 5–5),TA1535 (wild type)0–2.0 mM/plate + for TA1535 (+GST 5–5)– for TA1535 (wild type)Not determined Pegram et al. (1997); Thieret al. (1993)Reverse mutationS. typhimuriumTA100, TA100/NG-11d0, 30, 60, 130 mM/plate ++ for TA100+ for TA100/NG-11Not determined Graves et al. (1994a)Reverse mutationS. typhimuriumTA100,RSJ100cUp to 24,000 ppm + for TA100+ for RSJ100+ for TA100+ for RSJ100DeMarini et al. (1997)150Table 4-29. Results from in vitro genotoxicity assays of dichloromethane with bacteria, yeast, or fungiResultsAssay Test system Concentration(s)Without metabolicactivation–S9With metabolicactivation+S9 ReferenceForward mutationS. typhimuriumBA13 0–130μmol/plate +++ + Roldán-Arjona and Pueyo(1993)Gene mutationS. typhimuriumTA1535/pSK1002c,NM5004c0, 2.5, 5.0, 10, 20 mM + NM5004– TA1535/ pSK100 2Not determined Oda et al. (1996)ProphageinductionE. coliK-39 (λ) 10μL/plate + Not determined Osterman-Golkar et al.(1983)Reverse mutationE. coliWP2 uvrapKM1012- and 6-hour exposures to6,300, 12,500, 25,000, and50,000 ppm+ + Dillon et al. (1992)Forward mutationE. coliK12 0, 30, 60, 130 mM/plate – + Graves et al. (1994a)Forward mutationE. coliUvr+, UvrB–20,000 ppm + Not determined Zielenska et al. (1993)Fungi and yeastsMitoticsegregationAspergillus nidulansUp to 8,000 ppm + only at 4,000 ppm; nodose-response relationshipestablishedNot determined Crebelli et al. (1988)Gene conversionand recombinationSaccharomycescerevisiaeUp to 209 mM + Not determined Callen et al. (1980)abacterial strains that have GSH (e.g. TA100, TA 98)bbacterial strains that do not have GSH (e.g. TA1535)cbacterial strains engineered to have more GSH activity than wild typedbacterial strains engineered to have less GSH activity than wild type151

compared with exposure of strain TA100, which produ ces normal levels of GSH. Furthermore, 4919 this difference was not apparent when the culture m edium contained 1 mM 4920 GSH (Graves et al., 1994a). 4921 In contrast to strain TA100, S. typhimurium strains TA1535, TA1537, and TA1538 4922 (strains deficient in GSH) did not develop base-pai r mutations in response to dichloromethane 4923 exposure (Gocke et al., 1981; Osterman-Golkar et al ., 1983; Simula et al., 1993; Thier et al., 4924 1993; Pegram et al., 1997). However, when strain T A1535 was transfected with rat GST-T1, 4925 dichloromethane induced base-pair reverse mutations (DeMarini et al., 1997; Pegram et al., 4926 1997; Thier et al., 1993). A 60-fold higher concen tration of dichloromethane was needed to 4927 induce a response (i.e., a sixfold increase over ba ckground levels in reverse mutations) in S. 4928 typhimurium strain TA100 than in TA1535 transfected with rat G ST-T1 (DeMarini et al., 1997). 4929 This study also included several trihalomethanes; d ichloromethane was several fold less 4930 genotoxic than dibromochloromethane or bromoform, b ut was similar in potency to 4931 bromodichloromethane (DeMarini et al., 1997; Pegram et al., 1997). The authors suggest that 4932 these results support a role of GST-T1 in the mutag enicity of the trihalomethanes. 4933 The mutagenic effects of dichloromethane have also been examined in fungi and yeasts 4934 assays with both systems reporting positive results . Fungi assays were positive for mitotic 4935 segregation in Asperigillus ridulans (Crebelli et al., 1988) but there was not a dose re sponse 4936 relationship as only the 4,000 ppm dichloromethane exposure was positive (exposure up to 8,000 4937 ppm). A yeast assay was positive for gene convers ion and recombination in Saccharomyces 4938

cerevisiae for concentrations up to 209 mM (Callen et al., 19 80). 4939

4940 Mammalian assays 4941 In the in vitro mammalian system studies conducted with murine cell lines (Table 4-30), 4942 dichloromethane was negative for producing point mu tations in the mouse lymphoma L5178Y 4943 cell line (Thilagar et al., 1984), but was positive in producing single stranded DNA breaks in 4944 mouse Clara cells (Graves et al., 1995) and mouse h epatocytes (1994b). Given that exposure to 4945 dichloromethane results specifically in lung and li ver tumors, this pattern is not surprising. 4946 Additionally, GST is localized in the nucleus of he patocytes and lung cells in the mouse 4947 (Mainwaring et al., 1996), which would also increas e sensitivity of these particular cell fractions 4948 to genotoxic effects of dichloromethane. DNA singl e strand breaks (SSBs) were induced at 4949 lower concentrations in mouse hepatocytes (0.5 mM) than in rat hepatocytes (30 mM). The 4950 extent of DNA damage was shown to be reduced to the background level seen in control (no 4951 exposure) conditions by pretreating the cells with buthionine sulfoxime to deplete cellular levels 4952

of GSH and thus inhibit dichloromethane metabolism via the GST pathway (Graves et al. 1995; 4953 1994b). Similar results were seen in mouse lung Cl ara cells. Freshly isolated Clara cells from 4954 the lungs of B6C3F 1 mice also showed significantly increased, concentr ation-dependent amounts 4955

of DNA SSBs when incubated in vitro for 2 hours in the presence of 5–60 mM dichloromethane. 4956152

Pretreatment with buthionine sulphoximine before Cl ara-cell isolation or the presence of 4957 buthionine sulphoximine in the culture medium decre ased the amount of in vitro DNA damage 4958

induced. 4959 In a series of experiments with freshly isolated he patocytes from multiple species (Table 4960 4-30), DNA-protein cross-links were detected in hep atocytes of B6C3F 1 mice but not in 4961 hepatocytes of F344 rats, Syrian golden hamsters, o r three human subjects, following 2-hour in 4962 vitro exposure to concentrations ranging from 0.5–5 mM dichloromethane (Casanova et al., 4963 1997). Within the range of concentrations tested, DNA-protein cross-links in mouse hepatocytes 4964 appeared to increase with increasing concentration of dichloromethane. 4965 Negative results for dichloromethane were predomina ntly seen in in vitro test systems 4966

that used rat or hamster cell lines with low or no GST activity (Table 4-30). Several genotoxic 4967 endpoints, including DNA and protein synthesis (Gar rett and Lewtas, 1983), chromosomal 4968 aberrations or sister chromatid exchanges (Thilagar et al., 1984; Thilagar and Kumaroo, 1983; 4969 Jongen et al., 1981), unscheduled DNA synthesis (Th ilagar et al., 1984; Andrae and Wolff, 1983; 4970 Jongen et al., 1981), and mutations (Thilagar et al ., 1984; Jongen et al., 1981) were evaluated in 4971 these cell lines. In contrast, positive results (D NA-protein cross links and DNA SSBs) were 4972 observed when mouse liver cytosol was included in C hinese hamster ovary (CHO) cells (Graves 4973 et al., 1994b; Graves et al., 1995). Dichloromethan e also induced hypoxanthine-guanine 4974 phosphoribosyl transferase (HPRT) gene mutations in CHO cells when they were incubated with 4975 GST-competent mouse liver cytosol preparations (Gra ves et al., 1996). 4976 The instability of the S-(chloromethyl)glutathione- adducts presents considerable 4977 challenges to studies of these products (Hashmi et al., 1994). Kayser and Vuilleumier (2001), 4978 however, demonstrated the formation of DNA adducts with radiolabeled dichloromethane in calf 4979 thymus DNA in the presence of dichloromethane dehal ogenase/GST purified from a bacterial 4980 source ( Methylophilus sp. strain DM11) and GSH (Table 4-30). The type o f adduct could not be 4981 identified because of low yield, but it was determi ned that guanine was more actively 4982 incorporated than cytosine, adenine or thymine by a t least 2 fold in the presence of GST4983 activated dichloromethane, indicating a base specif icity for these adducts. Incubation of calf 4984 thymus DNA with formaldehyde and GSH, however, did not result in detectable DNA adduct 4985 formation. In another study, Marsch et al. (2004) further evaluated the presence of adducts in 4986 calf thymus DNA in the presence of dichloromethane and human (GSTT1-1), rat (GST 5-5) or 4987 bacterial (DM11) GST (Marsch et al., 2004). This s tudy found that all three enzymes yielded a 4988 similar pattern of adduct formation, forming primar ily with guanine and to a lesser extent with 4989 cytosine, adenine, and thymine (2-3 fold less than guanine), consistent with the results reported 4990 by Kayser and Vuilleumier (2001). High levels of g uanosine-specific adducts were also seen 4991 with S-(1-acetoxymethyl)glutathione, a compound tha t is structurally similar, but more stable, 4992 than S-(chloromethyl)glutathione (Marsch et al., 20 01). These findings indicate that the S4993 (chloromethyl)glutathione intermediate formed by GS H conjugation has mutagenic potential and 4994153

is likely responsible, at least in part, for the mu tagenic response observed following 4995 dichloromethane exposure. 4996 In studies with human cell lines or isolated cells, positive results were reported for sister 4997 chromatid exchanges and chromosomal aberrations (Th ilagar et al., 1994) and in the 4998 micronucleus test (Doherty et al., 1996). Negative results with human cells were seen in the 4999 unscheduled DNA synthesis assays (Perocco and Prodi , 1981; Jongen et al., 1981), DNA SSBs, 5000 and DNA-protein cross-links (Graves et al., 1995; C asanova et al., 1997). 5001 Dichloromethane-induced DNA damage (comet assay) wa s examined in primary cultures 5002 of human lung epithelial cells collected by brush b iopsy from four healthy volunteers (Landi et 5003 al., 2003). This study was designed to assess the genotoxicity of four thrihalomethanes 5004 (chloroform, bromodichloromethane, dibromochloromet hane and bromoform), with 5005

dichloromethane included because of its known activ ation by GST-T1. Two of the subjects were 5006

of the GST-T1 + genotype, and two were of the GST-T1

– genotype. 6 The cells had been frozen, 5007 and GST activity was not detected in the cultured c ells. DNA damage was reported to occur in 5008

the combined GST-T1

– samples (tail extent moment 7.1, 13.7 and 15.3 in the 10, 100 and 1000 5009 μM dichloromethane groups, respectively), but not i n the combined GST-T1 + samples (tail 5010 extent moment 8.1, 11.5 and 10.4 in the 10, 100 and 1000 μM dichloromethane groups, 5011 respectively). This pattern was not seen across th e individual samples, however, as only one 5012 sample exhibited a clear dose-response gradient. G iven the absence of GST activity, an analysis 5013 combining the four samples could provide a more inf ormative picture of the dose-response 5014 relation between dichloromethane (and the other com pounds) studied and DNA damage. For 5015 dichloromethane, values of 9.4, 7.6, 12.6, and 12.9 were seen in the 0, the 10, 100 and 1000 μM 5016 groups, respectively. This pattern was similar to that seen with chloroform (9.4, 6.9, 11.4, and 5017 12.7 in the 0, the 10, 100 and 1000 μM groups, resp ectively), but weaker than the pattern for 5018 bromoform (9.4, 12.5, 15.8, and 18.2 in the 0, the 10, 100 and 1000 μM groups, respectively), 5019 and much weaker than for bromodichloromethane (9.4, 25.2, 28.5, and 39.1 in the 0, the 10, 100 5020 and 1000 μM groups, respectively). 7 No dose-response gradient was seen with 5021 dibromochloromethane (9.4, 6.5, 8.1 and 8.0 in the 0, the 10, 100 and 1000 μM groups, 5022 respectively). This relative pattern is also seen in the estimated slopes (beta coefficient for the 5023 change in tail extent moment per unit increase in μ M concentration): 0.0, 0.003, 0.004., 0.006 5024 and 0.02 for dibromochloromethane, dichloromethane, chloroform, bromoform, and 5025 bromodichloromethane, respectively (statistical sig nificance not reported). 5026

6 Landi et al. (2003) did not clearly describe their treatment of GST-T1 +/- heterozygote genotypes; the EPA believes

it is most likely they were included in the pool fr om which the GST-T1 + samples were drawn. In addition, there is a discrepancy in the paper regarding this the coding of the GST-T1 genotypes. Samples A and C are noted to be the GST-T1

– samples in one part of the paper, and C and D are described as the GST-T1

– samples in another part of the

paper. 7 These values are based on the mean of the GST-T1 + and the GST-T1 – samples from Table 1 of Landi et al. (2003)154

A stronger and more consistent response was seen un der the same experimental 5027 conditions with bromodichloromethane, but dibromoch loromethane resulted in no increase in 5028 DNA damage in any of the donor cells at any concent ration tested. 5029 Several studies have examined patterns of mutations or DNA damage with 5030 dichloromethane and formaldehyde to assess the rela tive role of S-(chloromethyl)glutathione and 5031 formaldehyde in the observed genotoxicity. In a s tudy in CHO cells incubated with 5032 dichoromethane (0.3% plus mouse liver cytosol), 2.5 -fold increases in DNA-protein cross-links 5033 that are indicative of formaldehyde exposure were o bserved, compared with a 25-fold increase 5034 when 1 mM formaldehyde was added directly to cultur es. Both treatments induced a comparable 5035 degree of DNA SSBs (Graves and Green, 1996). In a subsequent study, Graves et al. (1996) 5036 compared the mutational spectra induced by dichloro methane to that induced by direct addition 5037 of formaldehyde or 1,2-dibromoethane (a chemical kn own to act through a glutathionyl 5038 conjugate metabolite) at the HPRT locus in CHO cell s. The mutations induced by 5039 dichloromethane and 1,2-dibromoethane were predomin antly GC to AT transitions, while all six 5040 formaldehyde-induced mutants sequenced were single base transversions. This provided further 5041 evidence that the S-(chloromethyl)glutathione inter mediate may be primarily responsible for 5042 dichloromethane genotoxicity. In contrast, Hu et al. (2006) found evidence of significant 5043 amounts of formaldehyde formation following dichlor omethane exposure in the cytosol of V79 5044 (hamster) cells transfected with the murine GSTT1 g ene compared to the parent cell line. In 5045 accordance with this, they observed concentration-d ependent increases in DNA-protein 5046 crosslinks in the GSTT1 transfected cells using the comet assay with and without proteinase K 5047

treatment that frees DNA from crosslinks and allows DNA migration. These findings are 5048 consistent with those by Casanova et al. (1997), wh o performed a comparison of the amounts of 5049 DNA-protein and RNA-formaldehyde crosslinks formed following dichloromethane exposure in 5050 hepatocytes isolated from mice, rats, hamsters, and human GSTT1 genetic variants. Only DNA5051 protein crosslinks were observed in mouse hepatocyt es, but RNA-formaldehyde crosslinks were 5052 found in all species, which were highest in the mou se hepatocytes, and were 4-, 7-, and 14-fold 5053 higher than rats, humans, and hamsters. These resu lts showed that human hepatocytes can 5054 metabolize dichloromethane to formaldehyde, resulti ng in RNA-formaldehyde crosslinks. In 5055 addition, the results indicate, that there is consi derable variation among species, and that the 5056 human variation in the GSTT1 gene can also affect t he amount of formaldehyde produced. The 5057 authors also noted that comparing results following ectopic addition of formaldehyde directly to 5058 cells with results following dichloromethane metabo lism in situ can be misleading, as the 5059 formaldehyde produced internally may reside in diff erent locations intracellularly, potentially 5060 affecting the capability of interacting with DNA. These results show that, while most studies 5061 indicate the importance of the S-(chloromethyl)glut athione intermediate in mediating genotoxic 5062 damage following dichloromethane exposure, DNA dama ge resulting from formaldehyde 5063

formation should also be considered. 5064155Table 4-30. Results from in vitro genotoxicity assays of dichloromethane with mammalian systems, by typeof testAssay Test system Concentrations Results ReferenceMousePoint mutation Mouse lymphomaL5178Y cellsNot provided NegativeThilagar et al. (1984)DNA SSBs byalkaline elutionMouse hepatocytes 0, 0.4, 3.0, 5.5 mM Positive at0.4 mM Graves et al. (1994b)DNA SSBs byalkaline elutionMouse Clara cells 0, 5, 10, 30, 60 mMPositive, but DNA damage was reduced by incubatingin the presence of GSH depletoryGraves et al. (1995)DNA-proteincross-linksMouse hepatocytes 0.5–5 mM PositiveCasanova et al. (1997)RatUnscheduledDNA synthesisRat hepatocytes Up to 16 mM(measured); 30 mM(nominal)NegativeAndrae and Wolff (1983)UnscheduledDNA synthesisRat hepatocytes Not provided Marginally positive Thilagar et al. (1984)DNA SSBs byalkaline elutionRat hepatocytes 0, 30, 90, 90 mM Positive at 30 mMGraves et al. (1994b)DNA-proteincross-linksRat hepatocytes 0.5–5 mM NegativeCasanova et al. (1997)Hamster with GST activity from mouseDNA-proteincross-linksChinese hamster ovarycells60 mM Positive with mouse liver cytosol (negative without)at much higher concentrations of dichloromethane (60mM) than formaldehyde (0.5–4 mM).Graves et al. (1994b)HPRTamutationanalysisChinese hamster ovarycells2,500 ppm Positive with mouse liver cytosol Graveset al. (1996)DNA SSBs andDNA-proteincross-linksChinese hamster ovarycells3,000 ppm (0.3%,volume per volume[v/v]) and 5,000 ppm(0.5%, v/v)Positive at concentration of 0.5% (v/v) for SSBs inpresence of mouse liver cytosol, but increase in DNAprotein cross-links marginal; formaldehyde (inabsence of mouse liver cytosol) was positive at 0.5mM for both DNA SSBs and DNA-protein crosslinks; Chinese hamster ovary cell cultures weresuspended.Graves and Green (1996)156Table 4-30. Results from in vitro genotoxicity assays of dichloromethane with mammalian systems, by typeof testAssay Test system Concentrations Results ReferenceDNA-proteincross-linksSyrian golden hamsterhepatocytes0.5–5 mM NegativeCasanova et al. (1997)Comet Assay V79 hamster cellstransfected with mouseGSTT12.5, 5, 10 mM A significant, dose-dependent increase in DNAdamage resulting from DNA-protein crosslinks inV79 cells transfected with mouse GSTT1 comparedto parental cellsHu et al. (2006)Hamster without GST activity from mouseForward mutation Chinese hamsterepithelial cells5,000, 10,000, 30,000,50,000 ppmNegativeJongen et al. (1981)UnscheduledDNA synthesisChinese hamsterepithelial cells5,000, 10,000, 30,000,50,000 ppmNegativeJongen et al. (1981)Sister chromatidexchangeChinese hamsterepithelial cells5,000, 10,000, 20,000,30,000, and 40,000 ppmWeak positive with or without rat-liver microsomalsystemJongen et al. (1981)ChromosomalaberrationsChinese hamster ovarycellsNot provided Positive, independent of rat liver S9Thilagar and Kumaroo(1983)Sister chromatidexchangeChinese hamster ovarycellsNot provided Negative with or without rat liver S9Thilagar and Kumaroo(1983)DNA and proteinsynthesisChinese hamster ovarycellsUp to 1,000μg/mL NegativeGarrett and Lewtas (1983)DNA SSBs byalkaline elutionHamster hepatocytes 0.4–90 mM NegativeGraves et al. (1995)CalfDNA Adducts Calf thymus DNA 50 mM Positive in thepresence of bacterial GST DM11 anddichloromethane dehalogenase. Adducts primarilyformed with the guanine residues.Kayser and Vuilleumier(2001)DNA Adducts Calf thymus DNA 0 -8.0 umol (0 – 60 mM)Positive in the presence of bacterial GST DM11, ratGST5-5, and human GSTT1-1. Adducts primarilyformed with the guanine residues.Marsch et al. (2004)HumanUnscheduledDNA synthesisHuman peripherallymphocytes250, 500, 1,000 ppm Negative with or without rat liver S9 Perocco and Prodi (1981)UnscheduledDNA synthesisPrimary humanfibroblast5,000, 10,000, 30,000,50,000 ppmNegativeJongen et al. (1981)157Table 4-30. Results from in vitro genotoxicity assays of dichloromethane with mammalian systems, by typeof testAssay Test system Concentrations Results ReferenceSister chromatidexchangeHuman peripherallymphocytesNot provided Weak positive Thilagar et al. (1984)ChromosomalaberrationsHuman peripherallymphocytesNot provided PositiveThilagar et al. (1984)DNA SSBs byalkaline elutionHuman hepatocytes Up to 120 mM Negative at concentrations between 5 and 120 mM Graves et al. (1995)Micronucleus test Human AHH-1,MCL-5, h2E1 cell linesUp to 10 mM Positive in all three cell lines Doherty et al. (1996)DNA-proteincross-linksMouse, rat, hamster,human hepatocytes0.5–5 mM NegativeCasanova et al. (1997)DNA damage bycomet assayPrimary human lungepithelial cells10, 100, 1,000μM Weak trend, independent of GST activity (GSTenzymatic activity not present in the cultured cells)Landi et al. (2003)aHPRT = hypoxanthine-guanine phosphoribosyl transferase5065158

4.5.1.2. In Vivo Genotoxicity Assays 5066 Genotoxicity findings in Drosophila melanogaster assays are mixed (Table 4-31). A 5067 study of gene mutation in D. melanogaster showed a marginal increase in sex-linked recessive 5068 deaths following oral exposure (Gocke et al., 1981) . An additional feeding study (Rodriguez5069 Arnaiz, 1998) reported a positive response in the s omatic w/w + assay. A third study of 5070 D. melanogaster (Kramers et al., 1991) found no evidence of increa sed sex-linked recessive 5071 deaths, somatic mutations, or recombinations follow ing exposure to airborne dichloromethane. 5072

5073 Table 4-31. Results from in vivo genotoxicity assa ys of dichloromethane in

insects

Assay Test system Doses Result ReferenceGene mutation (sexlinked recessive lethal) Drosophila 125, 620 mM Positive (feeding exposure) Gocke et al. (1981)

Gene mutation (sexlinked recessive lethal, somatic mutation and recombination) Drosophila 6 hours—1,850, 5,500 ppm 1 week—2,360, 4,660 ppm 2 weeks— 1,370, 2,360 ppm (all approximate) Negative (inhalation exposure) Kramers et al. (1991)

Somatic w/w+ assay Drosophila 50, 100, 250, 500 mM Positive (feeding exposure) Rodriguez-Arnaiz (1998)

5074 Some in vivo studies investigating certain genotoxi c endpoints in mice exposed to 5075 dichloromethane produced negative results (Table 4- 32). Unscheduled DNA synthesis was not 5076 induced in hepatocytes from mice (and rats) after 2 - or 6-hour inhalation exposure to 5077 concentrations that were carcinogenic in the NTP (1 986) mouse bioassay (Trueman and Ashby, 5078 1987) or other exposure routes (Lefevre and Ashby, 1989). Although positive results were not 5079 observed in the unscheduled DNA synthesis studies, it is generally recognized that this assay is 5080 not sensitive for detecting genotoxic chemicals (Ea stmond et al., 2009; Madle et al., 1994). 5081 Distinct, unequivocal cytogenetic effects (e.g., in duction of micronuclei, sister chromatid 5082 exchanges, or chromosome aberrations) were not cons istently found in bone marrow or 5083 erythrocytes in several studies of mice after acute oral exposures (Sheldon et al., 1987) or 5084 parenteral exposures (Westbrook-Collins et al., 199 0; Gocke et al., 1981). However, 5085 tumorigenic effects in mice are generally localized to the liver and lung (due to high GST 5086 activity) and therefore it is not surprising that g enotoxic effects were, for the most part, not 5087 observed in the bone marrow or erythrocytes (cell t ypes with minimal GST activity). Crebelli et 5088 al. (1999) stated that genotoxic effects induced by halogenated hydrocarbons (such as 5089 dichloromethane) are not very effective in inducing micronucleus formation in mouse bone 5090 marrow and a negative bone marrow micronucleus assa y should not offset the consistently 5091 positive in vitro results (Dearfield and Moore, 200 5) 5092 When genotoxic endpoints were examined in the cance r target tissues (liver and lung) in 5093 mice exposed to dichloromethane, positive results w ere consistently reported (Table 4-32). 5094159

These findings provide supporting evidence that GST -pathway metabolites may be key actors in 5095 the genotoxic effects and carcinogenic mode of acti on for dichloromethane. Increased sister 5096 chromatid exchanges were found in lung cells and pe ripheral lymphocytes from mice exposed by 5097 inhalation for 2 weeks to 8,000 ppm or for 12 weeks to 2,000 ppm (Allen et al., 1990). Under 5098 the same exposure conditions, increased chromosomal aberrations in lung and bone cells and 5099 micronuclei in peripheral red blood cells also were found (Allen et al., 1990). DNA-protein 5100 cross-links were detected in mouse hepatocytes, but not in lung cells, after a 3-day inhalation 5101 exposure to 4000 ppm (Casanova et al., 1992) and be tween 500 and 4000 ppm (Casanova et 5102 al.,1996). DNA damage, detected as increased DNA S SBs, was observed in liver and lung tissue 5103

of B6C3F 1 mice immediately following 3-hour exposures (Grave s et al., 1995). The DNA 5104 damage was not detectable 2 hours after in vivo exp osure, indicating that DNA repair occurs 5105 rapidly. Pretreatment of mice with buthionine sulp hoximine, a GSH depletor, caused a decrease, 5106 to levels seen in controls, in the amount of DNA da mage detected immediately after in vivo 5107 exposure in liver and lung tissue, indicating GSH i nvolvement in the genotoxic process. DNA 5108 damage (detected by the comet assay) was also repor ted in liver and lung tissues from male CD5109

1 mice sacrificed 24 hours after administration of a single oral dose of 1,720 mg/kg of 5110 dichloromethane (Sasaki et al., 1998). In this stu dy, DNA damage in lung and liver was not 5111 detected 3 hours after dose administration, and no DNA damage occurred at either time point in 5112 several other tissues in which a carcinogenic respo nse was not seen in chronic animal cancer 5113 bioassays (e.g., stomach, kidney, bone marrow). 5114

Formation of DNA adducts was evaluated in male and female B6C3F 1 mice as well as in 5115 male F344 rats (Watanabe et al., 2007). Animals we re administered 5 mg/kg, i.p., of 5116

radiolabeled dichloromethane and sacrificed at 1 or 8 hours after administration. The kidneys 5117

and livers were removed and the DNA was isolated fr om these tissues to evaluate formation of 5118 DNA adducts. At the administered dose, DNA adducts were not detected. 5119

Other studies in mice have looked for mutations in specific oncogenes (K- ras or H- ras ) 5120 (Devereux et al., 1993) or in a tumor suppressor ge ne ( p53 ) (Hegi et al., 1993) in liver or lung 5121 tumors from dichloromethane-exposed mice. These st udies have not demonstrated exposure5122 related patterns of mutations in these genes, altho ugh it should be noted that the statistical power 5123 of this analysis for the lung tumors is limited (di scussed further in sections 4.5.2 and 4.5.3). 5124

5125160Table 4-32. Results from in vivo genotoxicity assays of dichloromethane in miceAssay Test system Route and dose Duration Results ReferenceMicronucleus test Mouse bone marrow425, 850, or 1,700 mg/kg 2 doses Negative at all doses Gocke et al.(1981)Micronucleus test Mouse bone marrowGavage,1,250, 2,500, and4,000 mg/kgsingle doseNegative at all doses Sheldon et al.(1987)Micronucleus test Mouse peripheral redblood cellsInhalation 6 hr/day, 5 d/wk,0, 4,000, 8,000 ppm2 wk Positive at 4,000 and8,000 ppmAllen et al.(1990)Micronucleus test Mouse peripheral redblood cellsInhalation, 6 hr/day, 5 d/wk,0, 2,000 ppm12 weeks Positive at 2,000 ppm Allen et al.(1990)DNA synthesis Mouse liverGavage, 1,000 mg/kg;inhalation, 4,000 ppmsingle dose;2 hoursNegative in both oral andinhalation studiesLefevre andAshby (1989)Unscheduled DNA synthesis Mouse hepatocytesInhalation, 2,000 and4,000 ppm.2 or 6 hours Negative Trueman andAshby (1987)Sister chromatid exchange Mouse bone marrowIntraperitoneal, 100, 1,000,1,500, 2,000 mg/kgsingle doseNegative WestbrookCollins et al.(1990)Sister chromatid exchange Mouse bone marrow Subcutaneous, 0, 2,500,5,000 mg/kgsingle dose Negative at all doses Allen et al.(1990)Sister chromatid exchange Mouse lung cells andperipheral lymphocytesInhalation 6 hr/day, 5 d/wk,0, 4,000, 8,000 ppm2 weeks Positive at 8,000 ppm Allen et al.(1990)Sister chromatid exchange Mouse lung cellsInhalation 6 hr/day, 5 d/wk,0, 2,000 ppm12 weeks Positive at 2,000 ppm Allen et al.(1990)Chromosome aberrations Mouse bone marrowIntraperitoneal, 100, 1,000,1,500, 2,000 mg/kgsingle dose Negative WestbrookCollins et al.(1990)Chromosome aberrations Mouse bone marrow Subcutaneous, 0, 2,500,5,000 mg/kgsingle dose Negative Allen et al.(1990)161Table 4-32. Results from in vivo genotoxicity assays of dichloromethane in miceAssay Test system Route and dose Duration Results ReferenceChromosome aberrations Mouse lung and bonemarrow cellsInhalation, 6 hr/day, 5 d/wk,0, 4,000, 8,000 ppm2 weeks Positive at 8,000 ppm Allen et al.(1990)DNA-protein cross-links Mouse liver and lungcellsInhalation, 6 hr/day, 3 days,4,000 ppm3 days Positive in mouse liver cellsat 4,000 ppm; negative inmouse lung cellsCasanova et al.(1992)DNA-protein cross-links Mouse liver and lungcellsInhalation, 6 hr/day, 150,500, 1,500, 3,000,4,000 ppm3 days Positive in mouse liver cellsat 500–4,000 ppm; negativein mouse lung cellsCasanova et al.(1996)DNA single strand breaks byalkaline elutionMouse hepatocytesInhalation, 2,000 and4,000 ppm3 or 6 hours Positive at 4,000 ppm at3 and 6 hoursGraves et al.(1994b)DNA single strand breaks byalkaline elutionMouse liver and lunghomogenateLiver: inhalation, 2,000,4,000, 6,000, 8,000 ppmLung: inhalation, 1,000,2,000, 4,000, 6,000 ppm3 hours3 hoursLiver: positive at 4,000–8,000 ppmLung: positive at 2,000–4,000 ppmGraves et al.(1995)DNA damage by comet assay Mouse liver and lungcellsGavage, 1,720 mg/kg;organs harvested at 0(control), 3, and 24 hourssingle dose Positive only at 24 hoursafter dosingSasaki et al.(1998)DNA damage by comet assay Mouse stomach, urinarybladder, kidney, brain,bone marrowGavage, 1,720 mg/kg;organs harvested at 0(control), 3, and 24 hourssingle dose Negative 3 or 24 hr afterdosingSasaki et al.(1998)DNA adducts Mouse liver and kidneycellsIntraperitoneal, 5 mg/kg Single dose Negative Watanabe et al.(2007)Kras and Hras oncogenes Mouse liver and lungtumors0, 2000 ppm Up to 104 weeks No difference in mutationprofile between control anddichloromethane-inducedliver tumors; number ofspontaneous lung tumors(n=4) limits comparison atthis siteDevereux et al.,1993p53 tumor suppressor gene Mouse liver and lungtumors0, 2000 ppm Up to 104 weeks Loss of heterozygocityinfrequently seenHegi et al., 1993162

Results from in vivo studies in other mammals (i.e ., rats and hamsters) of hepatocyte 5126 sensitivity to dichloromethane induction of DNA SSB s (Table 4-33) are consistent with 5127 interspecies differences in the induction of liver tumors in the inhalation cancer bioassays. A 5128 gavage study in rats reported the presence of DNA S SBs with a dose of 1,275 mg/kg (Kitchin 5129 and Brown, 1989). The other available studies, how ever, did not find any genotoxicity following 5130 dichloromethane exposure. No increase in unschedul ed DNA synthesis in rat hepatocytes was 5131 seen following inhalation dichloromethane exposure of 2 to 6 hours at 2000 or 4000 ppm 5132 (Trueman and Ashby, 1987), exposure by gavage up to 1000 mg/kg (Trueman and Ashby, 1987), 5133 or intraperitoneal exposure of 400 mg/kg (Mirsalis et al., 1989). DNA adducts were not detected 5134 in the livers and kidneys of male F344 rats dosed w ith 5 mg/kg dichloromethane, i.p. (Watanabe 5135 et al., 2007). DNA SSBs were significantly increas ed in hepatocytes isolated from B6C3F 1 mice 5136 exposed to 4,831 ppm (4,000 ppm nominal) for 6 hour s but were not increased in hepatocytes 5137 from Sprague-Dawley rats exposed to 4,527 ppm (4,00 0 ppm nominal) for 6 hours (Graves et al., 5138 1994b). Results from in vivo interspecies comparis ons of dichloromethane induction of DNA5139 protein cross-links in hepatocytes (expected produc ts of the GSH pathway) are also consistent 5140 with the hypothesis that the mouse is more sensitiv e than other mammalian species due to greater 5141 activity of the GST pathway. DNA-protein cross-lin ks were formed in the liver of mice, but not 5142 hamsters, following in vivo exposure to air concent rations ranging from 500 to 4,000 ppm, 5143 6 hours/day for 3 days (Casanova et al., 1996). Th e absence of a genotoxic response in the rat 5144 and hamster is consistent with considerably lower G ST activity and therefore, these mammalian 5145 systems would be expected to be less sensitive at d etecting genotoxic effects than the studies 5146

conducted in mice. 5147 Table 4-34 compares results from studies of mice an d rats in which comparable tissue5148 specific endpoints were examined in in vivo genotox icity assays. Several of the endpoints that 5149 were positive in mice (e.g., sister chromatid excha nge, DNA-protein cross-links, comet assay) 5150

have not been examined in the rat. Unscheduled DNA synthesis has been demonstrated in 5151 mouse, but not in rat hepatocytes. In contrast to the positive results seen in mouse inhalation 5152 exposure studies, DNA SSB induction was not seen in rat inhalation studies but was seen in an 5153 oral gavage study. 5154 In summary, the available data provide evidence for mutagenicity of dichloromethane. 5155 Most of the in vitro bacterial assays with GST acti vity showed positive results when there was 5156 GST activity. Non-positive results were reported o nly in bacterial assays with low GST activity; 5157 in experiments where GST was added, positive result s were then observed. Evaluation of the in 5158

vitro mammalian studies also demonstrates consisten cy of the requirement for GST for 5159 observation of genotoxic effects. In rat and hamst er cell lines where GST activity is significantly 5160 less than mouse, primarily negative results were re ported following dichloromethane exposure. 5161 However, when mouse liver cytosol or transfected mo use GST were included in these same cell 5162 lines, mutagenic effects were reported after dichlo romethane exposure. In mouse cell lines, 5163163

positive results were obtained in Clara cells but n o effects were observed in a mouse lymphoma 5164 cell line, which is consistent with the absence of tumors in this site for mice. The results of in 5165 vivo mutagenicity in mice also provide support for the site-specificity of the observed tumors. 5166 Assays using mouse bone marrow were all negative. However, micronuclei and sister chromatid 5167 exchange tests in peripheral blood produced a posit ive response at high doses. With the 5168 exception of one study of unscheduled DNA synthesis in hepatocytes, numerous site-specific 5169 studies in either the liver or lung were also posit ive at various doses. These liver and lung 5170 studies included chromosomal aberrations, SSBs and sister chromatid exchanges, and DNA5171 protein cross-links and correspond to genotoxic and mutagenic effects associated with 5172 metabolites from the GST pathway. 5173

51741645175Table 4-33. Results from in vivo genotoxicity assays of dichloromethane in rats and hamstersAssay Test system Route and dose Duration Results ReferenceUnscheduled DNA synthesis Rat hepatocytesGavage, 100, 500,1,000 mg/kgLiver harvested 4 and12 hours after dosingNegative 4 or 12 hours afterdosingTrueman andAshby (1987)Unscheduled DNA synthesis Rat hepatocytesInhalation, 2 or 6 hours,2,000 and 4,000 ppm2 or 6 hours Negative at bothconcentrations and exposuredurationsTrueman andAshby (1987)Unscheduled DNA synthesis Rat hepatocytesIntraperitoneal, single dose,400 mg/kgSingle dose Negative 48 hours afterdosingMirsalis et al.(1989)DNA SSBs by alkaline elution Rat hepatocytesInhalation, 3 or 6 hours,2,000 and 4,000 ppm3 or 6 hours Negative at allconcentrations and timepointsGraves et al.(1994b)DNA SSBs by alkaline elution Rat liver homogenateGavage, 2 doses, 425 mg/kgand 1,275 mg/kg,administered 4 and 21 hoursbefore liver harvesting4 or 21 hours (timebetween dosing andliver harvesting)Positive at 1,275 mg/kg Kitchin andBrown (1989)DNA SSBs by alkaline elution Rat liver and lunghomogenateLiver: inhalation, 4,000,5,000 ppmLung: inhalation, 4,000 ppm3 hours3 hoursNegative for both liver andlung at all concentrationsGraves et al.(1995)DNA-protein cross-links Hamster liver and lungcellsInhalation, 6 hr/day, 500,1,500, 4,000 ppm3 days Negative at allconcentrationsCasanova et al.(1996)DNA adducts Rat liver and kidneycellsIntraperitoneal, 5 mg/kg Single dose Negative Watanabe et al.(2007)165Table 4-34. Comparison of in vivo dichloromethanegenotoxicity assays targeted to lung or liver cells, by speciesStudies in miceStudies in ratsAssay Test systemRoute, dose (duration) Results ReferenceTest systemRoute, dose (duration)Results ReferenceDNAsynthesisLiverGavage, 1,000 mg/kg;inhalation, 4,000 ppm(2 hours)Negative in oral andinhalation studiesLefevre andAshby(1989)No studiesUnscheduledDNAsynthesisHepatocytesInhalation, 2,000 and4,000 ppm.(2 or 6 hours)Negative Truemanand Ashby(1987)HepatocytesInhalation, 2,000 and4,000 ppm (2 or6 hours)Negative Trueman andAshby (1987)UnscheduledDNAsynthesisHepatocytesIntraperitoneal,400 mg/kgNegative Mirsalis et al.(1989)SisterchromatidexchangeLung cells Inhalation 6 hr/day,5 days/wk, 0, 4,000, 8,000ppm (2 weeks)Inhalation 6 hr/day,5 days/wk, 0, 2,000 ppm(12 weeks)Positive at8,000 ppmPositive at2,000 ppmAllen et al.(1990)No studiesChromosomeaberrationsLung cells Inhalation, 6 hr/day,5 days/wk, 0, 4,000, 8,000ppm (2 weeks)Positive at8,000 ppmAllen et al.(1990)No studiesDNA-proteincross-linksLiver andlung cellsInhalation, 6 hr/day,3 days, 4,000 ppm (3 days)Inhalation, 6 hr/day, 150,500, 1,500, 3,000,4,000 ppm (3 days)Positive in liver4,000 ppmPositive in liver at500–4,000 ppm;both studies negativein lungCasanova etal. (1992)No studiesDNA SSBsby alkalineelutionHepatocytesInhalation, 2,000 and4,000 ppm (3 or 6 hours)Positive at4,000 ppmGraves etal. (1994b)HepatocytesInhalation, 3 or 6 hours,2,000 and 4,000 ppmNegative at allconcentrationsand time pointsGraves et al.(1994b)DNA SSBsby alkalineelutionLiver andlunghomogenateLiver: inhalation, 2,000,4,000, 6,000, 8,000 ppm (3hours)Lung: inhalation, 1,000,2,000, 4,000, 6,000 ppm (3hours)Liver: Positive at4,000–8,000 ppmLung: Positive at2,000–4,000 ppmGraves etal. (1995)Liver andlunghomogenateLiver: inhalation, 4,000,5,000 ppmLung: inhalation,4,000 ppmNegative inliver and lung atallconcentrationsand time pointsGraves et al.(1995)DNA SSBsby alkalineelutionLiverhomogenateGavage, 425 mg/kg and1,275 mg/kgPositive at1,275 mg/kgKitchin andBrown (1989)166Table 4-34. Comparison of in vivo dichloromethanegenotoxicity assays targeted to lung or liver cells, by speciesStudies in miceStudies in ratsAssay Test systemRoute, dose (duration) Results ReferenceTest systemRoute, dose (duration)Results ReferenceDNA damageby cometassayLiver andlung cellsGavage, 1,720 mg/kg;organs harvested at 0(control), 3, and 24 hoursPositive only at 24hours after dosingSasaki et al.(1998)No studiesDNA adductsLiver andkidney cellsIntraperitoneal, 5 mg/kg NegativeWatanabe etal. (2007)Liver andkidney cellsIntraperitoneal, 5 mg/kgNegative Watanabe etal. (2007)5176167

4.5.2. Mechanistic Studies of Liver Effects 5177 One of the major target organs from dichloromethan e exposure is the liver, and several 5178 studies have focused on examining the potential mec hanisms producing liver tumors. This 5179 section summarizes the primary mechanistic studies that were conducted in order to examine the 5180 hepatic tumors produced by dichloromethane in mice. A parallel set of studies, discussed in the 5181 next section, focus on potential mechanisms that pr oduce lung tumors. Briefly, 5182 dichloromethane-induced liver tumors first appeared in mice after 52 weeks of exposure 5183 (Maronpot et al., 1995; Kari et al., 1993), which w as when tumors began to appear in control 5184 mice, indicating a similar time course in tumor for mation between treated and untreated groups. 5185 Onset of liver tumor formation is not preceded by l iver cell proliferation (Casanova et al., 1996; 5186 Foley et al., 1993). Further mechanistic studies w ere conducted to assay the tumor for 5187 significant changes in proto-oncogene activation an d tumor suppressor gene inactivation 5188 (Maronpot et al., 1995; Devereux et al., 1993; Hegi et al., 1993). A second subset of mechanistic 5189

studies was conducted to elucidate the reason that mice are the most sensitive species to liver 5190 tumors and if other species exhibited changes in li ver function (Thier et al., 1998; Reitz et al., 5191 1989). It was found that mice have the highest lev el of GST-T1 catalytic activity but that 5192 humans, rats, and hamsters among other species also metabolize dichloromethane in the liver to a 5193 GST conjugate. In contrast, there has been little research focusing on the mechanisms through 5194 which nonneoplastic hepatic effects (seen most stro ngly in the rat) are produced, and the role of 5195 the parent material, metabolites of the CYP2E1 path way, metabolites of the GST pathway, or 5196 some combination of parent material and metabolites is not known. 5197

5198

Liver tumor characterization studies 5199

Several studies have examined the time course of a ppearance of liver tumors in B6C3F 1 5200 mice exposed to 2,000 or 4,000 ppm and possible lin ks between hepatic nonneoplastic 5201 cytotoxicity, enhanced hepatic cell proliferation, and the development of liver tumors (Casanova 5202 et al., 1996; Maronpot et al., 1995; Foley et al., 1993; Kari et al., 1993). The studies provide no 5203 clear evidence for a sustained liver cell prolifera tion response to dichloromethane that can be 5204 linked to the development of dichloromethane-induce d liver tumors. Additionally, a few studies 5205

have examined if dichloromethane-induced liver tumo rs are the result of proto-oncogene 5206 activation and tumor suppressor gene inactivation ( Maronpot et al., 1995; Devereux et al., 1993; 5207 Hegi et al., 1993). 5208 Kari et al. (1993) (also summarized by Maronpot et al. [1995]) reported data from 5209 6 groups of 68 female B6C3F 1 mice exposed to six ”stop-exposure“ protocols of d iffering 5210 durations and sequences, with each exposure concent ration standardized at 2,000 ppm for 5211 6 hours per day, 5 days per week. The six stop-exp osure protocols were 26 weeks of exposure 5212 followed by 78 weeks without exposure, 78 weeks wit hout exposure followed by 26 weeks of 5213 exposure, 52 weeks without exposure followed by 52 weeks with exposure, 52 weeks exposed 5214168

followed by 52 weeks without exposure, 78 weeks exp osed followed by 26 weeks without 5215 exposure, and 26 weeks without exposure followed by 78 weeks with exposure. A control group 5216 (no exposure, 104 weeks duration) and a maximum exp osure (104 weeks duration) group were 5217 also included. Exposure for 26 weeks did not resul t in an increased incidence of liver tumors 5218 (adenomas or carcinomas). Respective percentages o f animals with liver tumors were 27% 5219 (18/67), 40% (27/67), and 34% (23/67) for the contr ols, early 26-week exposure and late 26- 5220 week exposure groups, respectively. Exposure to 2, 000 ppm for 52 weeks (followed by no 5221 exposure until 104 weeks), 78 weeks (either early o r late exposure periods), or 104 weeks 5222 produced increased incidence of mice with liver tum ors ( p < 0.05), but this increase was not seen 5223 in the 52-week late exposure group. Respective per centages of animals with liver tumors 5224 (adenomas or carcinomas combined) were 44% (28/64), 31% (21/67), 62% (42/68), 48% (32/67) 5225 and 69% (47/68) for the 52 (early exposure), 52 (la te exposure), 78 (early exposure), 78 (late 5226 exposure), and 104 week exposure periods, respectiv ely. With the 78 week exposures, the 5227

difference in the liver tumor incidence between the early and late exposure periods was 5228 statistically significant ( p < 0.01). A greater increase in multiplicity of li ver tumors was also 5229 seen with the early 78-week exposure period. These data suggest that 52 weeks of exposure was 5230 required to increase the incidence of liver tumors in mice, that early exposure was more effective 5231 than late exposure, and that the increased risk con tinued after cessation of exposure. 5232 Histopathologic examination of liver tissue at int erim killings at eight time periods (13, 5233 26, 52, 68, 75, 78, 83, or 91 weeks) of exposure to 2,000 ppm (n = 20 mice per killing) found no 5234 evidence of nonneoplastic cytotoxicity that precede d the appearance of proliferative neoplastic 5235

liver lesions. Incidences of mice with liver adeno mas or carcinomas were elevated (between 40– 5236 60%) at five of the six interim killings after 52 w eeks. The incidence rates at each time period 5237 were 0/20 (0%) at 13 weeks, 1/20 (5%) at 26 weeks, 8/20 (40%) at 52 weeks, 4/26 (15%) at 5238 68 weeks, 13/20 (65%) at 75 weeks, 12/19 (63%) at 7 8 weeks, 8/20 (40%) at 83 weeks, and 5239 20/30 (66%) at 91 weeks. The collected liver lesio n data identify no exposure-related increased 5240 incidence of nonneoplastic liver lesions that could be temporally linked to liver tumor 5241 development. Liver tumors first appeared at about the same time in control and exposed animals 5242 (52 weeks). 5243 Foley et al. (1993) examined indices of cell proli feration in livers of female B6C3F 1 mice 5244 exposed to 1,000, 2,000, 4,000, or 8,000 ppm dichlo romethane (6 hours/day, 5 days/week) for 1, 5245 2, 3, or 4 weeks or to 2,000 ppm for 13, 26, 52, or 78 weeks but found no evidence for sustained 5246 cell proliferation with prolonged exposure to dichl oromethane. To label liver cells in S phase, 5247 tritiated thymidine (1- to 4-week exposure protocol s) or bromodeoxyuridine (13- to 78-week 5248 protocols) was administered subcutaneously via an o smotic mini-pump for 6 days prior to 5249 killing. Labeled hepatocytes in liver sections (fr om 10 mice in each exposure/duration group) 5250 were counted to assess the number of cells in S-pha se per 1,000 cells. S-phase labeling indices 5251 in livers of exposed mice at most killings were equ ivalent to or less than those in control mice. 5252169

A transient increase in S-phase labeling index of a bout two- to fivefold over controls was 5253 observed at the 2-week killing of mice exposed to 1 ,000, 4,000, or 8,000 ppm. Because of the 5254 transient nature and small magnitude of the respons e, it is not expected to be of significance to 5255 the promotion of liver tumors in chronically expose d mice. Foley et al. (1993) also compared 5256 cell proliferation labeling indices in foci of cell ular alteration and non-affected liver regions in 5257 control and exposed mice but found no significant d ifference between control and exposed mice. 5258 S-phase labeling was accomplished by immunohistolog ic staining for proliferating cell nuclear 5259 antigen in liver sections from 24 control mice and 15 exposed mice, with livers showing foci of 5260 cellular alteration. In both control and exposed l ivers, the labeling index was about four- to 5261 fivefold higher in foci of cellular alteration than in surrounding unaffected liver tissue. 5262 In mice exposed to 2,000 ppm for 13–78 weeks, rela tive liver weights were statistically 5263 significantly elevated compared with controls; abou t 10% increased at 13 and 26 weeks and 5264

about 30–40% increased at 52 and 78 weeks. Histolo gic changes in liver sections of 2,000 ppm 5265 mice exposed for 13–78 weeks were restricted to hep atocellular hypertrophy (observed at all 5266 killing intervals) and preneoplastic (foci of cellu lar alteration) and neoplastic (adenoma and 5267 carcinoma) lesions. No signs of liver tissue degen eration were found. Adenoma and focus of 5268 alteration were first detected at 26 weeks (2/10 ve rsus 0/10 in controls). At 52 weeks, 4/10 5269 exposed mice had proliferative lesions (1 focus, 1 adenoma, and 2 carcinomas), compared with 5270 1/10 in controls (1 adenoma). At 78 weeks, 7/10 ex posed mice had proliferative lesions (2 foci, 5271 3 adenomas, 6 carcinomas) compared with 1/10 in con trols (1 adenoma). In summary, the 5272 results indicate that inhalation exposure to 2,000 ppm dichloromethane produced an increase 5273

incidence of liver tumors in female B6C3F 1 mice. No evidence was found for sustained cell 5274 proliferation or liver tissue degeneration in respo nse to dichloromethane exposure, but exposure 5275 was associated with relative liver weight increases and hepatocellular hypertrophy. 5276 Casanova et al. (1996) found no clear evidence of increased cell proliferation in the livers 5277

of male B6C3F 1 mice exposed to dichloromethane concentrations >1, 500 ppm 6 hours/day for 5278 3 days. Three or four groups of three mice were ex posed to 146, 498, 1,553, or 3,923 ppm 5279 unlabeled dichloromethane for 2 days and then expos ed to [ 14 C]-labeled dichloromethane for 5280 6 hours on the third day. Radiolabel incorporated into liver DNA deoxyribonucleosides was 5281 measured as an index of cell proliferation. Radiol abel incorporated into liver DNA 5282 deoxyribonucleosides increased approximately fivefo ld from 146 to 1,553 ppm, but further 5283 increases were not apparent at 3,923 ppm. (In cont rast, as described in section 4.5.3, radiolabel 5284 incorporation into lung DNA deoxyribonucleosides di splayed a 27-fold increase over this 5285 concentration range.) The small magnitude of the i ncrease in radiolabel incorporation into liver 5286 DNA deoxyribonucleosides with increasing exposure c oncentration suggests that little, if any, 5287 enhanced cell proliferation occurred in the liver i n response to dichloromethane exposure. 5288 Devereux et al. (1993) (also reported in Maronpot et al. [1995]) analyzed liver tumors in 5289

female B6C3F 1 mice for the presence of activated H- ras oncogenes. Fifty dichloromethane- 5290170

induced and 49 spontaneous liver tumors were screen ed for H- ras mutations. There was a 5291 relatively high frequency of activated H- ras among the nonexposed B6C3F 1 mice: 67% of the 5292 spontaneous tumors and 76% of the dichloromethane-i nduced tumors contained mutations in the 5293

H- ras gene. Overall, the mutation profile of the dichlo romethane-induced tumors did not 5294 significantly differ from the spontaneous tumors. 5295 Similarly, Hegi et al. (1993) analyzed the liver t umors from female B6C3F 1 mice for 5296 inactivation of the tumor suppressor genes, p53 and Rb-1 . Half of the liver tumors used for this 5297 study had an activated H- ras oncogene. Twenty liver tumors (15 carcinomas and 5 adenomas) 5298 were screened for loss of heterozygosity (LOH) on c hromosome 11 and 14, which is associated 5299 with malignant conversion of the p53 gene (chromosome 11) and the Rb-1 gene (chromosome 5300 14). Only one tumor out of 20 contained a LOH at c hromosome 14 and no dichloromethane5301

induced liver tumors contained a LOH at chromosome 11. 5302

5303

Liver metabolic studies 5304 As described in detail in section 3.3, GST-T1 enzym atic activity and distribution is 5305 variable among species, and there is also considera ble intraspecies variability among humans. In 5306 summary, Reitz et al. (1989) demonstrated a greater metabolic activity with respect to 5307

dichloromethane in livers of B6C3F 1 mice compare with F344 rats, Syrian golden hamster s, and 5308 humans. The rates of in vitro metabolism by the GS T pathway were about 4-, 12-, and 20-fold 5309 greater in B6C3F 1 mouse liver samples than in F344 rat, human, and S yrian golden hamster 5310 samples, respectively (Reitz et al., 1989). A more recent study characterized the 5311 dichloromethane metabolic capacity specifically of hepatic GST-T1 (Thier et al., 1998). 5312 Enzymatic activities of GST-T1 in liver from F344 r ats, B6C3F 1 mice, Syrian golden hamsters, 5313 and humans with three different GST-T1 phenotypes ( nonconjugators, low conjugators, high 5314 conjugators) showed the following order with dichlo romethane as a substrate: mouse >> rat > 5315 human high conjugators>human low conjugators > hams ter > human nonconjugators. 5316

5317 4.5.3. Mechanistic Studies of Lung Effects 5318 The finding of increased lung tumors in B6C3F 1 mice exposed to dichloromethane 5319 (Mennear et al., 1988; NTP, 1986) has stimulated a number of studies designed to examine the 5320 mechanism for dichloromethane-induced lung tumors i n this animal. The lung tumor mechanism 5321

studies were conducted with B6C3F 1 mice, and the frequency of lung tumors in control animals 5322 was very low. Time-course studies for lung tumor d evelopment were conducted, and it was 5323 found that the onset of lung tumor development was much shorter than liver tumors (Kari et al., 5324 1993) (reported in Maronpot et al. [1995]). As a r esult, it is hypothesized that a potential 5325 common mechanism independent of liver metabolism is producing tumors in the lung. As with 5326 the liver tumors, there were no significant increas es in mutations for the K- ras oncogene 5327 (Devereux et al., 1993) or the p53 and Rb-1 tumor suppressor genes (Hegi et al., 1993). 5328171

Additionally, the Clara cells, which are non-ciliar y secretory cells found in the primary 5329 bronchioles of the lung, are selectively targeted a fter dichloromethane exposure. Acute 5330 dichloromethane exposure produces Clara cell vacuol ization, which is not sustained with long5331 term exposure (Foster et al., 1992). There is a co rrelation between the acute effects on the Clara 5332 cell and the lung tumors from chronic exposure to d ichloromethane (Kari et al., 1993). 5333 However, the exact mechanism for producing these lu ng effects is not completely understood at 5334 this point. Provided below is a summary of the stu dies examining the potential mechanisms for 5335 producing lung tumors resulting from dichloromethan e exposure. 5336

5337 Lung tumor characterization studies 5338 Kari et al. (1993) (also summarized in Maronpot et al. [1995]) demonstrated that only 26 5339 weeks of exposure to 2,000 ppm was necessary to pro duce significantly increased incidence of 5340

female B6C3F 1 mice with lung tumors. In the six “stop-exposure” protocol experiment 5341 (26 weeks exposure followed by 78 weeks without exp osure, 78 weeks without exposure 5342 followed by 26 weeks exposure, 52 weeks without exp osure followed by 52 weeks with 5343 exposure, 52 weeks exposed followed by 52 weeks wit hout exposure, 78 weeks exposed 5344 followed by 26 weeks without exposure, and 26 weeks without exposure followed by 78 weeks 5345 with exposure), early but not late exposure for 26 or 52 weeks resulted in an increased incidence 5346 of animals with lung tumors (adenoma or carcinomas) . Respective percentages of animals with 5347 lung tumors were 7.5% (5/67), 31% (21/68), 4% (3/67 ), 63% (40/63), and 15% (10/67) for the 5348 controls, early 26-, late 26-, early 52-, and late 52-week exposure groups, respectively. With the 5349 78-week exposures, both the early and late exposure regimens produced an increased incidence 5350 of lung tumors compared with controls (56% [38/68] and 19% [13/68], respectively), compared 5351 with the incidence of 63% (42/67) seen in the group exposed for the full 104 weeks. Thus a 5352 plateauing of risk was seen, with similar incidence rates seen with the early 52-week, early 5353 78-week, and 104-week exposure periods. The differ ence in the lung tumor incidence between 5354 the early and late exposure periods of similar dura tion was statistically significant ( p < 0.01) for 5355 the 26-, 52-, and 78-week duration protocols. A gr eater increase in multiplicity of lung tumors 5356 was also seen with the early 78-week exposure perio d. As with the liver tumor data from the 5357 same series of experiments, these data suggest that early exposure was more effective than late 5358 exposure and that the increased risk continued afte r cessation of exposure. 5359 Histopathologic examination of lung tissue from mic e killed at 13, 26, 52, 68, 75, 78, 83, 5360 or 91 weeks of exposure to 2,000 ppm (n = 20 mice p er killing) found no evidence of 5361 nonneoplastic cytotoxicity that preceded the appear ance of proliferative neoplastic lung lesions. 5362 In contrast, incidences of mice with lung adenomas or carcinomas (combined) were elevated at 5363 interim killings ≥ 52 weeks; incidences for the interim killings of mi ce exposed to 2,000 ppm 5364 (6 hours per day, 5 days per week) between 13 and 9 1 weeks were 0/20 (0%) at 13 weeks, 5365 0/20 (9%) at 26 weeks, 6/20 (30%) at 52 weeks, 6/26 (23%) at 68 weeks, 8/20 (40%) at 5366172

75 weeks, 9/19 (47%) at 78 weeks, 10/20 (50%) at 83 weeks, and 14/30 (47%) at 91 weeks, 5367 respectively. Lung hyperplasia was found at an inc reased incidence only at 91 weeks, well after 5368 the 26- and 52-week periods that induced increased incidences of mice with lung tumors. 5369 Kanno et al. (1993) found no evidence for histolog ic changes or increased cell 5370 proliferation in lung tissue of female B6C3F 1 mice exposed to 2,000 or 8,000 ppm 5371 dichloromethane for 1, 2, 3, or 4 weeks, compared w ith control mice, or in mice exposed to 5372 2,000 ppm for 13 or 26 weeks. Osmotic mini-pumps w ere used to deliver tritiated thymidine and 5373 label cells undergoing replicative DNA synthesis ov er 6-day periods before killing. Labeled 5374 cells undergoing rapid DNA synthesis and cell proli feration were assessed in sections of 5375 proximal and terminal bronchioles and alveoli of lu ngs from groups of 5 mice exposed for 1– 5376 4 weeks or 10 mice exposed for 13 or 26 weeks. The re were no exposure-related histopathologic 5377 or labeling index changes in the alveoli, but lower labeling indices were found in the bronchiolar 5378 epithelium of exposed mice compared with controls. 5379 The combined results from the Kari et al. (1993) a nd Kanno et al. (1993) studies are 5380 consistent with the hypothesis that dichloromethane -induced lung tumors in B6C3F 1 mice are not 5381 preceded by overt cytotoxicity, enhanced and sustai ned cell proliferation, or hyperplasia in the 5382 lung. Two other studies (Casanova et al., 1996; Fo ster et al., 1992), however, have reported 5383 evidence for enhanced cell proliferation in lungs o f B6C3F 1 mice exposed for acute durations to 5384 airborne dichloromethane. Only one of these studie s (Foster et al., 1992), however, looked for 5385 sustained cell proliferation in the lung with prolo nged exposure. In agreement with the results 5386 from Kanno et al. (1993), no evidence was found for sustained cell proliferation in lungs with 5387 prolonged exposure to dichloromethane at concentrat ions demonstrated to induce lung tumors in 5388

mice. 5389 Casanova et al. (1996) detected evidence of increa sed cell proliferation in the lungs of 5390

male B6C3F 1 mice exposed to dichloromethane concentrations >1, 500 ppm 6 hours/day for 5391 3 days. Three or four groups of three mice were ex posed to 146, 498, 1,553, or 3,923 ppm 5392 unlabeled dichloromethane for 2 days and then expos ed to [ 14 C]-labeled dichloromethane for 5393 6 hours on the third day. Radiolabel incorporated into lung DNA deoxyribonucleosides (after 5394 removal of DNA-protein cross links containing radio labeled formaldehyde) was measured as an 5395 index of cell proliferation. Radiolabel incorporat ion into lung DNA deoxyribonucleosides 5396 increased with increasing exposure concentration, w ith the amount increasing by about 27-fold 5397 between 146 and 3,923 ppm. In hamsters that did no t develop tumors in response to chronic 5398 inhalation exposure to 3,500 ppm dichloromethane (B urek et al., 1984), no evidence for 5399 enhanced radiolabel incorporation into lung DNA deo xyribonucleosides was found following 5400 acute exposure (Casanova et al., 1996). 5401 Devereux et al. (1993) (also summarized in Maronpot et al. [1995]) analyzed lung tumors 5402

in female B6C3F 1 mice for the presence of activated K- ras oncogenes. Fifty-four 5403 dichloromethane-induced and 17 spontaneous lung tum ors (7 from the NTP [1986] study and 10 5404173

from a study in C57BL/6 × C34F1 mice reported by Ca ndrian et al. [1991]) were screened for K5405

ras mutations. Twenty percent of the dichloromethane-i nduced tumors and 24% of the 5406 spontaneous tumors contained mutations in the K- ras gene. Devereux et al. (1993) stated that 5407 there may be a significant difference in the incide nce of K- ras activation between spontaneous 5408 and dichloromethane-induced tumors. However, the s mall number of the spontaneous tumors 5409 that were available for the study limits the conclu sions that can be made from the results. 5410 Hegi et al. (1993) analyzed the lung tumors from f emale B6C3F 1 mice for inactivation of 5411 the tumor suppressor genes, p53 and Rb-1 . Forty-nine dichloromethane-induced lung 5412 carcinomas, five lung adenomas, and seven spontaneo us lung carcinomas were screened for 5413 LOH on mouse chromosome 11 and 14, which is associa ted with malignant conversion of the 5414 p53 gene (chromosome 11) and the Rb-1 gene (chromosome 14). Fourteen percent (n = 7) of the 5415 dichloromethane-induced lung carcinomas exhibited L OH at chromosome 11. No p53 mutations 5416 were detected in the seven spontaneous lung tumors or the five dichloromethane-induced lung 5417 adenomas. Only three dichloromethane-induced tumor s exhibited LOH at chromosome 14. The 5418

authors noted that inactivation of the p53 and Rb-1 tumor suppressor genes infrequently occur in 5419 lung and liver tumors. 5420

5421

Clara cell studies 5422 Foster et al. (1992) found enhanced cell proliferat ion in bronchiolar cells and, to a lesser 5423 degree, alveolar cells in the lungs of male B6C3F 1 mice exposed for acute durations (2, 5, 8, or 5424 9 days) to 4,000 ppm dichloromethane (6 hours/day, 5 days/week), but the response was less 5425 distinct after subchronic durations of exposure (89 , 92, or 93 days). To measure cell 5426 proliferation, mice (n = 5 per exposure-duration gr oup) were given subcutaneous doses of 5427 tritiated thymidine for five consecutive days prior to killing. Labeled cells in bronchiolar or 5428 alveolar epithelium in lung sections were counted t o assess the number of cells in S phase per 5429 1,000 cells. Counts of bronchiolar epithelium cell s in S phase in exposed mice sacrificed on 5430 days 2, 5, 8, and 9 were approximately 2-, 15-, 3-, and 5-fold higher, respectively, than those of 5431 unexposed mice at day 0 of the experiment. In expo sed mice sacrificed on days 89, 92, and 93, 5432 less than twofold increases in bronchiolar epitheli um cell labeling were observed. Increased cell 5433 labeling was found in alveolar epithelium only on d ay 8 (about seven- to eightfold increase) and 5434 day 9 (about fourfold increase). Vacuolation of th e Clara cells of the bronchiolar epithelium was 5435 observed on day 2 (scored as ++, majority of cells affected), day 9 (+++, virtually all the cells 5436 affected), and day 44 (+, moderate effect to cells) but was not apparent on days 5, 8, 40, 43, 89, 5437 92, or 93. No hyperplasia of the bronchiolar epith elium or changes to Type II alveolar epithelial 5438 cells were observed in the lungs of any of the expo sed mice at any time point. The appearance 5439 and disappearance of the Clara cell vacuolation wer e generally correlated with the appearance 5440 and disappearance of enhanced cell proliferation in the bronchiolar epithelium; enhanced cell 5441 proliferation was observed on days 2, 5, 8, and 9 ( along with appearance of Clara cell 5442174

vacuolation on days 2 and 9) but was not observed o n days 89, 92, and 93 when Clara cell 5443 lesions also were not observed. This suggests that cell proliferation was enhanced in response to 5444 Clara cell damage but was not sustained with repeat ed exposure to dichloromethane. 5445 Currently, a mechanistic connection has not been es tablished between the acute effects of 5446

dichloromethane on Clara cells in the bronchiolar e pithelium and the development of lung 5447

tumors in B6C3F 1 mice exposed by inhalation to concentrations ≥ 2,000 ppm dichloromethane 5448 for 2 years (NTP, 1986) or for 26 weeks followed by no exposure through 2 years (Maronpot et 5449 al., 1995; Kari et al., 1993). There appears to be a concordance between exposure concentrations 5450 inducing acute Clara cell vacuolation ( ≥ 2,000 ppm) and those inducing lung tumors 5451 ( ≥ 2,000 ppm). However, transient acute Clara cell va cuolation does not appear to progress to 5452 necrosis or lead to sustained cell proliferation (w hich could promote the growth of tumor5453 initiated cells) and appears to be dependent on CYP metabolism of dichloromethane (see the 5454 following paragraphs discussing pertinent findings reported by Foster et al. [1994, 1992]). In 5455 contrast, there is consistent and specific evidence for an association between the formation of 5456 DNA-reactive GST-pathway metabolites and the format ion of lung and liver tumors from 5457 inhalation exposure (see sections 4.5.2 and 4.7.3). 5458 Foster et al. (1992) noted that the appearance and disappearance of Clara cell vacuolation 5459 in mouse lungs showed concordance with temporal pat terns for immunologic staining for 5460 CYP2B1 and 2B2 levels in lung sections. A similar temporal pattern was reported for CYP2B1 5461 and 2B2 monooxygenase activities (ethoxycoumarin O- dealkylation or aldrin epoxidation) 5462 assayed in lung microsomes. When there was a marke d decrease in CYP2B1 and 2B2 staining 5463 (e.g., on day 5) or monooxygenase activities, the l esion was not present. Similarly, the 5464 appearance of the lesion was preceded (the day befo re) by the recovery of monooxygenase 5465 activities or immunologic staining close to control levels. These patterns suggested to Foster et 5466 al. (1992) that Clara cells may have developed tole rance to dichloromethane due to inactivation 5467 of a CYP isozyme. 5468 In subsequent studies, increased percentages of vac uolated bronchiolar epithelium cells 5469 were noted in mice exposed to 2,000 ppm (26.3 ± 6.7 %) or 4,000 ppm (64.8 ± 12.8%), but 5470

vacuolated cells were not observed in bronchiolar e pithelium of lung sections from control mice 5471 or mice exposed to 125, 250, 500, or 1,000 ppm (Fos ter et al., 1994). Pretreatment with the CYP 5472 inhibitor, piperonyl butoxide, counteracted the 2,0 00 ppm effect (2.4 ± 3.6% vacuolated cells), 5473 whereas GSH-depleted mice showed no statistically s ignificant change in percentage of 5474 vacuolated cells (32.7 ± 16.9%) compared with the m ean percentage in mice exposed to 5475 2,000 ppm without pretreatment. No consistent, sta tistically significant, exposure-related 5476 changes were found in cytosolic GST metabolic activ ities (with dichloromethane as substrate) or 5477 microsomal CYP monooxygenase activities (ethoxycoum arin O-dealkylation), but mean 5478 cytosolic levels of nonprotein sulfhydryl compounds were elevated in lungs of mice exposed to 5479 1,000 and 2,000 ppm (134.6 ± 17.1 and 146.4 ± 6.7 n mol/mg protein, respectively) compared 5480175

with control levels (109.5 ± 7.6 nmol/mg protein). Increased cell proliferation was found in 5481 cultured Clara cells isolated from 4,000 ppm expose d mice compared with nonexposed mice; 5482 respective values for percentage of cells in S phas e were 18.97 ± 1.18 and 2.02 ± 0.86% (Foster 5483 et al., 1994). 5484 Results from the studies by Foster et al. (1994, 1 992) indicate that 6-hour exposures of 5485

B6C3F 1 mice to dichloromethane concentrations ≥ 2,000 ppm caused transient Clara cell 5486 vacuolation in the bronchiolar epithelium, which wa s not consistently observed following 5487 repeated exposures. With repeated exposure to 4,00 0 ppm, the Clara cell vacuolation did not 5488 progress to necrosis, and no hyperplasia of the bro nchiolar epithelium was found after up to 5489 13 weeks of exposure. The transient Clara cell vac uolation was decreased by CYP inhibition 5490 with piperonyl butoxide and was unaffected by GSH d epletion, indicating that a CYP metabolite 5491

was involved. Clara cell vacuolation was not found after five consecutive daily 6-hour 5492 exposures to 4,000 ppm but reappeared after 2 days without exposure followed by two additional 5493 consecutive daily exposures (day 9). With repeated exposure, the lesion was detected at a 5494 diminished severity on day 44 (but was not found on day 40 or 43) and on day 93 (but was not 5495 found on day 89 or 92). The temporal pattern of Cl ara cell vacuolation with repeated exposure 5496 was reflected in the occurrence of transiently decr eased CYP metabolic activity after the 5497 appearance of vacuolation. Foster et al. (1994, 19 92) proposed that the diminishment of severity, 5498 or the disappearance, of the Clara cell vacuolation with repeated exposure was due to the 5499 development of a tolerance to dichloromethane, link ed with a decrease of CYP metabolism of 5500

dichloromethane. 5501

5502 4.5.4. Mechanistic Studies of Neurological Effects 5503 Several neurobehavioral studies (see section 4.4.3 for a complete summary) have 5504 demonstrated that dichloromethane exposure results in decreased spontaneous motor activity 5505 with pronounced lethargy at high concentrations (1, 000 ppm or greater). These effects, 5506

combined with the observation that dichloromethane impairs learning and memory (Alexeef and 5507 Kilgore, 1983) and affects production of evoked res ponses to sensory stimuli (Herr and Boyes, 5508 1997; Rebert et al., 1989), indicate that dichlorom ethane produces significant neurological 5509 effects. The mechanisms behind producing these cha nges have been examined by measuring 5510 changes in neurotransmitter levels and changes in n eurotransmitter localization. Specific brain 5511 regions (e.g., hippocampus, caudate nucleus, cerebe llum) were analyzed to determine if 5512 dichloromethane-induced behavioral effects, such as learning and memory (hippocampus, 5513 caudate nucleus) and movement (cerebellum), are res ulting from pathological changes in these 5514 regions. Changes in neurotransmitter levels were a lso monitored to see if there was any 5515 correlation in behavior and neurochemical changes. Summaries of these studies are provided 5516 below. It is not yet known if dichloromethane dire ctly interacts with neuronal receptors, as has 5517 been demonstrated for toluene and ethanol, two othe r solvents with neurobehavioral and 5518176

neurophysiological profiles that are similar to tho se of dichloromethane (for a review see Bowen 5519 et al. [2006]). 5520 Kanada et al. (1994) examined the effect of dichlor omethane on acetylcholine and 5521 catecholamines (dopamine, norepinephrine, serotonin ) and their metabolites in the midbrain, 5522 hypothalamus, hippocampus, and medulla from male Sp rague-Dawley rats (four to five per 5523 group). The rats were sacrificed 2 hours after a s ingle gavage dose of 0 or 534 mg/kg of 5524

undiluted dichloromethane. Administration of dichl oromethane significantly increased the 5525 concentration of acetylcholine in the hippocampus a nd increased dopamine and serotonin levels 5526 in the medulla. Dichloromethane decreased norepine phrine levels in the midbrain, and 5527 hypothalamus and serotonin levels were decreased in the hypothalamus. There was a trend 5528 toward decreased dopamine in the hypothalamus, but the variability between the animals was so 5529 high that the effect was not significant. The auth ors speculated that increased acetylcholine 5530 release from dichloromethane administration may be due to decreased acetylcholine release from 5531

the nerve terminals. It is unclear as to how these neurochemical changes could be correlated to 5532 the neurobehavioral changes observed after dichloro methane exposure. 5533 In a 2-week exposure study, male Wistar rats were e xposed to dichloromethane at 500 or 5534 1,000 ppm (6 hours/day, 5 days/week for 1 or 2 week s) or 1,000 ppm TWA (1 hour at 100 ppm, 5535 1 hour peak at 2,800 ppm, 1 hour at 100 ppm, repeat ed immediately, 5 days/week for 1 or 5536 2 weeks) (Savolainen et al., 1981). Brains were re moved from rats at the end of the study and 5537 analyzed. The 1,000 ppm TWA group displayed increa ses in cerebral RNA. Other changes 5538 noted for this group in the cerebrum included signi ficant increases in NADPH-diaphorase and 5539 succinate dehydrogenase activity. These changes su ggest increased neural activity to possibly 5540 offset the overall inhibitory effect of dichloromet hane in the CNS. It could also possibly explain 5541 why lethargic effects decrease with continued dichl oromethane exposure, and this result 5542 demonstrates a neuroprotective mechanism resulting from dichloromethane exposure. After a 5543 7-day withdrawal, RNA levels in the cerebrum were s ignificantly lower in the 1,000 ppm group. 5544 Succinate dehydrogenase levels remained lowered in the 1,000 ppm TWA group after the 7-day 5545 exposure-free period. 5546 Changes in brain catecholamine levels after a subac ute exposure to dichloromethane were 5547 evaluated using male Sprague-Dawley rats (Fuxe et a l., 1984). Rats were exposed to 70, 300, 5548 and 1,000 ppm dichloromethane, 6 hours/day for 3 co nsecutive days. At all exposures, there was 5549 a significant decrease of catecholamine concentrati ons in the posterior periventricular region of 5550 the hypothalamus. The impact of dichloromethane wa s also evaluated on the hypothalamic5551 pituitary gonadal axis. The hypothalamus regulates secretion of reproductive hormones, such as 5552 follicle-stimulating hormone and luteinizing hormon e. The levels of the hormone release were 5553 not significantly changed with dichloromethane expo sure. In the caudate nucleus, which is 5554 involved in memory processes, the catecholamine lev el initially increased (at 70 ppm) and then 5555 was lower (1,000 ppm) in comparison to the control. The study overall demonstrates significant 5556177

changes in catecholamine levels in the hypothalamus and caudate nucleus. Catecholamine level 5557 changes in the hypothalamus did not have any signif icant effect on hormonal release and 5558

decreased catecholamine levels in the caudate nucle us at higher exposures may lead to memory 5559 and learning impairment. 5560

A series of studies were conducted in male and fema le Mongolian gerbils exposed 5561 continuously to ≥ 210 ppm dichloromethane for 3 months, followed by a 4-month exposure-free 5562 period (Karlsson et al., 1987; Briving et al., 1986 ; Rosengren et al., 1986). Decreased DNA 5563 concentrations were noted in the hippocampus at bot h the 210 and 350 ppm exposures. At 5564 350 ppm, there was also decreased DNA concentration in the cerebellar hemispheres, indicating 5565 a decreased cell density in these regions probably due to cell loss (Rosengren et al., 1986). 5566 These findings indicate that the cerebellum, which is the section of the brain that regulates motor 5567 control, is a target for dichloromethane. In the s ame study, increased astroglial proteins were 5568 found in the frontal and sensory motor cerebral cor tex, which directly correlated to the 5569 astrogliosis that was observed in those areas. Up- regulation of these astroglial proteins is a good 5570 indicator of neuronal injury (Rosengren et al., 198 6). 5571 Karlsson et al. (1987) measured DNA concentrations in different regions of the gerbil 5572 brain. The total brain protein concentration per w et weight was not significantly different 5573 between dichloromethane-exposed and control animals . However, DNA concentrations per wet 5574 weight were significantly decreased in the hippocam pus after dichloromethane exposure. No 5575 other examined regions demonstrated significant cha nges in DNA concentrations after 5576 dichloromethane exposure. Therefore, this result i ndicates that the hippocampus, which plays a 5577 role in the formation of new memories, is another t arget for dichloromethane in the CNS. This 5578

selective DNA concentration decrease observed in th e hippocampus is a sign of neurotoxicity as 5579 noted by the authors and may possibly explain why s ome studies have noted memory and 5580 learning deficits with dichloromethane exposure. 5581 At a 210 ppm exposure, Briving et al. (1986) observ ed that dichloromethane decreased 5582 glutamate, γ -aminobutyric acid, and phosphoethanolamine levels in the frontal cortex, while 5583 glutamate and γ -aminobutyric acid were increased in the posterior cerebellar vermis. Increased 5584 levels of glutamate in the posterior cerebellar ver mis could reflect an activation of astrocytic glia, 5585 since glutamine synthetase is localized exclusively in astrocytes. 5586

5587

4.6. SYNTHESIS OF MAJOR NONCANCER EFFECTS 5588 4.6.1. Oral Exposures 5589

4.6.1.1. Summary of Human Data 5590 Information on noncancer effects in humans exposed orally to dichloromethane are 5591 restricted to case reports of neurological impairme nt, liver and kidney effects (as severe as organ 5592 failure), and gastrointestinal irritation in indivi duals who ingested amounts ranging from about 5593 25 to 300 mL (Chang et al., 1999; Hughes and Tracey , 1993). Neurological effects with these 5594178

individuals consisted of general CNS depressive sym ptoms, such as drowsiness, confusion, 5595 headache, and dizziness. Hemoglobinuria has been n oted as a kidney effect associated with 5596 ingestions. 5597

5598

4.6.1.2. Summary of Animal Data 5599 Acute oral or intraperitoneal administration of di chloromethane in animals has resulted in 5600 several significant effects. General activity and function were affected as evidenced by 5601 decreased neuromuscular activity (Moser et al., 199 5). Additionally, decreased sensorimotor 5602 function was detected through measurement of evoked potentials (Herr and Boyes, 1997) and by 5603 using the FOB (Moser et al., 1995). Neurochemical changes (e.g. acetylcholine, dopamine, 5604 norepinephrine, serotonin) were detected 2 hours af ter oral dosage of dichloromethane within 5605 specific parts of the brain. It should be noted th at all the acute effects that were observed after 5606 oral or intraperitoneal administration occurred wit hin 5 hours after dosage. No other significant 5607 organ effects were noted after a single acute oral exposure, but in oral pharmacokinetic studies it 5608 is known that dichloromethane is primarily distribu ted to the liver, lungs, and kidneys (Angelo et 5609 al., 1986a). 5610 Results from short-term, subchronic, and chronic or al toxicity studies in laboratory 5611

animals are summarized in Table 4-35. The data ind icate that rats may be more sensitive than 5612 mice to nonneoplastic liver effects from orally adm inistered dichloromethane, as evidenced by 5613 lower NOAELs and LOAELs, with more severe liver eff ects in rats. The most frequently 5614 observed liver effect was hepatocyte vacuolation, s een with drinking water exposure (90 days) in 5615

F344 rats at ≥ 166 mg/kg-day and B6C3F 1 mice at 586 mg/kg-day (Kirschman et al., 1986) and 5616 with gavage exposure (14 days) in CD-1 mice at 333 mg/kg-day (Condie et al., 1983). 5617 Hepatocyte degeneration or necrosis was observed in female F344 rats exposed in drinking water 5618 for 90 days to 1,469 mg/kg-day (Kirschman et al., 1 986) and in female F344 rats exposed by 5619 gavage for 14 days to 337 mg/kg-day (Berman et al., 1995), but was not seen in a 90-day 5620 drinking water study in B6C3F 1 mice exposed to doses as high as 2,030 mg/kg-day ( Kirschman 5621 et al., 1986). In the chronic-duration (2-year) st udy, liver effects were described as 5622 nonneoplastic foci and areas of alteration in F344 rats exposed to drinking water doses between 5623 50 and 250 mg/kg-day; an increased incidence of fat ty changes in the liver was also noted but the 5624 incidence of the latter was not provided (Serota et al., 1986a). These effects were considered to 5625 be nonneoplastic for several reasons. Serota et al ., (1986b) observed a dose-related increased 5626 incidence of 0, 65, 92, 97, 98 and 100% in male rat s and 51, 41, 73, 89, 91 and 85% in female 5627 rats for the 0, 5, 50, 125, 250 and 250 with recove ry groups, respectively. Evidence for liver 5628 tumors has been reported in female rats only. Spec ifically, evidence for liver tumors in rats 5629 includes a small number of hepatocellular carcinoma s observed in female rats at 50 and 250 5630 mg/kg-day, which reached statistical significance ( for trend and for individual pairwise 5631 comparisons) only with the combined grouping of neo plastic nodules and hepatocellular 5632179

carcinomas. In male rats, only one hepatocellular c arcinoma was observed in all of the exposure 5633 groups (compared to 4 in the controls), and the inc idence of neoplastic nodules and 5634 hepatocellular carcinomas was higher in controls (1 6%) than in any exposure group (16, 3, 0, 6, 5635 5 and 13% for the 0, 5, 50,125, 250 mg/kg-day and 2 50 with recovery groups, respectively). The 5636 authors (Serota et al., 1986a) did not elaborate on the characterization of the altered foci. 5637 However, the characterization of altered foci could range from a focal change in fat distribution 5638 (nonneoplastic effect) to enzyme altered foci which are generally considered a precursor to 5639 tumor formation (Goodman et al., 1994). Serota et al (1986a) reported an increased incidence of 5640 fatty change in the liver at doses of 50 mg/kg-day and higher, but the incidence was not reported. 5641 In addition, a 90-day study (Kirschman et al., 1986 ) demonstrated that increased fatty deposits 5642 were present in the hepatocyte vacuoles. Therefore , the altered foci (i.e. change in fat 5643 distribution) observed by Serota et al., (1986b) ma y represent a precursor to fatty liver changes 5644 which is considered a nonneoplastic effect. Taken together, the data support the conclusion that 5645 the altered foci were nonneoplastic. 5646 The NOAEL and LOAEL, 101 and 337 mg/kg-day, for alt ered neurological functions in 5647 female F344 rats (as reported by Moser et al. [1995 ]) were identical to those reported by Berman 5648 et al. (1995) for hepatocyte necrosis in female F34 4 rats. In the 90-day (Kirschman et al., 1986) 5649 and 104-week (Serota et al., 1986a, b) drinking wat er studies, no obvious clinical signs of 5650 neurological impairment were observed in rats or mi ce at exposure levels that induced liver 5651 effects (see Table 4-35), but this study did not in clude a standardized neurological testing 5652 battery. 5653180Table 4-35. NOAELs and LOAELs in selected animal studies involving oral exposure to dichloromethanefor short-term, subchronic, or chronic durationsNOAEL LOAELType of effect andexposure, reference Species and exposure detailsResults(mg/kg-day)Hepatic, 14-day gavageBerman et al. (1995) F344 rat, female, 8/dose group0, 34, 101, 337, 1,012 mg/kg-dayHepatocyte necrosis101 337Condie et al. (1983) CD-1 mouse, male, 5/group forhistological examinations; 8/group forblood urea nitrogen, serum creatinine,and serum glutamate-pyruvatetransaminase; 0, 133, 333, 665 mg/kgdayHepatocyte vacuolation (minimal to mild in 3/5) 133333Hepatic, 90-day drinking waterKirschman et al. (1986) F344 rat, male and female;15/sex/group;males 0, 166, 420, 1,200 mg/kg-dayfemales 0, 209, 607, 1,469 mg/kg-dayHepatic vacuolation (generalized, centrilobular, orperiportal, at lowest dose, in 10/15 males and 13/15 femalescompared with 1/15 males and 6/15 females in controls)Notidentified166Kirschman et al. (1986) B6C3F1mouse, male and female,males 0, 226, 587, 1,911 mg/kg-dayfemales 0, 231, 586, 2,030 mg/kg-dayHepatic vacuolation (increased severity of centrilobular fattychange in mid- and high-dose groups compared withcontrols)231 586Hepatic, 104-week drinking waterSerota et al. (1986a) F344 rat, male and female,0, 5, 50, 125, 250 mg/kg-dayLiver foci/areas of alteration (considered nonneoplastichistologic changes); fatty liver changes also seenat samedoses but incidence data not reported; no evidencethatincreased altered foci progresses to liver tumor formation5 50Serota et al. (1986b)Hazelton Laboratories(1983)B6C3F1mouse, male and female,0, 60, 125, 185, 250 mg/kg-daySome evidence of fatty liver; marginal increase inthe OilRed-O-positive material in the liver185 250Neurologic, 14 dayMoser et al. (1995) F344 rat, female,0, 34, 101, 337, 1,012 mg/kg-dayFOB 24 hours postexposure: altered autonomic,neuromuscular, and sensorimotor and excitability measures101 337ReproductiveGeneral Electric Co.(1976)Charles River CD rat, male and female,gavage for 90 d before mating (10 daysbetween last exposure and matingperiod); 0, 25, 75, 225 mg/kg-day; F1offspring received same treatment asparents for 90 daysReproductive performance of F0 and histologic examinationof tissues from F1 offspring225 Notidentified181Table 4-35. NOAELs and LOAELs in selected animal studies involving oral exposure to dichloromethanefor short-term, subchronic, or chronic durationsNOAEL LOAELType of effect andexposure, reference Species and exposure detailsResults(mg/kg-day)Raje et al. (1988) Swiss-Webster mouse, male, 0, 250,500 mg/kg (subcutaneous injection),3×per week, 4 weeks prior to matingwith nonexposed females (1 weekbetween last exposure and mating period)No statistically significant effects on testes, number oflitters, live fetuses/litter, percent dead fetuses/litter, percentresorbed/litter, or fertility index200 NotidentifiedDevelopmentalNarotsky and Kavlock(1995)F344 rat, pregnant female, gavage onGDs 6–19; 0, 338, 450 mg/kg-dayMaternal: weight gain depression 338 450Fetal: no effects on pup survival, resorptions, pupweight 450 NotidentifiedOther developmentalNo studies182

Results from a limited number of studies do not pr ovide evidence for effects on 5654 reproductive or developmental endpoints (Table 4-35 ). No effects on pup survival, resorptions, 5655 or pup weight were found following exposure of preg nant F344 rats to doses as high as 5656 450 mg/kg-day on GDs 6–19, a dose that depressed ma ternal weight gain (Narotsky and 5657 Kavlock, 1995), and no effects on reproductive perf ormance endpoints (fertility index, number 5658 of pups per litter, pup survival) were found in Cha rles River CD rats exposed for 90 days before 5659 mating to doses as high as 225 mg/kg-day. There ar e no oral exposure studies focusing on 5660 neurobehavioral effects or other developmental outc omes. 5661

5662 4.6.2. Inhalation Exposures 5663

4.6.2.1. Summary of Human Data 5664 As discussed in sections 4.1.3.1 and 4.1.3.2, acute inhalation exposure of humans to 5665

dichloromethane has been associated with cardiovasc ular impairments due to decreased oxygen 5666 availability from COHb formation and neurological i mpairment from interaction of 5667 dichloromethane with nervous system membranes. Res ults from studies of acutely exposed 5668 human subjects indicate that acute neurobehavioral deficits, measured, for example, by 5669 psychomotor tasks, tests of hand-eye coordination, visual evoked response changes, and auditory 5670 vigilance, may occur at concentrations >200 ppm wit h 4–8 hours of exposure (Bos et al., 2006; 5671 American Conference of Governmental Industrial Hygi enists [ACGIH], 2001; ATSDR, 2000; 5672 Cherry et al., 1983; Putz et al., 1979; Gamberale e t al., 1975; Winneke, 1974). 5673 The clinical and workplace studies of noncancer hea lth effects of chronic 5674 dichloromethane exposure have examined markers of d isease and specific clinical endpoints 5675 relating to cardiac, neurological disease, hepatic function, and reproductive health. As 5676 summarized in section 4.1.2.9, the limited availabl e data do not provide evidence of cardiac 5677 damage related to dichloromethane exposure in occup ationally exposed workers (Hearne and 5678 Pifer, 1999; Tomenson et al., 1997; Gibbs et al., 1 996; Lanes et al., 1993; Ott et al., 1983d; 5679 Cherry et al., 1981). Relatively little is known a bout the long-term neurological effects of 5680 chronic exposures, although there are studies that provide some evidence of an increased 5681 prevalence of neurological symptoms among workers w ith average exposures of 75–100 ppm 5682 (Cherry et al., 1981), long-term effects on some ne urological measures (i.e., possible detriments 5683 in attention and reaction time in complex tasks) in workers whose past exposures were in the 5684 100–200 ppm range (Lash et al., 1991), and an incre ased risk of suicide in worker cohort studies 5685 (Hearne and Pifer, 1999; Gibbs, 1992). Given the s uggestions from these studies and their 5686 limitations (particularly with respect to sample si ze and power considerations), the statement that 5687 there are no long-term neurological effects of chro nic exposures to dichloromethane cannot be 5688 made with confidence. With respect to markers of h epatic damage, three studies measured 5689 serum enzyme and bilirubin levels in workers expose d to dichloromethane (Soden, 1993; 5690 Kolodner et al., 1990; Ott et al., 1983c). There i s some evidence of increasing levels of serum 5691183

bilirubin with increasing dichloromethane exposure (Kolodner et al., 1990; Ott et al., 1983c), but 5692 there are no consistent patterns with respect to th e other hepatic enzymes examined (serum γ

5693 glutamyl transferase, serum AST, serum ALT). Thus these studies do not provide clear evidence 5694 of hepatic damage in dichloromethane exposed worker s to the extent that this damage could be 5695 detected by these serologic measures. 5696 Only limited, and somewhat indirect, evidence perta ining to immune-related effects of 5697

dichloromethane in humans is available. No risk ofthe broad category of infection- and parasite5698 related mortality was reported by Hearne and Pifer (1999), but there was some evidence of an 5699 increased risk of influenza and pneumonia-related m ortality at two cellulose triacetate fiber 5700 production work sites in Maryland and South Carolin a (Gibbs, 1992). 5701 Few studies have been conducted pertaining to repro ductive effects (i.e., spontaneous 5702 abortion, low birth weight, or oligospermia) of dic hloromethane exposure from workplace 5703 settings (Wells et al., 1989; Kelly, 1988; Taskinen et al., 1986) or environmental settings (Bell et 5704 al., 1991). Of these, the data pertaining to spont aneous abortion provide the strongest evidence 5705 of an adverse effect of dichloromethane exposure. The limitations of the only study (Taskinen et 5706 al., 1986) pertaining to this outcome, however, do not allow firm conclusions to be made 5707 regarding dichloromethane and risk of spontaneous a bortion in humans. 5708

5709

4.6.2.2. Summary of Animal Studies 5710 Acute and short-term (up to 7 days) inhalational ex posure to dichloromethane has 5711 resulted in neurological and hepatocellular changes . Several neurological-mediated parameters 5712 were reported, including decreased activity (Kjells trand et al., 1985; Weinstein et al., 1972; 5713 Heppel and Neal, 1944), impairment of learning and memory (Alexeef and Kilgore, 1983), and 5714 changes in responses to sensory stimuli (Rebert et al., 1989). Although learning and memory 5715 properties were impaired in one acute exposure (47, 000 ppm until loss of righting reflex), it 5716

should be noted that this effect has not been chara cterized by using other learning and memory 5717 tasks nor any other exposure paradigms. In a 3-day exposure to dichloromethane (70, 300, or 5718 1,000 ppm 6 hours/day), it was found that in the ra t brain there were changes in catecholamine 5719 (dopamine, serotonin, norepinephrine) in the hypoth alamus and caudate nucleus (Fuxe et al., 5720 1984). The catecholamine level changes did not aff ect hormonal release, which is a primary 5721 function of the hypothalamus. 5722 Another acute exposure study examined immunologica l response as measured by 5723 increased streptoccocal pneumonia-related mortality and decreased bactericidal activity of 5724 pulmonary macrophages in CD-1 mice following a sing le 3-hour exposure to dichloromethane at 5725 100 ppm (Aranyi et al., 1986). No effects were see n at 50 ppm. A 4-week inhalation exposure 5726 to 5,000 ppm dichloromethane in rats did not result in changes in immune response as measured 5727 by the sheep red blood cell assay (Warbrick et al., 2003). These studies suggest a localized, 5728 portal-of-entry effect within the lung, without evi dence of systemic immunosuppression. 5729184

Mouse hepatocytes showed balloon degeneration (diss ociation of polyribosomes and 5730 swelling of rough endoplasmic reticulum) within 12 hours of exposure to 5,000 ppm (Weinstein 5731 et al., 1972). A subacute exposure in Wistar rats to 500 ppm dichloromethane 6 hours/day for 6 5732 days resulted in increased hemochrome content in li ver microsomal CYP (Savolainen et al., 5733 1977). 5734 Results pertaining to liver, lung, and neurological effects from longer (>7 days) 5735 subchronic and chronic inhalation toxicity studies in laboratory animals are summarized in 5736 Table 4-36; reproductive and developmental studies are summarized in Table 4-37. 5737185Table 4-36. NOAELs and LOAELs in animal studies involving inhalation exposure to dichloromethane forsubchronic or chronic durations, hepatic, pulmonary, and neurologic effectsNOAEL LOAELType of effect andexposure period,referenceSpecies and exposure details Results ppmHepatic, subchronic (13–14 weeks)Haun et al. (1971) Beagle, female (n = 8);rhesus monkeys, females (n = 4);Sprague-Dawley rats, male (n = 20),ICR mice, females (n = 380)0, 1,000, 5,000 ppm (continuousexposure; 14 weeks)Fatty liver at 1,000 ppm in dogs, “borderline” liverchanges in monkey at 5,000 ppm, mottled liverchanges in rats at 5,000 ppmDecreased movement and lethargy at 1,000 ppm indogs, mice, and monkey.Not identifiedNot identified1,000Not identified1,000 dog1,000 monkey5,000 rat1,000 mouseHaun et al. (1972) Beagle (n = 16);Rhesus monkey (n = 4);Sprague-Dawley rats (n = 20),ICR mice (n = 20)0, 25, 100 ppm (continuous exposure;14 weeks)CYP levels decreased in liver microsomes in miceat 100 ppmIncreased fatty liver content at 25, 100 ppm in rats25;Not identifiedNot identified25100 dog25 monkey25 rat100 mouseLeuscher et al. (1984) Sprague-Dawley rat, male andfemale,(20/sex/group) - 0, 1,000 ppm (6hours/day; 90 days)Beagle, male and female (3/sex/group) -0, 5,000 ppmNo liver effects noted 1,0005,000Not identifiedNot identifiedNTP (1986) F344/N rat, male and female(10/sex/group)0, 525, 1,050, 2,100, 4,200, 8,400 ppm(6 hours/day, 5 days/week, 13 weeks)Decreased lipid:liver weight ratios at4,200 (females); 8,400 (males); decreased BW by23% and 11% in males and females at 8,400 ppmcompared with controls; one male and one femaledied at 8,400 ppm before the end of the study.4,200 8,400NTP (1986) B6C3F1mouse, male and female(10/sex/group)0, 525, 1,050, 2,100, 4,200, 8,400 ppm(6 hours/day, 5 days/week, 13 weeks)Hepatocyte centrilobular degeneration at 4,200females) and 8,400 (males); decreased lipid:liverweight ratios at 8,400 (females); at 8,400 ppm, 4/10males and 2/10 females died before end of study.2,100 4,200Hepatic, 2 years (6 hours/day, 5 day/week)Mennear et al. (1988);NTP (1986)F344/N rat, male and female0, 1,000, 2,000, 4,000 ppmHepatocyte vacuolation and necrosisHemosiderosis in liverRenal tubular degenerationNot identifiedNot identified1,0001,0001,0002,000186Table 4-36. NOAELs and LOAELs in animal studies involving inhalation exposure to dichloromethane forsubchronic or chronic durations, hepatic, pulmonary, and neurologic effectsNOAEL LOAELType of effect andexposure period,referenceSpecies and exposure details Results ppmMennear et al. (1988);NTP (1986)B6C3F1mouse, male and female0, 2,000, 4,000 ppmHepatocyte degenerationRenal tubule castsNot identifiedNot identified2,0002,000Burek et al. (1984) Syrian golden hamster, male andfemale0, 500, 1,500, 3,500 ppmNo effects on histologic, clinical chemistry,urinalytic, and hematologic variables no obviousclinical signs of toxicity3,500 Not identifiedBurek et al. (1984) Sprague-Dawley rat, male and female0, 500, 1,500, 3,500 ppmHepatocyte vacuolation (M and F)Hepatocyte necrosis (M only), no obvious clinicalsigns of toxicity)Not identified5005001,500Nitschke et al. (1988a) Sprague-Dawley rat, male and female0, 50, 200, 500 ppmHepatocyte vacuolation significantly increased infemales; non-significant increase in males at 500ppm (31% in controls and 40% in 500 ppm group).200 500Pulmonary, 13 weeks (6 hours/day, 5 days/week)NTP (1986) F344 rat, male and female0, 525, 1,050, 2,100, 4,200, 8,400 ppmForeign body pneumonia 4,200 8,400Foster et al. (1992) B6C3F1mouse, male and female0, 4,000 ppmClara cell vacuolation Not identified 4,000Neurological, 14 daysSavolainen et al. (1981) Wistar rats, male500, 1,000, 1,000 TWA (100 + 2,8001-hour peaksa) ppm (6 hours/day,5 days/week, 2 weeks)Increased RNA in cerebrum at 1,000 ppm;increased enzymatic activitiesbin cerebrum andcerebellum at 1,000 ppm TWA500 1,000 for brainRNAconcentration;1,000 TWA forbrain enzymaticactivityNeurological, 13–14 weeksMattsson et al. (1990) F344 rat, male and female0, 50, 200, 2,000 ppm(6 hours/day, 5 days/week)No exposure-related effects on an observationalbattery, hind-limb grip strength, a battery of evokedpotentials, or histology of brain, spinal cord,peripheral nerves; measured 64 hours postexposure2,000 Not identifiedHaun et al. (1971) Beagle dogs (female);Rhesus monkeys (female);Sprague-Dawley rats (male);ICR mice (females)0, 1,000, 5,000 ppm(continuous exposure)CNS depression most evident in dogs Not identifiedNot identified1,000Not identified1,000 dog1,000 monkey5,000 rat1,000 mouse187Table 4-36. NOAELs and LOAELs in animal studies involving inhalation exposure to dichloromethane forsubchronic or chronic durations, hepatic, pulmonary, and neurologic effectsNOAEL LOAELType of effect andexposure period,referenceSpecies and exposure details Results ppmKarlsson et al. (1987)Briving et al. (1986)Rosengren et al. (1986)Mongolian gerbils, male and female210, 350, 700 ppm (continuous exposure,followed by 4 month exposure-freeperiod)Astrogliosis in frontal and sensory motor cerebralcortex suggested by increases in astroglial proteins;cell loss in cerebellar regions; decreased DNA inhippocampus; neurochemical changes observed atall exposuresNot identified 210Thomas et al. (1972) ICR mice, female0, 25, 100 ppm, continuousIncreased spontaneous activity observed at 25 ppmbut not 100 ppmNot identified 25CoHb, 13-14 weeksHaun et al. (1972) Beagles (n = 16);Rhesus monkeys (n = 4);Sprague-Dawley rats (n = 20),ICR mice (n = 20)0, 25, 100 ppm (continuous exposure;14 weeks)CoHb levels significantly higher at 25, 100 ppmfor monkeys and 100 ppm for beaglesNot identified 25COHb, 2 years (6 hours/day, 5 day/week)Burek et al. (1984) Syrian golden hamster, male andfemale0, 500, 1,500, 3,500 ppmAbout 30% COHb in each exposed groupBurek et al. (1984) Sprague-Dawley rat, male and female0, 500, 1,500, 3,500 ppmAbout 12–14% COHb in each exposed groupNitschke et al. (1988a) Sprague-Dawley rat, male and female0, 50, 200, 500 ppmCOHb values at 2 years: about 2, 7, 13, 14%aEquivalent to 1,000 ppm TWA.bDecreased GSH,γ-aminobutyric acid, and phosphoethanolamine in frontal cortex; GSH andγ-aminobutyric acid increased in posterior cerebellar vermis.5738573957401885741Table 4-37. NOAELs and LOAELs in selected animal studies involving inhalation exposure todichloromethane, reproductive and developmental effectsNOAEL LOAELType of effect and exposureperiod, referenceSpecies and exposure details Results ppmReproductiveNitschke et al. (1988b) F344 rat, male and female,F0: 6 hr/d,5 d/wk for 14 wk before mating andGDs 0 to 21; F1: 6 hr/d, 5 d/wk,beginning PND 4 for 17 wk beforemating; 0, 100, 500, 1,500 ppmNo statistically significant effects on fertilityor litter size, neonatal survival, growth rates,or histopathologic lesions in F1 orF2 weanlings1,500 Not identifiedMennear et al. (1988); NTP(1986)B6C3F1mouse; 0, 2,000 or 4,000 ppm,6 hours/day, 5 days/week for 2 yearsTesticular atrophyOvarian atrophy (considered secondary tohepatic effects)2,000Not identified4,0002,000Raje et al. (1988) Swiss-Webster mouse, male, 2 hr/d,5 d/wk for 6 wk before mating withnonexposed females; 0, 100, 150,200 ppmNo statistically significant effects on testes,number of litters, live fetuses/litter, percentdead fetuses/litter, percent resorbed/litterFertility index was lower in 150 and 200 ppmgroups (80%) compared with controls and100 ppm groups (95%) (statisticalsignificance depends on test used).200100Not identified150DevelopmentalSchwetz et al. (1975) Swiss-Webster mouse, pregnantfemale,7 hr/d, GDs 6–15; 0, 1,250 ppmMaternal effects: 9–10% COHb; increasedabsolute, not relative, liver weight, increasedmaternal weight (11–15%).Fetal effects: increased litters with extracenter of ossification in sternumNot identified1,2501,250Not identifiedSchwetz et al. (1975) Sprague-Dawley rat, pregnantfemale,7 hr/d, GDs 6–15; 0, 1,250 ppmMaternal: 9–10% COHb; increased absolute,not relative, liver weightFetal: increased incidence of delayedossification of sternebraeNot identified1,2501,250Not identifiedOther developmentalBornschein et al. (1980); Hardinand Manson (1980)Long-Evans rat, female, 6 hr/d for 12–14 d before breeding and GDs 1–17;6 hr/d on GDs 1–15; 0, 4,500 ppmMaternal (both protocols): increased absoluteand relative liver weight (~10%)Fetal/offspring: decreased fetal BW (~10%);changed behavioral habituation to novelenvironments; no changes in gross, skeletal,or soft-tissue anomaliesNot identifiedNot identified4,5004,500189

Hepatic centrilobular degeneration was observed in several studies containing different 5742 species and inhalational exposures. This effect wa s observed in guinea pigs exposed to 5743 5,000 ppm (7 hours/day) for 6 months (Heppel et al. , 1944). Monkeys, rats, and mice 5744 continuously exposed (24 hours/day) to 5,000 ppm di chloromethane for 14 weeks also had 5745 increased centrilobular degeneration (Haun et al., 1972, 1971). This effect was also observed at 5746 lower exposures when mice were exposed to 4,200 ppm for 6 hours/day for 13 weeks (NTP, 5747 1986) and in dogs exposed to 1,000 ppm for 24 hours /day for up to 14 weeks (Haun et al., 1972, 5748 1971). 5749 Increased incidences of histologic hepatic lesions were not found in F344 rats exposed to 5750 4,200 or 8,400 ppm 6 hours/day for 13 weeks (NTP, 1 986) or in Sprague-Dawley rats exposed to 5751 10,000 ppm 6 hours/day for 90 days (Leuschner et al ., 1984). Hepatic lesions were also not 5752 observed in beagle dogs exposed to 5,000 ppm 6 hour s/day for 90 days (Leuschner et al., 1984) 5753 or in dogs, monkeys, rats, and mice exposed to 25 o r 100 ppm for 24 hours/day for up to 5754 14 weeks (Haun et al., 1972). Heppel et al. (1944) also demonstrated absence of hepatic lesions 5755 in unspecified strains of monkeys, rabbits, and rat s exposed to 10,000 ppm 4 hours/day for up to 5756 8 weeks and in unspecified strains of dogs, rabbits , and rats exposed to 5,000 ppm 7 hours/day 5757 for up to 6 months. 5758 Gross neurological impairments were observed in sev eral laboratory species with 5759 repeated exposure to 10,000 ppm for 4 hours/day for 8 weeks (Heppel et al., 1944) or to 1,000 or 5760 5,000 ppm for 24 hours/day for 14 weeks (Haun et al ., 1972, 1971). Dogs exposed to 5,000 ppm 5761 6 hours/day for 90 days showed slight sedation duri ng exposures, but Sprague-Dawley rats 5762 exposed to 10,000 ppm for 90 days did not (Leuschne r et al., 1984). In F344 rats exposed to 5763 concentrations up to 2,000 ppm 6 hours/day for 13 w eeks, no effects were observed on an 5764 observational battery, hind-limb grip strength, a b attery of evoked potentials, or histology of the 5765 brain, spinal cord, or peripheral nerves; these tes ts were conducted beginning 65 hours or more 5766 after the last exposure (Mattsson et al., 1990). 5767 Exposure-related nonneoplastic effects on the lungs reported in the subchronic studies 5768 were restricted to foreign body pneumonia in rats e xposed to 8,400 ppm 6 hours/day for 5769 13 weeks (NTP, 1986), Clara cell vacuolation in mic e exposed to 4,000 ppm 6 hours/day for 5770 13 weeks (Foster et al., 1992), and pulmonary conge stion in guinea pigs exposed to 5,000 ppm 5771 7 hours/day for 6 months (Heppel et al., 1944). 5772

The chronic duration inhalation studies were conduc ted at lower exposure levels than the 5773 short-term and subchronic studies and provide resul ts indicating that the liver is the most 5774 sensitive target for noncancer toxicity in rats and mice (Table 4-36). Life-time exposure was 5775 associated with hepatocyte vacuolation and necrosis in F344 rats exposed to 1,000 ppm 5776 6 hours/day (Mennear et al., 1988; NTP, 1986), hepa tocyte vacuolation in Sprague-Dawley rats 5777 exposed to 500 ppm 6 hours/day (Nitschke et al., 19 88a; Burek et al., 1984), and hepatocyte 5778 degeneration in B6C3F 1 mice exposed to 2,000 ppm 6 hours/day (lower conce ntrations were not 5779190

tested in mice) (Mennear et al., 1988; NTP, 1986). As shown in Tables 4-36 and 4-37, other 5780 effects observed include renal tubular degeneration s in F344 rats and B6C3F 1 mice at 2,000 ppm, 5781 testicular atrophy in B6C3F 1 mice at 4,000 ppm, and ovarian atrophy in B6C3F 1 mice at 5782 2,000 ppm (considered secondary to hepatic effects) . No exposure-related increased incidences 5783 of nonneoplastic lung lesions were found in any of the chronic studies (Table 4-36). 5784 In comparison to rats and mice, Syrian golden hamst ers are less sensitive to the chronic 5785 inhalation toxicity of dichloromethane. No exposur e-related changes were found in 5786 comprehensive sets of histologic, clinical chemistr y, urinalytic, and hematologic variables 5787 measured in hamsters exposed for 2 years to 500, 1, 500, or 3,500 ppm for 6 hours/day, with the 5788 exception that mean COHb percentages were about 30% in each of these groups compared with 5789 a mean value of about 3% for the controls (Burek et al., 1984). 5790 The reproductive and developmental studies are limi ting in terms of the exposure 5791 regimen used, with two of the developmental studies using only a single, relatively high daily 5792 exposure over the gestational period (1,250 ppm, GD 6-15 in Schwetz et al. [1975] and 4,500 5793 ppm, GD 1-17 in Hardin and Manson [1980] and Bornsc hein [1980]). No significant effects on 5794 reproductive performance variables were found in a two-generation reproduction assay with 5795 F344 rats exposed to concentrations as high as 1,50 0 ppm (Nitschke et al., 1988b). No effects on 5796 most of the measures of reproductive performance we re observed in male mice exposed to 5797 200 ppm for 2 hours/day for 6 weeks before mating t o nonexposed females. Fertility index was 5798 reduced in the 150 and 200 ppm groups, but the stat istical significance of this effect varied 5799 considerably depending on the statistical test used in this analysis (Raje et al., 1988). No adverse 5800 effects on fetal development were found following e xposure of pregnant Swiss-Webster mice or 5801 Sprague-Dawley rats to 1,250 ppm 6 hours/day on GDs 6–15 (Schwetz et al., 1975). Following 5802 exposure of female Long-Evans rats to 4,500 ppm (6 hours/day) for 14 days before breeding plus 5803 during gestation or during gestation alone, a 10% d ecrease in fetal BW and changed behavioral 5804 habituation of the offspring to novel environments were seen (Bornschein et al., 1980; Hardin 5805 and Manson, 1980). No exposure-related changes in gross, skeletal, or soft-tissue anomalies 5806

were found. 5807

5808

4.6.3. Mode of Action Information 5809

4.6.3.1. Mode of Action for Nonneoplastic Liver Effects 5810 Studies of chronically exposed rats, both by the o ral route and the inhalation route, 5811 identified liver changes as the most sensitive expo sure-related noncancer effect associated with 5812 exposure to dichloromethane (Tables 4-35 to 4-37). The liver changes included increased 5813 incidence of liver foci/areas of alteration and hep atocyte vacuolation in rats and degenerative 5814 liver effects in rats, guinea pigs, monkeys, and mi ce. 5815 The mode of action by which dichloromethane induce s these nonneoplastic hepatic 5816

effects is unknown. The determination of whether o r not these effects are due to the parent 5817191

material, metabolites of the CYP2E1 pathway, metabo lites of the GST pathway, or some 5818 combination of parent material and metabolites has not been elucidated. The available data 5819 indicate that rats may be more sensitive than mice to the noncancer hepatotoxicity, but a 5820 mechanistic explanation of this possible interspeci es difference is not currently available. 5821

5822

4.6.3.2. Mode of Action for Nonneoplastic Lung Effects 5823 Single 6-hour inhalation exposures to concentration s ≥ 2,000 ppm dichloromethane 5824 produced a transient vacuolation of Clara cells in the bronchiolar epithelium of B6C3F 1 mice. 5825 Vacuolization of the Clara cells disappeared or was diminished with repeated exposure and was 5826 correlated with subsequent transient diminishment o f CYP metabolic activity. CYP inhibition 5827 with piperonyl butoxide counteracted the vacuolatio n observed in the Clara cells (Foster et al., 5828 1994, 1992). With repeated exposure to 4,000 ppm ( up to 13 weeks), the Clara cell vacuolation 5829 did not appear to progress to necrosis, and no hype rplasia of the bronchiolar epithelium was 5830 found. Foster et al. (1994, 1992) proposed that th e diminished severity or disappearance of Clara 5831 cell vacuolation with repeated exposure was due to the development of tolerance to 5832 dichloromethane, linked with a transient decrease o f CYP metabolism of dichloromethane. The 5833 available data suggests that CYP metabolism of dich loromethane may be involved in the mode 5834

of action for the acute effects of dichloromethane on the bronchiolar epithelium of mice. 5835 Mode of action research attention on lung effects f rom chronic exposure to 5836 dichloromethane has focused on neoplastic effect; n onneoplastic lung effects have received 5837 relatively little attention. No exposure-related i ncreased incidences of nonneoplastic lung lesions 5838 (including epithelial hyperplasia) were found in an y of the chronic studies listed in Table 4-36, 5839 but chronic inhalation exposure of B6C3F 1 mice to concentrations ≥ 2,000 ppm has consistently 5840 been shown to induce lung tumors in several studies (Kari et al., 1993; NTP, 1986). In a study 5841 that included interim sacrifices at 13, 26, 52, 68, 75, 78, 83, and 91 weeks of B6C3F 1 mice 5842 exposed to 2,000 ppm, hyperplasia of lung epitheliu m (the only nonneoplastic lung lesion found) 5843 was found in only three of the eight interim sacrif ices (68, 78, and 91 weeks) and was only 5844 statistically significantly elevated at 91 weeks (5 /30 versus 0/15 in controls) (Kari et al., 1993). 5845

5846

4.6.3.3. Mode of Action for Neurological Effects 5847 Results from studies of acutely exposed human subje cts indicate that mild 5848 neurobehavioral deficits may occur at air concentra tions >200 ppm with 4–8 hours of exposure 5849 (Bos et al., 2006; ACGIH, 2001; ATSDR, 2000; Cherry et al., 1983; Putz et al., 1979; Gamberale 5850 et al., 1975; Winneke, 1974). Acute high-dose expo sures also resulted in gross neurological 5851 impairments in several laboratory species (Haun et al., 1972, 1971; Heppel et al., 1944). 5852 Exposure of F344 rats to concentrations up to 2,000 ppm 6 hours/day for 13 weeks produced no 5853 effects on an observational battery, hind-limb grip strength, a battery of evoked potentials, or 5854 histology of the brain, spinal cord, or peripheral nerves (Mattsson et al., 1990). However, oral 5855192

exposures have been shown to alter autonomic, neuro muscular, and sensorimotor functions have 5856 been observed in F344 rats exposed to gavage doses ≥ 337 mg/kg-day for 14 days (Moser et al., 5857 1995). 5858 Dichloromethane may be metabolized by the CYP2E1 e nzyme to CO (Guengerich, 1997; 5859 Hashmi et al., 1994; Gargas et al., 1986). Many of the acute human exposure studies evaluated if 5860 CO was the primary metabolite responsible for produ cing the CNS depressant effects observed 5861 during dichloromethane exposure. Overall, at lower exposures and acute durations, it appears 5862 that CO is the primary mediator of the neurobehavio ral effects. Putz et al. (1979) demonstrated 5863 that similar neurobehavioral deficits were present when an equivalent COHb blood level (and 5864 CO exposure) was achieved between CO and dichlorome thane exposures. Incidentally, after a 5865 longer duration, neurobehavioral deficits are more pronounced with dichloromethane exposure in 5866 comparison to CO exposure alone. This additional i ncrease in the CNS depressive effects is 5867 most likely due to the saturation of the CYP2E1 met abolic pathway. In humans, saturation of the 5868 CYP2E1 metabolic pathway was seen at approximately 400–500 ppm after a 1-hour exposure 5869 (Ott et al., 1983e). CYP2E1 pathway saturation wit h dichloromethane has also been noted in 5870 hamsters (Burek et al., 1984) and in rats (McKenna et al, 1982; Nitschke et al., 1988a). It is 5871 highly probable that initially CYP2E1 is metabolizi ng dichloromethane to CO, which results in 5872 the neurological effects. However, at higher conce ntrations (greater than 500 ppm) and for 5873 longer durations (greater than 3 hours), the CYP2E1 pathway is most likely saturated. As a 5874 result, either the remaining dichloromethane could be metabolized by the GST pathway or the 5875 parent compound is producing the effects itself. 5876 Once the CYP2E1 enzyme is saturated, it is unknown whether dichloromethane or a 5877 GST-T1 pathway metabolite (e.g., formaldehyde) medi ates the resulting neurological effects. 5878

Based on the available literature on other solvents , such as toluene and perchloroethylene (for a 5879 review see Bowen et al. [2006]), it can be hypothes ized that once the CYP2E1 enzyme is 5880 saturated dichloromethane or a GST metabolite may i nteract directly with excitatory and 5881 inhibitory receptors, such as the NMDA, GABA, dopam ine, and serotonin receptors among other 5882 targets, to produce the resulting neurobehavioral e ffects. This hypothesis is supported by the 5883 evidence that changes in relation to dichloromethan e exposures in glutamate, GABA, dopamine, 5884 serotonin, acetylcholine, and other neurotransmitte rs are found in the brain (Kanada et al., 1994; 5885 Briving et al., 1986; Fuxe et al., 1984). Addition ally, several neurobehavioral effects, such as 5886 decreased spontaneous motor activity, deficits in l earning and memory, and deficits in FOB 5887 parameters are similar to other more characterized solvents such as toluene. However, more 5888 comprehensive studies specifically designed to dete rmine the mode of action for 5889 dichloromethane-induced impairment of neurological functions have not been conducted. 5890

5891

4.6.3.4. Mode of Action for Reproductive and Developmental E ffects 5892193

No significant effects on reproductive performance variables were found in a two5893 generation reproduction assay with F344 rats expose d to concentrations as high as 1,500 ppm 5894 (Nitschke et al., 1988b), and no effects were seen on most of the measures of reproductive 5895 performance examined in a study of male mice expose d to 200 ppm for 2 hours/day for 6 weeks 5896 before mating to nonexposed females (Raje et al., 1 988). In the mouse study, fertility index 5897 (number of females impregnated divided by total num ber of females mated times 100) was 5898 reduced in the 150 and 200 ppm groups (Raje et al., 1988), but the statistical significance of this 5899 effect varied considerably depending on the statist ical test used in the analysis. Mechanistic 5900

studies of dichloromethane or its metabolites that would provide mode of action information on 5901 reproductive effects in the male are not available. 5902 The mode of action for developmental effects can be hypothesized to involve the 5903 CYP2E1 pathway and, specifically, the production of CO. CO is a known developmental 5904

neurotoxicant. Demonstrated effects include neurob ehavioral deficits and neurochemical 5905 changes (Giustino et al., 1999; Cagiano et al., 199 8; De Salvia et al., 1995; Fechter, 1987). In 5906 addition, placental transfer of dichloromethane has been demonstrated with inhalation exposure 5907 (Withey and Karpinski, 1985; Anders and Sunram, 198 2). Pups exposed in utero to high 5908 concentrations of dichloromethane (4,500 ppm) demon strated neurobehavioral-related changes 5909 in comparison to air-exposed animals (Bornschein et al., 1980). This observed effect coupled 5910 with the known developmental neurotoxicological eff ects produced by CO suggests that the 5911 CYP2E1 metabolic pathway is involved in producing o bserved and suspected 5912 neurodevelopmental effects. In humans, CYP2E1 acti vity in the brain occurs earlier in gestation 5913 than it does in the liver, with activity in the bra in seen in the first trimester (Johnsrud et al., 20 03; 5914 Brzezinski et al., 1999). Thus, the direct effects of dichloromethane in fetal circulation, as well 5915

as the effects of CO and the effects of the CYP2E1- related metabolism in the fetal liver and the 5916 fetal brain, may be relevant to the risk of develop mental effects in humans. Mechanistic studies 5917 of dichloromethane or its metabolites that would pr ovide mode of action information on other 5918 noted developmental effects such as delayed ossific ation (Schwetz et al., 1975) are not available. 5919

5920

4.6.3.5. Mode of Action for Immunotoxicity 5921 Evidence of a localized immunosuppressive effect i n the lung resulting from inhalation 5922 dichloromethane exposure was seen in an acute expos ure (3 hours, 100 ppm) study in CD-1 mice 5923 (Aranyi et al., 1986). The lung infectivity assay used in this study examined response to 5924 bacterial challenges (i.e., risk of streptococcal-p neumonia-related mortality and clearance of 5925 Klebsiella bacteria). The innate immune response p lays an important role in limiting the initial 5926 lung burden of bacteria through the activity of mac rophages, neutrophils, and dendritic cells , and 5927 alveolar macrophages are particularly important in the response to respiratory infections 5928 (Marriott and Dockrell, 2007). The adaptive respon se develops from several days up to several 5929 weeks following infection, so that an effective imm une response in a lung infectivity assay 5930194

requires multiple immune mechanisms and in particul ar cooperation of macrophages, 5931 neutrophils, and T cells along with the appropriate cytokines (Selgrade and Gilmour, 2006). 5932 Although immunosuppression in the Streptococcal and Klebsiella infectivity models has been 5933 reported in the acute exposure scenarios tested in Aranyi et al. (1986), mechanistic studies of 5934 dichloromethane or its metabolites that would provi de mode of action information on the 5935 immune system cells or function have not been perfo rmed. 5936

5937

4.7. EVALUATION OF CARCINOGENICITY 5938 4.7.1. Summary of Overall Weight of Evidence 5939 Following U.S. EPA (2005a) Guidelines for Carcinogen Risk Assessment

, 5940 dichloromethane is “likely to be carcinogenic in hu mans” by the inhalation and oral routes of 5941 exposure, based predominantly on evidence of carcin ogenicity at two sites in 2-year bioassays in 5942

B6C3F 1 mice (liver and lung tumors with inhalation exposu re in both sexes, liver tumors with 5943 drinking water exposure in males only). In additio n, evidence of a trend for increased risk of 5944 liver tumors (described as neoplastic nodule or hep atocellular carcinoma) was seen in female 5945 F344 rats exposed via drinking water (p < 0.01) (Se rota et al., 1986a) or inhalation (p = 0.08) 5946 (NTP, 1986). However, the potential malignant char acterization of the nodules was not 5947 described, and no trend was seen in the data limite d to hepatocellular carcinomas. Additional 5948 evidence of the tumorigenic potential of dichlorome thane comes from the observation of an 5949 increase in benign mammary tumors following inhalat ion exposure (NTP, 1986; Burek et al., 5950 1986b; Nitschke et al. 1988a). An inhalation study (exposures of 0, 50, 200, and 500 ppm) also 5951 reported the presence of another relatively rare tu mor in rats, astrocytoma or glioma (mixed glial 5952 cell) tumors (Nitschke et al., 1988a). This collec tion of studies in the rat does not provide 5953 evidence for a carcinogenic response that is as str ong as that seen in the mouse. Taken together, 5954 however, the rat data provide supporting evidence o f carcinogenicity. Studies in humans found 5955 some evidence linking occupational exposure to dich loromethane and increased risk for some 5956 specific cancers, including brain cancer (Hearne an d Pifer, 1999; Tomenson et al., 1997; 5957 Heineman et al., 1994) and liver cancer (Lanes et a l., 1993, 1990). 5958 The proposed mode of action for dichloromethane-in duced liver tumors is through a 5959 mutagenic mode of carcinogenic action. Mode of act ion data indicate that dichloromethane5960 induced DNA damage in cancer target tissues of mice involves DNA-reactive metabolites 5961 produced via a metabolic pathway initially catalyze d by GST-T1. Evidence of mutagenicity 5962 includes in vitro bacterial and mammalian assays as well as in vivo mammalian system assays, 5963 although mutational events in critical genes (tumor suppressor genes, oncogenes) leading to 5964 tumor initiation and tumor promotion have not been established. This metabolic pathway has 5965 been found in human tissues, albeit at lower activi ties than in mouse tissues; therefore, the cancer 5966

results in animals are considered relevant to human s. 5967

5968195

4.7.2. Synthesis of Human, Animal, and Other Suppo rting Evidence 5969 Section 4.1.2 reviewed the results, strengths, and limitations of epidemiological research 5970 of dichloromethane and cancer, including cohort and case-control studies. The available 5971 epidemiologic studies provide some evidence of an a ssociation between dichloromethane and 5972 brain cancer and liver cancer, but the available da ta are limited. 5973 Two small cohort studies with relatively good expo sure metrics and relatively long 5974 follow-up periods (mean over 25 years) reported an increased risk of brain cancer, with SMRs of 5975 1.45 (95% CI 0.40–3.72) in Tomenson et al. (1997) a nd 2.2 (95% CI 0.79–4.69) in Cohort 1 of 5976 Hearne and Pifer (1999). Cohort 1 is an inception cohort, following workers from the beginning 5977 of their employment, which is methodologically more robust than Cohort 2, which only included 5978 workers who were working between 1964 and 1970. Th ese observations are supported by the 5979 data from a case-control study of brain cancer that reported relatively strong trends with 5980 increasing probability, duration, and intensity mea sures of exposure but not with a cumulative 5981 exposure measure. This difference could reflect a relatively more valid measure of relevant 5982 exposures in the brain from the intensity measure, as suggested by the study in rats reported by 5983 Savolainen et al. (1981) in which dichloromethane l evels in the brain were much higher with a 5984 higher intensity exposure scenario compared with a constant exposure period with an equivalent 5985 TWA (see section 3.2). A statistically significant increased incidence of brain or CNS tumors 5986 has not been observed in any of the animal cancer b ioassays, but a 2-year study using relatively 5987 low exposure levels (0, 50, 200, and 500 ppm) in Sp rague-Dawley rats observed a total of six 5988 astrocytoma or glioma (mixed glial cell) tumors in the exposed groups (in females, the incidence 5989 was 0, 0, 0, and 2 in the 0, 50, 200, and 500 ppm e xposure groups; in males, the incidence was 0, 5990 1, 2, and 1 in the 0, 50, 200, and 500 ppm exposure groups; sample size of each group was 5991 70 rats). These tumors are exceedingly rare in rat s, and there are few examples of statistically 5992 significant trends in animal bioassays (Sills et al ., 1999). These cancers were not seen in two 5993 other studies in rats, both involving higher doses (1,000–4,000 ppm) (NTP, 1986; Burek et al., 5994 1984), or in a high dose (2,000–4,000 ppm) study in mice (NTP, 1986). 5995 With respect to epidemiologic studies of liver and biliary duct cancer, the highest 5996 exposure cohort, based in the Rock Hill, South Caro lina, triacetate fiber production plant, 5997 suggested an increased risk of liver cancer, with a n SMR of 2.98 (95% CI 0.81, 7.63) in the latest 5998 study update (Lanes et al., 1993). This observatio n was based on four cases; an earlier analysis 5999 in this cohort reported an SMR of 5.75 (95% CI 1.82 , 13.8), based on these same four cases but 6000 with a shorter follow-up period (and thus a lower n umber of expected cases) (Lanes et al., 1990). 6001 No other cohort study has reported an increased ris k of liver cancer mortality, although it should 6002 be noted that there is no other inception cohort st udy of a population with exposure levels similar 6003 to those of the Rock Hill plant, and no data from a case-control study of liver cancer are 6004 available pertaining to dichloromethane exposure. 6005196

The primary limitation of all of the available dic hloromethane cohort studies is the 6006 limited statistical power for the estimation of eff ects relating to relatively rare cancers (such as 6007 brain cancer, liver cancer, and leukemia). Limitat ions with respect to studies of other cancers 6008 can also be noted. With respect to breast cancer, the only cohort that included a significant 6009 percentage of women had limited exposure informatio n (analysis was based on a dichotomous 6010 exposure variable) and co-exposure to other solvent s (Blair et al., 1998). The only breast cancer 6011 case-control study available used death certificate data to classify disease and occupational 6012 exposure (Cantor et al., 1995), which is likely to result in significant misclassification; exposure 6013 misclassification in particular would be expected t o result in an attenuated measure of 6014 association (Rothman and Greenland, 1998). No stud ies of adult leukemia and dichloromethane 6015 exposure and only one study of childhood leukemia ( acute lymphoblastic leukemia) in relation to 6016 maternal occupational dichloromethane exposure were found. Thus, EPA views the 6017 epidemiologic data pertaining to breast cancer and leukemia as inadequate to assess carcinogenic 6018 potential. 6019 In addition to the epidemiologic studies, several dichloromethane cancer bioassays in 6020 animals are available. In the only oral exposure c ancer bioassay involving lifetime exposure, 6021

increases in incidence of liver adenomas and carcin omas were observed in male (trend p- value = 6022 0.058) but not female B6C3F 1 mice exposed for 2 years (Table 4-38) (Serota et a l., 1986b; 6023 Hazelton Laboratories, 1983). The authors conclude d that these increases were “within the 6024 normal fluctuation of this type of tumor incidence, ” noting that there was no dose-related trend 6025 and that most of the individual group paired tests were not statistically significant after use of a 6026 Bonferroni correction factor. [The trend p- value and pairwise test p- values were not given in the 6027 Serota et al. (1986b) paper but can be found in the full report (Hazelton Laboratories, 1983)]. 6028 However, the trend p -value for these results is of borderline statistic al significance and it may not 6029 be reasonable to apply a correction for multiple co mparisons given the lack of independence of 6030 the groups and given a specific focus on the liver as a target organ. In Syrian golden hamsters 6031 exposed to 500, 1,500, or 3,500 ppm for 2 years, no statistically significantly increased 6032 incidences of tumors were found in any tissues (Bur ek et al., 1984). 6033

6034197 6035

Table 4-38. Incidence of liver tumors in male B6C3 F 1 mice exposed

to dichloromethane in a 2-year oral exposure (drink ing water) study a

Estimated mean intake (mg/kg-day) a Controls 0 61 124 177 234

Number of male mice b 125 c 200 100 99 125 Trend

p- value d

Number of cancers (%) Hepatocellular adenoma or carcinoma Mortality-adjusted percent e Mortality-adjusted p- value e 24 (19) (22) 51 (26)

(29) P = 0.071 30 (30)

(34)

p = 0.023 31 (31)

(35)

p = 0.019 35 (28)

(32)

p = 0.036 0.058

a Target doses were 60, 125, 185, and 250 mg/kg-day f rom the lowest dose group (excluding controls) to the highest dose group, respectively.

b No significant increases in females were found, but incidence data were not reported.

c Two control groups combined. The incidence in cont rol groups 1 and 2 were 20 and 23%, respectively.

d Cochran-Armitage trend test (source: Hazelton Labo ratories [1983]).

e Mortality-adjusted percent calculated, based on num ber at risk, using Kaplan-Meier estimation, taking into account mortality losses; p- value for comparison with control group, using asym ptotic normal test (source: Hazelton Laboratories [1983]).

Sources: Serota et al. (1986b); Hazelton Laborator ies (1983). 6036 6037 In a similar study in F344 rats (Serota et al., 198 6a), no increased incidence of liver 6038 tumors was seen in male rats, and the pattern in fe male rats was characterized by a jagged 6039 stepped pattern of increasing incidence of hepatoce llular carcinoma or neoplastic nodule 6040 (Table 4-39). Information was not provided which w ould allow characterization of the nodules 6041 as benign or malignant. Statistically significant increases in incidences were observed in the 50 6042 and 250 mg/kg-day groups (incidence rates of 0, 3, 10, 3, and 14%, respectively, for the 0, 5, 50, 6043 125, and 250 mg/kg-day groups) and in the group exp osed to 250 mg/kg-day for 78 weeks 6044 followed by a 26-week period of no exposure (incide nce rate 10%). A similar pattern, but with 6045 more sparse data, was seen for hepatocellular carci nomas, with 2 incidences in the 50 mg/kg-day 6046 and 2 in the 250 mg/kg-day groups. The authors co ncluded that dichloromethane exposure did 6047

not result in an increased incidence of liver tumor s, because the increase was based on a low rate 6048 (0%) in the controls and because of a lack of monot onicity. 6049198 6050

Table 4-39. Incidences of liver tumors in male and female F344 rats exposed

to dichloromethane in drinking water for 2 years

Target dose (mg/kg-day)

Controls 0 a 5 50 125 250 Trend

p- value b 250 with

recovery c

Males Estimated mean intake (mg/kg-day) 0 6 52 125 23 5 232

n per group d 76 34 38 35 41 15 Number (%) with neoplastic nodules 9 (12) 1 (3) 0 (0) 2 (6) 1 (2) Not reported 2 (13)

Number (%) with hepatocellular carcinoma 3 (4) 0 (0) 0 (0) 0 (0) 1 (2) Not reported 0 (0)

Number (%) with neoplastic nodules and hepatocellular carcinoma 12 (16) 1 (3) 0 (0) 2 (6) 2 (5) Not reported 2 (13)

Females Estimated mean intake (mg/kg-day) 0 6 58 136 26 3 239

n per group d 67 29 41 38 34 20 Number (%) with neoplastic nodules 0 (0) 1 (3) 2 (5) 1 (3) 3 (9) 2 (10) e Number (%) with hepatocellular carcinoma 0 (0) 0 (0) 2 (5) 0 (0) 2 (6) Not reported 0 (0)

Number (%) with neoplastic nodules and hepatocellular carcinoma 0 (0) 1 (3) 4 (10) e 1 (3) 5 (14) e p < 0.01 2 (10) e

a Two control groups combined.

b Cochran-Armitage trend test was used for trend test of liver foci/areas of alteration. For tumor mort alityunadjusted analyses, a Cochran-Armitage trend test was used, and, for tumor mortality-adjusted analyse s, a tumor prevalence analytic method by Dinse and Lagakos (19 82) was used. Similar results were seen in these t wo analyses.

c Recovery group was exposed for 78 weeks and then ha d a 26-week period without dichloromethane exposure ; n = 17 for neoplastic lesions.

d n available at terminal sacrifice; starting with 13 5 controls (combining both control groups) and 85 p er sex per dose group except recovery group (n = 25); subtract ed 5, 10, and 20 per group (except for recovery gro up) sacrificed at 25, 52, and 78 weeks, respectively, a nd subtracted unscheduled deaths, which ranged from 5 to 19 per

group.

e Significantly (

p < 0.05) different from controls with Fisher’s exac t test, mortality-unadjusted and mortalityadjusted analyses.

Source: Serota et al. (1986a). 6051 6052 Another oral (gavage) exposure study in Sprague-Da wley rats and in Swiss mice provides 6053 limited data concerning cancer incidence because th e study was terminated early (at 64 weeks) 6054 due to high treatment-related mortality (Maltoni et al., 1988). Exposure groups included controls 6055 (olive oil), 100, or 500 mg/kg-day 4–5 days/week. High-dose female rats showed an increased 6056 incidence of malignant mammary tumors, mainly adeno carcinomas (8, 6, and 18% in the control, 6057 100, and 500 mg/kg dose groups, respectively), but the increase was not statistically significant. 6058 Data were not provided to allow an analysis account ing for differing mortality rates. A dose6059 related increase, although not statistically signif icant, in pulmonary adenomas was observed in 6060199

male mice (5, 12, and 18% in control, 100, and 500 mg/kg-day groups, respectively). When 6061 mortality was taken into account, high-dose male mi ce that died in the period ranging from 52 to 6062 78 weeks were reported to show a statistically sign ificantly ( p < 0.05) elevated incidence for 6063 pulmonary tumors (1/14, 4/21, and 7/24 in control, 100, and 500 mg/kg-day groups, 6064 respectively). Details of this analysis were not p rovided. EPA applied a Fisher’s exact test to 6065 these incidences and determined a p -value of 0.11 for the comparison of the 500 mg/kg-day 6066 group (7/24) to the controls (1/14). 6067 As discussed in section 4.2, repeated inhalation e xposure to concentrations of 2,000 or 6068 4,000 ppm dichloromethane produced increased incide nces of lung and liver tumors in B6C3F 1 6069 mice (Mennear et al., 1988; NTP, 1986). The incide nce of mortality-adjusted liver tumors across 6070 dose groups (0, 2,000, and 4,000 ppm) increased fro m 48 to 67 and 93%, respectively, in male 6071 mice (trend p- value = 0.013) and from 10 to 48 and 100% female mi ce (trend p- values <0.001) 6072 (Table 4-40). For lung tumors, the mortality-adjus ted incidence was 12, 74, and 100% in males 6073 and 11, 83, and 100% in females in the 0, 2,000, an d 4,000 ppm groups, respectively (trend p6074 values <0.001). Elevated incidences of lung and li ver tumors in B6C3F 1 mice were observed 6075 with 52 weeks of exposure to 2,000 ppm, and lung tu mors were also elevated by week 104 in 6076 mice exposed for only 26 weeks to 2,000 ppm, follow ed by 78 weeks without exposure 6077 (Maronpot et al., 1995; Kari et al., 1993). 6078200 6079 6080 Table 4-40. Incidences of selected neoplastic lesi ons in B6C3F 1 mice exposed

to dichloromethane by inhalation (6 hours/day, 5 da ys/week) for 2 years

Exposure (ppm) a 0 (Controls) 2,000 4,000

Sex and neoplastic lesion n (%) b (%) c n (%) b (%) c n (%) b (%) c Trend

p- value d

Males Liver—hepatocellular adenoma or carcinoma 22 (44) (48) 24 (49) (67) 33 e (67) (93) 0.013

Lung—bronchoalveolar adenoma or carcinoma 5 (10) (12) 27 e (54) (74) 40 e (80) (100) <0.001

Females Liver— hepatocellular adenoma or carcinoma 3 (6) (10) 16 e (33) (48) 40 e (83) (100) <0.001

Lung—bronchoalveolar adenoma or carcinoma 3 (6) (11) 30 e (63) (83) 41 e (85) (100) <0.001

a 2,000 ppm = 6,947 mg/m 3 , 4,000 ppm = 13,894 mg/m 3

. b Total sample size was 50 per sex and dose group. P ercentages based on the number of tissues examined microscopically per group; for male mice, 49 livers were examined in the 2,000 and 4,000 ppm groups; f or female mice, 48 liver and lungs were examined. For compar ison, incidence in historical controls reported in NTP (1986) were 28% for male liver tumors, 31% for male lung t umors, 5% for female liver tumors, and 10% for fema le lung

tumors.

c Mortality-adjusted percentage.

d Life-table trend test, as reported by NTP (1986).

e Life-table test comparison dose group with control <0.05, as reported by NTP (1986).

Sources: Mennear et al. (1988); NTP (1986). 6081 6082 Liver tumors are relatively rare in F344 rats, and a moderate trend of increasing incidence 6083 of what was described as neoplastic nodules or hepa tocellular carcinoma was seen in the females 6084 (trend p -value = 0.08) but not the males in the NTP (1986) study (Table 4-41). As with the rat 6085 oral exposure study by Serota et al., 91986a), thes e nodules were not characterized as benign or 6086 malignant. There was no evidence of an increasing trend in incidence when hepatocellular 6087 carcinomas only were considered. 6088 Female F344 rats exposed by inhalation to 2,000 or 4,000 ppm showed significantly 6089 increased incidences of benign mammary tumors (aden omas or fibroadenomas) (Table 4-41); the 6090 number of benign mammary tumors per animal also inc reased with dichloromethane exposure in 6091 studies in Sprague-Dawley rats at levels of 50–500 ppm (Nitschke et al., 1988a) and 500–3,500 6092 ppm (Burek et al., 1984) (Table 4-42). Male rats i n two of these studies (Nitscke et al., 1988a ; 6093 NTP, 1986) also exhibited a low rate of sarcoma or fibrosarcoma in mammary gland or 6094 subcutaneous tissue around the mammary gland. 6095

6096201Table 4-41. Incidences of selected neoplastic lesions in F344/N rats exposed to dichloromethane by inhalation(6 hours/day, 5 days/week) for 2 yearsExposure (ppm)a0 (Controls) 1,000 2,000 4,000Sex and neoplastic lesion n(%)b(%)cn(%)b(%)cn(%)b(%)cn(%)b(%)cTrendp-valuedMalesLiver—Neoplastic nodule or hepatocellular carcinoma2(4)(10)3(6)(13)4(8)(19)1(2)(6)0.55Liver—hepatocellular carcinoma 2(4)(10)1(2)(4)2(4)(10)1(2)(6)nrLung—bronchoalveolar adenoma or carcinoma 11 (2)2 (4)1(2)Mammary glandAdenoma, adenocarcinoma, or carcinomaSubcutaneous tissue fibroma or sarcomaFibroadenomaMammary gland or subcutaneous tissue adenoma,fibroadenoma, fibroma, or sarcoma0101(0)(2)(0)(2)(6)(0)(6)0101(0)(2)(0)(2)(6)(0)(6)0224(0)(4)(4)(8)(9)(12)(21)1519d(2)(10)(2)(18)(23)(8)(49)0.008< 0.001<0.001FemalesLiver—neoplastic nodule or hepatocellular carcinoma2(4)(7)1(2)(2)4(8)(14)5(10)(20)0.08Liver—hepatocellular carcinoma 0 (0)(0)0 (0)(0)1(2)(4)0 (0)(0)nrLung—bronchoalveolar adenoma or carcinoma 1(2)1 (2)0(0)0(0)Mammary glandAdenocarcinoma or carcinomaAdenoma, adenocarcinoma, or carcinomaFibroadenomaMammary gland adenoma, fibroadenoma, oradenocarcinoma1156(2)(2)(10)(12)(16)(18)2211d13(4)(4)(22)(26)(41)(44)2213d14d(4)(4)(26)(28)(44)(45)0122d23e(0)(2)(44)(46)(79)(86)<0.001<0.001a1,000 ppm = 3,474 mg/m3, 2,000 ppm = 6,947 mg/m3, 4,000 ppm = 13,894 mg/m3.bTotal sample size was 50 per sex and dose group. Percentages based on the number of tissues examinedmicroscopically per group; for male rats, 49 liverswere examined in the 2,000 and 4,000 ppm groups; for females, only 48 liver and lungs and 49 mammary glands were microscopically examined in the 2,000and 4,000 ppm groups. For comparison, incidence inhistorical controls reported in NTP (1986) were 1%for female liver tumors and 16% for femalemammary fibroadenomas.cMortality-adjusted percentage.dLife-table trend test, as reported by NTP (1986).nr = not reported.eLife-table test comparison dose group with control<0.05, as reported by NTP (1986).Sources: Mennear et al. (1988); NTP (1986).60976098202

Table 4-42. Incidences of mammary gland tumors in two studies of male and female Sprague-Dawley rats exposed to dichlorometha ne by inhalation (6 hours/day, 5 days/week) for 2 years

Exposure (ppm) a

Study, lesion Controls 0 50 200 500 Late 500 b Early 500 b 1,500 3,500 Nitschke et al. (1988a) Males—n per group 57 65 59 64 c c c c Number (%) with Mammary gland tumors Adenocarcinoma or carcinoma Fibroadenoma Fibroma Fibrosarcoma Undifferentiated sarcoma Fibroma, fibrosarcoma, or undifferentiated sarcoma d 0 (0) 2 (4) 6 (11) 0 (0) 0 (0) 6 (11) 0 (0)

0 (0) 1 (6) 1 (6) 2 (4) 4 (6) 0 (0)

2 (3) 6 (11) 1 (6) 0 (0) 7 (12) 0 (0) 2 (3) 10 (16) 0 (0) 0 (0) 10 (16)

Females—n per group 69 69 69 69 25 25 c c Number (%) with mammary gland Mammary gland tumors Adenocarcinoma or carcinoma Adenoma Fibroadenoma Fibroma Fibrosarcoma 6 (9) 1 (1) 51 (74) 0 (0) 1 (1) 5 (7)

1 (1) 57 (83) 1 (1) 0 (0) 4 (6)

2 (3) 60 (87) 0 (0) 0 (0) 4 (6) 1 (1) 55 (80) 1 (1) 0 (0) 3 (12) 2 (8) 22 (88) 1 (4) 0 (0) 2 (8) 0 (0) 23 (92) 1 (1) 0 (0) Number with benign tumors e 52 (74) 58 (83) 61(87) f 55 (79) 23 (92) 23 (92)Number of benign tumors per tumorbearing rat e 2.0 2.3 2.2 2.7 2.2 2.6

Burek et al. (1984) Males n per group 92 c c 95 c c 96 97 Number (%) with benign tumors 7 (8) 3 (3) 7 (7) 14 (14) Total number of benign tumors 8 6 11 17 Number of tumors per tumor-bearing

rat g 1.1 2.0 1.6 1.2

Females n per group 96 c c 95 c c 96 97 Number (%) with benign tumors 79 (82) 81 (85) 80 (83) 83 (86) Total number of benign tumors 165 218 245 28 7 Number of tumors per tumor-bearing

rat f 2.1 2.7 3.1 3.5

a 50 ppm = 174 mg/m 3 , 200 ppm = 695 mg/m 3 , 500 ppm = 1,737 mg/m 3 , 1,500 ppm = 5,210 mg/m 3 , 3,500 ppm = 12,158 mg/m 3

. b Late 500 = no exposure for first 12 months followed by 500 ppm for last 12 months; early 500 = 500 ppm for first 12 months followed by no exposure for last 12 months.

c No data for this exposure level in this study.

d EPA summed across these tumor types, assuming no ov erlap.

e In historical controls, percent with benign tumors reported as 79–82% and number per tumor-bearing rat was 2.1.

f Significantly (

p ≤ 0.05) higher than control incidence by Fisher’s ex act test (Nitschke et al., 1988a).

g Calculated by EPA.

Sources: Nitschke et al. (1988a); Burek et al. (19 84b). 6099203

Supporting evidence for the carcinogenicity of dic hloromethane comes from the results 6100 of genotoxicity and mode of action studies discusse d in section 4.5. A mutagenic mode of 6101 carcinogenic action for dichloromethane involves me tabolic activation by GST, as evidenced by 6102 several observations, including the enhancement of dichloromethane mutagenic activity in 6103 normally unresponsive S. typhimurium strain TA1535 after it is transfected with the gen e for rat 6104 GST-T1 (DeMarini et al., 1997; Thier et al., 1993); increased HPRT gene mutations and DNA 6105 damage (DNA SSBs) in CHO cells when they are incuba ted with dichloromethane in the 6106 presence of mouse liver cytosol preparations rich i n GST enzymatic activities (Graves and 6107 Green, 1996; Graves et al., 1996, 1994b); the detec tion of DNA damage (DNA SSBs) in liver 6108 and lung tissue of B6C3F 1 mice immediately following 6-hour inhalation expos ure to 6109 dichloromethane (2,000–8,000 ppm); and a suppressio n of the DNA damage when mice were 6110 pretreated with buthionine sulphoximine, a GSH depl etor (Graves et al., 1995). 6111

Additional data from several studies indicate that dichloromethane genotoxicity is 6112 expressed in cancer target tissues in mice followin g in vivo exposure. Increased sister chromatid 6113 exchanges were observed in lung cells of B6C3F 1 mice after 90 days of inhalation exposure to 6114 2,000 ppm or 10 days of exposure to 4,000 or 8,000 ppm (Allen et al., 1990). DNA damage 6115 (comet assay) was detected in liver and lung tissue (but not stomach, kidney, brain, or bone 6116 marrow) 24 hours after oral administration of 1,720 mg/kg dichloromethane to CD-1 mice 6117 (Sasaki et al., 1998). DNA-protein cross-links wer e observed in the liver of B6C3F 1 mice but 6118 not hamsters, following inhalation exposure to conc entrations ranging from 500 to 4,000 ppm 6119 6 hours/day for 3 days (Casanova et al., 1996, 1992 ). Much less is known about genotoxicity in 6120 the liver in rats. Studies of single-strand DNA br eaks in rat hepatocytes or liver homogenate 6121 were negative, with inhalation exposures up to 5,00 0 ppm for 3 hours (Graves et al., 1995, 6122 1994b), but positive results were seen in a high-do se gavage study (1,275 mg/kg) (Kitchin and 6123 Brown, 1989). Few other specific types of genotoxi city endpoints (e.g., sister chromatid 6124 exchange, DNA-protein cross-links) have been studie d in the rat liver. 6125 Since there are limited data on mutagenic events f ollowing oral exposure, EPA conducted 6126 a pharmacokinetic analysis to evaluate how comparab le the internal doses to the liver in the oral 6127 bioassay (Serota et al., 1986b; Hazelton Laboratori es, 1983) were to the internal doses to the 6128 liver in the inhalation bioassay (Mennear et al., 1 988; NTP, 1986). The PBTK model of Marino 6129 et al. (2006) predicted that the average daily amou nt of dichloromethane metabolized via GST 6130 per liter of liver was about 14-fold lower in mice exposed to the highest dose of 244 mg/kg-day 6131 in the drinking water bioassay than in mice exposed to the lowest inhalation exposure of 2,000 6132 ppm, inducing liver tumors (Table 4-43). Thus, the lower incidence of liver tumors induced by 6133 oral doses of 244 mg/kg-day, compared with the high er incidence induced by inhalation 6134 exposure to 2,000 ppm, is consistent with the predi cted lower liver dose of GST metabolites (and 6135 hence lower probability of DNA modification) with o ral exposure. 6136

6137204 6138 Table 4-43. Comparison of internal dose metrics in inhalation and oral exposure scenarios, in male mic e and rats

Internal exposure in liver (mg metabolized through GST pathway/L liver tissue/day) a

Male

External dose Mouse Rat Inhalation (ppm)

2,000 2,364 1,502 4,000 4,972 3,111

Oral (mg/kg-day) b

61 17.5 77.0

124 63.3 233.5

174 112.0 385.4

234 169.5 589.8

a Mouse values derived by EPA from the PBTK model of Marino et al. (2006); rat values derived from EPA based on the modified PBTK model o f Andersen et al. (1991) (see Appendix C for model details).

b Actual doses administered to mice (Serota et al., 1986a); BWs not given for males and females, so simulation results only provided for on e gender. 6139 6140

6141 4.7.3. Mode of Action Information 6142

4.7.3.1. Hypothesized Mode of Action 6143 The hypothesized mode of action for dichloromethane-induced tumors is through a 6144 mutagenic mode of carcinogenic action. Specificall y, the data indicate that dichloromethane is 6145 metabolized by GST to reactive metabolites that ind uce mutations in DNA leading to 6146 carcinogenicity. Much of the experimental mode of action research has focused on the liver and 6147 lung, the sites of tumor formation in chronic bioas says (Mennear et al., 1988; NTP, 1986; Serota 6148 et al., 1986b, Hazleton Laboratories, 1983). The m ode of action is potentially relevant to other 6149 sites, particularly those in which GST-T1 is expres sed, such as mammary tissue (Lehmann and 6150 Wagner, 2008) and the brain (Juronen et al., 1996). 6151 Support for the importance of GST in the hypothesi zed mutagenic mode of action has 6152

been demonstrated in in vitro bacterial and mammali an assays as well as in vivo mammalian 6153 system assays. Dichloromethane is consistently mut agenic in S. typhimurium strains with GST 6154 capability, but did not produce mutagenic effects i n non-GST S. typhimurium strains 6155 (summarized in Section 4.5.1.1 and Table 4-29). I n vitro mammalian cell studies (see Table 4- 6156 30) have consistently demonstrated genotoxic effect s in CHO cell lines when a mouse liver 6157 cytosol fraction was exogenously added and in mouse Clara cells; positive responses were seen 6158 in studies measuring DNA-protein crosslinks, HPRT m utation analysis, and DNA SSBs. Other 6159205

studies have demonstrated DNA adducts with dichloro methane exposure in calf thymus DNA in 6160 the presence of bacterial GST DM11. Negative result s were seen in most of the other in vitro cell 6161 studies using rat hepatocytes or CHO cells without mouse liver cytosol incubation. These 6162

studies were conducted in cell lines where GST acti vity is considerably lower than in mouse cell 6163 lines and therefore these results are not unexpecte d. 6164

In studies with human cell lines or isolated cells , positive results were reported for sister 6165 chromatid exchanges, chromosomal aberrations, and i n the micronucleus test. In vivo studies in 6166 mice (Section 4.5.1.2 and Table 4-32) consistently showed genotoxic effects following 6167 dichloromethane exposure in the liver and lung, whe re tumors are observed. Other organs in the 6168 mouse were evaluated and mutagenic changes were not consistenly observed. The specificity of 6169 the observed effects support the hypothesized mode of action since these positive mutagenic 6170 responses are seen in organs where tumor formation occurs (i.e., liver and lung) rather than in 6171

areas that were not the site of tumors in the mouse bioassays (e.g., stomach, bladder, kidney). In 6172 vivo genotoxicity studies in rats and hamsters (the other test systems used, see Table 4-33) were 6173 predominantly non-positive. However, rats and hams ters have considerably lower GST activity 6174 than the mouse and may be less sensitive to dichlor omethane-induced genotoxic effects. 6175 In vivo binding of S-(chloromethyl)glutathione, di chloromethane’s reactive GST 6176 metabolite,to DNA was not demonstrated in one study in rats and mice using a relatively low 6177 dose (5 mg/kg). The reactivity of the postulated D NA-reactive species and the instability of the 6178 derived adducts presents considerable challenges to the ability to provide direct evidence of 6179

adduct formation. Thus this lack of in vivo eviden ce of S-(chloromethyl)glutathione binding to 6180 DNA does not in itself represent a basis for invali dating the proposed mode of action. 6181

6182

4.7.3.1.1. Experimental support for the hypothesized mode of action 6183 Strength, consistency, and specificity of associati on 6184 It is hypothesized that mutagenic events lead to t he development of liver and lung tumors 6185 following dichloromethane exposure. Several observa tions from experimental studies support the 6186 mutagenicity of dichloromethane and the key role of GST metabolism and the formation of 6187 DNA-reactive GST-pathway metabolites. The GST path way produces two metabolites of 6188 dichloromethane, S-(chloromethyl)glutathione and fo rmaldehyde, which are potentially reactive 6189

with DNA and other cell macromolecules. Enhanced d ichloromethane genotoxicity in bacterial 6190 and mammalian in vitro assays with the introduction of GST metabolic capacity provides support 6191 that GST metabolism and metabolites are involved (D eMarini et al., 1997; Graves and Green, 6192 1996; Graves et al., 1996, 1995, 1994b; Thier et al ., 1993). 6193 In bacterial strains where GST activity was not pr esent (e.g., TA1535, TA1538), 6194 mutagenic effects were not reported following dichl oromethane exposure (Gocke et al., 1981; 6195 Osterman-Golkar et al., 1983; Simula et al., 1993; Oda et al., 1996). Further tests of GST6196 dependent mutagenicity were evaluated by transfecti ng GST into non-GST bacterial strains or 6197206

decreasing GST activity in GST bacterial strains (e .g. TA100). When GSTT1-1 was cloned into 6198 bacterial strain TA1535, dichloromethane treatment resulted in reverse mutations in this new 6199

GST + TA1535 strain and these mutations were independent of rat S9 metabolic activation (Their 6200 et al., 1993; Pegram et al., 1997; DeMarini et al., 1997). Similarly, TA100/NG-11, a bacterial 6201 strain with decreased GST activity in comparison to the wild-type TA100 strain, showed 6202 significantly decreased mutagenicity (reverse mutat ions) following dichloromethane treatment 6203 (Graves et al., 1994a). 6204 In vitro mammalian genotoxicity studies also suppo rt the importance of the GST pathway 6205 in relation to the positive effects observed follow ing dichloromethane exposure. Positive results 6206 in the in vitro assays were limited to experiments with the presence of GST in the cell system. 6207 When mouse liver cytosol was added to hamster cell lines, dichloromethane induced DNA6208 protein crosslinks, DNA SSBs, and HPRT gene mutatio ns (Graves et al., 1994b; Graves et al., 6209 1996; Graves and Green, 1996). Additionally, in mo use Clara cells (GST is localized in the lung 6210 cells of mice), DNA SSBs were reported following di chloromethane treatment and the extent of 6211 DNA damage was significantly decreased when the cel ls were pretreated with a glutathione 6212 depletor (Graves et al., 1995). Other studies eval uating similar genotoxic endpoints in rat or 6213

CHO cells without modification of the low GST activ ity in the test system generally reported no 6214 evidence of genotoxic events (Graves et al., 1995; Andrae and Wolff, 1983; Jongen et al., 1981; 6215 Garrett and Lewis, 1983; Thilagar and Kumaroo, 1983 ). A study evaluating the genotoxic 6216 effects of dichloromethane (up to 6 mM) in freshly isolated mouse, rat, hamster, and human 6217 hepatocytes provides additional supporting evidence of the influence of GST activity on 6218 mutagenicity (Casanova et al., 1997). Positive res ults were only observed in hepatocytes from 6219

B6C3F 1 mice; the interspecies variability in effects corr elated proportionally with the enhanced 6220 GST metabolic capacity in mice (Reitz et al., 1989) . In studies with human cell lines or isolated 6221 cells, positive results were reported for sister ch romatid exchanges, chromosomal aberrations, 6222 DNA damage, and in the micronucleus test. Negative results were obtained with human cells in 6223 unscheduled DNA synthesis assays (Perocco and Prodi , 1981; Jongen et al., 1981) and 6224 dichloromethane was not demonstrated to be genotoxi c in studies of human hepatocytes (Graves 6225 et al., 1995; Casanova et al., 1997). 6226 Two of three in vivo genotoxicity studies in insec ts reported positive results. 6227 Genotoxicity was observed in Drosophila for the gen e mutation assay (Gocke et al., 1981) and 6228 the somatic assay (Rodriguez-Arnaiz, 1998) when dic hloromethane was administered through 6229 the food. When Drosophila were exposed to dichloro methane via inhalation, genotoxic effects 6230 were negative as measured through gene mutation ass ays (sex-linked recessive lethal, somatic 6231 mutation and recombination) (Kramers et al., 1991). 6232 In vivo genotoxicity studies reported DNA-protein cross links, DNA SSBs, chromosomal 6233 aberrations, and sister chromatid exchanges in live r cells of B6C3F 1 mice following acute 6234 inhalation exposure to concentrations producing liv er tumors with chronic exposure (Casanova et 6235207

al., 1996, 1992; Graves et al., 1995, 1994b). The formation of DNA SSBs was suppressed when 6236 the mice were pretreated with a GSH depletor (Grave s et al., 1995), providing additional support 6237

for the involvement of GST metabolism. Increased s ister chromatid exchanges and 6238 chromosomal aberrations were found in the lungs in mice exposed to dichloromethane for 6239 2 weeks to 8,000 ppm or for 12 weeks to 2,000 ppm. In this study, however, there was evidence 6240 of damage at other sites, too: sister chromatid exc hanges were also seen in peripheral 6241 lymphocytes, chromosomal aberrations were seen in b one marrow and micronuclei were seen in 6242 in peripheral red blood cells under the same exposu re protocol (Allen et al., 1990). As was seen 6243 in the liver, DNA SSBs were seen in lungs of B6C3F 1 mice following acute inhalation exposure 6244 to concentrations producing lung tumors with chroni c exposure, and this effect was suppressed 6245 with pretreatment with a GSH depletor, buthionine s ulfoximine (Graves et al., 1995). Other 6246 studies of sister chromatid exchange (Allen et al., 1990) or DNA damage detected by the comet 6247 assay (Sasaki et al., 1998) also provide evidence o f genotoxic effects specifically in lung cells in 6248 mice. These in vivo mammaliam genotoxicity studies demonstrate site-specific effects 6249 correlating to the dichloromethane-induced tumors i n animals. Additional evidence for site 6250 specificity comes from a study in which DNA damage (detected by the comet assay) was 6251 enhanced in liver tissue, but not stomach, kidney, brain, or bone marrow, 24 hours after oral 6252 administration of 1,720 mg/kg dichloromethane to CD -1 mice (Sasaki et al., 1998). 6253 DNA reaction products (e.g., DNA adducts) produced by GST metabolites, such as S6254 (chloromethyl)glutathione, have not been identified in in vivo studies (Watanabet et al., 2007). 6255 The authors speculated that these results are due t o the instability of the reaction products 6256 (Hashmi et al., 1994). However, adducts with nuc leosides have been observed in in vitro 6257 studies, when DNA was treated with S-(1-acetoxymeth yl)glutathione, a compound structurally 6258 similar to S-(chloromethyl)glutathione and in calf thymus DNA in the presence of 6259 dichloromethane, but not formaldehyde, and GST (Mar sch et al., 2004; Kayser and Vuilleumier, 6260 2001). These findings indicate that the S-(chlorom ethyl)glutathione intermediate formed by 6261 GSH conjugation has mutagenic potential and is like ly responsible, at least in part, for the 6262 mutagenic response observed following dichlorometha ne exposure. However, other studies (Hu 6263 et al., 2006; Casanova et al., 1996) provide eviden ce of formaldehyde-related DNA-protein 6264 cross-links in relation to dichloromethane exposure . These results show that, while most studies 6265 indicate the importance of the S-(chloromethyl)glut athione intermediate in mediating genotoxic 6266 damage following dichloromethane exposure, DNA dama ge resulting from formaldehyde 6267

formation should also be considered. 6268 Mutagenic data in critical genes leading to the in itiation of dichloromethane-induced liver 6269 or lung tumors are not available. In vivo assays e valuating mutations in tumor suppressor genes 6270 and oncogenes reported similar frequencies of activ ated H- ras genes and inactivation of the 6271

tumor suppressor genes, p53 and Rb-1 in the liver tumors seen in the nonexposed and 6272 dichloromethane-exposed B6C3F 1 mice (Devereaux et al., 1993; Hegi et al., 1993). There were 6273208

too few lung tumors (n=4) in controls to provide a conclusive comparison of mutation patterns 6274 between exposed and nonexposed tumors. 6275

6276 Dose-response concordance 6277 Statistically significant increases in liver tumor incidences in male and female (2,000 and 6278 4,000 ppm) mice were observed in the inhalation bio assay in B6C3F 1 mice (NTP, 1986). 6279 Several studies provide evidence of an association between mutagenic events mediated by GST6280 pathway metabolites with the exposure levels induci ng liver tumors in B6C3F 1 in this study, and 6281 concentration dependent increases in genotoxicity h ave been observed in in vitro and in vivo 6282

assays. 6283 In vitro mammalian genotoxicity studies were posit ive and demonstrated a dose-response 6284 relationship for DNA protein crosslinks, DNA single stranded breaks, and DNA damage as 6285 measured by the comet assay at concentrations rangi ng from 2.5 to 60 mM when mouse liver 6286 cytosol was added or if mouse GSTT1-1 was transfect ed into hamster cell lines (Graves et al., 6287 1994b; Graves et al., 1996; Hu et al., 2006). In m ouse hepatocytes, DNA protein crosslinks were 6288 observed following dichloromethane exposures rangin g between 0.5 – 6.0 mM (Casanova et al., 6289 1997). DNA-protein cross-links were detected in m ouse hepatocytes incubated with 1.9 mM 6290 dichloromethane (Casanova et al., 1997), a concentr ation chosen based on its correspondence to 6291

the TWA liver concentration of dichloromethane that was predicted by the Andersen et al. 6292 (1987) PBTK model for mice exposed by inhalation to 4,000 ppm for 6 hours (a dose that 6293

resulted in increased liver tumor incidence in the two-year bioassay reported by NTP, 1986). 6294

Consistent with the relative lack of liver tumor re sponses in Syrian golden hamsters (Burek et al., 6295 1984) and F344 rats (NTP, 1986) with chronic exposu re to 3,500 or 4,000 ppm, hepatocytes 6296

from these strains of animals did not form detectab le DNA-protein cross-links when incubated 6297 with 1.9 mM dichloromethane (Casanova et al., 1997)

. ................................................ .................. 6298 DNA-protein cross-links were not detected in livers of mice exposed to 146 ppm 6299 6 hours/day for 3 days, but a concentration-depende nt increase in DNA-protein cross-links was 6300 observed in DNA from livers of mice exposed to seve ral concentrations between 500 and 6301 4,000 ppm (Casanova et al., 1996). Following expos ure under similar conditions (concentrations 6302 of 498, 1,553, or 3,923 ppm), DNA-protein cross-lin ks were not detected in the livers of Syrian 6303 golden hamsters, a species that did not develop tum ors after chronic inhalation exposure to 6304 dichloromethane (Casanova et al., 1996, 1992). Inc reased DNA SSBs were detected in liver 6305

tissue of B6C3F 1 mice immediately following a 6-hour inhalation exp osure to dichloromethane 6306 at concentrations ranging from 2,000 to 8,000 ppm ( Graves et al., 1995), and in mouse 6307 hepatocytes after a 3-hour exposure to 4000 (but no t 2000) ppm (Graves et al., 1994b). 6308 Statistically significant increases in the inciden ce of lung tumors were observed in the 6309 inhalation chronic bioassay in male and female B6C3 F 1 mice exposed to 2,000 or 4,000 ppm 6310 dichloromethane (Mennear et al., 1988; NTP, 1986). Evidence of mutagenicity at these exposure 6311209

levels comes from two inhalation studies (Graves et al., 1995; Allen et al., 1990). Increased 6312 DNA SSBs were detected in lung tissue of B6C3F 1 mice immediately following a 6-hour 6313 inhalation exposure to dichloromethane at concentra tions ranging from 2,000 to 8,000 ppm 6314 (Graves et al., 1995). In the study by Allen et al . (1990), increased presence of sister chromatid 6315 exchanges was observed in mouse lung cells followin g a 12-week exposures at 2000 ppm; 6316 shorter durations of exposure (2 weeks) were positi ve for measures of sister chromatid exchange 6317 and chromosome aberrations at 8000 ppm, but not at 2000 or 4000 ppm. 6318

DNA adducts were observed and increased with dose i n an in vitro preparation of calf 6319 thymus DNA when treated with dichloromethane (5 – 6 0 mM) and bacterial, rat, or human GST 6320 (Marsch et al., 2004). 6321

6322 Temporal relationship 6323 Dichloromethane-induced liver and lung tumors first appeared in mice after 52 weeks of 6324 exposure (Maronpot et al., 1995; Kari et al., 1993) . The detection of DNA-protein cross-links in 6325

the livers of B6C3F 1 mice following short-term inhalation exposures to dichloromethane 6326 concentrations that induced tumors with chronic exp osure (Casanova et al., 1996, 1992) provides 6327 temporal support for the proposed mutagenic mode of action. Additional supporting evidence 6328

comes from observations that increased levels of DN A SSBs were detected in the liver and lungs 6329

of B6C3F 1 mice immediately following 3-hour inhalation expos ure to 2,000–8,000 ppm 6330 dichloromethane (Graves et al., 1995; 1994b). Sin gle dose and inhalation exposure studies of 6 6331 hours or less did not detect an effect on DNA synth esis (Lefevre and Ashby, 1989) or 6332 unscheduled DNA synthesis (Trueman and Ashby, 1987) in mouse liver cells. 6333

6334 Biological plausibility and coherence 6335 Bioactivation of a parent compound into a mutagenic metabolite resulting in cancer is a 6336 plausible mode of action carcinogenicity in humans and is a generally accepted mode of action. 6337 Dichloromethane-induced carcinogenicity is hypothes ized to be due to metabolism of the parent 6338 compound by the GST pathway (GST-T1) to a metabolit e that is tumorigenic. The GST 6339 metabolite, S-(chloromethyl)glutathione, formed fro m dichloromethane, has been characterized 6340 as labile and highly reactive through in vitro evaluation of dichloromethane metabolism in 6341 hepatocytes using 13 C-NMR techniques (Hashmi et al., 1994) and through an enzyme digestion 6342 assay using calf thymus DNA and GST-T1 enzyme (Mars ch et al., 2004). The hypothesis that 6343 the formation of a mutagenic metabolite is a prelim inary step resulting in carcinogenicity is 6344 based on evidence that malignant tumors are primari ly located in areas where dichloromethane is 6345 highly metabolized by GST-T1, such as the liver and the lung, and on mutagenicity studies 6346 indicating the importance of the GST pathway and th at the lung and liver are more prone to 6347 mutagenic effects of dichloromethane (Sasaki et al ., 1998; Casanova et al., 1996, 1992; Graves 6348 et al., 1995, 1994b). The site selectivity of the mutagenicity in liver and lung tissue as evidenced 6349210

by several studies suggests that the GST reactive m etabolite remains in the tissue where it is 6350 formed. Collectively, the studies support the hypo thesis that dichloromethane-mediated 6351 carcinogenicity results from a GST metabolite that produces selective DNA damage in the 6352 tissues where the metabolite is formed, but this hy pothesis is based, in part, on assumptions 6353 regarding metabolite clearance and reactivity. DNA damage in the liver and lung as well as the 6354 increased incidence of tumor formation resulting fr om dichloromethane exposure indicates 6355 coherence of the mutagenic and carcinogenic effects and is evidence supporting a mutagenic 6356

mode of action. 6357 Differences in GST activity in mice compared with o ther species, and the interspecies 6358 variability in genotoxic effects corresponding to i nterspecies variability in tumor response 6359 support the mode of action hypothesis. DNA SSBs we re not detected in liver or lung cells in rats 6360 exposed to similar inhalation exposures that induce strand breaks in mice (Graves et al., 1995; 6361 Graves et al., 1994b), and were detected at much lo wer in vitro concentrations in isolated 6362 hepatocytes from B6C3F 1 mice (0.4 mM) than in hepatocytes from Alpk:APfSD rats (30 mM) 6363 (Graves et al., 1995, figure 3). The difference in susceptibility to carcinogenic response between 6364 mice and rats likely reflects differences in GST me tabolism. Toxicokinetic studies indicate that, 6365 with increasing exposure levels, increasing amounts of dichloromethane are metabolized via 6366

GST metabolism. 6367

6368

4.7.3.1.2. Other possible modes of action for liver or lung tu mors in rodents. 6369 Data are not available to support other possible mo des of action for the liver and lung 6370

tumors in rodents. Efforts to observe sustained ce ll proliferation in liver following 6371 dichloromethane exposure of B6C3F 1 mice have been unsuccessful. Groups of female B6C 3F 1 6372 mice were exposed to 0 or 2,000 ppm dichloromethane 6 hours/day, 5 days/week for up to 78 6373 weeks did not exhibit enhanced cell proliferation i n the liver when assessed at various intervals 6374 during exposure (Foley et al., 1993). . 6375 Indices of enhanced cell proliferation have been me asured in the lungs of male B6C3F 1 6376 mice following acute duration exposure at concentra tions of about 1,500, 2,500, or 4,000 ppm 6377 dichloromethane (6 hours/day for 2 days) but not at exposure concentrations of 150 or 500 ppm 6378 and not in lungs of Syrian golden hamsters exposed to concentrations up to 4,000 ppm 6379 (Casanova et al., 1996). Earlier studies showed so mewhat consistent findings in that the 6380 numbers of bronchiolar cells undergoing DNA synthes is (thymidine incorporation labeling) were 6381 markedly increased (about 6- to 15-fold) in bronchi olar cells of B6C3F 1 mice exposed to 6382 4,000 ppm dichloromethane 6 hours/day on days 5, 8, and 9 of exposure, but no evidence for 6383 increased cell proliferation was found after 89, 92 , or 93 days of exposure (Foster et al., 1992). 6384 The results suggest that enhanced cell proliferatio n is not sustained in the lung with longer-term 6385 exposure to dichloromethane concentrations associat ed with lung tumor development in mice, 6386211

and that this mode of tumor promotion is not import ant in the development of dichloromethane6387 induced lung tumors. 6388

6389

4.7.3.2. General Conclusions About the Mode of Action for Tu mors in Rodents and 6390

Relevance to Humans 6391 The mode of action for dichloromethane is hypothes ized to involve mutagenicity via 6392

reactive metabolites. Mechanistic evidence indicat es that dichloromethane-induced DNA 6393 damage in cancer target tissues of mice involves DN A-reactive metabolites produced via a 6394 metabolic pathway initially catalyzed by GST. Alth ough mutational events in critical genes 6395 leading to tumor initiation have not been establish ed, evidence supporting a mutagenic mode of 6396 action includes the identification of mutagenic res ponse (reverse mutations) in short-term 6397 bacterial assays (with microsomal activation) and i nduced DNA-protein cross-links and DNA 6398 SSBs in mammalian cell assays. There are numerous positive in vivo genotoxicity studies 6399 specifically examining responses in the liver and/o r lung; these studies included evidence of 6400 chromosomal aberrations, SSBs and sister chromatid exchanges, and DNA-protein cross-links. 6401 The negative in vivo genotoxicity assays are genera lly those that were based on a micronucleus 6402 test using mouse bone marrow, which is expected as halogenated hydrocarbons (such as 6403 dichloromethane) are not very effective in this typ e of assay (Crebelli et al., 1999; Dearfield and 6404 Moore, 2005). 6405

6406 Is the hypothesized mode of action sufficiently sup ported in test animals ? 6407 Consistent and specific evidence for the associatio n between the formation of DNA6408 reactive GST-pathway metabolites and the formation of liver and lung tumors from inhalation 6409 includes (1) enhanced GST metabolic capacity in the liver and lung and enhanced localization of 6410 GST-T1 in hepatic cell nuclei in B6C3F 1 mice, compared with rats and hamsters, which do no t 6411 show strong tumor responses to chronic inhalation e xposure; (2) the detection of DNA-protein 6412 cross-links, or DNA SSBs in livers and lungs of B6C 3F 1 mice following acute inhalation 6413 exposure to concentrations that produce tumors with chronic exposure; (3) suppression of the 6414 formation of DNA SSBs in livers and lungs of B6C3F 1 mice pretreated with a GSH depletor; (4) 6415 the inability to detect DNA-protein cross-links or DNA SSBs in livers or lungs of similarly 6416 exposed rats or hamsters (5) detection of DNA SSBs at much lower in vitro concentrations in 6417 isolated hepatocytes from B6C3F 1 mice than in hepatocytes from Alpk:APfSD rats; (6) dose6418 response concordance and a temporal relationship fo r the formation of DNA-protein cross-links 6419 and DNA SSBs with the formation of liver and lung t umors in B6C3F 1 mice exposed to 6420 dichloromethane; (7) the detection of increased sis ter chromatid exchanges in lung cells from 6421 CD-1 mice exposed by inhalation to dichloromethane; and (8) enhancement of dichloromethane 6422 genotoxicity in bacterial and mammalian in vitro as says with the introduction of GST metabolic 6423212

capacity. However, mutations in critical genes lin ked to initiation of tumor cells have not been 6424

identified. 6425 The much weaker carcinogenic response in the liver of rats and mice to chronic drinking 6426 water exposure (Serota et al., 1986a, b) than that noted in mice exposed by inhalation (Kari et al., 6427 1993; NTP, 1986) is correlated with much smaller am ounts of GST metabolites produced in the 6428 liver under the exposure conditions of the oral bio assay than in the inhalation bioassay (Andersen 6429 et al., 1987). 6430

6431 In conclusion, there is sufficient evidence support ing a mutagenic mode of action and to 6432

establish the involvement of GST metabolism in the lung and liver carcinogenicity of 6433

dichloromethane in mice. 6434

6435 Is the hypothesized mode of action relevant to huma ns ? 6436 The postulated mode of action, that dichloromethan e is metabolized by GST to reactive 6437 metabolites that induce mutations in DNA leading to carcinogenicity is possible in humans. 6438 Mutagenicity as a mode of action for carcinogenicit y in humans is generally accepted and is a 6439 biologically plausible mechanism for tumor inductio n. The toxicokinetic and toxicodynamic 6440 processes that would enable reactive metabolites to produce mutations in animal models are 6441 biologically plausible in humans. Furthermore, the detection of the GST pathway in human 6442 tissues indicates that the hypothesized mode of act ion involving reactive metabolites from this 6443 pathway, S-(chloromethyl)glutathione and formaldehy de, is relevant to humans. 6444 Another factor that may play a role in the apparen t species differences in carcinogenicity 6445 resulting from dichloromethane exposure is species differences in intracellular localization of 6446 GST-T1 (Sherratt et al., 2002; Mainwaring et al., 1 996). In mouse liver tissue, GST-T1 appears 6447 to be localized in the nuclei of hepatocytes and bi le-duct epithelium, but rat liver does not show 6448 preferential nuclear localization of GST-T1. In hu man liver tissue, some hepatocytes show 6449

nuclear localization of GST-T1 and others show loca lization in cytoplasm. Nuclear production 6450 of S-(chloromethyl)glutathione catalyzed by GST-T1 in the nucleus is more likely than 6451 cytoplasmic production to lead to DNA alkylation. The finding of some nuclear localization of 6452 GST-T1 in human liver tissue supports the relevance of the hypothesized mode of action to 6453

humans. 6454 Comparisons in mice, rats, humans, and hamsters of GST enzyme activity in liver and 6455 lung tissues have indicated the following rank orde r: mice > rats > or ≈ humans > hamsters 6456 (Reitz et al., 1989; Thier et al., 1998). This rel ative ranking in GST activity corresponds to the 6457 rank order of the strength of the association betwe en inhalation exposure to dichloromethane and 6458 liver tumors in long-term cancer bioassays with mic e, rats, and hamsters. This relative ranking 6459 does not preclude the relevance of the hypothesized mode of action to humans, however. 6460

6461213

Which populations or lifestages can be particularly susceptible to the hypothesized mode of 6462

action ? 6463 As discussed in section 3.3, a polymorphism of the GST-T1 gene is present in humans. 6464 People with two functional copies of the gene (+/+) readily conjugate GSH to dichloromethane. 6465 Individuals having only one working copy of the gen e (+/–) display relatively decreased 6466 conjugation ability. Individuals with no functiona l copy of the gene (–/–) do not express active 6467 GST-T1 protein and do not metabolize dichloromethan e via a GST-related pathway (Thier et al., 6468 1998). Thus the GST-T1 +/+ (wild-type) genotype would be considered to be the more “at risk” 6469 population; this subgroup represents approximately 30% of the U.S. population (Haber et al., 6470 2002) but would be expected to be more common among Caucasians and African-Americans 6471 than among Asians (Raimondi et al., 2006; Garte et al., 2001; Nelson et al., 1995) (see Table 6472 3-3). 6473 According to the Supplemental Guidance for Assessing Susceptibility from Early-Life 6474 Exposure to Carcinogens (U.S. EPA, 2005b), children exposed to carcinogens with a mutagenic 6475 mode of action are assumed to have increased early- life susceptibility. The Supplemental 6476

Guidance (U.S. EPA, 2005b) recommends the application of ag e-dependent adjustment factors 6477 (ADAFs) for carcinogens that act through a mutageni c mode of action. Although the database is 6478 lacking in vivo evidence of specific mutagenic even ts following chronic exposure to 6479 dichloromethane, the weight of the available eviden ce indicates that dichloromethane is acting 6480 through a mutagenic mode of carcinogenic action. A pplication of ADAFs is recommended for 6481 both the oral and inhalation routes of exposure whe n risks are assessed that are associated with 6482 early-life exposure. 6483

6484

4.8. SUSCEPTIBLE POPULATIONS AND LIFE STAGES 6485 4.8.1. Possible Childhood Susceptibility 6486 In humans, hepatic CYP2E1 begins to be expressed in the second trimester (Johnsrud et 6487 al., 2003), increases significantly in the third tr imester, and continues to increase during the first 6488 year of life (Hines, 2007; Johnsrud et al., 2003; T reluyer et al., 1996; Vieira et al., 1996). In the 6489 fetal brain, however, CYP2E1 activity is seen as ea rly as GD 50, with increasing levels seen until 6490 at least the end of the first trimester (Brzezinski et al., 1999). Neurobehavioral effects of 6491 dichloromethane are seen with acute exposures in ad ults, and the available data regarding 6492 neurological symptom history and standardized testi ng suggest the possibility of longer-term 6493 effects. The relatively high activity of CYP2E1 in the brain compared to the liver of the 6494 developing human fetus raises the potential for neu rodevelopmental effects from 6495 dichloromethane exposure. Results from a developme ntal toxicity study in rats also raise 6496 concern for possible neurodevelopmental effects. D ecreased offspring weight at birth and 6497 changed behavioral habituation of the offspring to novel environments were seen following 6498 exposure of adult Long-Evans rats to 4,500 ppm for 14 days prior to mating and during gestation 6499214

(or during gestation alone) (Bornschein et al., 198 0; Hardin and Manson, 1980). In the only 6500 other animal study examining possible early-life su sceptibility to dichloromethane toxicity, 6501 Alexeef and Kilgore (1983) found that exposure of y oung male mice to approximately 6502 47,000 ppm for about 20 seconds significantly impai red the ability to learn, using a passive6503 avoidance conditioning task. Three-week-old mice w ere more affected than 5- or 8-week-old 6504 mice. The broad issue of childhood susceptibility to chronic neurobehavioral effects of early life 6505 exposure represents a data gap in the understanding of the health effects of dichloromethane. 6506 The relatively low CYP2E1 activity in the liver of infants would tend to shift metabolism 6507 of dichloromethane to the GST pathway. This shift c ould affect cancer risk, given the evidence 6508 of genotoxicity through this metabolic pathway. Ho wever, the available data in humans are not 6509 sufficient to address the question of whether in ut ero or early life exposures represent a period of 6510 increased susceptibility to potential carcinogenic effects of dichloromethane. A threefold 6511 increased risk of childhood leukemia (acute lymphob lastic leukemia) was seen in relation to 6512 maternal occupational exposure in the year before a nd during pregnancy in one population-based 6513 case-control study (OR 3.22 [95% CI 0.88, 11.7]) fo r ratings of “probable or definite” exposure 6514 compared with possible or no exposure (Infante-Riva rd et al., 2005). The estimates for 6515 categories based on concentration and frequency wer e similar, but there was no evidence for an 6516 increasing risk with increasing exposure level. 6517 Experiments comparing cancer responses from early-l ife exposures with those from adult 6518 exposures are not available for F344 rats or B6C3F 1 mice, the strains of animals in which 6519 carcinogenic responses to dichloromethane have been observed (mammary gland tumors in F344 6520 rats and liver and lung tumors in B6C3F 1 mice exposed by inhalation; liver tumors in female 6521

F344 rats and male B6C3F 1 mice exposed via drinking water). Animal data eva luating the effect 6522 of age on the susceptibility to dichloromethane car cinogenicity are restricted to a bioassay in 6523 which 54 pregnant Sprague-Dawley rats were exposed, starting on GD 12, to 100 ppm 6524 dichloromethane 4 hours/day, 5 days/week for 7 week s, followed by 7 hours/day, 5 days/week 6525 for 97 weeks (Maltoni et al., 1988). Groups of 60 male and 69 female newborns continued to be 6526 exposed after birth to 60 ppm dichloromethane 4 hou rs/day, 5 days/week for 7 weeks, followed 6527 by exposure 7 hours/day, 5 days/week for 97 weeks. Additional groups of 60 male and 6528 70 female newborns were exposed after birth to 60 p pm dichloromethane 4 hours/day, 6529 5 days/week for 7 weeks and then for 7 hours/day, 5 days/week for 8 weeks. Endpoints 6530 monitored included clinical signs, BW, and full nec ropsy at sacrifice (when spontaneous death 6531 occurred). For each animal sacrificed, histopathol ogic examinations were performed on the 6532 following organs: brain and cerebellum, zymbal glan ds, interscapular brown fat, salivary glands, 6533 tongue, thymus and mediastinal lymph nodes, lungs, liver, kidneys, adrenals, spleen, pancreas, 6534 esophagus, stomach, intestine, bladder, uterus, gon ads, and any other organs with gross lesions. 6535 There was no significant effect of exposure to dich loromethane on the incidence of benign or 6536 malignant tumors among adults or the progeny. The results provide no evidence that Sprague- 6537215

Dawley rats would be more sensitive to potential ca rcinogenic activity of dichloromethane 6538 during early life stages. Further conclusions from these results are precluded because the study 6539 included only one exposure level, which was below t he maximum tolerated dose for adult 6540 Sprague-Dawley rats. 6541

6542

4.8.2. Possible Gender Differences 6543

The limited data available from studies in humans do not indicate that there are large 6544 differences by gender in sensitivity to cardiovascu lar, neurologic, cancer, or other effects; studies 6545 have not been conducted specifically to examine thi s question and so do not provide information 6546 pertaining to smaller or more subtle differences. The available animal studies similarly do not 6547 establish whether either gender may be more suscept ible to the toxic effects of dichloromethane. 6548 Studies of the carcinogenic effects of dichlorometh ane, either by inhalation or by the oral route, 6549 have not suggested an increased susceptibility of e ither male or female animals. 6550

6551 4.8.3. Other Susceptible Populations 6552 As discussed in section 3.3, a polymorphism exists within the GST-T1 gene in humans, 6553 resulting in individuals with diminished, or a lack of, ability to conjugate GSH to 6554 dichloromethane. While the possible effects of thi s polymorphism on the toxicity of 6555 dichloromethane have not been directly demonstrated , it can be inferred from the proposed mode 6556 of action that a decrease in the GST-T1 metabolic p athway would result in a decreased 6557 generation of reactive metabolites and a decrease i n any chronic effects mediated through those 6558 metabolites (Jonsson and Johanson, 2001; El-Masri e t al., 1999). 6559 Interindividual variation in the ability to metabol ize dichloromethane via GST-T1 is 6560 associated with genetic polymorphisms in humans. E stimated U.S. population prevalence of 6561 nonconjugators (–/– at the GST-T1 locus) is about 20%, but higher prevalences (47–64%) have 6562 been reported for Asians (Raimondi et al., 2006; Ha ber et al., 2002; Garte et al., 2001; Nelson et 6563 al., 1995). Although nonconjugators are expected t o have negligible extra risk for 6564 dichloromethane-induced cancer, the U.S. prevalence s for low (+/– at the GST-T1 locus) and 6565 high (+/+) conjugators have been estimated at 48 an d 32%, respectively (Haber et al., 2002). 6566 The liver and kidney are the most enriched tissues in GST-T1, but evidence is available for the 6567 presence of GST-T1 in other tissues, including the brain and lung, at lower levels (Sherratt et al., 6568 2002, 1997). 6569 Individuals may vary in their ability to metaboliz e dichloromethane through the CYP2E1 6570 pathway. Individuals with decreased CYP2E1 activit y may experience decreased generation of 6571

CO and an increased level of GST-related metabolite s, following exposure to dichloromethane, 6572 which may result in increased susceptibility to the chronic effects of dichloromethane from GST6573 related metabolites. Conversely, individuals with higher CYP2E1 activity may experience 6574 relatively increased generation of CO at a given di chloromethane exposure level and, therefore, 6575216

may be more susceptible to the acute CO-related tox icity or other chronic effects of 6576

dichloromethane. Several studies indicate a three- to sevenfold variability in CYP2E1 activity 6577 among humans, as assessed by various types of measu rements among “healthy” volunteers 6578 (Sweeney et al., 2004; Haufroid et al., 2003; Lipsc omb et al., 2003; Bernauer et al., 2000; Lucas 6579 et al., 2001, 1999; Kim et al., 1995; Shimada et al ., 1994). This variability is incorporated into 6580

the PBTK models for dichloromethane. Factors that may induce or inhibit CYP2E1 activity 6581 (e.g., obesity, alcohol use, diabetes) or co-exposu res (i.e., to various solvents or medications) 6582 (Lucas et al., 1999) may result in greater variatio n within segments of the population. This 6583 variation in CYP2E1 activity may result in earlier saturation of this pathway and greater 6584 exposure to the parent compound, which would be of particular relevance to neurological effects. 6585

6586217 5. DOSE-RESPONSE ASSESSMENTS 6587 6588 6589 5.1. ORAL REFERENCE DOSE (RfD) 6590 5.1.1. Choice of Principal Study and Critical Effe ct—with Rationale and Justification 6591 As discussed in section 4.6.1, human data for oral exposures to dichloromethane are 6592 limited to case reports involving intentional (i.e. , suicidal) or accidental, acute ingestion 6593 exposures (Chang et al., 1999; Hughes and Tracey, 1 993). Reported effects reflect frank toxicity 6594 from very high doses, such as marked CNS depression , injury to the gastrointestinal tract, liver 6595 and kidney failure, coma, and death. No studies of human chronic oral exposures are available. 6596 In the absence of adequate studies evaluating possi ble health effects in humans repeatedly 6597 exposed to dichloromethane via the oral route, the results from the chronic laboratory animal 6598

studies are assumed to be relevant to humans. 6599 The database of laboratory animal oral exposure st udies includes 90-day (Kirschman et 6600 al., 1986) and 2-year drinking water toxicity studi es in F344 rats (Serota et al., 1986a) and 6601

B6C3F 1 mice (Serota et al., 1986b). A reproductive study exposed Charles River CD rats via 6602 gavage before mating (General Electric Co., 1976), and a developmental study exposed F344 rats 6603 via gavage during GDs 6–19 (Narotsky and Kavlock, 1 995). A 14-day gavage study examined 6604 neurotoxicity in F344 rats (Moser et al., 1995). 6605 Hepatic effects (hepatic vacuolation, nonneoplasti c liver foci) are the critical dose6606 dependent noncancer effects associated with oral ex posure to dichloromethane (see Table 4-35). 6607 The 90-day drinking water toxicity study in F344 ra ts (Kirschman et al., 1986) reported 6608 significant increases in hepatocyte vacuolation and necrosis in animals dosed between 166 – 6609 1200 mg/kg-day (males) or 200 – 1469 mg/kg-day (fem ales). These doses were used to develop 6610 dosing levels for the 104-week drinking water study (Serota et al., 1986a). The 104-week 6611 drinking water study of F344 rats (Serota et al., 1 986a) provides adequate data to describe dose6612 response relationships for nonneoplastic liver lesi ons from chronic oral exposure to 6613 dichloromethane (e.g., includes four exposure level s and a control group). In this study, rats 6614 dosed at 50 mg/kg-day or higher in both sexes had i ncreased fatty livers, but quantitative data 6615 were not provided by the authors. Liver lesions, d escribed as foci or areas of cellular alteration, 6616 were also seen in this study in the same dose group s in which the fatty changes had occurred. A 6617 limitation of this study is that Serota et al (1986 a) did not describe the evaluation of the altered 6618 foci in detail. However, increases in altered foci did not correspond to tumor rate incidences in 6619 either male or female rats. Instead, the altered f oci correlated more closely to fatty liver 6620 incidence changes for both sexes in the rats. Alt ered foci could range from a focal fatty change 6621 (nonneoplastic) to an enzymatic altered foci change (neoplastic) (Goodman et al., 1994). Several 6622 lines of evidence were considered in determining wh ether the lesions should be characterized as 6623 nonneoplastic or neoplastic: 1) There is a congruen ce between the incidence of this lesion and 6624218

the incidence of the fatty liver in the study by Se rota et al. (1986a); 2) At higher doses, 6625 hepatocyte vacuolation and hepatocyte necrosis were seen (Kirschman et al., 1986; Berman et 6626 al., 1995); and 3) there is no clear indication tha t these altered foci progress to liver tumors since 6627

the rate of increased foci did not correlate with l iver tumor increases in either male or female 6628 rats. Based on these observations, the altered foc i were determined to be more likely to be 6629 representative of a focal fatty change (nonneoplast ic) than a neoplastic event. 6630 The LOAELs for nonneoplastic liver lesions in rode nts following repeated oral exposure 6631 (50–586 mg/kg-day) (Table 4-35) are in the same ran ge or below the NOAELs of 225 mg/kg-day 6632 for reproductive performance in Charles River CD ra ts exposed for 90 days before mating 6633 (General Electric Co., 1976) and 450 mg/kg-day for developmental toxicity in pregnant F344 6634 rats exposed during gestation (Narotsky and Kavlock , 1995). The LOAEL (337 mg/kg-day) and 6635 NOAEL (101 mg/kg-day) for mild neurological impairm ent in a 14-day gavage exposure study 6636 of F344 rats (Moser et al., 1995) indicates that th e threshold for neurological effects may be 6637

similar to the threshold for liver effects. A limi tation of the Moser et al. (1995) study, however, 6638

is that the observed effects were limited to measur es taken within 4 hours of exposure. 6639 The subchronic (i.e., 90-day or less study) data w ere not considered in the selection of a 6640 principal study for deriving the chronic RfD becaus e the database contains reliable dose-response 6641 data from a chronic study at lower doses than the 9 0-day study (Kirschman et al., 1986) 6642 (conducted to provide data pertaining to relevant d oses to use in the chronic study). The data 6643 from the subchronic studies are, however, used to c orroborate the findings in the chronic studies 6644 with respect to relevant endpoints (i.e., hepatic a nd neurological effects). The neurotoxicity 6645 study was not selected as the principal study due t o the limited measurements to inform the 6646 chronic exposure to dichloromethane. The rat rathe r than the mouse chronic bioassay (Serota et 6647 al., 1986a) was selected as the principal study for the RfD because of the consistent evidence that 6648 rats may be more sensitive than mice to nonneoplast ic liver effects from orally administered 6649 dichloromethane; available rat LOAELs for nonneopla stic liver lesions are lower than mouse 6650 LOAELs (see Table 4-35). Figure 5-1 is an exposure -response array that presents NOAELs, 6651 LOAELs, and the dose range tested, corresponding to selected health effects from the short-term 6652 (neurotoxicological) and subchronic studies, and fr om the chronic, reproductive, and 6653 developmental toxicity studies that were evaluated for use in the derivation of the RfD. 6654219Figure 5-1. Exposure Response Array for Oral Exposure to Dichloromethane.02004006008001000120014002000

Dose (mg/kg-day)Hepatocytevacuolation (F344rat) - Kirschman etal. (1986)Hepatocytevacuolation(B6C3F1 malemice) -Kirschman etal. (1986)Nonneoplasticliver foci (F344rat, male andfemale) - Serotaet al. (1986a)Fatty liver(B6C3F1mice, maleandfemale) -Serota etal. (1986b)Neurologic,FunctionalObservationalBattery (F344rat, female) -Moser et al.(1995)ReproductivePerformance(CD rat, maleand female) -GeneralElectric Co.(1976)Maternal weightgain (F344 rat,pregnant female)- Narotsky andKavlock (1995)FetalToxicity(F344 rat) -Narotskyand Kavlock(1995)SUBCHRONIC HEPATICCHRONIC HEPATICNEUROTOX REPRODUCTIVE AND DEVELOPMENTALNOAELLOAELThe vertical lines =range of exposures instudy.Closed dots ( ) =exposure concentrationsused in studyReproductiveorgans;performance(male SwissWebster ) -Raje et al.(1988)66556656Figure 5-1. Exposure response array for oral exposure to dichloromethane6657220

5.1.2. Derivation Process for Noncancer Reference Values 6658 The toxicity values (oral RfD and inhalation RfC) for noncancer endpoints were derived 6659 by using rat and human PBTK models to calculate int ernal doses in rats from experimental 6660 exposures and extrapolate points of departure to hu man equivalent exposures. Figure 5-2 6661 illustrates the process of using the PBTK models fo r toxicity value derivation. The process for 6662 the RfD and RfC is summarized below, using the exam ple of a noncancer liver effect. 6663

A deterministic PBTK model for dichloromethane in rats was first used to convert rat 6664 drinking water or inhalation exposures to values of an internal liver dose metric (see Appendix C 6665 for details of the rat PBTK model). Available model s in EPA benchmark dose (BMD) software 6666 (BMDS) version 2.0 were then fit to the liver lesio n incidence data and internal liver dose data 6667 for rats, and BMD 10 s and their lower 95% confidence limits associated with a 10% extra risk 6668 (BMDL 10 ) were calculated from each of the models. Adequac y of model fit was assessed by 6669

overall χ 2 goodness of fit (

p -value > 0.10) and examination of residuals, partic ularly in the region 6670 of the benchmark response (BMR). The choice of bes t-fitting model was based on the lowest 6671 Akaike’s Information Criterion [AIC] among models w ith adequate fits (U.S. EPA, 2000b). 6672 The use of a PBTK model can replace the use of the BW 0.75 scaling factor to account for 6673 interspecies differences in toxicokinetics. The de cision with respect to use of a scaling factor 6674 depends on the dose metric that is used. Where PBT K models predict the concentration (in 6675 particular, the AUC) of the proximate causative age nt, a scaling factor to account for interspecies 6676 differences is not typically used. That is, it is assumed that if the time-averaged (or steady-state) 6677 concentration of the proximate causative agent pred icted by the PBTK model in the target tissue 6678 is the same in the test species as in humans, and t he test species was exposed for an equivalent 6679 portion of its lifetime (2 years in rats and mice b eing equivalent to a 70-year lifetime in humans), 6680 then the resulting risks in the two species are the same. However, when the PBTK model 6681 predicts the rate of production of the agent, rathe r than its concentration, then a BW 0.75 scaling 6682 factor may be appropriate, depending on what is kno wn or expected regarding the rate of 6683 clearance of the agent or metabolite of interest. Two different scenarios can be considered. If 6684 the metabolite formed is considered to be highly re active, then it can be assumed that the rate of 6685 clearance (i.e., disappearance due to local reactiv ity) for this metabolite, per volume tissue, is 6686 equal in rodents and humans. Thus, in that situati on, as with the AUC dose metric, no BW 0.75 6687 scaling factor is necessary, although differences i n tissue volume fraction in humans versus rats 6688 (as occurs for liver) should be and are accounted f or by the PBTK model. However, if the 6689 metabolite is not highly reactive, then it is expec ted that interspecies differences in clearance or 6690 removal of the toxic metabolite follow the generall y assumed BW 0.75 scaling for rates of 6691 metabolism and blood circulation. In this case, or in situations in which the reactivity or rate of 6692 removal of the metabolite has not been established, it is appropriate to use a scaling factor, based 6693 on BW ratios, to account for this difference. In t he case of the noncancer liver effects of 6694 dichloromethane, very limited information is availa ble on the mechanism(s) involved in creating 6695221Rodent DoseResponse DataPBTKModelEstimates of RodentInternal DoseRodent Internal BMDL1095% Lower Bound Estimate of InternalDose Associated with a 10% responseBMDModelingProbabilisticHuman PBTK Model(What administered doses willproduce a BMDL10in apopulation?)Monte Carlo Sampling fromDistributions of Human ModelParametersDistribution of Human Equivalent AdministeredDoses (mg/kg) or Inhalation Concentrations (mg/m3)(Points of Departure)1st5thDivide by Uncertainty Factorsfor Interspecies ToxicodynamicVariability, HumanToxicodynamic Variability andDatabase Deficiencies )Recommend lower percentile (e.g.,1st) to protect sensitive individualsOral Reference Doses orInhalation ReferencesConcentrationsBMDBMDL10ScalingFactorHuman InternalBMDL106696Figure 5-2. Process for deriving noncancer oral RfDs and inhalation RfCs using rodent and human PBTK6697models.6698222

the type of hepatic damage seen. The dose metric u sed in the PBTK modeling is a rate of 6699 metabolism, rather than the concentration of putati ve toxic metabolites, and the clearance of 6700 these metabolites may be slower per volume tissue i n the human compared with the rat. Thus the 6701

rat internal dose metric for noncancer effects was adjusted by dividing by a pharmacokinetic 6702 scaling factor to obtain a human-equivalent interna l BMDL 10 6703 A probabilistic PBTK model for dichloromethane in humans, adapted from the model 6704 of David et al. (2006) as described in Appendix B, was then used to calculate distributions of 6705 chronic exposures associated with the human equival ent internal BMDL 10 , based on the 6706 responses in rats. Parameters in the human PBTK mo del are distributions that incorporate 6707

information about dichloromethane toxicokinetic and physiological variability and 6708 uncertainty among humans, incorporating information about both the CYP2E1 and GST-T1 6709 metabolic pathways (see Table 3-9 and Appendix B). Monte Carlo sampling was performed 6710 in which each human model parameter was defined by a value randomly drawn from each 6711 respective parameter distribution. The model was t hen executed by using the human internal 6712

BMDL 10 as input, and the resulting human equivalent admin istered dose or human equivalent 6713 concentration (HEC) was recorded. This process was repeated for 10,000 iterations to 6714 generate a distribution of human equivalent adminis tered doses or concentrations. 6715 The parameter statistics reported by David et al. ( 2006) include both the inter-individual 6716 variability that would have been elucidated by the Bayesian analysis (variation between mean 6717

values for each individual for which data were avai lable) and uncertainty in those values. Since 6718 EPA’s objective is to account for both population v ariability and parameter uncertainty, 6719 however, these statistics were primarily used as pu blished in David et al. (2006) (exceptions 6720 discussed in Appendix B) to define population distr ibutions. Assuming that these parameters are 6721 distributed independently, ignoring the covariance that was likely represented in the actual 6722 posterior chains, will tend to over-estimate the ov erall range of parameters and hence distribution 6723 of dose metrics in the population, compared to what one would obtain if the covariance were 6724 explicitly included. Thus if the covariance (i.e., the variance-covariance matrix) for the set of 6725 parameters had been reported by David et al., it co uld have been used to narrow the predicted 6726 distribution of internal doses, or equivalent appli ed doses. Lacking such information the 6727 approach used will not under-estimate risk or over- estimate lower bounds on human equivalent 6728 exposure levels. 6729 From these distributions of human equivalent admin istered doses (or concentrations), 6730 candidate RfDs or RfCs were derived by dividing the first percentile value (point of 6731 departure) by uncertainty factors (UFs) to account for uncertainty about potential interspecies 6732 toxicodynamic variability, human toxicodynamic vari ability, and database deficiencies. The 6733 first percentile was chosen because it allowed gene ration of a stable estimate for the lower 6734 end of the distribution while being protective of t he overall human population, including 6735223

sensitive individuals. Choosing this lower point r eplaces the use of an additional UF to 6736 account for human toxicokinetic variability. 6737

6738

5.1.3. Evaluation of Dose Metrics for Use in Nonca ncer Reference Value Derivations 6739 There are no data to support the role of a specific metabolite in the development of 6740

the noncancer liver lesions seen in oral and inhala tion exposure studies. Four dose metrics 6741 were examined as potential metrics for the internal dose of interest: rate of hepatic 6742 metabolism through the CYP pathway, rate of hepatic metabolism through the GST pathway, 6743 the combined rate of hepatic metabolism through the CYP and GST pathways, and the 6744 concentration (area under the curve, AUC) of dichlo romethane, the parent compound, in the 6745 liver. The dose-response patterns for each of these metrics in the oral study in rats (Serota et 6746 al., 1986a) and in two inhalation studies in rats ( Nitscke et al., 1988a; Burek et al., 1984) 6747 were examined for fit and congruence. 6748 Using the oral exposure data, only one of the seven models, the log-logistic model, 6749 produced an adequately fit (p > 0.10) for the GST m etabolism metric and the 6750 dichloromethane AUC metrics. Adequate model fit was seen in all of the models using the 6751 CYP dose metric with the oral data, and using the G ST, CYP, and AUC dose metrics for the 6752

inhalation data. 6753 A limitation in using the GST metric can be observe d when comparing the oral and 6754 inhalation responses at various exposure levels. At 200 ppm, where the GST metric is 6755 predicted by the PBTK model to be 93 mg metabolism/ L liver/day, no liver effects were 6756 seen. In contrast, liver responses were elevated a t an oral dose of 50 mg/kg-day, where the 6757 GST metric is predicted to be 60 mg metabolism/L li ver/day (see Tables 5-1 and 5-5, 6758 respectively, for the oral and inhalation internal metrics). Thus the liver GST metric 6759 produces an inconsistency in the dose-response rela tionship, with very different responses 6760 observed depending on the route of exposure. A sim ilar inconsistency occurs with the AUC 6761 metric. These differences are not observed, howeve r, when using the CYP metric. At the 6762 200 ppm inhalation exposure, where no hepatoxicity was observed, the CYP metric is 6763 predicted to be 660 mg/L liver/day. This internal CYP metabolism metric is less than that 6764 predicted for the oral dose for the 50 mg/kg-day gr oup (i.e., 872 mg metabolism/L liver/day), 6765 in which liver effects were observed. Thus, the C YP internal metric is consistent with the 6766 observed responses seen in the oral and inhalation exposure studies. 6767

The GST metabolism and the AUC dose metrics did n ot present reasonable choices 6768 based on model fit and consistency of response acro ss studies at comparable dose levels. 6769 Given these results, the combination of hepatic met abolism through the GST and the CYP 6770 pathways would not be expected to result in an impr ovement to a metric based only on CYP 6771

metabolism. The CYP-metabolism dose metric is the most consistent with the data. This 6772 metric was selected for the subsequent RfD and RfC derivations. The lack of information on 6773224

mechanisms with respect to noncancer health effects represents data gaps in the 6774 understanding of the health effects of dichlorometh ane. 6775

6776 5.1.4. Methods of Analysis—Including Models (PBTK, BMD, etc.) 6777

PBTK models for dichloromethane in rats were descr ibed previously in section 3.5. 6778 From the evaluation described in Appendix C, a modi fied model of Andersen et al. (1991) 6779

was selected for the calculation of internal dosime try of ingested dichloromethane in the rats 6780 in the principal study (Serota et al., 1986a). 6781 PBTK model simulations of the drinking water study of Serota et al. (1986a) (Table 6782 5-1) were performed to calculate average lifetime d aily internal liver doses in male and 6783 female F344 rats. In the absence of data for group - and sex-specific BWs, reference values 6784

were used for male and female F344 rats in chronic studies (U.S. EPA, 1988a). The mode of 6785 action by which dichloromethane induces nonneoplast ic liver effects in rodents has not 6786

received research attention to determine the role o f the parent material, metabolites of the 6787 CYP2E1 pathway, metabolites of the GST pathway, or some combination of parent material 6788

and metabolites. In the absence of this kind of kn owledge, and considering the pattern of 6789 response seen in the oral and inhalation studies (a s described in section 5.1.3.), an internal 6790

dose metric based on the amount of dichloromethane metabolized via the CYP pathway in 6791 the liver (mg dichloromethane metabolized via CYP p athway per liter liver per day) was 6792 used. Figure 5-3 shows the comparison between oral external and internal doses, using this 6793

dose metric for the rat and for the human. 6794225 6795 6796 Table 5-1. Incidence data for nonneoplastic liver lesions and internal liver doses, based on various metrics, in male and female F344 rats exposed to dichloromethane in drinking water for 2 years (Sero ta et al., 1986a)

Rat internal liver dose b

Sex Nominal

(actual) daily intake (mg/kg-day) Rat liver

lesion incidence a CYP GST GST and CYP Parent

AUC 0 (0) 52/76 (68%) 0 0 0 0 5 (6) 22/34 (65%) 133.9 2.1 136.1 0. 5 50 (52) 35/38 (92%) c 872.7 58. 8 931.4 13. 1 125 (125) 34/35 (97%) c 1,433.1 236.0 1,669.1 52. 6 Male (BW = 380 g)

250 (235) 40/41 (98%) c 1,868.6 561.5 2,430.0 125.0

0 (0) 34/67 (51%) 0 0 0 0 5 (6) 12/29 (41%) 134.5 2.1 136.6 0. 4 50 (58) 30/41 (73%) c 97 7.8 66. 0 1,04 3.8 12.6 125 (136) 34/38 (89%) c 1577. 0 258.7 1,835.7 49.5 Female (BW = 229 g )

250 (263) 31/34 (91%) c 2070.0 642.4 2,712.3 122.9

a Liver foci/areas of cellular alteration; number aff ected divided by total sample size.

b Internal doses were estimated using a rat PBTK mode l from simulations of actual daily doses reported by the study authors. CYP dose is in units of mg dichloromethane metabolized via CYP pathway/L tissue/day; GST dose is in units of mg di chloromethane metabolized via GST pathway/L tissue/day.; GST and CYP dose is in units of mg di chloromethane metabolized via CYP and GST pathways/L tissue/day; and Parent AUC dose is in u nits of mg dichloromethane*hrs)/L tissue.

c Significantly (

p < 0.05) different from control with Fisher’s exact test.

Source: Serota et al., 1986a. 6797 6798226 6799

6800 6801 Figure 5-3. PBTK model-derived internal doses (mg dichloromethane metabolized via the 6802 CYP pathway per liter liver per day) in rats and hu mans and their associated external 6803 exposures (mg/kg-day), used for the derivation of R fDs. Six simulated daily drinking water 6804 episodes are described by Reitz et al. (1997). The human metabolism rates were estimated 6805 using a computational sample of 1000 individuals pe r dose, including random samples of the 6806 three GST-T1 polymorphisms (+/+, +/-, -/-) in the c urrent U.S. population based on data from 6807 Haber et al. (2002). Since a different set of samp les was used for each dose, some stochasticity 6808 is evident as the human points (values) do not fall on smooth curves. 6809 6810 The seven dichotomous dose-response models in BMDS version 2.0 were fit to the rat 6811

liver lesion incidence data and PBTK model-derived internal dose data to derive a rat internal 6812

BMD 10 , and corresponding BMDL 10 , associated with 10% extra risk (Table 5-2). The quantal 6813 model is identical to the one-stage multistage mode l and so is not included in this set of models. 6814 A BMR of 10% was selected because, in the absence o f information regarding the magnitude of 6815 change in a response that is thought to be minimall y biologically significant, a BMR of 10% is 6816 generally recommended, since it provides a consiste nt basis of comparison across assessments. 6817 There are no additional data to suggest that the cr itical response has a greater sensitivity that 6818

would warrant a lower BMR. The male rats exhibited a greater sensitivity compared to the 6819 female rats (based on lower BMDL 10 values for all of the models examined) and thus th e male 6820

data are used as the basis for the RfD derivation. The logistic model was the best fitting model 6821 for the male incidence data, based on AIC value amo ng models with adequate fit (U.S. 6822 EPA,2000b). Modeling results are shown in detail in Appendix D-1). 6823(mg DCM metabolized/L liver/day)227 6824

Table 5-2. BMD modeling results for incidence of noncancer liver l esions in male and female F344 rats exposed to dichloromet hane in drinking water for 2 years, based liver-specific CYP metabol ism dose metric (mg dichloromethane metabolism via CYP pathway per lite r liver tissue per day)

Sex and model a BMD 10 BMDL 10 χ 2 goodness of fit

p -value AIC

Males Gamma a 151.73 48.93 0.62 185.33 Logistic 85.17 61.78 0.75 183.61 Log-logistic a 213.73 37.06 0.83 184.79 Multistage (1) a 68.62 47.58 0.71 183.74

Probit 98.87 75.49 0.69 183.81 Log-probit a 197.65 77.56 0.81 184.84

Weibull a 117.29 48.39 0.57 185.49

Females Gamma a 336.38 98.70 0.52 233.07 Logistic 169.77 134.87 0.59 231.70 Log-logistic a 404.87 101.15 0.60 232.80 Multistage (1) a 123.59 91.46 0.47 232.32

Probit 179.59 146.27 0.59 231.70 Log-probit a 400.95 173.57 0.60 232.80

Weibull a 283.24 97.31 0.47 233.27

a These models in EPA BMDS version 2.0 were fit to th e rat dose-response data shown in Table 5-1 by using internal dose metrics calculated with the rat PBTK model. Details of the models are as follows: Gamma and Weibull models re strict power ≥ 1; Log-logistic and Log-probit models restrict to slope >1, multistage model restrict betas ≥ 0; lowest degree polynomial with an adequate fit is reported (degree of polynomial noted in parentheses). Bolded model is the best-fitting model in the most sensitive sex (males), which is used in the RfD derivation.

Source: Serota et al. (1986a).

. 6825 The BMDL 10 from the logistic model was used as the point of d eparture for the RfD 6826 calculations (Table 5-3). This rat internal dose m etric for noncancer effects was adjusted to 6827 obtain a human-equivalent internal BMDL 10 by dividing by a pharmacokinetic scaling factor 6828 based on a ratio of BWs (BW human /BW rat ) 0.25 = 4.09). This scaling factor was used because the 6829 metric is a rate of metabolism, rather than the con centration of putative toxic metabolites, and the 6830 clearance of these metabolites may be slower per vo lume tissue in the human compared with the 6831 rat (that is, total rate of removal may scale as BW 0.75 , while tissue volume scales as BW 1 ). 6832 The human PBTK model (adapted from David et al. [2 006], as described in Appendix B), 6833 using Monte Carlo sampling techniques, was used to calculate quantiles of human equivalent 6834 administered oral daily doses (in mg/kg-day) associ ated with the internal BMDL 10 values 6835228

(Table 5-3), as described above in section 5.1.2. The human model used parameter values 6836 derived from Monte Carlo sampling of probability di stributions for each parameter, including 6837 MCMC-derived distributions for the metabolic parame ters for the metabolism through the 6838 CYP2E1 pathway (V max and K m ) and a distribution of GST metabolic rate constant s that is 6839 weighted to reflect the estimated frequency of GST- T1 genotypes (20% GST-T1 –/– , 48% GST6840

T1 +/– , and 32% GST-T1 +/+ ) in the current U.S. population, based on data fro m Haber et al. 6841 (2002). All simulations also included a distributi on of CYP activity based on data from 6842 Lipscomb et al. (2003). The drinking water exposures were comprised of six discrete drinking 6843 water episodes for specified times and percentages of total daily intake (Reitz et al., 1997). The 6844 mean and two lower points on the distributions of h uman equivalent administered daily doses 6845 derived from the Serota et al. (1986a) data for mal e rats, using the BMDL 10 from the logistic 6846 model , are shown in Table 5-3. 6847

6848

Table 5-3. RfD for dichloromethane based on PBTK m odel-derived probability distributions of human drinking water e xposures extrapolated from nonneoplastic liver lesion incide nce data for male rats exposed via drinking water for 2 years, based on liver-specific CYP metabolism dose metric (mg dichloromethane meta bolized via CYP pathway per liter liver tissue per day)

Human equivalent dose (mg/kg-day) d

Model a Rat

internal BMDL 10 b Human

internal BMDL 10 c 1 st

percenti le 5 th

percentile Mean Human

RfD (mg/m 3 ) e

Logistic 61.78 15.11 0.214 0.252 0.395 7 x 10 -3

a Based on the best-fitting model from Table 5-2.

b Rat dichloromethane PBTK model-derived internal liv er dose associated with the lower bound on 10% extra risk for developing li ver foci/areas of cellular alteration.

c Human dichloromethane internal liver dose, derived by dividing the rat internal BMDL 10 by a scaling factor of 4.09 [(BW human /BW rat ) 0.25 ] to account for potential interspecies pharmacokinetic differences in the cle arance of metabolites.

d PBTK model-derived distributions of daily average d ichloromethane drinking water doses predicted by the PBTK model to yield an internal dose in humans equal to the dichloromethane internal BMDL 10 , e Human RfD, based on male rat data, derived by divid ing the 1 st percentile of human equivalent dose value by a total UF of 30: 3 (10 0.5 ) for possible toxicodynamic differences between species, 3 (10 0.5 ) for variability in human toxicodynamic response, and 3 (10 0.5 ) for database deficiencies. The 1 st percentile point of departure is a stable estimate of the lower end of the distribution. Use of this value in the lower tail replaces use of a UF for human toxicokinetic variability.

Source: Serota et al. (1986a). 6849229 6850 5.1.5. RfD Derivation—Including Application of Uncertainty Factors (UFs) 6851

The 1 st percentile point of departure is a stable estimate of the lower end of the 6852

distribution. Use of this value associated with a sensitive human population addresses the 6853 uncertainty associated with human toxicokinetic var iability. To derive the candidate RfD based 6854 on data from male rats, the first percentile value of the distribution of human equivalent 6855

administered dose associated with the male rat-deri ved BMDL 10 was divided by a composite UF 6856 of 30 (3 [10 0.5 ] to account for uncertainty about interspecies tox icodynamic equivalence, 3 [10 0.5 ] 6857 to account for uncertainty about toxicodynamic vari ability in humans, and 3 [10 0.5 ] for database 6858 deficiencies) (Table 5-3). The resulting RfD recom mended for dichloromethane is 7 × 10 –3 6859 mg/kg-day. 6860 In deriving this RfD, factors for the following are as of uncertainty were considered: 6861

• Uncertainty in extrapolating from laboratory animal s to humans (UF A ). The use of 6862 PBTK models to extrapolate internal doses from rats to humans reduces toxicokinetic 6863 uncertainty in extrapolating from the rat liver les ion data but does not account for the 6864 possibility that humans may be more sensitive than rats to dichloromethane due to 6865 toxicodynamic differences. A UF of 3 (10 0.5 ) to account for this toxicodynamic 6866 uncertainty was used, as shown in Table 5-3. 6867

6868

• Uncertainty about variation from average humans to sensitive humans (UF H ). The 6869 probabilistic human PBTK model used in this assessm ent incorporates the best available 6870 information about variability in toxicokinetic disp osition of dichloromethane in humans 6871 but does not account for humans who may be sensitiv e due to toxicodynamic factors. 6872 Thus, a UF of 3 (10 0.5 ) was applied to account for possible toxicodynamic differences in 6873

sensitive humans. 6874

6875

• Uncertainty in extrapolating from LOAELs to NOAELs (UF L ). A UF for extrapolation 6876 from a LOAEL to a NOAEL was not applied because BMD modeling was used to 6877 determine the POD, and this factor was addressed a s one of the considerations in 6878 selecting the BMR. The BMR was selected based on th e assumption that it represents a 6879 minimum biologically significant change. 6880

6881

• Uncertainty in extrapolating from subchronic to chr onic durations (UF S ). The derived 6882

RfD is based on results from a chronic-duration dri nking water toxicity study. No cross6883 duration UF is necessary. 6884

6885

• Uncertainty reflecting incompleteness of the overal l database (UF D ). The oral database 6886

for dichloromethane includes well-conducted lifetim e drinking water studies in rats 6887230

(Serota et al., 1986a) and mice (Serota et al., 198 6b) and a supporting subchronic study in 6888 rats and mice (Kirschman et al., 1986). These stud ies provided dose-response data for 6889 the hepatic effects of dichloromethane. The databa se also includes one-generation oral 6890 reproductive toxicity (General Electric Co., 1976) and developmental toxicity (Narotsky 6891 and Kavlock, 1995) studies that found no reproducti ve or developmental effects at dose 6892 levels in the range of doses associated with liver lesions. A two-generation oral exposure 6893 study is not available; however, a two-generation i nhalation exposure study by Nitschke 6894 et al. (1988a) reported no effect on fertility inde x, litter size, neonatal survival, growth 6895 rates, or histopathologic lesions at exposures of ≥ 100 ppm. However, there have been no 6896 oral exposure studies that evaluated neurobehaviora l effects in offspring. This is a 6897 relevant endpoint given the increase in blood CO (a known developmental neurotoxicant) 6898 that occurs through the CYP2E1 metabolic pathway fo r dichloromethane after oral and 6899 inhalation exposures. There are no oral exposure s tudies that include functional immune 6900 assays; however, there is a 4-week inhalation study of potential systemic immunotoxicity 6901 that found no effect of dichloromethane exposure at concentrations up to 5,000 ppm on 6902 the antibody response to sheep red blood cells (War brick et al., 2003). The Warbrick et 6903 al. (2003) data suggest that systemic immunosuppres sion is not a concern for 6904 dichloromethane exposure. Route-specific local imm unosuppression from acute 6905 inhalation exposure in CD1 mice was seen in Aryani et al. (1986). The findings from 6906 Aryani et al. (1986) were considered to be portal-o f-entry effects involving local 6907 immunosuppression within the lung (Streptococcus an d Klebsiella infectivity models) 6908 and unlikely to be observed following oral exposure . Because of concern regarding the 6909 adequacy of available data pertaining to possible n eurodevelopmental toxicity and the 6910 lack of a two-generation reproductive study, a UF D of 3 was applied. 6911

6912

5.1.6. Previous RfD Assessment 6913 The previous IRIS assessment derived an RfD of 0.0 6 mg/kg-day based on the NOAELs 6914 of 5.85 and 6.47 mg/kg-day for nonneoplastic liver toxicity (foci/areas of cellular alteration) in 6915 male and female rats, respectively, in a 2-year dri nking water study (Serota et al., 1986a). The 6916

LOAELs associated with these NOAELs were 52.58 and 58.32 mg/kg-day for males and 6917 females, respectively. The RfD of 0.06 mg/kg-day w as derived by dividing the average NOAEL 6918 of 6 mg/kg-day (for male and female rats) by a UF o f 100 (10 for intraspecies variability and 10 6919 for interspecies variability). 6920

6921

5.1.7. RfD Comparison Information 6922 Use of the mean value (3.95

 ×

 10 –1 mg/kg-day) of the human equivalent administered 6923

dose distribution instead of the 1 st percentile, with an additional UF of 3 (10 0.5 ) to account for 6924 human toxicokinetic variability, would yield an RfD of 4 ×

 10 –3 mg/kg-day. 6925231

Additional comparisons between the derived RfD and values developed from other 6926 endpoints or data sets using NOAEL/LOAEL methods ar e shown in Table 5-4 and Figure 5-4. 6927 NOAELs were used as comparison points of departure and were not scaled allometrically. The 6928 point of departure for three endpoints (Serota et a l., 1986a; Moser et al., 1995; Narotsky and 6929 Kavlock) are presented in Table 5-4. For Serota et al. (1986a), this is based on BMD modeling 6930 of a 10% increase in liver lesions using internal liver dose metric (mg dichloromethane 6931 metabolism via CYP pathway per liter liver tissue p er day) derived from a rat PBTK model. 6932 After an allometric scaling factor of 4.09 was appl ied, the human internal BMDL 10 was 15.11 6933 mg/kg-day. A probabilistic human PBTK model adapte d from David et al. (2006) was used to 6934 generate a distribution of human equivalent doses f rom the human internal BMDL 10 and the first 6935 percentile of this distribution was used as the poi nt of departure. For other studies, POD is based 6936

on the lowest non-control dose in which no effect w as seen. These NOAELs that were used as 6937 points of departure and were not scaled allometrica lly 69382326939Table 5-4. Potential points of departure with applied UFs and resulting candidate RfDsUncertainty Factors AppliedbEndpointPODa(mg/kg-day)POD Type andDescriptionTotalUF UFAUFHUFLUFSUFDCandidateRfD(mg/kg-day) ReferenceNonneoplasticliver foci, malerats61.78BMD; 10% increasein incidence of liverlesion30 3 3 1 1 37 x 10-3Serota et al.(1986a)Neurologicalchanges (FOB),female rats101NOAEL; No effect atPOD, approximatedoubling of severityscore of neuromuscularand sensorimotordomains3,000 10 10 1 10 33.4×10–2Moser et al.(1995)Maternal weightgain, female rats338NOAEL; No effect atPOD, approximate 33%decrease in weight gainseen at next dose300 10 10 1 1 3 1.1Narotsky andKavlock(1995)aPOD = point of departure. .bUFA=uncertainty in extrapolating from laboratory animals to humans, UFH= uncertainty about variation from average humans to sensitive humans, UFL=uncertainty about extrapolating from LOAEL to NOAEL, UFS= uncertainty in extrapolating from subchronic tochronic durations, and UFD= uncertaintyreflecting incompleteness of the overall database.A UF for extrapolation from a LOAEL to NOAEL (UFL) was not used for any of these studies. For theSerota et al., (1986a) study, the use of the firstpercentile of the human equivalent dose distribution as the point of departure replaces the use of a UFHforhuman toxicokinetic variability.Bolded value is the basis for the RfD of 7 × 10–3mg/kg-day.23369406941Figure 5-4. Comparison of candidate RfDs derived from selected point of departures for endpoints6942presented in Table 5-4.69436944234

5.2. INHALATION REFERENCE CONCENTRATION (RfC) 6945 5.2.1. Choice of Principal Study and Critical Effe ct—with Rationale and Justification 6946 Figure 5-5 includes exposure-response arrays from some of the human studies that were 6947

evaluated for use in the derivation of the RfC. Se veral acute-duration controlled exposure 6948 studies (section 4.1.2.2) and cross-sectional occup ational studies (sections 4.1.2.3 and 4.1.2.4) in 6949 humans are available that show neurological effects from dichloromethane exposure. These 6950 effects include an increase in prevalence of neurol ogical symptoms among workers (Cherry et 6951 al., 1981) and possible detriments in attention and reaction time in complex tasks among retired 6952 workers (Lash et al., 1991). However, these studie s have inadequate power for the detection of 6953 effects with an acceptable level of precision. In addition, the Cherry et al. (1981) study is limited 6954 by the definition and documentation of neurological symptom history, and the Lash et al. (1991) 6955 study has exposure measurements from 1974–1986, but the work histories of exposed workers 6956 go back to the 1940s. Ott et al. (1983c) reported an increase in serum bilirubin among exposed 6957 workers, but there was no association seen with res pect to the other hepatic enzymes examined 6958 (serum γ

-glutamyl transferase, serum AST, serum ALT), and n o evidence of hepatic effects was 6959 seen in a later study of the same cohort (Soden, 19 93). Because of these limitations, these 6960 human studies of chronic exposures do not serve as an adequate basis for RfC derivation. As 6961 discussed in section 5.2.6, however, the quantitati ve measures of neurological function from 6962 Cherry et al. (1983) were used to derive a comparat ive RfC. 6963 The database of experimental animal dichloromethan e inhalation studies includes 6964 numerous 90-day and 2-year studies, with data on he patic, pulmonary, and neurological effects, 6965 (see Table 4-36) and reproductive and developmental studies (Table 4-37) (see summary in 6966 Section 4.6.2). NOAELs, LOAELs, and the dose range tested corresponding to selected health 6967 effects from the chronic studies are shown in Figur e 5-5, and effects seen in subchronic, 6968 reproductive, and developmental studies are shown i n Figure 5-6. The subchronic (i.e., 90-day 6969 or less study) data were not considered in the sele ction of a principal study for deriving the RfC 6970 because the database contains reliable dose-respons e data from the chronic study at lower doses 6971 than the 90-day study. The data from the subchroni c studies are, however, used to corroborate 6972 the findings with respect to relevant endpoints (i. e., hepatic and neurological effects). 6973235110100100010000Hepatocytevacuolation,SpragueDawley rat(Nitschke etal., 1988a)Hepatocytenecrosis,SpragueDawley rat(Burek etal., 1984)Hepatocytevacuolation,male andfemaleSpragueDawley rat(Burek etal., 1984)Hepatocytevacuolation,necrosis,hemosiderosisin liver, maleand femaleF344 rat(Mennear et al.,1988; NTP,1986)Renal tubulardegeneration,male andfemale F344rat (Mennearet al., 1988;NTP, 1986)Hepatocytedegeneration,male andfemaleB6C3F1 mice(Mennear etal., 1988;NTP, 1986)Renaltubule casts,mlae andfemaleB6C3F1mice(Mennear etal., 1988;NTP, 1986)Changesin CNSmeasures,males(Lash etal., 1991)Chronicneurologicaleffects,males(Cherry etal., 1983)Cardiac –ST segmentdepression,males (Ottet al.,1983c)NOAELLOAELThe vertical lines = range ofexposures in study.Closed dots ( ) = exposureconcentrations used in studyRAT MOUSE HUMAN

Concentration (ppm)697469756976Figure 5-5. Exposure response array for chronic (animal) or occupational (human) inhalation exposureto6977dichloromethane (log Y axis)697823669796980Figure 5-6. Exposure response array for subacute to subchronic inhalation exposure to dichloromethane(log Y6981axis)6982237

Hepatic effects (hepatic vacuolation and necrosis, hemosiderosis, hepatocyte 6983 degeneration) are the critical dose-dependent nonca ncer effects associated with inhalation 6984 exposure to dichloromethane. These effects were se en in mice (Mennear et al., 1988; NTP, 6985 1986) and rats (Mennear et al., 1988; Nitschke et a l., 1988a; NTP, 1986; Burek et al., 1984), but 6986 not in Syrian golden hamsters (Burek et al., 1984). Inhalation bioassays with Sprague-Dawley 6987

rats identified the lowest inhalation LOAEL for non neoplastic liver lesions in the database: 500 6988 ppm (6 hours/day, 5 days/week for 2 years) (Nitschk e et al., 1988a; Burek et al., 1984), and 6989 Nitschke et al. (1988a) identified a NOAEL of 200 p pm in female rats. Based on the results 6990 reviewed above, nonneoplastic liver lesions (specif ically, hepatic vacuolation) in rats are 6991

identified as the critical noncancer effect from ch ronic dichloromethane inhalation in animals. 6992 Because Nitschke et al. (1988a) examined a range of exposures that included doses at the low 6993 end of the rangecompared with the range examined in Burek et al. (1984), the former study was 6994 selected as the principal study for derivation of a chronic inhalation RfC. 6995 Reproductive performance (e.g., as assessed by num ber of litters, resorption rate, fetal 6996 survival, and growth) was not affected in two gener ations of F344 rats exposed to up to 6997 1,500 ppm for 14 or 17 weeks before mating of the F 0 and F1 generations, respectively 6998 (Nitschke et al., 1988b) or in a study of Swiss-Web ster mice or Sprague-Dawley rats exposed to 6999 1,250 ppm on GDs 6–15 (Schwetz et al., 1975). A de crease in fertility index was seen in the 150 7000 and 200 ppm groups in a study of male Swiss-Webster mice exposed via inhalation for 6 weeks 7001 prior to mating (Raje et al., 1988), but the statis tical significance of this effect varied 7002 considerably depending on the statistical test used in this analysis. Two types of developmental 7003 effects, decreased offspring weight at birth and ch anged behavioral habituation of the offspring 7004 to novel environments, were seen in Long-Evans rats following exposure to 4,500 ppm for 7005 14 days prior to mating and during gestation (or du ring gestation alone) (Bornschein et al., 1980; 7006 Hardin and Manson, 1980). This dose was the only e xposure dose used in this study. Schwetz et 7007 al. (1975) did not observe an adverse effect on gro ss development or soft tissue abnormalities in 7008 a study involving exposure to 1,250 ppm on GD 6 in Swiss-Webster mice or Sprague-Dawley 7009 rats, but an increase in delayed ossification of th e sternebrae was seen. 7010 Neurological impairment was not seen in lifetime r odent bioassays involving exposure to 7011

airborne dichloromethane concentrations of ≤ 2,000 ppm in F344 rats (Mennear et al., 1988; 7012 NTP, 1986), ≤ 3,500 ppm in Sprague-Dawley rats (Nitschke et al., 1988a; Burek et al., 1984), or 7013

≤ 4,000 ppm in B6C3F 1 mice (Mennear et al., 1988; NTP, 1986). It should be noted, however, 7014

that these studies did not include standardized neu rological or neurobehavioral testing. The sole 7015 subchronic or chronic study in which neurobehaviora l batteries were utilized found no effects in 7016 an observational battery, a test of hind-limb grip strength, a battery of evoked potentials, or 7017 brain, spinal cord, or peripheral nerve histology i n F344 rats exposed to concentrations up to 7018 2,000 ppm for 13 weeks, with the tests performed be ginning 65 hours after the last exposure 7019 (Mattsson et al., 1990). 7020238

Other effects associated with lifetime inhalation exposure to dichloromethane include 7021 renal tubular degeneration in F344 rats exposed to ≥ 2,000 ppm, testicular atrophy in male 7022

B6C3F 1 mice exposed to 4,000 ppm, and ovarian atrophy in female B6C3F 1 mice exposed to 7023

≥ 2,000 ppm (Mennear et al., 1988; NTP, 1986). No ef fects on histologic, clinical chemistry, 7024 urinalysis, or hematologic variables were found in Syrian golden hamsters exposed to 7025 concentrations up to 3,500 ppm for 2 years, with th e exception that the mean COHb percentage 7026 of exposed hamsters was about 30%, compared with va lues of about 3% in controls (Burek et al., 7027 1984). 7028

7029

5.2.2. Derivation Process for Reference Concentrat ion Values 7030 The derivation process used for the RfC parallels t he process described in section 5.1.2 7031 on the RfD derivation; consideration of dose metric s was described in section 5.1.3. As was 7032 noted in the RfD discussion, the mechanistic issues with respect to noncancer health effects 7033 represents data gaps in the understanding of the he alth effects of dichloromethane. 7034

7035 5.2.3. Methods of Analysis—Including Models (PBTK, BMD, etc.) 7036 The modified rat PBTK model of Andersen et al. (19 91), described in Appendix C and 7037 also used in the derivation of the RfD (Figure 5-2) , was used for calculating internal dosimetry of 7038 inhaled dichloromethane in Sprague-Dawley rats. Si mulations of 6 hours/day, 5 days/week 7039 inhalation exposures used in the Nitschke et al. (1 988a) study were performed to calculate 7040 average daily internal liver doses (Table 5-5). In the absence of data for group- and sex-specific 7041 BWs, reference values for male and female Sprague-D awley rats in chronic studies were used 7042 (U.S. EPA, 1988a). 7043239 7044

Table 5-5. Incidence data for nonneoplastic liver lesions (hep atic vacuolation) and

internal liver doses, based on various metrics, in female Sprague-Dawley rats exposed to dichloromethane via inhalation for 2 years (Nitschk e et al., 1988a)

Rat internal liver dose b

Sex Exposure

(ppm) Liver lesion

incidence a CYP GST GST and CYP Parent

AUC

Male 0 22/70 (31) 50 Not reported 200 Not reported Not modeled because results for middle two doses we re not reported

500 28/70 (40) 0 41/70 (59%) 0 0 0 0 50 42/70 (60%) 280.3 6.3 286.6 1.2 200 41/70 (58%) 656.5 93.2 749.7 17.8 Female (BW = 229 g ) 500 53/70 (76%) c 772.6 359.0 1,131.6 68.7

a Number affected divided by total sample size.

b Internal doses were estimated using a rat PBTK mode l using exposures reported by study authors (50 ppm = 174 mg/m 3 , 200 ppm = 695 mg/m 3 , and 500 ppm = 1737 mg/m 3 ) and are weighted-average daily values for 1 week of exposure @ 6 h/day, 5 day/week. CYP dose is in units of mg dichloromethane metabolized via CYP pathway/L tissue/day; GST dose is in units of mg di chloromethane metabolized via GST pathway/L tissue/ day.; GST and CYP dose is in units of mg dichloromethane metabolized via CYP and GST pathways/L tissue/day; and Parent AUC dose is in units of mg dichlorometh ane*hrs)/L tissue.

c Significantly (

p < 0.05) different from control with Fisher’s exact test.

Source: Nitschke et al., 1988a. 7045 As described in section 5.1.2, the internal dose m etric used was based on total hepatic 7046 metabolism through the CYP2E1 pathway (mg dichlorom ethane metabolized via CYP 7047 pathway/L liver/day). Figure 5-7 shows the comparison between inhalation external and internal 7048 doses, using this dose metric for the rat and the h uman. 7049240 7050

7051 7052 7053 Figure 5-7. PBTK model-derived internal doses (mg dichloromethane metabolized 7054 via the CYP pathway/L liver/day) in rats and humans versus external exposures 7055 (ppm). Average daily doses were calculated from simulated rat exposures of 6 hours/day, 5 7056 days/week, while simulated human exposures were con tinuous. The human metabolism rates 7057 were estimated using a computational sample of 1000 individuals per dose, including random 7058 samples of the three GST-T1 polymorphisms (+/+, +/- , -/-) in the current U.S. population based 7059 on data from Haber et al. (2002). Since a differen t set of samples was used for each dose, some 7060 stochasticity is evident as the human points (value s) do not fall on smooth curves. 7061 7062 The seven dichotomous dose-response models availabl e in EPA BMDS version 2.0 were 7063

fit to the female rat liver lesion incidence of Nit schke et al. (1988a) and PBTK model-derived 7064

internal dose data to derive rat internal BMD 10 and the associated BMDL 10 values (Table 5-6). 7065 The quantal model is identical to the one-stage mul tistage model; therefore, it is not included in 7066

this set of models. A BMR of 10% was selected beca use, in the absence of information 7067 regarding the magnitude of change in a response tha t is thought to be minimally biologically 7068 significant, a BMR of 10% is generally recommended, as it provides a consistent basis of 7069 comparison across assessments. There are no additi onal data to suggest that the critical response 7070 has a greater sensitivity that would warrant a lowe r BMR. The log-probit model was the best 7071 fitting model for the female incidence data, based on AIC value among models with adequate fit. 7072 Modeling results are shown in detail in Appendix D- 2). 7073

7074(mg DCM metabolized/L liver/day)241 7075 Table 5-6. BMD modeling results for incidence of n oncancer liver

lesions in female Sprague-Dawley rats exposed to di chloromethane by inhalation for 2 years, based on liver specific CYP metabolism metric (mg dichloromethane metabolized via CYP pathway per liter liver tissue per day)

Model a BMD 10 BMDL 10 χ 2 goodness of fit

p -value AIC Gamma a 614.27 225.96 0.48 367.22 Logistic 274.58 150.43 0.14 369.77 Log-logistic a 697.90 499.42 0.94 365.90 Multistage (3) a 506.94 153.13 0.25 368.53 Probit 275.49 152.52 0.14 369.75 Log-probit a 728.96 523.94 0.98 365.82 Weibull a 706.45 487.45 0.95 365.87

a These models in EPA BMDS version 2.0 were fit to th e rat dose-response data shown in Table 5-5 by using internal dose metrics calculated with the rat PBTK model. Gamma and Weibull models restrict power ≥ 1; Log-logistic and Log-probit models restrict to s lope >1, multistage model restrict betas ≥ 0; lowest degree polynomial with an adequate fit re ported (degree of polynomial in parentheses). Bolded model is the best-fitting model in the most sensitive sex (males), which is used in the RfC derivation.

Source: Nitschke et al., (1988a). 7076 7077 As with the RfD derivation, the human-equivalent in ternal BMDL 10 was obtained by 7078 dividing this rat internal dose metric by a pharmac okinetic scaling factor based on a ratio of BWs 7079 (scaling factor = 4.09) (Table 5-7). This scaling factor was used because the metric is a rate of 7080 metabolism rather than the concentration of putativ e toxic metabolites, and the clearance of these 7081 metabolites may be slower per volume tissue in the human compared with the rat. A probabilistic 7082 PBTK model for dichloromethane in humans, adapted f rom the model of David et al. (2006) as 7083 described in Appendix B, was then used with Monte Carlo sampling to calculat e distributions of 7084 chronic HECs (in units of mg/m 3 ) associated with the internal BMDL 10 , based on the responses 7085 in female Sprague-Dawley rats. Estimated mean, fir st, and fifth percentiles of this distribution 7086

are shown in Table 5-7. 7087242 7088

Table 5-7. Inhalation RfC for dichloromethane base d on PBTK modelderived probability distributions of human inhalati on exposure extrapolated from nonneoplastic liver lesion data f or female rats exposed via inhalation for 2 years, based on liver-specific CYP metabolism dose metric (mg dichloromethane metabolized via CYP path way per liter liver tissue per day)

HEC (mg/m 3 ) d

Model a Rat

internal BMDL 10 b Human

internal BMDL 10 c 1 st

percentile 5 th

percentile Mean Human RfC

(mg/m 3 ) e

Log-probit 523.94 128.10 16.63 20.89 47.36 0.2

a Based on the best-fitting model from Table 5-6. b Rat dichloromethane PBTK model-derived internal liv er dose associated with lower bound on 10% extra risk for developing hepatocyte vacuola tion.

c Human dichloromethane internal liver dose, derived by dividing the rat internal BMDL 10 by a scaling factor of 4.09 [(BW human /BW rat ) 0.25 ] to account for potential interspecies pharmacokinetic differences in the clearance of met abolites.

d PBTK model-derived distributions of long-term, dail y average airborne dichloromethane concentrations predicted by the PBTK model to yield an internal dose in humans equal to the dichloromethane internal BMDL 10

. e Human candidate RfC, based on female rat data, deri ved by dividing the 1 st percentile of HEC values by a total UF of 100: 3 (10 0.5 ) for possible toxicodynamic differences between species, 3 (10 0.5 ) for variability in human toxicodynamic response, and 10 for database deficiencies. The 1 st percentile point of departure is a stable estimate of the lower end of the distribution. Use of this value in the lower tail replaces use of a UF for human toxicokinetic variability.243 7089 7090 5.2.4. RfC Derivation—Including Application of Unc ertainty Factors (UFs) 7091

The 1 st percentile point of departure is a stable estimate of the lower end of the 7092

distribution. Use of this value associated with a sensitive human population addresses the 7093 uncertainty associated with human toxicokinetic var iability. The RfC was calculated by dividing 7094 the first percentile of the HEC distribution in Tab le 5-7 by a composite UF of 100 (3 [10 0.5 ] to 7095 account for uncertainty about interspecies toxicody namic equivalence, 3 [10 0.5 ] to account for 7096 uncertainty about toxicodynamic variability in huma ns, and 10 for database deficiencies). The 7097 resulting RfC was 0.2 mg/m 3 based on liver lesions in female Sprague-Dawley ra ts in Nitschke et 7098 al. (1988a). In deriving this RfC, factors for the following areas of uncertainty were considered: 7099

7100

• Uncertainty in extrapolating from laboratory animal s to humans (UF A ). The use of 7101 PBTK models to extrapolate internal doses from rats to humans reduces toxicokinetic 7102 uncertainty in extrapolating from the rat liver les ion data but does not account for the 7103 possibility that humans may be more sensitive than rats to dichloromethane due to 7104 toxicodynamic differences. A UF of 3 (10 0.5 ) to account for this toxicodynamic 7105 uncertainty was applied, as shown previously in Tab le 5-7. 7106

7107

• Uncertainty about variation in human toxicokinetics (UF H ). The probabilistic human 7108 PBTK model used in this assessment incorporates the best available information about 7109 variability in toxicokinetic disposition of dichlor omethane in humans but does not 7110 account for humans who may be sensitive due to toxi codynamic factors. Thus, a UF of 7111 3 (10 0.5 ) was applied to account for possible toxicodynamic differences in sensitive 7112

humans. 7113

7114 Uncertainty in extrapolating from LOAELs to NOAEL ( UF L ). A UF for extrapolation from a 7115 LOAEL to a NOAEL was not applied because BMD modeli ng was used to determine the POD, 7116

and this factor was addressed as one of the conside rations in selecting the BMR. The BMR was 7117 selected based on the assumption that it represents a minimum biologically significant change. 7118

• Uncertainty in extrapolating from subchronic to chr onic durations (UF S ). The derived 7119

RfD is based on results from a chronic-duration dri nking water toxicity study. No cross7120 duration UF is necessary. 7121

7122

• Uncertainty reflecting incompleteness of the overal l database (UF D ). A UF of 10 was 7123

selected to address the deficiencies in the dichlor omethane toxicity database. The 7124

inhalation database for dichloromethane includes se veral well-conducted chronic 7125 inhalation studies. In these chronic exposure stud ies, the liver was identified as the most 7126 sensitive noncancer target organ in rats (Nitschke et al., 1988a; NTP, 1986; Burek et al., 7127244

1984). The critical effect of hepatocyte vacuolati on was corroborated in the two principal 7128 studies (Nitschke et al., 1988a; Burek et al., 1984 ), which identified 500 ppm as the 7129

lowest inhalation LOAEL for noncancer liver lesions . Gross signs of neurologic 7130 impairment were not seen in lifetime rodent inhalat ion bioassays for dichloromethane at 7131 exposure levels up to 4,000 ppm (see section 4.2.2. 2 for references), and no exposure7132

related effects were observed in an observational b attery, a test of hind-limb grip 7133 strength, a battery of evoked potentials, or histol ogic examinations of nervous tissues in 7134 F344 rats exposed to dichloromethane concentrations as high as 2,000 ppm (Mattson et 7135 al., 1990). A two-generation reproductive study in F344 rats reported no effect on 7136 fertility index, litter size, neonatal survival, gr owth rates, or histopathologic lesions at 7137

exposures ≥ 100 ppm dichloromethane (Nitschke et al., 1988b). Fertility index (measured 7138 by number of unexposed females impregnated by expos ed males per total number of 7139 unexposed females mated) was reduced following inha lation exposure of male mice to 7140 150 and 200 ppm dichloromethane for 2 hours/day for 6 weeks, but the statistical 7141 significance of this effect varied considerably dep ending on the statistical test used in this 7142 analysis. (Raje et al., 1988). The available devel opmental studies include single-dose 7143 studies that use relatively high exposure concentra tions (1,250 ppm in Schwetz et al. 7144 [1975]; 4,500 ppm in Hardin and Manson [1980]; and 4,500 ppm in Bornschein et 7145 al.[1980]). In one of the single-dose studies, dec reased offspring weight at birth and 7146 changed behavioral habituation of the offspring to novel environments were seen 7147 following exposure of adult Long-Evans rats to 4,50 0 ppm for 14 days prior to mating 7148 and during gestation (or during gestation alone) (B ornschein et al., 1980; Hardin and 7149 Manson, 1980). CO, a known developmental neurotoxi cant, is produced through the 7150 CYP2E1 metabolic pathway for dichloromethane. Schw etz et al. (1975) reported 7151 increased concentrations (~10% higher compared with controls) in maternal blood COHb 7152 levels in mice and rats exposed during GDs 6–15. A chronic exposure study in F344 rats 7153 reported a dose-related increase in blood COHb in f emales exposed to 50, 200, and 500 7154 ppm, beginning with the first measure taken after 6 months of exposure (Nitschke et al., 7155 1988a). The increase was seen at the lowest exposu re group (50 ppm). Anders and 7156 Sunram (1982) reported elevated CO levels in matern al and fetal blood in rats following 7157 exposure to 500 ppm for 1 hour on GD 21; levels wer e similar in the maternal and fetal 7158 samples. Placental transfer of dichloromethane was also seen, although levels were lower 7159 in the fetus. The results from the single dose dev elopmental toxicity study in rats 7160 (Bornschein et al., 1980; Hardin and Manson, 1980), in addition to the known increase in 7161 CO, the placental transfer of dichloromethane, and the relatively high activity of CYP2E1 7162 in the brain compared to the liver of the developin g human fetus (Hines, 2007; Brzezinski 7163 et al., 1999; Johnsrud et al., 2003), raise uncerta inty regarding possible 7164 neurodevelopmental toxicity from gestational exposu re to inhaled dichloromethane. In 7165245

addition, Aranyi et al. (1986) demonstrated evidenc e of immunosuppression following a 7166 single 100 ppm dichloromethane exposure for three h ours in CD-1 mice. This study used 7167 a functional immune assay that is directly relevant to humans (i.e., increased risk of 7168 Streptococcal pneumonia-related mortality and decre ased clearance of Klebsiella 7169 bacteria). No effects were seen with 50 ppm exposu re for either 1 or 5 days. Systemic 7170 immunosuppression was not seen in a 4-week, 5,000 p pm inhalation exposure study, 7171 measuring the antibody response to sheep red blood cells in Sprague-Dawley rats 7172 (Warbrick et al., 2003). These studies suggest a l ocalized, portal-of-entry effect within 7173 the lung rather than a systemic immunosuppression. Therefore, in consideration of the 7174 entire database for dichloromethane, a database UF of 10 was selected. This UF accounts 7175 for the lack of neurodevelopmental toxicity studies and developmental toxicity studies at 7176

low doses. 7177

7178

5.2.5. Previous RfC Assessment 7179 No RfC was derived in the previous IRIS assessment

. 7180

7181 5.2.6. RfC Comparison Information 7182 A candidate RfC, based on a different approach to accounting for human toxicokinetic 7183 variability is similar to thederived RfC of 0.2 mg/ m 3 . Use of the mean value on the HEC 7184 distribution (47.36), with an additional UF of 3 (1 0 0.5 ) to account for human toxicokinetic 7185 variability, would yield an RfC of 0.2 mg/m 3

. 7186 For an additional comparison, an RfC was derived ba sed on neurological endpoints from 7187 human occupational exposures. Cherry et al. (1983) compared 56 exposed and 36 unexposed 7188 workers at an acetate film manufacturing plant for dichloromethane inhalation exposure, blood 7189 levels of dichloromethane, subjective self-reportin g of general health, and two objective, 7190 quantitative measurements of neurological function (digit symbol substitution and simple 7191 reaction time). The exposed and unexposed individu als were matched to within 3 years of age. 7192 The measured dichloromethane concentrations from pe rsonal breathing zone sampling of the 7193 exposed workers ranged from 28 to 173 ppm. No infor mation on exposure duration was given, 7194 and Cherry et al. (1983) did not indicate if the ex posure measurements were indicative of 7195 historical exposure levels. There were no signific ant differences between exposed and 7196 unexposed workers in subjective or objective measur ements collected at the beginning of the 7197 work shift on a Monday (after 2 nonworking days). Exposed workers showed a slightly slower 7198 (but not significant) score than the control worker s on a reaction time test, but the scores did not 7199 deteriorate during the shift. These findings sugge st that repeated inhalation exposures in the 7200 range of 28–173 ppm do not result in significant ef fects, but the actual duration of exposure of 7201

the workers is uncertain. In the absence of data f or the mean exposure levels, the exposure range 7202 midpoint of 101 ppm serves as a NOAEL for chronic n eurological effects from dichloromethane 7203246

exposure. Thus, a candidate RfC of 3.5 mg/m 3 was derived by dividing the NOAEL of 351 7204 mg/m 3 (101 ppm) by a composite UF of 100. A UF of 10 wa s applied to account for potentially 7205 susceptible individuals in the absence of quantitat ive information on the variability of 7206 neurological response to dichloromethane in the hum an population. A UF of 10 was applied for 7207 database deficiencies. The duration of exposures o f acetate film workers (Cherry et al., 1983) 7208 was not reported, and a limited number of endpoints was evaluated. Further, definitive 7209 neurological batteries were not administered in chr onic-duration animal bioassays. 7210 Another candidate RfC was developed by using the n eurological data from the study by 7211 potential long-term CNS effects in a study of retir ed aircraft maintenance workers (Lash et al., 7212 1991). Retired aircraft maintenance workers, ages 55-75 years, employed in at least one of 7213 14 targeted jobs (e.g., paint strippers) with dichl oromethane exposure for 6 or more years 7214 between 1970 and 1984 (n = 25) were compared to a l ike group of workers without 7215 dichloromethane exposure (n = 21). From 1974 to 19 86, when 155 measurements for 7216 dichloromethane exposure were made, mean breathing zone TWAs ranged from 82 to 236 ppm 7217 and averaged 225 ppm for painters and 100 ppm for m echanics; information on exposure levels 7218 prior to this time was not provided. The evaluatio n included several standard neurological tests, 7219 including physiological measurement of odor and col or vision senses, auditory response 7220 potential, hand grip strength, measures of reaction time (simple, choice, and complex), short7221 term visual memory and visual retention, attention, and spatial ability. The exposed group had a 7222 higher score on verbal memory tasks (effect size ap proximately 0.45, p = 0.11) and lower score 7223 on attention tasks (effect size approximately –0.55

, p = 0.08) and complex reaction time (effect 7224 size approximately –0.40, p = 0.18) compared with the control group. None of these differences 7225 were statistically significant. Given the sample s ize, however, the power to detect a statistically 7226 significant difference between the groups was very low (i.e., approximately 0.30 for an effect 7227 size of 0.40 using a two-tailed alpha of 0.05) (Coh en, 1987), and these results cannot be taken as 7228 evidence of no effect. An estimated exposure level from the study can be generated from the 7229 midpoint value from the exposure range (82–236 ppm; mean = 159 ppm), converted to 552 7230 mg/m 3 . If these results are viewed as a LOAEL and this estimated mean exposure level of 552 7231 mg/m 3 was used, a composite UF of 1,000 would be applied for interspecies toxicodynamics 7232 (10), extrapolation from a LOAEL to a NOAEL (10), a nd database uncertainties (10), resulting 7233 in an RfC of 0.55 mg/m 3

. 7234

The value of the candidate RfC based on the data f rom Cherry et al. (1983), 3.5 mg/m 3 , is 7235 approximately 15-fold higher, and the value of the candidate RfC based on the data from Lash et 7236 al. (1991), 0.55 mg/m 3 is approximately three times higher than thederive d RfC of 0.2 mg/m 3

, 7237

based on liver lesions in rats. The animal-derived RfC is preferable to the human-derived RfC 7238 because of the uncertainties about the exposure dur ations for the workers in the Cherry et al. 7239 (1983) study and uncertainties regarding the exposu res and effect sizes in Lash et al. (1991), and 7240

because the RfC based on the rat data is more healt h protective. 7241247

Additional comparisons among the RfC and candidate values developed from other 7242 endpoints or data sets, using NOAEL/LOAEL methods, are shown in Table 5-8 and Figure 5-8. 7243248Table 5-8. Potential points of departure with applied UFs and resulting candidate RfCsUFscEndpointPoint ofdeparture(mg/m3)aPOD Type andDescriptionbTotalUFUFAUFHUFLUFSUFDRfC (mg/m3)ReferenceHepatocyte vacuolation,female rat523BMD, 10% increasein incidence of liverlesion100 3 3 1 1 10 0.2Nitschke et al.(1988a)Renal tubular degeneration;NOAEL, male rat620 NOAEL 1000 3 10 1 1 10 2.07Mennear et al.(1988); NTP (1986)Reproductive - fertilityindex; NOAEL, malemouse20.7No effect at POD,16% decrease infertility index seen atLOAEL dose300 3 10 1 1 10 0.071 Raje et al. (1988)Increased infectionsusceptibility (mortalityrisk), female mouse15.5 NOAEL 3,000 3 10 1 10 10 0.005 Aranyi et al. (1986)Increased IgM production,male and female rat17,366 NOAEL 3,000 3 10 1 10 10 1.03Warbrick et al.(2003)Chronic CNS effects,human male351 NOAEL 100 1 10 1 1 3 3.51 Cherry et al. (1983)CNS changes, human male 552 LOAEL 1,000 1 10 10 1 30.55 Lash et al. (1991)aPOD = point of departure. For Nitschke et al. (1988a), this is based on BMD modeling of a 10% increase in liver lesions using internal liver dose metric (mgdichloromethane metabolism via CYP pathway per liter liver tissue per day) derived from a rat PBTK model. After an allometric scaling factor of 4.09 wasapplied, the human internal BMDL10was 128 mg/m3. A probabilistic human PBTK model adapted from David et al. (2006) was used to generate a distributionof human equivalent concentrations from the human internal BMDL10and the first percentile of this distribution wasused as the point of departure. For otherrodent studies, the NOAEL or LOAEL concentration, in mg/m3, was adjusted to a continuous exposure taking intoaccount hours per day and days per week ofexposure. This adjusted exposure was then converted to an HEC by multiplying the value by a dosimetric adjustment factor (DAF). Blood:air partitioncoefficients were 8.24 for humans, 19.8 for rats, and 23 for mice. Since the blood:air partition coefficients for both the mice and rats were greater than forhumans, a DAF of 1 is recommended and was used. NOAELs or LOAELs were used as points of departure inhuman studies since the concentrations werealready human exposures.bExtra risk defined for incidence data as (Incidence1– Incidence0)/(1-Incidence0), where 1 = dose at observed increased and 0 = background incidencecUFA= uncertainty in extrapolating from laboratory animals to humans, UFH= uncertainty about variation from average humans to sensitive humans, UFL=uncertainty about extrapolating from LOAEL to NOAEL, and UFD= uncertainty reflecting incompleteness of the overall database. A UF extrapolating fromsubchronic to chronic durations (UFS) was not used for any of these studies.Bolded value is the basis of the RfC of 0.2 mg/m3.2497244Figure 5-8. Comparison of candidate RfCs derived from selected point of departures for endpoints presented in7245Table 5-8.7246250

5.3. UNCERTAINTIES IN THE ORAL REFERENCE DOSE AND INHALATION 7247

REFERENCE CONCENTRATION 7248

Risk assessments need to include a discussion of un certainties associated with the derived 7249 toxicity values. For dichloromethane, uncertaintie s related to inter- and intraspecies differences in 7250 toxicodynamics and database deficiencies are treate d quantitatively via the UF approach (U.S. EPA, 7251 1994b). Uncertainties in the toxicokinetic differe nces of dichloromethane between species and within 7252 humans are reduced by application of the PBTK model s for rats and humans. These and other areas of 7253 uncertainty of the derived RfD and RfC are discusse d below. 7254

7255 Adequacy of database for derivation of RfD and RfC 7256 As summarized in sections 4.6.1.1 and 4.6.2.1, data from the available human studies on the 7257 health effects from occupational inhalation exposur es provide some, but not conclusive, evidence of 7258 long-term health consequences of chronic dichlorome thane exposure, specifically with respect to 7259 neurologic and hepatic damage. These data are not adequate for derivation of an RfD or RfC. However, 7260 a broad range of animal toxicology data is availabl e for the hazard assessment of dichloromethane, as 7261 described in chapter 4. The database of oral (Tabl e 4-35) and inhalation (Tables 4-36 and 4-37) toxic ity 7262 studies includes numerous chronic, subchronic, acut e, reproductive, and developmental studies. Liver 7263 toxicity in multiple rodent species is consistently identified as the most sensitive noncancer effect from 7264 oral and inhalation exposure to dichloromethane. I n addition to the oral and inhalation toxicity data

, 7265 there are numerous studies describing the toxicokin etics of dichloromethane. Consideration of the 7266 available dose-response data to determine an estima te of oral exposure that is likely to be without an 7267 appreciable risk of adverse noncancer health effect s over a lifetime has led to the selection of nonca ncer 7268 liver lesions in the 2-year drinking water study in F344 rats (Serota et al., 1986a) as the critical e ffect 7269 and principal study for deriving the RfD for dichlo romethane. The critical effect selected for the 7270 derivation of the chronic RfC is also hepatic lesio ns; two different studies in Sprague-Dawley rats 7271 (Nitschke et al.,1988a; Burek et al., 1984), spanni ng overlapping exposures, reported data on hepatic 7272 vacuolation, and the lower exposure study was chose n as the principal study (Nitschke et al.,1988a). 7273 A critical data uncertainty was identified for neur odevelopmental effects. Animal bioassays 7274 have not identified gross or microscopic effects on neural tissues from long-term exposures or single 7275 (Schwetz et al., 1975) or multigenerational (Nitsch ke et al., 1988b) developmental toxicity studies. 7276 However, behavioral changes were observed in pups b orn to rats exposed to high levels (4,500 ppm) of 7277 dichloromethane (Bornschein et al., 1980; Hardin an d Manson, 1980); lower exposures were not 7278 examined in this study. Uncertainty exists as to t he development of neurological effects from lower 7279 gestational exposures in animals or humans. In add ition, a critical data uncertainty has been identif ied 7280 that relates to potential immunotoxicity, specifica lly immunosuppression seen as a localized portal–of

7281 entry effect within the lung with an acute inhalati on exposure. The lack of data on immune effects fr om 7282251

longer-term exposure represents a significant data gap and is of particular importance because of the 7283 potential importance of immunosuppression with resp ect to response to infections and tumor 7284 surveillance. The weight of evidence for nonneopla stic effects in humans and animals suggests that th e 7285 development of liver lesions is the most sensitive effect, with a UF applied because of the lack of 7286 neurodevelopmental studies and, for the RfC, the un certainty regarding immunotoxicity. 7287

7288 Dose-response modeling 7289 The selection of the BMD model(s) for the quantitat ion of the RfD and RfC does not lead to 7290 significant uncertainty in estimating the point of departure. It should be noted, however, that a lev el of 7291 uncertainty is inherent given the lack of data in t he region of the BMR. 7292

7293 Interspecies extrapolation of dosimetry and risk 7294 The extrapolation of internal dichloromethane dosi metry from nonneoplastic rat responses to 7295 human risk was accomplished using PBTK models for d ichloromethane in rats and humans. 7296 Uncertainties in rat and human dosimetry used for R fD and RfC derivation can arise from uncertainties 7297 in the PBTK models to accurately simulate the toxic okinetics of dichloromethane for animals under 7298 bioassay conditions and humans experiencing relativ ely low, chronic environmental exposures. 7299 There is uncertainty associated with the pharmacok inetic data used for model parameter 7300

estimation and structure validation. The data are primarily measurements of parent dichloromethane 7301 kinetics (e.g., blood or closed-chamber air concent rations over time), rather than measurements of 7302 metabolite levels which can be unambiguously attrib uted to one of the two principal metabolic pathways 7303 (GST and CYP). For the mouse model in particular, only parent dichloromethane data were used, 7304 though exhaled amounts of CO 2 and CO are available. Marino et al. (2006) did in clude data from mice 7305 pre-treated with trans -1,2-dichloroethylene (tDCE), a specific CYP 2E1 in hibitor, but the authors 7306

assumed without verification that 100% of the CYP 2E1 activity was eliminated by the inhi bitor when 7307 using those data. In contrast, Mathews et al. (199 7) found that pretreatment of F344 rats by tDCE (1 00 7308 mg/kg ip) only yielded 65% inhibition of CYP 2E1. If a significant fraction of the CYP 2E1 activity 7309 was not eliminated in the dichloromethane experimen ts, then that activity is erroneously assigned to t he 7310 GST pathway in the parameter estimation Marino et a l. (2006). 7311 In addition to the possibility of incomplete inhibi tion of CYP 2E1 effecting the data 7312 interpretation, the Michaelis-Menten rate equation used in all of the published PBTK models for 7313 dichloromethane, including that of Marino et al. (2 006), has in fact not been shown to accurately 7314

describe the CYP 2E1-mediated metabolism of dichlor omethane in the relevant concentration range. 7315 While Michaelis-Menten kinetics usually describe CY P-mediated oxidation data quite well, the 7316 approach of Marion et al. (2006) implicitly assumes that any metabolism notdescribed by the Michaelis7317 Menten equation is GST-mediated. If pathway-specif ic metabolite data were used to define or bound 7318252

the ratio of GST to CYP metabolism, the resulting e stimates would be less sensitive to what otherwise 7319 might be small errors in the CYP rate equation. Bu t in the modeling of the mouse data by Marino et al

. 7320 (2006), the fraction of total metabolism assigned t o the CYP pathway depends quite strongly on the 7321 assumed form of the CYP rate equation, along with t he assumption of 100% inhibition by tDCE. In 7322 EPA’s modeling of the rat in vivo PK data, using th e same model structure and equations, a set of 7323 parameter values could not be found which described both the parent dichloromethane kinetics and the 7324 total amount of CO exhaled at both high and low exp osure levels; in particular see panel C of Figure C -3 7325 and note discrepancy between model and 50 mg/kg dat a. That the model does not describe well the 7326 dose-dependent shift in metabolism shown by those C O data suggests that the dose-dependence of the 7327 CYP Michaelis-Menten rate-equation is not adequate. As will be shown, an alternative equation for 7328 CYP kinetics may fit the existing dichloromethane d ata better than Michaelis-Menten kinetics, with the 7329 result that a higher portion of total dichlorometha ne metabolism would be interpreted as being CYP7330 mediated. Thus there is uncertainty in the choice of equation for the CYP pathway, which leads to 7331 uncertainty in the estimated GST:CYP metabolic rati o, upon which current risk predictions are based. 7332 The potential error in assuming Michaelis-Menten ki netics for CYP-mediated oxidation of 7333 dichloromethane is reinforced by examining the in v itro oxidative (i.e., CYP-specific) kinetics of 7334 dichloromethane reported by Reitz et al. (1989). W hen extrapolated from in vitro to in vivo, the 7335

apparent

values of the oxidative saturation constant, Km, i dentified by Reitz et al. (1989) for mice, rats, 7336 and humans are over 2 orders of magnitude greater t han those obtained from in vivo PBTK modeling. 7337 Part of the explanation for this apparent discrepan cy lies in the disparate concentration ranges 7338 investigated: Reitz et al. (1989) used much higher dichloromethane concentrations in vitro than those 7339 observed in or predicted for the various in vivo ph armacokinetic studies. In particular, the oxidatio n of 7340

dichloromethane could involve two oxidative processes, one with a high affinity (low Km) 7341 corresponding to the nonlinearity observed in vivo and one with a low affinity (high Km) corresponding 7342 to the nonlinearity observed in vitro. Further, th e low-affinity process would have nearly linear kin etics 7343 in the exposure range used for the in vivo dosimetr y studies and hence be difficult to distinguish fro m 7344 GST-mediated metabolism unless pathway-specific met abolite data are used. One can hypothesize that 7345 this second oxidative process is not inhibited by t DCE and hence corresponds to the 35% of oxidative 7346

metabolism which was observed to remain in rats aft er tDCE treatment by Mathews et al. (1997). 7347 The data of Reitz et al (1989) could simply indicat e a second CYP with low-affinity 7348 dichloromethane activity. However that possibility is contradicted by the results of Kim and Kim (199 6) 7349 who observed that another CYP 2E1-specific inhibito r, disulfiram, completely abolished 7350

dichloromethane-induced increases on COHb in rats. Another possible explanation which would 7351 support the findings observed in Kim and Kim (1996) as well as Reitz et al (1989) and the various in 7352 vivo data is that a number of CYPs exhibit “atypica l” kinetics, not described by the classic Michaelis

7353 Menten equation, consistent with the enzymes having dual binding sites as proposed by Korzekwa et al 7354253

(1988). (Korzekwa et al. (1988) demonstrated atypi cal kinetics for several CYP-isozyme/substrate 7355 pairs, but not specifically for CYP 2E1.) Figure 5 -9 shows kinetic model fits to the in vitro mouse 7356

dichloromethane oxidation kinetic data of Reitz et al. (1989), after expressing those data on a per gr am 7357

of liver basis. Both the standard Michaelis-Menten kinetic equation (solid line) and the dual-binding 7358 equation (dashed line) given by Korzekwa et al. (19 88) are shown. In particular, the high-affinity (l ow) 7359 Km for the dual-binding equation was set equal to t hat obtained by Marino et al. (2006) from their 7360 PBTK modeling. This figure shows that the dual bin ding model is not only consistent with the apparent 7361 high-affinity saturation obtained from in vivo PBTK modeling (Km of Marino et al. (2006)), but also 7362 with the apparent low-affinity (high Km) data of Re itz et al. (1989), and describes those in vitro dat a 7363 better than the standard Michaelis-Menten equation. (Reitz et al. (1989) used classic Lineweaver-Burk 7364 plots to display their kinetic data; i.e., 1/reacti on rate vs. 1/concentration. The systematic discre pancy 7365

between their data and Michaelis-Menten kinetics ev ident in Figure 5-9 is much less obvious with that 7366 scaling, which likely explains why they made no not e of it.) 7367

7368

0 0.5 1 1.5 2 2.5 3

0 100 200 300 400 500 [DCM] (mg/L)CYP rxn rate (mg/h/g liver)

Reitz et al. (1989) data

Michaelis-Menten kinetics

Dual-binding CYP kinetics

7369 Figure 5-9. Comparison of dichloromethane oxidatio n rate data with alternate 7370 kinetic models. Dichloromethane (DCM) oxidation dat a obtained with mouse liver 7371 microsomes by Reitz et al. (1989) (points), express ed on a per gram of liver basis, 7372 are shown with a fitted Michaelis-Menten equation ( solid line) or a fitted dual7373 binding-site equation as described by Korzekwa et a l. (1988) (dashed line), where 7374 the high affinity saturation constant of the dual-b inding-site equation set equal to 7375 the mean Km determined for mice via PBTK modeling b y Marino et al. (2006). The 7376 Km for the Michaelis-Menten equation (108 mg/L) is inconsistent with the in vivo 7377254

DCM dosimetry data, while the in vitro data shown h ere are inconsistent with the 7378 Km estimated in vivo (0.42 mg/L) if that equation i s used. 7379 7380 In summary regarding model equations, the current P BTK model used the standard Michaelis7381 Menten equation to describe CYP 2E1-catalyzed oxida tion of small volatile organic compounds. 7382 Analysis of the dichloromethane (pharmaco)kinetic d ata and evaluation of the inconsistencies describe 7383 above suggest that an alternate equation, which wou ld impact risk predictions, may better represent CY P 7384 2E1-induced oxidation of dichloromethane. However, this hypothesis requires further laboratory 7385 testing, for example, by measuring dichloromethane oxidation in a bacterial expression system where 7386 only CYP 2E1 is expressed over a concentration rang e sufficient to firmly distinguish between the two 7387 kinetic forms indicated in the figure above. Until such experiments are conducted, the existing PBTK 7388

model remains the best available science for dose- and hence risk-extrapolation from rodents to humans

. 7389 Still, this model structure uncertainty implies unc ertainty in the quantitative results obtained with the 7390 model. Analysis of the GST-mediated metabolism of dichloromethane measured by Reitz et al. (1988) 7391

shows that those results are within a factor of thr ee of the GST kinetic parameters used in the curren t 7392 PBTK model, indicating that the any error in the GS T:CYP balance is no greater than that, a reasonable 7393 level of uncertainty. 7394 One other component of quantitative uncertainty ari ses in examining the results of the Bayesian 7395 modeling for the human PBTK model of David et al. ( 2006). The authors reported Bayesian posterior 7396 statistics for the population average of each fitte d parameter when calibration was performed either w ith 7397 specific published data sets or the entire combined data set. While one would generally expect that t he 7398

values obtained from the combined data set should b e a weighted average of the values from individual 7399 data sets, the population mean for the liver GST ac tivity (coefficient), K FC , was 0.852 while the values 7400 from the individual data sets ranged from 1.92-34.0 kg 0.3 /h. 7401 A clarification provided by D. Marino (personal com munication) 8 is that the parameter bounds 7402 stated in the text of David et al. (2006) were only applied for the analysis of the DiVincenzo and Kap lan 7403 (1981) and the combined data set. But according to the text and distribution prior statistics specifi ed, the 7404 upper bound for K FC would have been 12 kg 0.3 /h (mean + 2.5 standard deviations (SDs), with mean = 2 7405 and SD = mean*CV = 2*2 = 4). The data of Andersen et al. (1991) were not used in the combined 7406 analysis because only group average values were ava ilable from that source, rather than individual dat a. 7407 Since the remaining study-specific mean K FC values were 7.95, 5.87, 34.0, and 1.92, with CVs o f less 7408 than 2, it seems unlikely that application of this upper bound would result in a value of K FC of only 7409 0.852 kg 0.3 /h. Given that there had been convergence problems with the combined data set when 7410 parameter values were unbounded, it is possible tha t convergence had not actually been reached after 7411

8 Email from Dale Marino to Glinda Cooper dated Apri l 25, 2007.255

parameter bounds were introduced, and a higher valu e for K FC would have been obtained had the chain 7412 been continued longer. 7413 Since the numerical average of the mean K FC values for the four data sets included in the 7414 combined data set was 12.4 and the upper bound was 12, the impact of using an intermediate value of 7415

K FC , specifically the DiVincenzo and Kaplan value of 5 .87 kg 0.3 /h was explored. Changing only the K FC 7416

is not realistic since the dichloromethane data eff ectively define total metabolism (sum of CYP and GS T 7417 pathways) and there is naturally a negative correla tion between the predicted CYP metabolic rate and 7418 the GST metabolic rate, required to describe this t otal. Therefore, it would be inconsistent with the 7419

dichloromethane data to increase K FC without adjusting the CYP metabolic rate downward, and likewise 7420 all other parameters. The distributions for all of the fitted parameters were rescaled by the ratio o f the 7421 mean for DiVincenzo and Kaplan (1981) to the mean f or the combined data set (e.g., the distribution fo r 7422

K FC was multiplied by 5.87/0.852, the ratio of the two posterior means). The resulting predicted upper7423 bound (95 th and 99 th percentile) GST metabolism rates for a fixed level of exposure (1 μ g/m 3 inhaled 7424 concentration or 1 mg/kg/day oral exposure) in the GST-T1 +/+ population increased by more than a 7425 factor of 10. (For inhalation exposure the mean va lue also increased by over 10-fold, but for oral 7426 exposure the mean increased by only 2-fold.) Since the majority of metabolism occurs via the CYP 7427 pathway at these low levels, there is not a proport ionate (i.e., over 10-fold) decrease in that rate, but 7428 HEC and HED calculations increased by 10-30% for th e mixed GST-T1 population, depending on the 7429 route of exposure and distribution statistic compar ed. Thus the impact of this model uncertainty appe ars 7430 to be relatively small for the noncancer assessment , but quite large for the cancer assessment. 7431

The dose metric used in the models is the rate of m etabolism to a putative toxic metabolite, 7432 rather than the concentration average or area under the concentration curve of the metabolite, so the 7433 model specifically fails to account for rodent-huma n differences in clearance or removal of the toxic 7434 metabolite. A scaling factor based on BW ratios, w as used to account for this difference. 7435

The rat model was modified and utilized in a deter ministic manner. Data were not available to 7436 perform a hierarchical Bayesian calibration in the rat. Thus, uncertainties in the rat model predicti ons 7437 had to be assessed qualitatively. To address these uncertainties, a sensitivity analysis was conducte d to 7438 determine which model parameters most influence the predictions for a given dose metric and exposure 7439

scenario. 7440 Sensitivity is a measure of the degree to which a g iven model output variable (i.e., dose metric) 7441 is influenced by perturbation in the value of model parameters. The approach implemented was a 7442 univariate analysis in which the value of an indivi dual model parameter was perturbed by an amount 7443 ( ∆ ), in the forward and reverse direction (i.e., an i ncrease and decrease from the nominal value), and t he 7444 change in the output variable was determined. Sens itivity coefficients were calculated as follows: 7445

7446256 ) ( ) ( ) ( ) ( '

x f x

x x f x x f x f

⋅

∆ − ∆ +

≈ (Eq. 5-1) 7447

7448

where x

is the model parameter f(x) is the output variable, ∆ x is the perturbation of the parameter from 7449 the nominal value, and f’(x) is the sensitivity coefficient. In equation 5-1, the sensitivity coefficients are 7450

scaled to the nominal value of x and f(x) to eliminate the potential effect of units of expr ession. 7451 Therefore, the sensitivity coefficient is a measure of the proportional (unitless) change in the outpu t 7452 variable produced by proportional change in the par ameter value. Parameters that have higher 7453 sensitivity coefficients have greater influence on the output variable. They are considered more 7454 sensitive than parameters with lower values. The r esults of the sensitivity analysis are useful for 7455 assessing uncertainty in model predictions, based o n the level of confidence or uncertainty in the mod el 7456 parameter(s) to which the dose metric is most sensi tive. 7457 Sensitivity coefficients for the noncancer dose me tric (mg dichloromethane metabolized via 7458 CYP-mediated pathway/L liver/day), were determined for each of the model parameters; a similar 7459 analysis was also done for a metric based on the GS T-mediated pathway. Sensitivity analyses for both 7460 oral and inhalation exposures were performed. The exposure conditions were set to be near or just 7461 below the lowest bioassay exposure resulting in sig nificant increases in the critical effect. 7462 For the CYP-mediated metabolism from oral exposure, the VLC and VSC (liver volume and 7463 slowly perfused tissue volume, respectively) parame ters exert the largest influence (Figure 5-10). Th e 7464 high influence of these two parameters was due to t he fact that the dose metric is a tissue-specific r ate of 7465 metabolism, the majority of CYP metabolism is attri buted to the liver, and that changes in liver volum e 7466 have a greater impact on the total CYP metabolism t hat the individual Vmax value. For inhalation 7467 exposures VMAXC, in addition to VLC and VSC have th e highest sensitivity coefficients (Figure 5-11). 7468 The physiological parameters (VLC and VSC) are know n with a high degree of confidence (Brown et 7469 al., 1997). V MAXC for the rat was estimated by fitting to the PK dat a as described in Chapter 3 and 7470 Appendix C, subject to model structure/equation unc ertainties as detailed above, and hence is known 7471 with less certainty than the physiological paramete rs. That total exhaled CO, which is a specific to the 7472 CYP pathway, is within 50% of measured levels (Fig. C-8, panel C), however, provides a similar level 7473 of confidence in the balance between CYP and GST pa thways predicted by the rat PBTK model. 7474257 Oral exposure-1.25-1-0.75-0.5-0.2500.250.50.7511.25 QCC VPR VLC VSC PB VMAXC KA A2 KFCParameter

Normalized sensitivity coefficient CYP

GST

7475 Figure 5-10. Sensitivity coefficients for long-ter m mass CYP- and GST-mediated 7476 metabolites per liver volume from a daily drinking water concentration of 10 mg/L in rats. 7477 KFC = GST-mediated metabolism rate; A2 = proporti on of liver GST metabolism attributed to 7478 the lung; KA = oral absorption rate from gut; VMAXC = CYP-mediated maximum rate of 7479 metabolism; PB = blood:air partition coefficient; V SC = slowly perfused tissue volume; VLC = 7480 liver volume; VPR = Ventilation perfusion ratio ; Q CC = cardiac output constant. 7481 7482

Inhalation exposure-1.25-1-0.75-0.5-0.2500.250.50.7511.25 QCC VPR VLC VSC PB VMAXC A2 KFCParameter

Normalized sensitivity coefficient CYP

GST

7483 Figure 5-11. Sensitivity coefficients for long-ter m mass CYP- and GST-mediated 7484 metabolites per liver volume from a long-term avera ge daily inhalation concentration of 7485 500 ppm in rats. (KA is not included since it has no impact on inhalation dosimetry.) KFC 7486 = GST-mediated metabolism rate; A2 = proportion o f liver GST metabolism attributed to the 7487 lung; VMAXC = CYP-mediated maximum rate of metaboli sm; PB = blood:air partition 7488 coefficient; VSC = slowly perfused tissue volume; V LC = liver volume; VPR = Ventilation 7489 perfusion ratio ; QCC = cardiac output constant. 7490 7491258

In summary, the uncertainties associated with use o f the rat PBTK model should not markedly 7492

affect the values of the RfD and RfC based on the m etrics considered. An additional uncertainty resul ts 7493 from the lack of knowledge concerning the most rele vant dose metric (e.g., a specific metabolite) for the 7494 non-cancer endpoints considered. This basic resear ch question represents a data gap. This uncertaint y 7495 was addressed by considering different dose metrics (CYP metabolism alone, GST metabolism alone, 7496 sum of GST and CYP, and the AUC of the parent compo und). The GST metabolism and the AUC dose 7497 metrics did not present reasonable choices based on model fit and consistency of response across studi es 7498 at comparable dose levels. Given these results, th e combination of hepatic metabolism through the GST 7499 and the CYP pathways would not be expected to resul t in an improvement to a metric based only on 7500

CYP metabolism. The CYP-metabolism dose metric see ms to be most consistent with the data., and so 7501

is the metric chosen for the RfD and RfC derivation s. 7502

7503 Sensitive human populations 7504 The potential for sensitivity to dichloromethane i n a portion of the human population due to 7505 pharmacokinetic differences was addressed quantitat ively by using a human probabilistic PBTK model, 7506 as modified by the EPA, to generate distributions o f human exposures likely to result in a specified 7507

internal BMDL 10 . The model and resulting distributions take into account the known non-chemical7508 specific variability in human physiology as well as total variability and uncertainty in dichlorometha ne7509 specific metabolic capability. The first percentil e values of the distributions of human equivalent d oses 7510 (Table 5-3) and HECs (Tables 5-7) served as points of departure for candidate RfDs and RfCs, 7511 respectively, to protect toxicokinetically sensitiv e individuals. Selection of the first percentile a llows 7512 generation of a stable estimate for the lower end o f the distribution. The mean value of the human 7513 equivalent oral dose in Table 5-3 was about two-fol d higher than the corresponding first percentile 7514 values, and the mean value of human equivalent inha lation concentration in Table 5-7 was 7515 approximately three-fold higher than the first perc entile value. The internal dose metric in the analyses 7516 described in these tables was the mg dichloromethan e metabolized via the CYP pathway per liter liver 7517 per day, and thus the comparisons of the first perc entile and mean values give estimates of the amount of 7518 variability in the population to metabolize dichlor omethane by the CYP metabolic pathways on a liver7519 specific basis. The mean: 1 st percentile ratios for these distributions is attri buted to the dependence of 7520 the dose metric on hepatic blood flow rate (metabol ism being flow limited). This blood flow is expect ed 7521 to be highly and tightly correlated with liver volu me, resulting in very similar delivery of 7522 dichloromethane per volume liver across the populat ion. While the mean: 1 st percentile ratios for the 7523

oral distribution is less than the default intra-hu man toxicokinetic UF of 3, it is quite similar to t hat 7524 obtained by Sweeney et al. (2003) for acrylonitrile , where an extensive sensitivity analysis indicated a 7525

99 th percentile:mean ratio of less than 2.2 among sever al internal dose metrics. The population7526 structured distributions for physiological paramete rs and broadened distributions for metabolic 7527259

parameters used here provide a good degree of confi dence that the population variability has not been 7528

under-estimated. 7529

The internal dose metric used in the RfD and RfC de rivations was based on the rate of CYP 7530 metabolism. GST-T1 polymorphisms could affect this rate, as the GST-T1 null genotype would be 7531 expected to result in an increase in the metabolism through the CYP pathway, resulting in a greater 7532 sensitivity to a CYP-related effect. The effect of GST variability on the RfD and RfC values was 7533 examined by comparing results obtained specifically for the GST-T1 null genotype to those obtained for 7534 the population of mixed genotypes. The values for human equivalent doses and HECs were very similar 7535 for these two groups (e.g., mean HEC 47.36 and 47.4 9 for the mixed and the GST-T1 -/- null genotypes, 7536 respectively; 1 st percentile HEC 16.63 and 16.69 for the mixed and t he GST-T1 -/- null genotypes, 7537 respectively), and use of this population would not result in a change in the recommended RfD or RfC. 7538 As a further level of sensitivity analysis, we comp ared model predictions of the human 7539 equivalent dose, as listed in Table 5-3, for the ge neral population (estimates covered 0.5- to 80-year -old 7540 male and female individuals) to three subpopulation s: 1-year-old children (males and females), 70-year

7541 old men, and 70-year-old women. For the general po pulation and each subpopulation a Monte Carlo 7542 simulation representing 10,000 individuals was cond ucted, and histograms of the resulting distribution 7543 of human equivalent administered doses are shown in Figure 5-12, with corresponding statistics in Tabl e 7544 5-9. All groups used in these comparisons were lim ited to the GST-T1 -/-

. 7545 The results shown above for differences in human eq uivalent dose values in different populations 7546 are qualitatively what would be expected: a relativ ely broad distribution for the general population w ith 7547 specific populations representing narrower componen ts of that distribution. There are some difference s 7548 between men and women at 70 years of age, but neith er of these would be greatly misrepresented by the 7549 general population estimate. While 1-year-old chil dren represent more of a distinct tail in the gener al 7550 population, in this case the distribution of human equivalent concentrations in the general estimate i s 7551

lower than that seen in what would otherwise be con sidered a more sensitive population. This 7552 difference most likely results from the higher spec ific respiration rate in children versus adults, wh ich 7553 allows them to eliminate more of orally ingested di chloromethane by exhalation, leading to lower 7554

internal metabolized doses 7555260 0 1 2 3 4 5 6 7 8 9 10

0.1 0.3 0.5 0.7 0.9 1.1

Human equivalent applied dose (mg/kg-day)Frequency density (fraction/unit dose)

General 70 yo Male 70 yo Female 1 yo Child Human equivalent applied dose distributions

7556 Figure 5-12. Frequency density of human equivalent applied doses in specific populations 7557 in comparison to a general population (0.5- to 80-y ear-old males and females) estimate for 7558 an internal dose of 15.1 mg dichloromethane metabol ized by CYP per liter liver per day; all 7559 groups were restricted to the GST-T1 -/- population). 7560 7561 7562

Table 5-9. Statistical characteristics of human equivalent applied doses in specific populations of the GST-T1 -/- group

Human equivalent applied dose (mg/kg-day) a Population Mean 5 th percentile 1 st percentile All ages b 3.95 x 10 -1 2.52 x 10 -1 2.14 x 10 -1 1-year-old children 6.34 x 10 -1 4.87 x 10 -1 4.54 x 10 -1 70-year-old men 3.18 x 10 -1 2.48 x 10 -1 2.29 x 10 -1 70-year-old women 2.64 x 10 -1 2.03 x 10 -1 1.85 x 10 -1

a Exposure levels predicted to result in 15.1 mg dich loromethane metabolized via CYP pathway per liter liver per day (based on mean BMDL 10 across acceptable models from Table 5-3).

b 0.5- to 80-year-old males and females. 7563

. 7564 A similar comparison was made for inhalation HEC va lues, as shown in Figure 5-13 and Table 7565 5-10. For HEC values, the distributions for 70-yea r-old men and women are both virtually 7566 indistinguishable from the general population, and while 1-year-old children are clearly distinct, the y are 7567261

less different than in the human equivalent adminis tered dose comparison and in this case are more 7568 sensitive than the population in general. As descr ibed in detail in Appendix B, the allometric alveol ar 7569 ventilation constant, QAlvC, is about 28 L/hour-kg 0.75 in a 1-year-old child but averages around 14 7570 L/hour-kg 0.75 in an adult. Combining this with the difference b etween a BW of 10 kg in that child and 7571 70 kg in an “average” adult, the respiration rate p er kg BW is about threefold higher in the child ver sus 7572 adult. As noted above, for oral exposures this lea ds to faster elimination by respiration in children

, 7573 while for inhalation exposures it leads to higher u ptake for a given air concentration. 7574 The lack of difference in elderly adults versus the general population in HEC values is likely due 7575 to the fact that the rate of exposure and rates of metabolism (the latter being the key dose metric) b oth 7576

scale as BW 0.75 , with the scaling coefficients being either simila r (respiration) or identical (metabolism) 7577 among adults, who comprise the majority of the popu lation, while for oral exposures the exposure rate is 7578

normalized to total BW and scales as BW 1 , while elimination routes increase as BW 0.75 . Moreover, oral 7579 exposures are simulated as occurring in a series of bolus exposures (drinking episodes) during the day

, 7580 and the higher body-fat content occurring in the el derly (see Appendix B) means that such a dose that 7581 might saturate metabolism and therefore have a high er fraction exhaled in a leaner individual will ten d 7582 to be more sequestered in fat and slowly released, resulting in a higher fraction metabolized (less 7583 saturation of metabolism) in a more obese individua l. The difference among adults of different ages f or 7584 dosimetry from oral ingestion (bolus exposure) will be greater than the difference for inhalation 7585 exposures. More careful examination of Figure 5-13 shows that the distribution for 70-year-old women, 7586 for whom the fat fraction is estimated to be greate st, has a lower peak and higher upper tail than for the 7587 general population. So the physiological differenc es do have some impact that is qualitatively consis tent 7588 with what is seen from oral exposure, given the mec hanistic considerations described here. But the 7589 impact of those differences is far less for inhalat ion exposure. 7590

7591262 0 0.02 0.04 0.06 0.08 0.1 0.12 0.14 0.16 0.18 0.2

0 10 20 30 40

Human equivalent concentration (ppm)Frequency density (fraction/ppm)

General population 1 year old 70 yo male 70 yo female Human equivalent concentration distributions

7592 Figure 5-13. Frequency density of HECs in specific populations in comparison to a 7593 general population (0.5- to 80-year-old males and f emales) estimate for an internal 7594 dose of 128.1 mg dichloromethane metabolized by CYP per liter liver per day ; all 7595 groups restricted to the GST-T1 -/- population). 7596 7597

Table 5-10. Statistical characteristics of HECs in specific populations of the GST-T1 -/- group

HEC (mg/m 3 ) a Population Mean 5 th percentile 1 st percentile All ages b 47.4 20.9 16.6 1-year-old children 24.7 14.4 12.1 70-year-old men 46.4 21.7 17.8 70-year-old women 50.0 22.2 18.0

a Exposure levels predicted to result in 128.1 mg dic hloromethane metabolized via CYP pathway per liter liver per day (based on mean BMDL 10 across acceptable models from Table 5-7).

b 0.5- to 80-year-old males and females. 7598 7599 No data are available regarding toxicodynamic diffe rences within a human population. 7600 Therefore, a UF of 3 for possible differences in hu man toxicodynamic responses is intended to be 7601 protective for sensitive individuals. 7602

7603

5.4. CANCER ASSESSMENT 7604263

5.4.1. Cancer Oral Slope Factor 7605

5.4.1.1. Choice of Study/Data—with Rationale and Justificati on 7606 No human data are available for the quantification of potential neoplastic effects from oral 7607 exposures to dichloromethane. In the only chronic (2-year) oral exposure cancer bioassay, significant 7608

increases in the incidence of liver adenomas and ca rcinomas was observed in male, but not female, 7609

B6C3F 1 mice exposed by drinking water, with incidence rat es of 19, 26, 30, 31, and 28% in groups with 7610 estimated mean intakes of 0, 61, 124, 177, and 234 mg/kg-day, respectively (trend p- value = 0.058) (see 7611 Table 4-38 for group comparisons) (Serota et al., 1 986b; Hazelton Laboratories, 1983). Evidence of a 7612 trend for increased risk of liver tumors (described as neoplastic nodule or hepatocellular carcinoma) was 7613 seen in female F344 rats, but not males, exposed vi a drinking water (p < 0.01) (Serota et al., 1986a). 7614 However, the potential malignant characterization o f the nodules was not described, and no trend was 7615 seen in the data limited to hepatocellular carcinom as. The derivation of the cancer oral slope factor 7616 (OSF) is based on the male mouse data (Serota et al ., 1986b; Hazelton Laboratories, 1983) because of 7617 their greater sensitivity compared to female mice a nd to male and female rats. The study authors 7618

concluded that these increases were “within the nor mal fluctuation of this type of tumor incidence”. 7619 However, the trend p -value for these results is of borderline statistic al significance and it may not be 7620 reasonable to apply a correction for multiple compa risons given the lack of independence of the groups 7621 and given a specific focus on the liver as a target organ. The development of liver tumors in B6C3F 1 7622 mice is associated with metabolite production in th is tissue via the GST metabolic pathway (section 7623 4.7.3), a pathway that also exists in humans. Model ing intake, metabolism, and elimination of 7624

dichloromethane in mice and humans is feasible. Th us, it is reasonable to apply the best available 7625 PBTK models to estimate equivalent internal doses i n mice and humans. 7626

7627

5.4.1.2. Derivation of Oral Slope Factor 7628

In a manner similar to the derivation of the nonca ncer toxicity values, PBTK models for 7629

dichloromethane in mice and humans were used in the derivation of toxicity values (cancer OSF and 7630 IUR) for cancer endpoints based on lung (for inhala tion) and liver (for oral and inhalation) tumor dat a in 7631 the mouse (Figure 5-14). A deterministic PBTK mode l for dichloromethane in mice was first used to 7632 convert mouse drinking water or inhalation exposure s to long-term daily average values of internal lun g7633 specific GST metabolism (GST metabolism in lung/lun g volume) or liver-specific GST metabolism 7634 (GST metabolism in liver/liver volume). The choice of this dose metric was made based on data 7635 pertaining to the mechanism(s) involved in the carc inogenic response, specifically data supporting the 7636 involvement of a GST metabolite(s). The evidence pe rtaining to the GST pathway is discussed in section 7637 4.7, and includes the enhanced genotoxicity seen in bacterial and mammalian in vitro assays with the 7638 introduction of GST metabolic capacity (Graves et a l., 1994a) and the suppression by pretreatment with 7639264

a GSH depletory of the production of DNA SSBs seen in acute inhalation exposure to dichloromethane 7640 in mice (Graves et al., 1995). 7641

7642265Recommend valueDistribution of Human CancerOral Slope Factors orInhalation Unit Risks95th99thMonte CarloSampling fromDistributions ofHuman PBTKModel ParametersDistribution ofHumanInternalDoses from Unit Oral Doses(1mg/kg) or InhalationConcentrations (1ug/m3)ProbabilisticHuman PBTKModelMultiply Human Tumor Risk FactorBy Distribution of Human InternalUnit DosesRodent DoseResponse DataRodentPBTK ModelEstimates of RodentInternal DoseBMDModelingBMDBMDLRodent Internal BMDL1095% Lower Bound Estimate of InternalDose Associated with a 10% response(0.1/Rodent BMDL10)Rodent Tumor Risk Factor(internal dose)-1Human Tumor Risk Factor(internal dose)-1ScalingFactor+Apply Age-Dependent Adjustment Factors(ADAFs) for early life exposure76437644Figure 5-14. Process for deriving cancer OSFs andIURs by using rodent and human PBTK models.7645266

The multistage cancer model (using BMDS version 2. 0) was then fit to the tumor 7646 incidence data and internal dose data for rodents, and BMD 10 and associated BMDL 10 values (for 7647 a BMR of 10% extra risk) were calculated. A probab ilistic PBTK model for dichloromethane in 7648 humans, adapted from David et al. (2006) (see Appen dix B), was used with Monte Carlo 7649 sampling to calculate distributions of internal lun g or liver doses associated with chronic unit 7650 oral (1 mg/kg-day) or inhalation (1 μ g/m 3 ) exposures. The resulting distribution of human 7651 internal doses was multiplied by a human internal d ose tumor risk factor (in units of reciprocal 7652 internal dose) to generate a distribution of OSFs o r IURs associated with a chronic unit oral or 7653 inhalation exposure, respectively. 7654 The parameter statistics reported by David et al. ( 2006) include both the inter-individual 7655 variability that would have been elucidated by the Bayesian analysis (variation between mean 7656

values for each individual for which data were avai lable) and uncertainty in those values. Since 7657 EPA’s objective is to account for both population v ariability and parameter uncertainty, 7658 however, these statistics were primarily as-is (exc eptions discussed in Appendix B) to define 7659 population distributions. Assuming that these para meters are distributed independently, ignoring 7660 the covariance that was likely represented in the a ctual posterior chains, will tend to over7661 estimate the overall range of parameters and hence distribution of dose metrics in the population, 7662 compared to what one would obtain if the covariance were explicitly included. Thus if the 7663 covariance (i.e., the variance-covariance matrix) f or the set of parameters had been reported by 7664 David et al., it could have been used to narrow the predicted distribution of internal doses, or 7665 equivalent applied doses. Lacking such information the approach used will not under-estimate 7666 risk or over-estimate lower bounds on human equival ent exposure levels. 7667

7668

5.4.1.3. Dose-Response Data 7669

Data for liver tumors in male B6C3F 1 mice following exposure to dichloromethane in 7670 drinking water were used to develop oral cancer slo pe factors (Serota et al., 1986b; Hazelton 7671 Laboratories, 1983). Significant increases in inci dence of liver adenomas and carcinomas were 7672 observed in male, but not female, B6C3F 1 mice exposed for 2 years (Table 5-11). No signifi cant 7673

decreases in survival were observed in the treated groups of either sex compared with controls. 7674 The at-risk study populations (represented by the d enominators in the incidence data) were 7675 determined by excluding all animals dying prior to 52 weeks. 7676

7677

5.4.1.4. Dose Conversion and Extrapolation Methods: Cancer O ral Slope Factor 7678

Dose conversion 7679 The mouse PBTK model of Marino et al. (2006) was b ased on the PBTK model for 7680 dichloromethane by Andersen et al. (1987), which wa s modified to include dichloromethane 7681 metabolism in the lung compartment and kinetics of CO and COHb (Andersen et al., 1991). For 7682 the mouse, physiological parameters and partition c oefficients were adjusted to match those 7683267

reported in Andersen et al. (1991, 1987) and Clewel l et al. (1993), respectively, while QCC, 7684 VPR, and metabolic parameter distribution mean valu es were derived via MCMC model 7685 calibration reported by Marino et al. (2006) (Appen dix B). The model of Marino et al. (2006) 7686 was used to simulate daily drinking water exposures comprising six discrete drinking water 7687 episodes for specified times and percentage of tota l daily intake (Reitz et al., 1997) and to 7688 calculate average lifetime daily internal doses for the male mouse data shown in Table 5-11. A 7689 first-order oral uptake rate constant (k a ) of 5 hours –1 was taken from Reitz et al. (1997) to 7690 describe the uptake of dichloromethane from the gas trointestinal tract to the liver. Study-specific 7691 BWs were not available, so reference BWs for male B 6C3F 1 mice in chronic studies (U.S. EPA, 7692 1988a) were used. Based on evidence that metabolit es of dichloromethane produced via the 7693 GST pathway are primarily responsible for dichlorom ethane carcinogenicity in mouse liver 7694 (summarized in section 4.7.3) and the assumption th at these metabolites are sufficiently reactive 7695 that they do not have substantial distribution outs ide the liver, the recommended selected internal 7696 dose metric for liver tumors was daily mass of dich loromethane metabolized via the GST 7697 pathway per unit volume of liver (Table 5-11). Figu re 5-15 shows the comparison between 7698

internal and external doses in the liver in mice an d humans. The whole-body metabolism metric 7699

was also examined. This metric would be more relev ant under a scenario of slowly cleared 7700 metabolites that undergo general circulation. 7701

7702

Table 5-11. Incidence data for liver tumors and in ternal liver doses,

based on GST metabolism dose metrics, in male B6C3F 1 mice exposed to dichloromethane in drinking water for 2 years

Sex Nominal (actual)

daily intake (mg/kg-day) Mouse liver

tumor incidence a Mouse internal

liver metabolism dose b Mouse whole body

metabolism dose c 0 (0) 24/125 (19%) 0 0 60 (61) 51/199 (26%) 17.5 0.73 125 (124) 30/99 (30%) d 63.3 2.65 185 (177) 31/98 (32%) d 112.0 4.68 Male (BW = 37.3 g)

250 (234) 35/123 (28%) d 169.5 7.1

a Hepatocellular carcinoma or adenoma, combined. Mic e dying prior to 52 weeks were excluded from the denominators. Cochran-Armitage trend p -value = 0.058.

b mg dichloromethane metabolized via GST pathway/L li ver/day. Internal doses were estimated from simulations of actual daily doses reported by the study authors.

c Based on the sum of dichloromethane metabolized via the GST pathway in the lung plus the liver, normalized to total BW (i.e., [lung GST meta bolism (mg/day) + liver GST metabolism (mg/day)]/kg BW. Units = mg dichloromethane metabo lized via GST pathway in lung and liver/kg-day.

d Significantly (

p ≤ 0.05) different from control incidence by Fisher’s exact test performed by Syracuse Research Corporation.

Sources: Serota et al., 1986b; Hazelton Laboratori es, 1983 7703268 

### Liver GST dose for oral exposure

1 10 100 1000

10 100 1000 Dose (mg/kg/day) Mouse

Human mixed

Human +/-

Human +/+Metabolic rate (mg metabolized/L liver/day)

7704 Figure 5-15. PBTK model-derived internal doses (mg dichloromethane metabolized via 7705 the GST pathway per liter liver per day) in mice an d humans and their associated 7706 external exposures (mg/kg-day) used for the derivat ion of cancer OSFs based on liver 7707 tumors in mice. Six simulated drinking water episodes are described by Reitz et al. (1997). 7708 The human metabolism rates were estimated using a c omputational sample of 1000 individuals 7709 per dose, including random samples of the three GST -T1 polymorphisms (+/+, +/-, -/-; “Human 7710 mixed” curve) or samples restricted to the GST +/+ or +/- populations, in the current U.S. 7711 population based on data from Haber et al. (2002). Since a different set of samples was used 7712 for each dose, some stochasticity is evident as the human points (values) do not fall on smooth 7713 curves. Error bars indicate the range of 5th-95th percentile for the sub-populations sampled at 7714 select concentrations. 7715 7716 Dose-response modeling and extrapolation 7717 The multistage dose-response model was fit to the mouse liver tumor incidence and 7718

PBTK model-derived internal dose data to derive mou se internal BMD 10 and BMDL 10 associated 7719 with 10% extra risk (Table 5-12). Different polyno mial models, and models dropping dose 7720 groups starting with the highest dose group were co mpared based on adequacy of model fit as 7721 assessed by overall χ 2 goodness of fit (

p- value > 0.10) and examination of residuals, particu larly 7722 in the region of the benchmark response (BMR). A ppendix E-1 provides details of the BMD 7723 modeling results. The mouse liver tumor risk facto r (extra risk per unit internal dose) was 7724 calculated by dividing 0.1 by the mouse BMDL 10 for liver tumors. 7725

7726

7727269 Table 5-12. BMD modeling results and tumor risk factors for int ernal dose metric associated with 10% extra risk for liver tumors in male B6C3F 1 mice exposed to dichloromethane in drinking water for 2 years, base d on liver-specific GST metabolism and whole body GST metabolism dose metrics

Tumor Risk Factor e

Internal dose metric BMDS

model b χ 2 goodness of fit p- value Mouse

BMD 10 c Mouse

BMDL 10 c Allometricscaled human BMDL 10 d Scaling = 1.0 Allometricscaled

Liverspecific MS (1,1) 0.56 73.0 39.6 5.66 2.53 × 10 –3 1.77 × 10 –2

Whole-body MS (1,1) 0.56 3.05 1.65 0.24 -- 4.24 × 10 –1

a Liver specific dose units = mg dichloromethane meta bolized via GST pathway per liter tissue per day; Whole-body dose units = mg dichloromethane metabol ized via GST pathway in lung and liver/kgday)

b The multistage (MS) model in EPA BMDS version 2.0 w as fit to the mouse dose-response data shown in Table 5-11 using internal dose metrics cal culated with the mouse PBTK model. Numbers in parentheses indicate (1) the number of dose grou ps dropped in order to obtain an adequate fit and (2) the degree polynomial of the model.

c BMD 10 and BMDL 10 refer to the BMD-model-predicted mouse internal and its 95% lower confidence limit, associated with a 10% extra risk for the inc idence of tumors.

d Mouse BMDL 10 divided by (BW human /BW mouse ) 0.25 = 7.

e Dichloromethane tumor risk factor (extra risk per u nit internal dose) derived by dividing the BMR (0.1) by the mouse BMDL 10 and by the allometric-scaled human BMDL 10 ,, for the scaling = 1.0 and allometric-scaled risk factors, respectively. 7728 7729 Linear extrapolation from the internal human BMDL 10 values (0.1/BMDL 10 ) was used to 7730 derive oral risk factors for liver tumors, based on tumor responses in male mice. Proposed key 7731 events for dichloromethane carcinogenesis are discu ssed in sections 4.7 and 5.4.1.1. The linear 7732 low-dose extrapolation approach for agents with a m utagenic mode of action was selected. 7733 Application of allometric scaling factor 7734

As discussed in section 4.7 and summarized in 5.4. 1.2, several lines of evidence point to 7735 the involvement of the GST metabolic pathway in the carcinogenic response seen in 7736 dichloromethane. The role of specific metabolites has not been firmly established, however. 7737 S-(chloromethyl)-glutathione is an intermediate to the production of formaldehyde through this 7738 pathway (Hashmi et al., 1994). Formation of free h ydrogen ion is also hypothesized, although 7739 no direct evidence supporting this has been present ed. The pattern of HPRT gene mutations seen 7740 in CHO cells incubated with GST-complete mouse live r cytosol preparations suggest that 7741 S-(chloromethyl)glutathione, rather than formaldehy de, is responsible for the mutagenic effects 7742 associated with dichloromethane (Graves et al., 199 6). DNA reaction products (e.g., DNA 7743 adducts) produced by S-(chloromethyl)glutathione ha ve not been quantified, possibly due to 7744 potential instability of these compounds (Guengeric h et al., 2003; Hashmi et al., 1994). 7745 The question of the role of specific metabolites, a nd particularly how these metabolites 7746 are transformed or removed is a key question affect ing the choice of a scaling factor to be used in 7747 conjunction with the internal dose metric based on rate of GST metabolism. If the key metabolite 7748 is established and is known to be sufficiently reac tive to not spread in systemic circulation, then 7749270

it can be assumed that 1) the level of reactivity a nd rate of clearance (i.e., disappearance due to 7750 local reactivity) for this metabolite, per volume t issue, is equal in rodents and humans; and (2) 7751 risk is proportional to the long-term daily average concentration of the metabolite. Under these 7752 assumptions, rodent internal BMDL 10 values based on tissue-specific dichloromethane 7753 metabolism require no allometric scaling to account for toxicodynamic differences and predict 7754 the corresponding level of human risk as a function of the metric (i.e., the scaling factor in Figure 7755 5-14 was equal to 1.0). (A single metabolite is re ferenced, but the same argument holds in 7756 general for more than one metabolite). Under this scenario and assumptions, humans and 7757 rodents with the same long-term daily average metab olite formation per volume tissue (e.g., 7758 equal internal BMDL 10 ) should both experience the same long-term average concentration of the 7759 metabolite when the metabolite is highly reactive a nd hence experience the same extra risk. 7760 Although the evidence points to a specific metabol ic pathway and to site-specific actions 7761 resulting from a reactive metabolite that does not escape the tissue in which it is formed, some 7762 assumptions remain concerning this hypothesis. Spec ifically, the active metabolite(s) have not 7763 been established, and data pertaining to the reacti vity or clearance rate of these metabolite(s) are 7764 lacking. Quantitative measurements of adducts of i nterest or of the half life of relevant 7765 compounds in humans and in mice are not available. To address the uncertainties in the 7766 available data it may be appropriate to use a scali ng factor that addresses the possibility that the 7767 rate of clearance for the metabolite is limited by processes that are known to scale allometrically, 7768 such as blood perfusion or enzyme activity. This c ase would result in use of a mouse:human 7769 dose-rate scaling factor of (BW human /BW mouse ) 0.25 = 7 to adjust the mouse-based BMDL 10 values 7770 downward. Using this internal dose metric (liver-sp ecific metabolism with allometric scaling), 7771 equivalent rodent and human internal BMDL 10 values result in a human liver tumor risk factor 7772 (0.1/BMDL 10 ) that is assumed equal to that for the mouse, give n a 70-year lifetime exposure. 7773

Another alternative that can be used is based on an allometrically-scaled whole-body 7774 metabolism metric. In this case, less weight is gi ven to the evidence of site-specificity, as this 7775 metric allows for systemic circulation of the relev ant metabolites. 7776 The cancer toxicity values derived using each of t hese metrics and scaling factors (i.e., 7777 liver-specific metabolism with and without allometr ic-scaling and the whole-body metabolism 7778 metric) are presented in the following tables. Con sidering the lack of data pertaining to 7779

clearance rates or the actual AUC of the active car cinogenic metabolite(s) in mice and humans, 7780 the OSF recommended by the EPA is based on the allo metrically-scaled tissue-specific GST 7781

metabolism rate dose metric. 7782

7783 Calculation of OSFs 7784 The human PBTK model adapted from David et al. (20 06) (see Appendix B), using 7785 Monte Carlo sampling techniques, was used to calcul ate distributions of human internal dose 7786 metrics of daily mass of dichloromethane metabolize d via the liver-specific GST pathway per 7787271

unit volume of liver resulting from a long-term ave rage daily drinking water dose of 1 mg/kg 7788 dichloromethane. In another analysis of whole body metabolism, a dose metric based on the 7789 total metabolites formed in liver and lungs via GST metabolism per BW was used. The human 7790 model used parameter values derived from Monte Carl o sampling of probability distributions for 7791 each parameter, including MCMC-derived distribution s for the metabolic parameters (David et 7792 al., 2006). The drinking water exposures comprised six discrete drinking-water episodes for 7793 specified times and percentage of total daily intak e (Reitz et al., 1997) (Appendix B). 7794

The distribution of cancer OSFs shown in Table 5-13 were derived by multiplying the 7795 human oral liver tumor risk factors by the respecti ve distributions of human average daily 7796 internal doses resulting from chronic, unit oral ex posures of 1 mg/kg-day dichloromethane. 7797 Because adjustments for interindividual variability are not generally used or recommended in 7798 cancer risk analysis, the mean slope factor was sel ected as the recommended value to be used in 7799 deterministic risk assessments; other values at the upper end of the distribution are also 7800 presented. 7801

7802 Consideration of Sensitive Human Subpopulations 7803 An important issue in the derivation process used b y EPA, pertaining to the use of the 7804 human PBTK model, stems from the assumption regardi ng the population for which the 7805 derivation should be applied. The inclusion of the GST-T1 null subpopulation in effect dilutes 7806 the risk that would be experienced by those who car ry a GST-T1 allele, by averaging in non7807 responders (i.e., the GST-T1 –/– genotype). Thus, the cancer OSF was derived speci fically for 7808 carriers of the GST-T1 homozygous positive (+/+) ge notype, that is the population that would be 7809 expected to be most sensitive to the carcinogenic e ffects of dichloromethane given the GST7810

related dose metric under consideration. In additi on, cancer values derived for a population 7811 reflecting the estimated frequency of GST-T1 genoty pes in the current U.S. population (20% 7812

GST-T1 –/– , 48% GST-T1 +/– , and 32% GST-T1 +/+ , the “mixed” population) are also presented. 7813

All simulations also included a distribution of CYP activity based on data from Lipscomb et al. 7814 (2003). 7815272Table 5-13.Cancer OSFs for dichloromethane based on PBTK model-derived internal liver doses in B6C3F1 mice exposedvia drinking water for 2 years, based on liver-specific GST metabolism and whole body metabolism dosemetrics, bypopulation genotypeDistribution of human internaldichloromethanedoses from 1 mg/kg-day exposuredResulting candidate humanOSFe(mg/kg-day)–1Internal dosemetric andscaling factoraPopulationgenotypebHuman tumorrisk factorcMean95thpercentile99thpercentileMean95thpercentile99thpercentileGST-T1+/+1.77 × 10–20.80 × 10–11.91 × 10–12.89 × 10–11.4 × 10–33.4 × 10–35.1 × 10–3Liver-specific,allometric-scaledMixed 1.77 × 10–20.45 × 10–11.39 × 10–12.24 × 10–18.0 × 10–42.5 × 10–34.0 × 10–3GST-T1+/+2.53 × 10–30.80 × 10–11.91 × 10–12.89 × 10–12.0 × 10–44.8 × 10–47.3 × 10–4Liver-specific,scaling = 1.0Mixed 2.53 × 10–30.45 × 10–11.39 × 10–12.24 × 10–11.2 × 10–43.5 × 10–45.7 × 10–4GST-T1+/+4.24 × 10–11.90 × 10–34.60 × 10–37.20 × 10–38.1 × 10–42.0 × 10–33.1 × 10–3Whole-body ,allometric-scaledMixed 4.24 × 10–11.08 × 10–33.40 × 10–35.49 × 10–34.6 × 10–41.4 × 10–32.3 × 10-3aLiver specific dose units = mg dichloromethane metabolized via GST pathway per liter tissue per day; Whole-body dose units = mgdichloromethane metabolized via GST pathway in lungand liver/kg-day.bGST-T1+/+= homozygous, full enzyme activity; mixed = population reflecting estimated frequency of genotypes incurrent U.S. population:20% GST-T–/–, 48% GST-T1+/–, and 32% GST-T1+/+(Haber et al., 2002).cDichloromethane tumor risk factor (extra risk per unit internal dose per day) derived by dividing theBMR (0.1) by the allometric-scaledhuman BMDL10and the mouse BMDL10for the allometric-scaled and scaling = 1.0 risk factors, respectively (from Table 5-12).dMean, 95th, and 99thpercentile of the human PBTK model-derived probability distribution of daily average internal dichloromethane doseresulting from chronic oral exposure of 1 mg/kg-day.eDerived by multiplying the dichloromethane tumor risk factor by the PBTK model-derived probabilistic internal doses from daily exposure to1 mg/kg-day.78167817273

5.4.1.5. Oral Cancer Slope Factor 7818

The recommended cancer OSF for dichloromethane is 1 × 10 –3 (mg/kg-day) –1 (rounded 7819

from 1.4 × 10 –3 ) , and is based on liver tumor responses in male B6C 3F 1 mice exposed to 7820 dichloromethane in drinking water for 2 years (Sero ta et al., 1986b; Hazelton Laboratories, 7821 1983). The OSF was derived by using a tissue-specif ic GST metabolism dose metric, with 7822 allometric scaling, for the population that is pres umed to have the greatest sensitivity (the GST7823

T1 +/+ genotype). The application of ADAFs to the cancer OSF is recommended and is described 7824

in section 5.4.4. 7825

7826

5.4.1.6. Alternative Derivation Based on Route-to-Route Extr apolation 7827 For comparison, alternative cancer OSFs were deriv ed via route-to-route extrapolations 7828

from the data for liver tumors in male and female B 6C3F 1 mice exposed by inhalation for 2 years 7829 (Mennear et al., 1988; NTP, 1986). This derivation , shown in Table 5-14, uses the cancer IUR 7830 derived in section 5.4.2.4 (see Table 5-19 for thes e IUR values) and the distribution of human 7831 internal dichloromethane exposures from 1 mg/kg-day exposure using the tissue-specific GST 7832 metabolism dose metric (mg dichloromethane metaboli zed via the GST pathway per liter liver 7833 per day). The oral cancer slope factor values base d on the route-to-route extrapolations from 7834 liver tumors in mice exposed by inhalation (Table 5 -14) are about one order of magnitude lower 7835 than those based on the liver tumor responses in mi ce exposed via drinking water. 7836274Table 5-14.Alternative route-to-route cancer OSFs for dichloromethane extrapolated from male B6C3F1mouseinhalation liver tumor incidence data using a tissue-specific GST metabolism dose metric, by population genotypeDistribution of human internaldichloromethanedoses from 1 mg/kg-day exposurecResulting candidate humanOSFd(mg/kg-day)–1Internal dose metricand scaling factorPopulationGenotypeaHumantumor riskfactorbMean95thpercentile99thpercentileMean95thpercentile99thpercentileGST-T1+/+1.29 × 10–30.80 × 10–11.91 × 10–12.89 × 10–11.0 × 10–42.5 × 10–43.7 × 10–4Liver-specific,allometric-scaledMixed 1.29 × 10–30.45 × 10–11.39 × 10–12.24 × 10–15.8 × 10–51.8 × 10–42.9 × 10–4GST-T1+/+1.84× 10–40.80 × 10–11.91 × 10–12.89 × 10–11.5 × 10–53.5 × 10–55.3 × 10–5Liver-specific,scaling = 1.0Mixed 1.84× 10–40.45 × 10–11.39 × 10–12.24 × 10–18.3 × 10–62.6 × 10–54.1 × 10–5GST-T1+/+3.03 × 10–21.90 × 10–34.60 × 10–37.20 × 10–35.8 × 10–51.4 × 10–42.2 × 10–4Whole-bodymetabolismMixed 3.03 × 10–21.08 × 10–33.40 × 10–35.49 × 10–33.3 × 10–51.0 × 10–41.7 × 10–4aGST-T1+/+= homozygous, full enzyme activity; mixed = population reflecting estimated frequency of genotypes incurrent U.S. population:20% GST-T–/–, 48% GST-T1+/–, and 32% GST-T1+/+(Haber et al., 2002).bDichloromethane tumor risk factor (extra risk per milligram dichloromethane metabolized via GST pathway per liter tissue per day) derivedby dividing the BMR (0.1) by the allometric-scaledhuman BMDL10and the mouse BMDL10for the allometric-scaled and scaling = 1.0 riskfactors, respectively (from inhalation unit risk data, Table 5-19).cMean, 95th, and 99thpercentile of the human PBTK model-derived probability distribution of daily average internal dichloromethane dose(milligrams dichloromethane metabolized via GST pathway per liter tissue per day) resulting from chronic oral exposure of 1 mg/kg-day.dDerived by multiplying the dichloromethane tumor risk factor by the PBTK model-derived probabilistic internal doses from daily exposureto 1 mg/kg-day.7837275

5.4.1.7. Alternative Based On Administered Dose 7838 One comparison that can be made is with an alterna tive OSF based on liver tumors in 7839 mice, using the external concentrations of dichloro methane in the mouse as converted to human 7840 equivalent doses and then applying this by using BM D modeling to obtain the BMDL 10 and 7841 resulting oral cancer risk. Mouse bioassay exposur es were adjusted to human equivalent doses 7842

as follows: 7843

7844 human equivalent dose = 7845 (nominal daily intake ÷ BW scaling factor) × daily exposure adjustment factor 7846

7847 where BW scaling factor = (BW human /BW mouse ) 0.25 = 7 7848

and 7849 daily exposure adjustment factor = 5/7 7850 7851 The human equivalent administered doses for the 0, 60, 125, 185, and 250 mg/kg-day 7852 dose groups used in the liver tumor analysis (Table 5-11) (from Serota et al. [1986b]) were 0, 7853 6.12, 12.75, 18.87,and 25.51 mg/kg-day, respectivel y. The BMD modeling and OSF derived 7854

from these values are shown in Table 5-15. The res ulting OSF, based on the liver tumors in the 7855 mouse, is approximately one order of magnitude high er than the current recommended value 7856 obtained by using the mouse and human PBTK models. 7857

7858

Table 5-15. Cancer OSF based on a human BMDL 10 using administered dose for liver tumors in male B6C3F 1 mice exposed to dichloromethane in drinking water for 2 years

Sex, tumor type BMDS model a χ 2 goodness of fit

p- value Human

BMD 10 c Human

BMDL 10 c Cancer

OSF d (mg/kg-day) –1

Male, liver MS (0,1) 0.55 19.4 10.4 1.0 × 10 –2

a The multistage (MS) model in EPA BMDS version 2.0 w as fit to the mouse liver tumor data shown in Table 5-11. The human equivalent doses for the 0, 60, 125, 185, and 250 mg/kg-day dose groups used in the liver tumor analysis were 0, 6.12, 12.75, 18.87 , and 25.51 mg/kg-day, respectively, based on application of BW scaling factor = (BW human /BW mouse ) 0.25 = 7 and adjusting for daily exposure by multiplying by 5/7 days. Numbers in parentheses indicate (1) the number of d ose groups dropped in order to obtain an adequate fit, starting with the highes t do se group, and (2) the degree polynomial of the mode l.

c BMD 10 and BMDL 10 refer to the BMD-model-predicted human equivalent administered dose (mg/kg-day) and its 95% lower confidence limit, associated with a 10% extra risk for the incidence of tumors.

d Cancer OSF (risk per mg/kg-day) = 0.1/human BMDL 10 . 7859 7860 The administered dose methodology can be considered equivalent to using a single7861 compartment, whole-body model of dichloromethane, w here the internal dose metric is the AUC 7862

of dichloromethane itself and clearance of dichloro methane scales from mice to humans as 7863

BW 0.75 . The estimates based on the PBTK model, in contra st, use the rate of metabolism of 7864276

dichloromethane (GST) as the metric. Another diffe rence is that the administered dose 7865 methodology does not account in any way for the GST polymorphism and so might be considered 7866 as representing the general/mixed-GST-genotype popu lation rather than the +/+ subpopulation. 7867

7868

5.4.1.8. Previous IRIS Assessment: Cancer Oral Slope Factor 7869 The previous IRIS assessment derived a cancer OSF of 7.5 × 10 –3 (mg/kg-day) –1 by the 7870 application of the multistage model to combined inc idence of hepatocellular adenomas, 7871 carcinomas, and neoplastic nodules from two studies . These were the 2-year drinking water 7872 study of dichloromethane in B6C3F 1 mice by the Hazelton Laboratories (1983) and the 2 -year 7873 inhalation study of dichloromethane in B6C3F 1 mice by NTP (1986). The slope factor was the 7874 arithmetic mean of two candidate slope factors, 1.2 × 10 –2 (mg/kg-day) –1 (Hazelton Laboratories, 7875 1983) and 2.6 × 10 –3 (mg/kg-day) –1 (NTP, 1986). Since the NTP (1986) animal data wer e from 7876 inhalation exposures, the estimated inhaled doses w ere calculated for mice and humans 7877 (assuming near complete uptake into lung tissues an d blood) and converted to administered 7878 doses in units of mg/kg-day. Assumed inhalation ra tes of 0.0407 and 20 m 3 /day were used for 7879 mice and humans, respectively. No adjustments were made for species differences in 7880

metabolism or toxicokinetics. 7881

7882

7883 5.4.1.9. Comparison of Cancer Oral Slope Factors Using Diffe rent Methodologies 7884 Cancer OSFs derived using different dose metrics an d assumptions are summarized in 7885

Table 5-16. The recommended OSF of 1 × 10 –3 per mg/kg-day (rounded to one significant digit) 7886 is based on a tissue-specific GST-internal dose met ric, with allometric scaling (= 7), because of 7887 some uncertainty regarding the rate of clearance of the relevant metabolite(s) formed via the 7888 GST pathway. The value derived specifically for th e GST-T1 +/+ population is recommended to 7889 provide protection for the population that is hypot hesized to be most sensitive to the carcinogenic 7890

effect. The values based on the GST-T1 +/+ group are approximately two-fold higher than those 7891 for the full population for the dose metrics used i n this assessment (Table 5-16). Within a 7892 genotype population, the values of the OSF among mo st of the various dose metrics vary by 7893 about one to two orders of magnitude. 78942777895Table 5-16. Comparison of OSFs derived using various assumptions and metrics, based on tumors in malemicePopulationaDose metricSpecies,sex TumorScalingfactorMean OSF(mg/kg-day)–1Source(Table)GST-T1+/+Tissue-specific GST-metabolism ratebMouse, male Liver 7.0 1.4 × 10–3Table 5-13Tissue-specific GST-metabolism rate Mouse, male Liver 1.0 2.0 × 10–4Table 5-13Whole-body GST metabolism rate Mouse, male Liver 7.0 8.1 × 10–4Table 5-13Route-to-route extrapolation, tissue-specific metabolism Mouse, male Liver 7.0 1.0 × 10–4Table 5-14Route-to-route extrapolation, tissue-specific metabolism Mouse, male Liver 1.0 1.5 × 10–5Table 5-14Route-to-route extrapolation, whole-body metabolism Mouse, male Liver 7.0 5.8 × 10–5Table 5-14Mixed Tissue-specific GST-metabolism ratebMouse, male Liver 7.0 8.0 × 10–4Table 5-13Tissue-specific GST-metabolism rate Mouse, male Liver 1.0 1.2 × 10–4Table 5-13Whole-body GST metabolism rate Mouse, male Liver 7.0 4.6 × 10–4Table 5-13Route-to-route extrapolation, tissue-specific metabolism Mouse, male Liver 7.0 5.8 × 10–5Table 5-14Route-to-route extrapolation, tissue-specific metabolism Mouse, male Liver 1.0 8.3 × 10–6Table 5-14Route-to-route extrapolation, whole-body metabolism Mouse, male Liver 7.0 3.3 × 10–5Table 5-14Applied dose (human equivalent dose) Mouse, male Liver1.0×10–2Table 5-151995 IRIS assessment Mouse, male Liver 7.5 × 10–3aGST-T1+/+= homozygous, full enzyme activity; Mixed = genotypes based on a population reflecting the estimatedfrequency of genotypes in the currentU.S. population: 20% GST-T1–/–, 48% GST-T1+/–, and 32% GST-T1+/+(Haber et al., 2002).Bolded value is the basis for the recommended OSF of 1×10–3per mg/kg-day.7896278 7897 5.4.2. Cancer Inhalation Unit Risk 7898

5.4.2.1. Choice of Study/Data—with Rationale and Justificati on 7899 As discussed in section 4.7, results from several cohort mortality studies of workers 7900 repeatedly exposed to dichloromethane and several c ase-control studies provide some supporting 7901 evidence of carcinogenicity in humans, specifically with respect to liver and brain cancer. 7902 However, the epidemiologic studies do not provide a dequate data to estimate exposure-response 7903 relationships for dichloromethane exposure and thes e cancers. 7904 Results from several bioassays provide sufficient e vidence of the carcinogenicity of 7905 dichloromethane in mice and rats exposed by inhalat ion, as well as adequate data to describe 7906 dose-response relationships. As discussed in secti on 4.7.2, repeated inhalation exposure to 7907 concentrations of 2,000 or 4,000 ppm dichloromethan e produced increased incidences of lung 7908

and liver tumors in male and female B6C3F 1 mice (Maronpot et al., 1995; Foley et al., 1993; 7909 Kari et al., 1993; Mennear et al., 1988; NTP, 1986) . A weaker trend was seen with respect to 7910 liver tumor incidence (described as neoplastic nodu les or hepatic carcinomas) in female rats, but 7911 this trend was not seen when limited to hepatic car cinomas (NTP, 1986). A statistically 7912 significant increased incidence of brain tumors has not been observed in any of the animal cancer 7913 bioassays, but a 2-year study using relatively low exposure levels (0, 50, 200, and 500 ppm) in 7914 Sprague-Dawley rats observed a total of six astrocy toma or glioma (mixed glial cell) tumors 7915 (combining males and females) in the exposed groups (Nitschke et al., 1988a). These tumors are 7916 exceedingly rare in rats, and there are few example s of statistically significant trends in animal 7917 bioassays (Sills et al., 1999). Male and female F3 44 rats exposed by inhalation to 2,000 or 4,000 7918 ppm showed significantly increased incidences of be nign mammary tumors (adenomas or 7919 fibroadenomas) and the male rats also exhibited a l ow rate of sarcoma or fibrosarcoma in 7920 mammary gland or subcutaneous tissue around the mam mary gland (NTP, 1986). 7921 The NTP inhalation study in B6C3F 1 mice (NTP, 1986) was used to derive an IUR for 7922 dichloromethane because of the completeness of the data, adequate sample size, and clear dose 7923 response with respect to liver and lung tumors. Th e liver tumor incidence in male mice 7924 increased from 44% in controls to 66% in the highes t dose group; in females the incidence of this 7925 tumor rose from 6 to 83%. For lung tumors, the inc idence rose from 10 to 80% in males and 7926 from 6 to 85% in females. Compelling evidence exis ts for the role of GST-mediated metabolism 7927 of dichloromethane in carcinogenicity in mice (sect ion 4.7.3), and both mice and humans possess 7928 this metabolic pathway. Modeling intake, metabolis m, and elimination of dichloromethane in 7929 mice and humans is feasible. Thus, it is reasonabl e to apply the best available PBTK models to 7930 estimate equivalent internal doses in mice and huma ns. 7931 The mammary tumor data from the NTP (1986) study w as also used to derive a 7932 comparative IUR. However, the toxicokinetic or mec hanistic events that might lead to 7933 mammary gland tumor development in rats are unknown , and so a clear choice of the optimal 7934279

internal dose metric could not be made. Thus, this derivation is based on the average daily AUC 7935

for dichloromethane in blood. The role of CYP- or GST-mediated metabolism in the mammary 7936 gland is uncertain, although both GST-T1 (Lehmann a nd Wagner, 2008) and CYP2E1 (El-Rayes 7937 et al., 2003; Hellmold et al., 1998) expressions ha ve been detected in human mammary tissue. It 7938 is also possible that some metabolites enter system ic circulation from the liver and lung, where 7939 they are formed. 7940 The female rat liver cancer data from the NTP (198 6) inhalation study was not used to 7941

derive an IUR because the trend was weaker than tha t seen in the mouse (incidence increased 7942 from 4% in controls to 10% in the highest dose grou p, trend p = 0.08), and because the effect 7943 categorization included neoplastic nodule or hepato cellular carcinoma. The brain tumor data 7944 seen in the Nitschke et al. (1988a) study in Spragu e-Dawley rats were not used to develop an 7945

IUR because of the low incidence of this rare tumor (a total of four astrocytoma or glioma 7946 tumors in exposed males and two in exposed females) . The mechanistic issues with respect to 7947 mammary tumors and health effects issues with respe ct to brain tumors represent data gaps in the 7948 understanding of the health effects of dichlorometh ane and relevance of the rat data to humans. 7949

7950

5.4.2.2. Derivation of the Cancer Inhalation Unit Risk 7951 The derivation of the IUR parallels the process des cribed in section 5.4.1.2 for the cancer 7952 OSF. Since modeling metabolism and elimination kin etics of dichloromethane in mice and 7953 humans is feasible, it is reasonable to apply the b est available PBTK models to determine 7954 equivalent target organ doses in mice and humans. 7955

7956

5.4.2.3. Dose-Response Data 7957 Data for liver and lung tumors in male and female B6C3F 1 mice following exposure to 7958 airborne dichloromethane were used to develop IURs for dichloromethane (Mennear et al., 1988; 7959 NTP, 1986). As discussed in section 5.4.1.8, the l iver tumor dose-response data were also the 7960 basis of an OSF, derived by route-to-route extrapol ation using the PBTK models, to compare 7961 with an OSF based on liver tumor data in mice expos ed to dichloromethane in drinking water 7962 (Serota et al., 1986b). In the NTP (1986) study, s ignificant increases in incidence of liver and 7963 lung adenomas and carcinomas were observed in both sexes of B6C3F 1 mice exposed 7964 6 hours/day, 5 days/week for 2 years (Table 5-17). Since significant decreases in survival were 7965 observed in the treated groups of both sexes, the a t-risk study populations (represented by the 7966 denominators in the incidence data) were determined by excluding all animals dying prior 52 7967

weeks. 7968280 7969

5.4.2.4. Dose Conversion and Extrapolation Methods: Cancer I nhalation Unit Risk 7970

Dose conversion 7971 The PBTK model of Marino et al. (2006) for dichloro methane in the mouse was used to 7972 simulate inhalation exposures of 6 hours/day, 5 day s/week (Mennear et al., 1988; NTP, 1986) 7973 and to calculate long-term daily average internal d oses. Study-, group-, and sex-specific mean 7974

BWs were used. Based on evidence that metabolites of dichloromethane produced via the GST 7975 pathway are primarily responsible for dichlorometha ne carcinogenicity in mouse liver and lung 7976 (summarized in section 4.7.3), and the assumption t hat these metabolites are sufficiently reactive 7977 that they do not have substantial distribution outs ide these tissues, the recommended selected 7978 internal dose metric for liver tumors and lung tumo rs were long-term average daily mass of 7979 dichloromethane metabolized via the GST pathway per unit volume of liver and lung, 7980 respectively (Table 5-17). Figure 5-16 show the co mparison between inhalation external and 7981 internal doses in the liver and lung, respectively, using this dose metric for the mouse and for the 7982 human. A whole-body metabolism metric was also exa mined. This metric would be more 7983 relevant under a scenario of slowly cleared metabol ites that undergo general circulation 7984

7985

Table 5-17. Incidence data for liver and lung tumors and intern al doses, based on GST

metabolism dose metrics, in male and female B6C3F 1 mice exposed to dichloromethane via inhalation for 2 years

Sex, tumor type BW (g) External dichloromethane concentration (ppm) Mouse

tumor incidence Mouse internal

tissue dose a Mouse whole body

metabolism dose b

-- 0 22/50 (44%) e 0 0 Male, liver c

34.0 2,000 24/47 (51%) 2,363.7 100.2

32.0 4,000 33/47 (70%) 4,972.2 210.7

-- 0 5/50 (10%) e 0 0 34.0 2,000 27/47 (55%) 475.0 100.2 Male, lung d

32.0 4,000 40/47 (85%) 992.2 210.7

-- 0 3/47 (6%) e 0 0 Female, liver c 30.0 2,000 16/46 (35%) 2,453.2 104.0 29.0 4,000 40/46 (87%) 5,120.0 217.0281 Table 5-17. Incidence data for liver and lung tumors and intern al doses, based on GST metabolism dose metrics, in male and female B6C3F 1 mice exposed to dichloromethane via inhalation for 2 years

Sex, tumor type BW (g) External dichloromethane concentration (ppm) Mouse

tumor incidence Mouse internal

tissue dose a Mouse whole body

metabolism dose b

-- 0 3/45 (6%) e 0 0 30.0 2,000 30/46 (65%) 493.0 104.0 Female, lung d

29.0 4,000 41/46 (89%) 1,021.8 217.0

a For liver tumors: mg dichloromethane metabolized vi a GST pathway/L liver tissue/day from 6 hours per d ay, 5 days per week exposure; for lung tumors: mg dichlor omethane metabolized via GST pathway/L lung tissue/ day from 6 hours per day, 5 days per week exposure.

b Based on the sum of dichloromethane metabolized via the GST pathway in the lung plus the liver, normal ized to total BW (i.e., [lung GST metabolism (mg/day) + liv er GST metabolism (mg/day)]/kg BW). Units = mg dichloromethane metabolized via GST pathway in lung and liver/kg-day.

c Hepatocellular carcinoma or adenoma. Mice dying pr ior to 52 weeks were excluded from the denominators

. d Bronchoalveolar carcinoma or adenoma. Mice dying pr ior to 52 weeks were excluded from the denominators

. e Statistically significant increasing trend (by inci dental and life-table tests; p ≤ 0.01).

Sources: Mennear et al., 1988; NTP, 1986 7986

7987

A. 7988

7989 7990 B. 7991282 7992 Figure 5-16. PBTK model-derived internal doses (mg dichloromethane metabolized via the 7993 GST pathways per liter tissue per day) for liver (A ) and lung (B) in mice and humans, and 7994 their associated external exposures (ppm), used for the derivation of cancer inhalation unit 7995 risks. Average daily doses were calculated from simulated mouse exposures of 6 hours/day, 5 7996 days/week, while simulated human exposures were con tinuous. The GST metabolism rate in each 7997 simulated human population was obtained by generati ng 1000 random samples from each population 7998 (ages 0.5-80 years, males and females) for each exp osure level, and calculating the average GST 7999 metabolic rate for each sample. 8000 8001 Dose-response modeling and extrapolation 8002 The multistage dose-response model was fit to the m ouse tumor incidence and PBTK 8003

model-derived internal dose data to derive mouse in ternal BMD 10 and BMDL 10 values associated 8004 with 10% extra risk (Table 5-18). Different polyno mial models and models dropping dose 8005 groups starting with the highest dose group were co mpared based on adequacy of model fit as 8006 assessed by overall χ 2 goodness of fit (

p -value > 0.10) and examination of residuals, partic ularly 8007 in the region of the BMR (U.S. EPA, 2000c). Appendi x E-2 provides details of the BMD 8008 modeling results for the male. The mouse liver and lung tumor risk factors (extra risk per unit 8009 internal dose) were calculated by dividing 0.1 by t he mouse BMDL 10 for liver and lung tumors, 8010 respectively. 8011 Linear extrapolation from the internal BMDL 10 (0.1/BMDL 10 ) was used to derive 8012 inhalation risk factors for lung and liver tumors i n male and female mice (Table 5-18). As 8013 discussed in section 4.7, the linear low-dose extra polation approach for agents with a mutagenic 8014

mode of action was selected. 8015

8016

80172838018Table 5-18.BMD modeling results and tumor risk factors associated with 10% extra risk for liver and lung tumors inmale and female B6C3F1mice exposed by inhalation to dichloromethane for2 years, based on liver-specific GSTmetabolism and whole body GST metabolism dose metricsTumor Risk FactoreInternal dosemetricaBMDSmodelbχ2goodness of fitp-valueMouseBMD10cMouseBMDL10cAllometricscaled humanBMDL10dScaling = 1.0AllometricscaledMale, liver MS (0,1) 0.40 913.9 544.4 77.8 1.84 × 10–41.29 × 10–3Male, lung MS (0,1) 0.64 61.7 48.6 7.0 2.06 × 10–31.44 × 10–2Female, liver MS (0,2) 0.53 1224.1 659.7 94.2 1.52× 10–41.06 × 10–3Liver-specificFemale, lung MS (0,1) 0.87 51.2 40.7 5.8 2.46 × 10–31.72 × 10–2Male, liverMS (0,1)0.40 38.7 23.1 3.3--3.03 × 10–2Male, lungMS (0,1)0.66 13.1 10.3 1.5--6.80 × 10–2Female, liverMS (0,2)0.53 51.9 28.0 4.0--2.50 × 10–2Whole bodyFemale, lungMS (0,1)0.88 10.8 8.6 1.2--8.14 × 10–2aLiver specific dose units = mg dichloromethane metabolized via GST pathway per liter tissue per day; Whole-body dose units = mgdichloromethane metabolized via GST pathway in lungand liver/kg-day)bThe multistage (MS) model in EPA BMDS version 2.0 was fit to the mouse dose-response data shown in Table 5-17 using internal dosemetrics calculated with the mouse PBTK model. Numbers in parentheses indicate (1) the number of dosegroups dropped in order toobtain an adequate fit and (2) the degree polynomial of the model.cBMD10and BMDL10refer to the BMD-model-predicted mouse internal dose and its 95% lower confidence limit, associated with a 10%extra risk for the incidence of tumors.dMouse BMDL10divided by (BWhuman/BWmouse)0.25= 7.eDichloromethane tumor risk factor (extra risk per unit internal dose) derived by dividing the BMR (0.1) by the mouse BMDL10and by theallometric-scaled human BMDL10,, for the scaling = 1.0 and allometric-scaled riskfactors, respectively.801980208021284

Application of allometric scaling factor 8022 As discussed in section 5.4.1.4., the choice of a scaling factor is based on the question of 8023 the role of specific metabolites, and particularly how these metabolites are transformed or 8024 removed. If the key metabolite is established and is known to be sufficiently reactive to not 8025 spread in systemic circulation, then it can be assu med that 1) the level of reactivity and rate of 8026 clearance (i.e., disappearance due to local reactiv ity) for this metabolite, per volume tissue, is 8027 equal in rodents and humans; and (2) risk is propor tional to the long-term daily average 8028

concentration of the metabolite. Under these assum ptions, rodent internal BMDL 10 values based 8029 on tissue-specific dichloromethane metabolism requi re no allometric scaling to account for 8030 toxicodynamic differences and predict the correspon ding level of human risk as a function of the 8031 metric (i.e., the scaling factor in Figure 5-14 was equal to 1.0). (A single metabolite is 8032 referenced, but the same argument holds in general for more than one metabolite). Under this 8033 scenario and assumptions, humans and rodents with t he same long-term daily average metabolite 8034 formation per volume tissue (e.g., equal internal B MDL 10 ) should both experience the same 8035 long-term average concentration of the metabolite w hen the metabolite is highly reactive and 8036 hence experience the same extra risk. However, some uncertainties remain concerning the 8037 hypothesized role of a highly reactive metabolite i n the carcinogenic effects seen with 8038 dichloromethane. The active metabolite(s) have no t been established, and data pertaining to the 8039 reactivity or clearance rate of these metabolite(s) are lacking. For example, quantitative 8040

measurements of adducts of interest or of the half life of relevant compounds in humans and in 8041

mice are not available. To address these uncertai nties, use of scaling factor that addresses the 8042 possibility that the rate of clearance for the meta bolite is limited by processes that are known to 8043 scale allometrically, such as blood perfusion or en zyme activity, may be appropriate.. This case 8044

would result in use of a mouse:human dose-rate scal ing factor of (BW human /BW mouse ) 0.25 = 7 to 8045 adjust the mouse-based BMDL 10 values downward. Using this internal dose metric ( liver8046 specific metabolism with allometric scaling), equiv alent rodent and human internal BMDL 10 8047 values result in a human liver tumor risk factor (0 .1/BMDL 10 ) that is assumed equal to that for 8048 the mouse, given a 70-year lifetime exposure. Anot her alternative that can be used is based on 8049 an allometrically-scaled whole-body metabolism metr ic. In this case, less weight is given to the 8050 evidence of site-specificity of the effects. As wi th the OSF derivations, the cancer toxicity 8051 values derived using each of these metrics and scal ing factors (i.e., liver-specific metabolism 8052 with and without allometric-scaling and the whole-b ody metabolism metric) are presented in the 8053 following tables. Considering the lack of data per taining to clearance rates or the actual AUC of 8054 the active carcinogenic metabolite(s) in mice and h umans, the IUR recommended by the EPA are 8055 based on the allometrically-scaled tissue-specific GST metabolism rate dose metric. 8056

8057 Calculation of IURs 8058285

A probabilistic PBTK model for dichloromethane in humans adapted from David et al. 8059 (2006) (see Appendix B) was used with Monte Carlo s ampling to calculate distributions of 8060 internal lung, liver, or blood doses associated wit h chronic unit inhalation (1 μ g/m 3 ) exposures. 8061

The data on which the model is based indicate that relationship between exposure and internal 8062

dose is linear at low doses. Parameters in the hum an PBTK model developed by David et al. 8063 (2006) are distributions that incorporate informati on about dichloromethane toxicokinetic 8064 variability and uncertainty among humans. Monte Ca rlo sampling was performed in which each 8065 human model parameter was defined by a value random ly drawn from each respective parameter 8066 distribution. The model was then executed by using the external unit exposure as input, and the 8067 resulting human equivalent internal dose was record ed. This process was repeated for 10,000 8068 iterations to generate a distribution of human inte rnal doses. 8069 The resulting distribution of IURs shown in Table 5 -19 were derived by multiplying the 8070 human internal dose tumor risk factor (in units of reciprocal internal dose) by the respective 8071 distributions of human average daily internal dose resulting from a chronic unit inhalation 8072 exposure of 1 μ g/m 3 dichloromethane. Table 5-19 presents the analysis using the male data. 8073 Analyses based on the female data produced very sim ilar results, and are summarized in 8074 Appendix F. Because adjustments for interindividual variability are not generally used or 8075 recommended in cancer risk analysis, the mean slope factor was selected as the recommended 8076 value to be used in deterministic risk assessments; other values at the upper end of the 8077 distribution are also presented. As with the oral cancer slope factor derivation, the cancer IUR is 8078 derived for a population composed entirely of carri ers of the GST-T1 homozygous positive 8079 genotype (the group that would be expected to be mo st sensitive to the carcinogenic effects of 8080 dichloromethane), and a population reflecting the e stimated frequency of GST-T1 genotypes in 8081 the current U.S. population (20% GST-T1 –/– , 48% GST-T1 +/– , and 32% GST-T1 +/+ , the “mixed” 8082 population). All simulations also included a distr ibution of CYP activity, based on data from 8083 Lipscomb et al. (2003). 8084

8085

80862868087Table 5-19.IURs for dichloromethane based on PBTK model-derived internal liver and lung doses in B6C3F1malemice exposedvia inhalation for 2 years, based on liver-specificGST metabolism and whole body metabolism dose metrics, by populationgenotypeDistribution of human internaldichloromethane doses from 1μg/m3exposuredResulting candidate humanIURe(μg/m3)–1Internal dose metricand scaling factoraPopulationgenotypebTumortypeHumantumor riskfactorcMean95thpercentile99thpercentileMean95thpercentile99thpercentileGST-T1+/+Liver 1.29 × 10–35.64 × 10–61.56 × 10–52.60 × 10–57.3 × 10–92.0 × 10–83.3 × 10–8GST-T1+/+Lung 1.44 × 10–23.31 × 10–78.55 × 10–71.34 × 10–64.8 × 10–91.2 × 10–81.9 × 10–8Mixed Liver 1.29 × 10–32.62 × 10–68.65 × 10–61.45 × 10–53.4 × 10–91.1 × 10–81.9 × 10–8Tissue-specific,allometric-scaledMixed Lung 1.44 × 10–21.81 × 10–75.67 × 10–79.84 × 10–72.6 × 10–98.2 × 10–91.4 × 10–8GST-T1+/+Liver 1.84 × 10–45.64 × 10–61.56 × 10–52.60 × 10–51.0 × 10–92.9 × 10–94.8 × 10–9GST-T1+/+Lung 2.06 × 10–33.31 × 10–78.55 × 10–71.34 × 10–66.8 × 10–101.8 × 10–92.8 × 10–9Mixed Liver 1.84 × 10–42.62 × 10–68.65 × 10–61.45 × 10–54.8 × 10–101.6 × 10–92.7 × 10–9Tissue-specific,scaling = 1.0Mixed Lung 2.06 × 10–31.81 × 10–75.67 × 10–79.84 × 10–73.7 × 10–101.2 × 10–92.0 × 10–9GST-T1+/+Liver 3.03 × 10–21.53 × 10–74.87 × 10–79.20 × 10–74.6 × 10–91.5 × 10–82.8 × 10–8GST-T1+/+Lung 6.80 × 10–21.53 × 10–74.87 × 10–79.20 × 10–71.0 × 10–83.3 × 10–86.3 × 10–8Mixed Liver 3.03 × 10–28.76 × 10–83.20 × 10–76.76 × 10–72.7 × 10–99.7 × 10–92.1 × 10–8Whole-body,allometric-scaledMixed Lung 6.80 × 10–28.76 × 10–83.20 × 10–76.76 × 10–76.0 × 10–92.2 × 10–84.6 × 10–8aTissue specific dose units = mg dichloromethane metabolized via GST pathway per liter tissue (liver orlung, respectively, for liver and lungtumors) per day; Whole-body dose units = mg dichloromethane metabolized via GST pathway in lung and liver/kg-day.bGST-T1+/+= homozygous, full enzyme activity;); mixed = population reflecting estimated frequency of genotypesin current U.S. population: 20% GST-T–/–, 48% GST-T1+/–, and 32% GST-T1+/+(Haber et al., 2002).cDichloromethane tumor risk factor (extra risk per unit internal dose) derived by dividing the BMR (0.1) by the allometric-scaled human BMDL10or by themouse BMDL10(from Table 5-18) for the allometric-scaled and scaling = 1.0 risk factors, respectively.dMean, 95th, and 99thpercentile of the human PBTK model-derived probability distribution of daily average internal dichloromethane dose resulting fromchronic exposure to 1μg/m3(0.00029 ppm).eDerived by multiplying the dichloromethane tumor risk factor by the PBTK model-derived probabilistic internal doses from daily exposure to 1μg/m38088287

5.4.2.5. Cancer Inhalation Unit Risk 8089

The recommended cancer IURs are 7 × 10 –9 ( μ g/m 3 ) –1 and 5 × 10 –9 ( μ g/m 3 ) –1 for the 8090 development of liver and lung cancer, respectively, based on the mean for the GST-T1 +/+ 8091 population (the group with the greatest presumed se nsitivity). These values are based on male 8092

B6C3F 1 mice, using a tissue-specific GST metabolism dose metric, with allometric scaling 8093 (Table 5-19). Risk estimates were slightly higher for liver tumors and essentially equivalent for 8094 lung tumors in males compared to females (Appendix F), so the estimates for males were 8095

selected for the candidate values. 8096

8097 Consideration of combined risk (summing risk across tumors) 8098 With two significant tumor sites, focusing on the more sensitive response may 8099

underestimate the overall cancer risk associated wi th exposure to this chemical. Following the 8100 recommendations of the National Research Council (N RC, 1994) and the Guidelines for 8101 Carcinogen Risk Assessment (U.S. EPA, 2005a), an upper bound on total risk wa s estimated in 8102 order to gain some understanding of the total risk from multiple tumor sites in the selected data 8103

set. Note that this estimate of overall risk descr ibes the risk of developing either tumor type, not 8104 just the risk of developing both simultaneously. 8105 NRC (1994) stated that an approach based on counts of animals with one or more tumors 8106 (or tumor-bearing animals) would tend to underestim ate overall risk when tumor types occur 8107 independently and that an approach based on combini ng the risk estimates from each separate 8108 tumor type should be used. For dichloromethane, th ere is no reason to expect that the occurrence 8109 of one tumor type depends on the incidence of the o ther, given the association of different dose 8110 metrics with each tumor response. Therefore it app ears reasonable to assume that the two tumor 8111 types occur independently. However, simply summing upper limit risks may result in an 8112

overestimation of overall of combined risk because of the statistical issues with respect to 8113 summing variances of distributions. An additional challenge results from the use of different 8114 internal dose metrics for different tumors, as is t he case with the dose metrics based on tissue8115 specific metabolism. Statistical methods based on a common metric can not be used with the 8116 tissue-specific metabolism metric used in these der ivations. 8117 An alternative approach is to derive an upper bound on the combined risk estimates by 8118 summing central tendency risks and calculating a po oled SD by using BMD 10 and BMDL 10 8119 values for liver and lung tumors. The SD associate d with the IUR for each tumor site is 8120 calculated as the difference between 95th percentil es of the distribution for upper bound and 8121 maximum likelihood estimate IURs (based on either f emale or male mouse tumor risk factors), 8122 divided by 1.645 (the relevant t statistic, assuming normal distributions of summed quantities). 8123 Variances for each tumor site are the squares of th e SDs. Pooled variance and SD are defined as 8124 the sum of variances for lung and liver tumors and the square root of that sum, respectively. 8125 Finally, the upper bound on the combined lung and l iver cancer risk is determined by multiplying 8126288

the cumulative SD by 1.645 and adding it to the sum med central tendency IURs. The 8127 calculations of these upper bound estimates for com bined liver and lung tumor risks are shown in 8128

Table 5-20. 8129 Using this approach and the male mouse-derived ris k factors, the combined human 8130 equivalent IUR values for both tumor types is 1 × 1 0 –8 ( μ g/m 3 ) –1 (rounded from 1.1 × 10 –8 ) in 8131 the most sensitive (GST-T1 +/+ ) population. This is the recommended inhalation c ancer unit risk 8132

value to be used in deterministic risk assessments for chronic exposure to dichloromethane. The 8133 corresponding value for a population with the frequ ency distribution of GST-T1 genotypes 8134 currently found in the U.S. population is 6 × 10 –9 ( μ g/m 3 ) –1

. 81352898136Table 5-20. Upper bound estimates of combined human IURs for liver and lung tumors resulting from lifetimeexposure to 1μg/m3dichloromethane, based on liver-specific GST metabolism and whole body metabolism dosemetrics, by population genotypeInternal dosemetric and scalingfactoraPopulationgenotypebTumor siteUpperbound IURcCentraltendencyIURdVariance oftissue-specifictumor riskeCombinedtumor riskSDfUpper bound oncombined tumorriskg(μg/m3)–1Liver 7.3 × 10–94.3 × 10–93.17 × 10–18Lung 4.8 × 10–93.8 × 10–93.75 × 10–19Tissue-specific,allometric-scaledGST-T1+/+Liver or lung 8.1 × 10–91.9 × 10–91.1 × 10–8Liver 3.4 × 10–92.0 × 10–96.84 × 10–19Lung 2.6 × 10–92.1 × 10–91.12 × 10–19MixedLiver or lung 4.1 × 10–98.9 × 10–105.5 × 10–9Liver 1.0 × 10–96.2 × 10–106.48 × 10–20Lung 6.8 × 10–105.4 × 10–107.72 × 10–21GST-T1+/+Liver or lung 1.2 × 10–92.7 × 10–101.6 × 10–9Liver 4.8 × 10–102.9 × 10–101.40 × 10–20Lung 3.7 × 10–102.9 × 10–102.31 × 10–21Tissue-specific,scaling = 1.0MixedLiver or lung 5.8 × 10–101.3 × 10–107.9 × 10–10Liver 4.6 × 10–92.8 × 10–91.29× 10–18Lung 1.0 × 10–88.2 × 10–91.84 × 10–18GST-T1+/+Liver or lung 1.1 × 10–81.8 × 10–91.4 × 10–8Liver 2.7 × 10–91.6 × 10–94.23 × 10–19Lung 6.0 × 10–94.7 × 10–96.04 × 10–19Whole-body,allometric-scaledMixedLiver or lung 6.3 × 10–91.0 × 10–97.9 × 10–9aTissue specific dose units = mg dichloromethane metabolized via GST pathway per liter tissue (liver orlung, respectively, for liver andlung tumors) per day; Whole-body dose units = mg dichloromethane metabolized via GST pathway in lung and liver/kg-day.bGST-T1+/+= homozygous, full enzyme activity;); mixed = population reflecting estimated frequency of genotypesin current U.S. population: 20%GST-T–/–, 48% GST-T1+/–, and 32% GST-T1+/+(Haber et al., 2002).cEstimated at the human equivalent BMDL10(0.1/BMDL10) (see Table 5-18).dEstimated at the human equivalent BMD10(0.1/BMD) (see Table 5-18).eCalculated as the square of the difference of the upper bound and central tendency IURs divided by thetstatistic, 1.645.rCalculated as the square root of the sum of the variances for liver and lung tumors.gCalculated as the product of the cumulative tumor risk SD and thetstatistic, 1.645, added to the sum of central tendency IURs.8137290

5.4.2.6. Comparative Derivation Based on Rat Mamma ry Tumor Data 8138 Mammary gland tumor data from male and female F344 rats following an inhalation 8139 exposure to dichloromethane were considered in deve lopment of a comparative IUR for 8140 dichloromethane (Mennear et al., 1988; NTP, 1986). In both the male and female rats, there 8141 were significant increases in the incidence of aden omas, fibroadenomas, or fibromas in or near 8142 the mammary gland. These were characterized as ben ign tumors in the NTP report (NTP, 1986). 8143 Increased numbers of benign mammary tumors per anim al in exposed groups were also seen in 8144 two studies of Sprague-Dawley rats (Nitschke et al. , 1988a; Burek et al., 1984). An oral 8145 (gavage) study in Sprague-Dawley rats reported an i ncreased incidence of malignant mammary 8146 tumors, mainly adenocarcinomas (8, 6, and 18% in th e control, 100, and 500 mg/kg dose groups, 8147 respectively), but the increase was not statistical ly significant. Data were not provided to allow 8148 an analysis that accounts for differing mortality r ates (Maltoni et al., 1988). There are 8149 considerably more uncertainties regarding the inter pretation of these data with respect to 8150 carcinogenic risk compared with the data pertaining to liver and lung tumors. The trends were 8151 driven in large part by benign tumors; adenocarcino mas and carcinomas were seen only in the 8152 females, with incidences of 1, 2, 2, and 0 in the 0 , 1000, 2000 and 4000 ppm exposure groups, 8153 respectively. There are little data to guide the c hoice of relevant dose metric, and the 8154 genotoxicity and mechanistic studies have not inclu ded mammary tissue. For these reasons, the 8155 analysis and the calculation of the comparative IUR based on rat mammary tumor data are 8156 presented in Appendix G. The IUR based on the femal e rat data was 1 × 10 –7 ( μ g/m 3 ) –1

. 8157

8158

5.4.2.7. Alternative Based on Administered Concentration 8159 Another comparison that can be made is with an alt ernative IUR based on liver and lung 8160 tumors in mice, using the external concentrations o f dichloromethane in the mouse studies as 8161 converted to HECs, and then applying this using BMD modeling to obtain the BMDL 10 and 8162 resulting IUR. Mouse bioassay exposures were adjus ted to HECs as follows: 8163

• Adjusting to continuous exposure: External concentr ation ADJ = External 8164

concentration × (6 hours/24 hours) × (5 days/7 days) 8165

• Concentrations in mg/m 3 = concentrations in ppm × 84.93/24.45. 8166

• [H b/g ] A /[H b/g ] H = the ratio of blood:gas (air) partition coefficie nts in animals and 8167 humans. Because the partition coefficient for mice (23.0) is higher than for humans 8168 (9.7), a value of 1.0 was used, as per U.S. EPA (19 94b) guidance. 8169

8170 Thus, 8171

HECs = External concentration ADJ × [H b/g ] A /[H b/g ] H = External concentration ADJ × 1 8172

8173291

The HECs (mg/m 3 ) for the 0, 2,000, and 4,000 ppm exposure groups w ere 0, 1,241, and 8174 2,481 mg/m 3 , respectively. The BMD modeling and IURs derived from these values, in 8175 conjunction with the liver and lung tumor data from Table 5-17 (NTP, 1986), are shown 8176 in Table 5-21. The resulting IURs, based on the li ver or lung tumors in the mouse, are 8177 approximately one order of magnitude higher than th e currently recommended value 8178 obtained by using the mouse and human PBTK models. 8179

8180

8181

Table 5-21. Inhalation units risks based on human BMDL 10 values using

administered concentration for liver and lung tumor s in B6C3F 1 mice exposed by inhalation to dichloromethane for 2 year s

Sex, tumor type BMDS model a χ 2 goodness of fit

p -value BMD 10 b BMDL 10 b Inhalation unit risk c ( μ g/m 3 ) –1 Male, liver MS (0,1) 0.44 456.17 272.01 3.7 × 10 –7 Male, lung MS (0,1) 0.21 137.51 108.27 9.2 × 10 –7 Female, liver MS (0,2) 0.37 602.67 344.26 2.9 × 10 –7

Female, lung MS (0,1) 0.76 126.68 100.78 9.9 × 10 –7

a The multistage (MS) model in EPA BMDS version 2.0 w as fit to each of the four sets of mouse dose-respo nse data shown in Table 5-17. The HEC used in these mo dels for the 0, 2,000,and 4,000 ppm exposure groups were 0, 1,241, and 2,481 mg/m 3 , respectively. Numbers in parentheses indicate (1 ) the number of dose groups dropped in order to obtain an adequate fit, and (2) the lowest degree polynomial of the model showing an adequate fit.

c BMD 10 and BMDL 10 refer to the BMD-model-predicted HECs (mg dichloro methane per cubic meter), and its 95% lower confidence limit, associated with a 10% e xtra risk for the incidence of tumors.

d IUR, (risk/mg-m 3 ) = 0.1/human BMDL 10 .

Sources: Mennear et al. (1988); NTP (1986). 8182 8183 The difference between the administered concentrat ion methodology and PBTK-based 8184 approaches depends on two key differences: the use of a dichloromethane-metabolite dose8185 metric, rather than dichloromethane AUC, and the fa ct that the rate of dichloromethane 8186

conversion to that metabolite is estimated in human s by using human data rather than default 8187 allometric scaling (BW 0.75 ). In addition, the administered concentration met hodology does not 8188 account in any way for the GST polymorphism and so might be considered as representing the 8189 general/mixed-GST-genotype population rather than t he +/+ subpopulation. 8190 8191

5.4.2.8. Previous IRIS Assessment: Cancer Inhalation Unit Ri sk 8192 The IUR in the previous IRIS assessment was determ ined from the combined incidence 8193 of liver and lung adenomas and carcinomas in B6C3F 1 mice exposed to dichloromethane for 8194 2 years by NTP (1986). A value of 4.7 × 10 – 7 ( μ g/m 3 ) – 1 was derived by the application of a 8195

modified version of the PBTK model of Andersen et a l. (1987), which incorporated the 8196292

pharmacokinetics and metabolism of dichloromethane. Internal dose estimates, based on 8197 dichloromethane metabolism via the GST pathway, wer e used and corrected for differences in 8198 interspecies sensitivity by applying an interspecie s scaling factor of 12.7, which was based on 8199 dose equivalence adjusted to BW to the 2/3 power, t o the human risks (Rhomberg, 1995; U.S. 8200 EPA, 1987a). 8201

5.4.2.8. Comparison of Cancer Inhalation Unit Risk Using Dif ferent Methodologies 8202 In this assessment, cancer IURs derived by using d ifferent dose metrics and assumptions 8203 were examined, as summarized in Table 5-22. The re commended IUR value of 1 × 10 – 8 8204 ( μ g/m 3 ) –1 is based on a tissue-specific GST-internal dose me tric, with allometric scaling, because 8205 of the evidence for the involvement of highly react ive metabolites formed via the GST pathway. 8206 The value derived specifically for the GST-T1 +/+ population is recommended to provide 8207 protection for the population that is hypothesized to be most sensitive to the carcinogenic effect. 8208

The values based on the GST-T1 +/+ group are approximately two to fivefold higher tha n those for 8209 the full population, for all dose metrics used in t his assessment. Within a genotype population, 8210 the values of the IUR among the various dose metric s vary by about one to two orders of 8211 magnitude. 82122938213Table 5-22. Comparison of IURs derived by using various assumptions and metricsPopulationaDose metricSpecies, sex, Tumor typeScalingfactorIURb(μg/m3)–1Source(Table)GST-T1+/+Tissue-specific GST-metabolism ratecMouse, male Liver and lung7.0 1.1 × 10–8Table 5-20GST-T1+/+Tissue-specific GST-metabolism rate Mouse, male Liver 7.0 7.3 × 10–9Table 5-19GST-T1+/+Tissue-specific GST-metabolism rate Mouse, male Lung 7.0 4.8 × 10–9Table 5-19GST-T1+/+Tissue-specific GST-metabolism rate Mouse, male Liver and lung 1.0 1.6 × 10–9Table 5-20GST-T1+/+Tissue-specific GST-metabolism rate Mouse, male Liver 1.0 1.0 × 10–9Table 5-19GST-T1+/+Tissue-specific GST-metabolism rate Mouse, male Lung 1.0 6.8 × 10–10Table 5-19GST-T1+/+Whole-body GST metabolism rate Mouse, male Liverand lung 7.0 1.4 × 10–8Table 5-20GST-T1+/+Whole-body GST metabolism rate Mouse, male Liver7.0 4.6 × 10–9Table 5-19GST-T1+/+Whole-body GST metabolism rate Mouse, male Lung 7.0 1.0 × 10–8Table 5-19Mixed Tissue-specific GST-metabolism rate Mouse, male Liver and lung 7.0 5.5 × 10–9Table 5-20Mixed Tissue-specific GST-metabolism rate Mouse, male Liver 7.0 3.4 × 10–9Table 5-19Mixed Tissue-specific GST-metabolism rate Mouse, male Lung 7.0 2.6 × 10–9Table 5-19Mixed Tissue-specific GST-metabolism rate Mouse, male Liver and lung 1.0 7.9 × 10–10Table 5-20Mixed Tissue-specific GST-metabolism rate Mouse, male Liver 1.0 4.8 × 10–10Table 5-19Mixed Tissue-specific GST-metabolism rate Mouse, male Lung 1.0 3.7 × 10–10Table 5-19Mixed Whole-body GST metabolism rate Mouse, male Liver and lung 7.0 7.9 × 10–9Table 5-20Mixed Whole-body GST metabolism rate Mouse, male Liver 7.0 2.7 × 10–9Table 5-19Mixed Whole-body GST metabolism rate Mouse, male Lung 7.0 6.0 × 10–9Table 5-19Administered concentration (HEC) Mouse, male Liver3.7 × 10–7Table 5-21Administered concentration (HEC)Mouse, male Lung9.2 × 10–7Table 5-211995 IRIS assessmentcMouse, male Liver and lung 12.7 4.7 × 10–7aGST-T1+/+= homozygous, full enzyme activity; Mixed = genotypes based on a population reflecting the estimatedfrequency of genotypes in the currentU.S. population: 20% GST-T1–/–, 48% GST-T1+/–, and 32% GST-T1+/+(Haber et al., 2002).bBased on mean value of the derived distributionsBolded value is the basis for the recommended IUR of 1× 10–8μg/m3per mg/kg-day.294

5.4.3. Differences Between Current Assessment and Previous IRIS PBTK-based 8214

Assessment 8215 To better understand the changes in assessment ris k predictions between previous EPA 8216 evaluations and the current assessment, the differe nces in PBTK model parameters for the 8217

B6C3F 1 mouse were evaluated. Values that differed signif icantly between the model version 8218 used previously and that of Marino et al. (2006), a long with derived group parameters that lend 8219 further insight, are shown in Table 5-23. 8220 The tissue:air partition coefficients in Table 5-2 3 show that, while several of the blood:air 8221 partition coefficients appear to differ significant ly between the two models, the corresponding 8222 blood:air partition coefficient does not, and it is the latter that can be more indicative of long8223 term equilibration between the tissue (tissue group ) and air. Thus, the partition coefficients, the 8224 ones that most significantly differ are the blood:a ir and liver:air partition coefficients that are, 8225 respectively, 2.8- and 2.6-fold higher in the curre nt version. The increased blood:air partition 8226 coefficient results in a tendency for simulated blo od concentrations to rise more quickly and 8227 reach higher values, other parameters being equal. The significantly increased QCC and VPR 8228 contribute even more to this difference, resulting in an even faster rise to steady state during 8229 inhalation exposure simulations, though also more r apid delivery to the liver (decreasing blood8230 flow limitation of hepatic metabolism) and more rap id exhalation. The increased liver:air 8231 partition coefficient leads to higher predicted liv er concentrations (again, other parameters being 8232 equal) and hence higher rates of metabolism. 8233 For metabolism, a much reduced oxidative metabolis m is seen, which at low doses 8234 depends on V max c/K m . The revised hepatic metabolism is over 40% lower and the total of lung + 8235 liver metabolism is 50% lower than previously used. This lower rate of metabolism means that 8236 far less of parent dichloromethane will be removed through metabolism, and hence predicted 8237 blood concentrations will be higher still, relative to the impact of changes in partition coefficient, 8238 QCC, and VPR, as noted above. 8239 The result of having higher predicted blood and li ver dichloromethane concentrations is 8240 that, while the GSH-pathway metabolic constant, k f C , is virtually the same for the mouse liver in 8241 both cases, the much higher concentration of dichlo romethane available will lead to a much 8242 higher predicted rate of metabolism via this pathwa y. For the lung, since the lung:liver ratio 8243 (A2) is 43% higher in the model of Marino et al. (2 006), the relative increase will be even 8244

greater. 8245

Because the revised rate of GST metabolism in mice was estimated by using data with 8246 CYP2E1 inhibited by a suicide inhibitor, there is c onsiderable confidence in the relative rate of 8247 metabolism by these two pathways, and the GST pathw ay in particular. The partition 8248 coefficients used by Marino et al. (2006) are as me asured by Clewell et al. (1993) and expected 8249

to be at least as reliable as those used in the 199 5 assessment. Considering that the revised 8250 PBTK model does an excellent job of reproducing clo sed-chamber gas uptake data that were not 8251295

available for calibration of the 1987 model, as wel l blood concentrations after intravenous 8252 injection, we can have fairly high confidence in it s predictions. 8253 The net result of these model changes is that, und er mouse bioassay conditions, the 8254 predicted dose metrics for liver and lung cancer, i .e., GST-mediated metabolism, are higher than 8255 those obtained with the previous model, resulting i n a lower risk estimated per unit of dose. 8256

8257 Table 5-23. Comparison of key B6C3F 1 mouse parameters differing

between prior and current PBTK model application

Parameter a Marino et al. (2006); mean values as applied (posterior) U.S. EPA (1988b, 1987a, b) Partition coefficients PB blood/air

PF fat/blood PF·PB (fat/blood)·(blood/air) = fat/air

PL liver/blood PL·PB (liver/blood)·(blood/air) = liver/air

PLu lung (tissue)/blood PLu·PB (lung/blood)·(blood/air) = lung/air

PR rapidly perfused/blood PR·PB rapidly perfused/air

PS slowly perfused/blood PS·PB slowly perfused/air 23

5.1 117.3

1.6 36.8

0.46 10.6

0.52 12.0

0.44 10.1 8.29

14.5 120.2

1.71 14.2

1.71 14.2

1.71 14.2

0.96 7.96 Flow rates QCC cardiac output (L/hour/kg 0.74 ) VPR ventilation:perfusion ratio 24.2 1.45 14.3 1.0 Metabolism parameters V max c maximum CYP metabolic rate (mg/hour/kg 0.7 ) K m CYP affinity (mg/L) V maxC /K m (L/hour/kg 0.7 )

A1 ratio of lung V max c to liver V max c Total lung + liver V max c/K m

k f C first-order GST metabolic rate constant (kg 0.3 /hr) A2 ratio of lung k f C to liver k f C Total lung + liver k f

C 9.27 0.574 16.1

0.207 19.5

1.41 0.196 1.69 11.1 0.396 28

0.416 39.7

1.46 0.137 1.66

a Parameters not listed differed by less than 10% bet ween versions. See Table 3-5 and associated text f or details. 8258 8259 The other piece of the PBTK-based risk estimation is the human model. In updating the 8260 parameter estimates for the human model (see Append ix B for details), the oxidative metabolism 8261

V max c/K m approximately doubled, which leads to lower predic ted blood concentrations of 8262 dichloromethane available for metabolism by GST. I n addition, the liver GST was reduced by 8263 almost 60%, and the lung:liver GST ratio decreased by almost fivefold, for a net change in lung 8264296

GST of over 90%. Given the larger human data set a vailable to David et al. (2006) and the 8265 sophisticated Bayesian analysis used to recalibrate the model, the expectation is that these values 8266

are more reliable than the values used in the 1995 IRIS assessment. 8267 Since actual rates of metabolism at a given exposu re level also depend on respiration rate 8268 and blood flows, these changes in metabolic paramet ers do not completely determine the relative 8269 (predicted) dosimetry. But the difference in cance r risk predictions between the current and 8270 previous assessments is primarily explained by the overall prediction of higher GST-mediated 8271 dosimetry in the mouse (at bioassay conditions) and lower human GST metabolism (due in part 8272 to greater predicted clearance of dichloromethane b y oxidative metabolism). In addition to these 8273 changes in PBTK parameters, the reduction of scalin g factor from 12.7 to 7 is a significant factor 8274 in the overall change from the previous assessments

. 8275

8276

5.4.4. Application of Age-Dependent Adjustment Factors (AD AFs) 8277

The available dichloromethane studies do not includ e an evaluation of early-life 8278 susceptibility to dichloromethane cancer risk. In the absence of this type of data and if a 8279 chemical follows a mutagenic mode of action for car cinogenicity, like dichloromethane, the 8280 Supplemental Guidance for Assessing Susceptibility from Early-Life Exposure to Carcinogens 8281 (U.S. EPA, 2005b) recommends that ADAFs be applied to the cancer values. Since the OSF of 1 8282

× 10 –3 (mg/kg-day) –1 and the IUR of 1 × 10 –8 ( μ g/m 3 ) –1 were calculated from adult 8283 dichloromethane exposures, early-life cancer suscep tibility has not been accounted for in these 8284 values and ADAFs need to be applied in combination with exposure informations when 8285 estimating cancer risks that include early-life exp osures. Sample calculations that incorporate 8286 ADAFs into the cancer risks are presented in subseq uent sections. Additional examples of 8287 evaluations of cancer risks incorporating early-lif e exposure are provided in section 6 of the 8288 Supplemental Guidance (U.S. EPA, 2005b). 8289

In the Supplemental Guidance (U.S. EPA, 2005b), ADAFs are established for three age 8290 groups. An ADAF of 10 is applied for age groups le ss than 2 years, 3 is applied for ages 2 to 8291 <16 years, and 1 is applied for 16 years and above (U.S. EPA, 2005b). The 10- and 3-fold 8292 adjustments in cancer values are combined with age- specific exposure estimates, when early-life 8293 exposure considerations need to be included in canc er risk estimates. The most current 8294 information on usage of ADAFs can be found at www.epa.gov/cancerguidelines . For estimation 8295 of risk, the Supplemental Guidance (U.S. EPA, 2005b) recommends obtaining and using a ge8296 specific values for exposure and cancer potency. I n the absence of age-specific values cancer 8297 potency values, as is true for dichloromethane, age -specific cancer values are estimated by using 8298 the appropriate ADAFs. Using this process, a cance r risk is derived for each age group. The 8299 risks are summed across the age groups to get the t otal cancer risk for the age-exposure period of 8300

interest. 8301

8302297 8303

5.4.4.1. Application of ADAFs in Oral Exposure Scenarios 8304 Sample calculations incorporating the use of ADAFs are presented for three exposure 8305

duration scenarios. These scenarios include full l ifetime exposure (assuming a 70-year lifespan), 8306 and two 30-year exposures from ages 0–30 and ages 2 0–50. A constant dichloromethane 8307 exposure of 1 mg/kg-day was assumed for each scenar io. 8308 Table 5-24 lists the four factors (ADAFs, OSF, assu med exposure, and duration 8309 adjustment) that are needed to calculate the partia l cancer risk based on the early age-specific 8310 group. The partial cancer risk for each age group is the product of the four factors in columns 2– 8311 5. Therefore, the partial cancer risk following da ily dichloromethane oral exposure in the age 8312 group 0 to <2 years is the product of the values in columns 2–5 or 10 × (1 × 10 –3 ) × 1 × 2/70 = 8313

2.9 × 10 –4 . The partial risks that are listed in the last co lumn of Table 5-24 are added together to 8314 get the total risk. Thus, a 70-year (lifetime) ris k estimate for continuous exposure to 1 mg/kg8315 day dichloromethane is 1.7 × 10 –3 per mg/kg-day, which is adjusted for early-life su sceptibility 8316 and assumes a 70-year lifetime and constant exposur e across age groups. 8317

8318 Table 5-24. Application of ADAFs to dichloromethan e cancer risk

following a lifetime (70-year) oral exposure

Age group (years) ADAF Unit risk(per mg/kgday) Exposure

concentration (mg/kg-day) Duration

adjustment Partial risk

0 – <2 10 1 × 10 –3 1 2 years/ 70 years 2.9 × 10 –4

2 – <16 3 1 × 10 –3 1 14 years/ 70 years 6.0 × 10 –4

≥ 16 1 1 × 10 –3 1 54 years/ 70 years 7.7 × 10 –4

Total risk 1.7 × 10 –3 8319 8320 In calculating the cancer risk for a 30-year const ant exposure to dichloromethane at an 8321 exposure level of 1 mg/kg-day from ages 0–30, the d uration adjustments would be 2/70, 14/70, 8322 and 14/70 and the partial risks for the three age g roups would be 2.9 × 10 –4 , 6.0 × 10 –4 , and 2.0 × 8323

10 –4

, which would result in a total risk estimate of 1.1 × 10 –3

. 8324 In calculating the cancer risk for a 30-year const ant exposure to dichloromethane at an 8325 exposure level of 1 mg/kg-day from ages 20–50, the duration adjustments would be 0/70, 0/70, 8326 and 30/70. The partial risks for the three groups are 0, 0, and 4.3 × 10 –4 , which would result in a 8327

total risk estimate of 4.3 × 10 –4

. 8328

8329

5.4.4.2. Application of ADAFs in Inhalation Exposure Scenar ios 8330298

Sample calculations incorporating the use of ADAFs are presented for three exposure 8331 duration scenarios involving inhalation exposure. These scenarios include full lifetime exposure 8332 (assuming a 70-year lifespan) and two 30-year expos ures from ages 0–30 and ages 20–50. A 8333 constant dichloromethane inhalation exposure of 1 μ g/m 3 was assumed for each scenario. 8334 Similar to the oral exposure scenarios presented in section 5.4.4.1, Table 5-25 lists the 8335 four factors (ADAFs, unit risk, assumed exposure, a nd duration adjustment) that are needed to 8336 calculate the partial cancer risk based on the earl y age-specific group. The partial cancer risk for 8337 each age group is the product of the four factors i n columns 2–5. Therefore, the partial cancer 8338 risk following daily dichloromethane inhalation exp osure in the age group 0 to <2 years is the 8339 product of the values in columns 2–5 or 10 × (1 × 10 –8 ) × 1 × 2/70 = 2.9 × 10 –9 . The partial 8340

risks that are listed in the last column of Table 5 -25 are added together to get the total risk. Thus

, 8341 a 70-year (lifetime) risk estimate for continuous e xposure to 1 μ g/m 3 dichloromethane is 1.8 × 8342

10 –8 per μ g/m 3 , which is adjusted for early-life susceptibility a nd assumes a 70-year lifetime and 8343

constant exposure across age groups. 8344

8345 Table 5-25. Application of ADAFs to dichloromethan e cancer risk

following a lifetime (70-year) inhalation exposure

Age group (years) ADAF Unit risk(per mg/kgday) Exposure

concentration (mg/kg-day) Duration

adjustment Partial risk

0 – <2 10 1 × 10 –8 1 2 years/ 70 years 2.9 × 10 –9

2 – <16 3 1 × 10 –8 1 14 years/ 70 years 6.0 × 10 –9

≥ 16 1 1 × 10 –8 1 54 years/ 70 years 7.7 × 10 –9

Total risk 1.7 × 10 –8 8346 8347 In calculating the cancer risk for a 30-year consta nt exposure to dichloromethane at a 8348

level of 1 μ g/m 3 from ages 0–30, the duration adjustments would be 2/70, 14/70, and 14/70, and 8349 the partial risks for the three age groups are 2.9 × 10 –9 , 6.0 × 10 –9 , and 2.0 × 10 –9 . These partial 8350

risks result in a total risk estimate of 1.1 × 10 –8

. 8351 In calculating the cancer risk for a 30-year consta nt exposure to dichloromethane at a 8352

level of 1 μ g/m 3 from ages 20–50, the duration adjustments would be 0/70, 0/70, and 30/70, and 8353 the partial risks for the three age groups are 0, 0 , and 4.3 × 10 –9 , resulting in a total risk estimate 8354

of 4.3 × 10 –9

. 8355

8356

5.4.5. Uncertainties in Cancer Risk Values 8357

The derivation of cancer risk values often involve s a number of uncertainties in the 8358 extrapolation of dose-response data in animals to c ancer risks in human populations. Several 8359 types of uncertainty have been quantitatively integ rated into the derivation of the recommended 8360299

OSFs and IURs for dichloromethane, while others are qualitatively considered. Table 5-26 and 8361 the ensuing discussion summarize the principal unce rtainties identified, their possible effects on 8362 the cancer risk values, decisions made in the deriv ations, and justifications for the decisions. 8363

8364 Table 5-26. Summary of uncertainty in the derivati on of cancer risk

values for dichloromethane

Consideration/ approach Impact on cancer

risk value Decision Justification

Selection of data set Selection of an

alternative data set or target organ could change the recommended cancer risk values. Select Serota et al. (1986b) and NTP (1986) as principal studies to derive recommended liver and lung cancer risks for humans from responses in mice. The NTP (1986) inhalation bioassay with mice provides the strongest cancer responses (liver and lung tumors) and the best dose-response data in the animal database. The Serota et al. (1986b) mouse drinking water study provides the best oral dose-response data for liver tumors. Dichloromethane carcinogenicity appears to be mediated by a metabolic pathway that is also present in humans (i.e., the GST pathway). In combination with the animal results, epidemiological studies provide evidence of increased risks for liver and biliary duct tract cancer but are limited by a number of factors discussed in sections 4.1.3.6 and 4.1.3.7.

Selection of target organ Selection of a target

organ could change the recommended cancer risk values. Examine cancer risk

values based on alternative tumor responses (mammary gland tumors in rats); identification of potential brain cancer risk as a data gap. Inhalation cancer risk values based on mammary tumors in rats are about one order of magnitude higher than risk values based on liver or lung tumors in mice, but the evidence for mammary gland tumors from dichloromethane exposure is less consistent than evidence for liver and lung

tumors.

Selection of extrapolation approach Selection of

extrapolation approach could change the recommended cancer risk values. Examine cancer risk values based on alternative approaches. Oral cancer risk values based on route-toroute extrapolation from the NTP (1986) inhalation mouse bioassay were about one order of magnitude lower than values based on liver tumors in orally exposed mice (Serota et al., 1986b) (see Table 5-16) but are inherently less certain than the values based on oral exposure due to the influence of route of exposure on toxicokinetics.

Selection of dose metric Selection of dose

metric could change the recommended cancer risk values. Evidence of GST

involvement supports focus on this pathway. Cancer risk estimates based on alternative (tissue-specific versus whole-body) metrics examined. Inhalation and oral liver cancer risk values derived using a tissue-specific GST metabolism dose metric were slightlyhigher than values derived using a wholebody GST metabolism dose metric; for lung tumors, the reverse pattern is seen. The values based on liver or lung tumors using the tissue-specific GST metabolism are recommended based on the evidence of site locality of effects300

Table 5-26. Summary of uncertainty in the derivati on of cancer risk values for dichloromethane

Consideration/ approach Impact on cancer

risk value Decision Justification Dose-response modeling Human risk values

could increase or decrease, depending on fits of alternative models Use multistage doseresponse model to derive a BMD. The multistage model has biological support and is the model most consistently used in EPA cancer assessments.

Low-dose extrapolation Human risk values

would be expected to decrease with the application of nonlinear tumorresponses in lowdose regions of doseresponse curves. Use linear extrapolation of risk in low-dose region. Linear extrapolation from the human tumor risk factors was used to derive cancer risk values for oral and inhalation exposures. The linear low-dose extrapolation approach for agents with a mutagenic mode of action was selected.

Interspecies extrapolation of dosimetry and risk Alternative values for

PBTK modelparameters and crossspecies scaling factor could increase or decrease human cancer risk values. Use PBTK model and allometric scaling for the primary dose metric. Application of rodent and human PBTK models reduced uncertainty on cancer risk values due to interspecies differences in toxicokinetics. Examination of impact of different values for key parameters in human model, and sensitivity analysis of rodent PBTK model parameters identified influential metabolic parameters for which little or no experimental data exist (see Interspecies Extrapolation of Dosimetry of Risk section, below).

Sensitive subpopulations Differences in CYP

and GST metabolic rates could change cancer risk values. CYP variability

incorporated in the PBTK model; separate risk estimates generated for the presumed most sensitive (GST-T1 +/+ ) genotype No data are available to determine the range of human toxicodynamic variability/sensitivity, including whether children are more sensitive than adults. The toxicokinetic effect of the GST-T1 polymorphism is included in the human PBTK model, as are other sources of variability in GST and CYP metabolic parameters. 8365 8366 Data selections for derivation of IUR and OSF 8367 The database of animal bioassays identifies the li ver and lung as the most sensitive target 8368 organs for dichloromethane-induced tumor developmen t. These effects demonstrate a dose8369 response relationship in mice exposed orally (liver only) or by inhalation (liver and lung). 8370 Statistically significant increases in benign mamma ry gland tumors were observed in one study 8371 of F344 rats exposed by inhalation to 2,000 or 4,00 0 ppm (Mennear et al., 1988; NTP, 1986), 8372 and evidence for a tumorigenic mammary gland respon se in Sprague-Dawley rats was limited to 8373 increased numbers of benign mammary tumors per anim al at levels of 50–500 ppm (Nitschke et 8374 al., 1988a) or 500–3,500 ppm (Burek et al., 1984). An oral (gavage) study in female Sprague8375 Dawley rats reported an increased incidence of mali gnant mammary tumors, mainly 8376 adenocarcinomas (8, 6, and 18% in the control, 100, and 500 mg/kg dose groups, respectively), 8377301

but the increase was not statistically significant. Data were not provided to allow an analysis that 8378 accounts for differing mortality rates (Maltoni et al., 1988). The toxicokinetic or mechanistic 8379 events that might lead to mammary gland tumor devel opment in rats are unknown, although 8380 CYP2E1 (El-Rayes et al., 2003; Hellmold et al., 199 8) and GST-T1 expression has been detected 8381 in human mammary tissue (Lehmann and Wagner, 2008). Rare CNS tumors were observed in 8382 one study in rats, a study spanning a relatively lo w range of exposures (0–500 ppm). These 8383 cancers were not seen in two other studies (NTP, 19 86; Burek et al., 1984) in rats, both involving 8384 higher doses (1,000–4,000 ppm), or in a similar hig h-dose study (NTP, 1986) in mice. The 8385 relative rarity of the tumors precludes the use of the low-dose exposure study in a quantitative 8386 dose-response assessment. The in vivo genotoxicity and mechanistic data in rodents provide a 8387 detailed sequence of steps from generation of react ive metabolites to mutagenic effects, such as 8388 DNA-protein cross-links and DNA strand breaks. Fur ther, the toxicokinetic pathways implicated 8389 in production of the putative carcinogenic metaboli tes in animals also exist in humans. Thus, 8390 there is high confidence that the dose-response dat a for liver and lung cancer in mice represents 8391 the best data currently available for derivation of human cancer risks. A more complete 8392 understanding of the carcinogenic potential of dich loromethane would be achieved by addressing 8393 data gaps identified with respect to issues regardi ng potential risk and mechanisms relating to 8394 brain cancer and mammary tumors. The available epi demiologic studies provide some evidence 8395

of an association between dichloromethane and brain cancer (see Section 4.1.3.7.1). The 8396 available epidemiologic studies do not provide an a dequate basis for the evaluation of the role of 8397

dichloromethane in breast cancer because there are currently no cohort studies with adequate 8398 statistical power and no case-control studies with adequate exposure methodology to examine 8399 this relationship (see section 4.1.3.7.6) 8400

8401 Target organ 8402 The liver and lung tumor incidence from chronic exp osure biassays provide clear 8403 evidence of the carcinogenic potential of dichlorom ethane exposure. The biassays are supported 8404 by a substantial literature of genotoxicity and mec hanistic studies (summarized in section 4.5). 8405 The evidence for mammary gland tumors from dichloro methane exposure is based primarily on 8406 observations of benign tumors in rats with inhalati on exposure (NTP, 1986). Derivation of 8407 cancer potenciy values based on these data are pres ented in Appendix G. The potential brain 8408 cancer risk, suggested by a limited number of these relatively rare tumors in both animal and 8409 human studies, is identified as a data gap which wo uld benefit from additional research. 8410

8411 Extrapolation approach 8412 A route-to-route extrapolation from the NTP (1986) inhalation mouse bioassay was used 8413 to develop an oral cancer slope value. This value is inherently less certain than the values based 8414 on oral exposure due to the influence of route of e xposure on toxicokinetics. 8415302 8416

Dose Metric 8417 There is considerable data supporting the role of G ST-related metabolism of 8418 dichloromethane in carcinogenicity, as described in sections 4.5.1 and 4.7. Pretreatment of mice 8419 with buthionine sulphoximine, a GSH depletor, cause d a decrease, to levels seen in controls, in 8420 the amount of DNA damage detected immediately after in vivo exposure in liver and lung tissue 8421 (Graves et al., 1995). Although the results of Land i et al. (2003) indicate that GST activity is not 8422 needed for the observation of DNA damage by the com et assay from some trihalomethanes (e.g., 8423 bromodichloromethane), the results for dichlorometh ane were much weaker and of uncertain 8424 significance. 8425

8426 Dose-response modeling 8427 Because of the adequacy of the fit of the multista ge model to the data, little modeling 8428 uncertainty would be expected to be introduced by t he choice of this model. Application of the 8429 multistage model allowed for estimation of a point of departure in the lower region of exposure 8430

for observable cancer effects. 8431

8432 Low-dose extrapolation 8433 The mode of action is a key consideration in deter mining how risks should be estimated 8434 for low-dose exposure. The in vitro and in vivo ge notoxicity data suggest that mutagenicity is 8435 the most plausible mode of action, although key mut agenic events in the development of liver or 8436 lung tumors have not been identified. No data are available that provide an adequate rationale 8437 for choosing a nonlinear dose response in the low-d ose region. Because a mutagenic mode of 8438 action is most plausible, a linear-low-dose extrapo lation approach was used to estimate OSFs and 8439

IURs. 8440

8441 Interspecies extrapolation of dosimetry and risk 8442 Target organ dosimetry for neoplastic mouse respons es and estimation of equivalent 8443 internal human doses were accomplished using PBTK m odels for dichloromethane in mice and 8444 humans. Uncertainty in the ability of the PBTK mod els to estimate animal and human internal 8445 doses from lifetime bioassay low-level environmenta l exposures may affect the confidence in the 8446 cancer risk extrapolated from animal data. Uncerta inties in the mouse and human model 8447 parameter values were integrated quantitatively int o parameter estimation by utilizing 8448 hierarchical Bayesian methods to calibrate the mode ls at the population level (David et al., 2006; 8449 Marino et al., 2006). The use of Monte Carlo sampl ing to define human model parameter 8450

distributions allowed for derivation of human distr ibutions of dosimetry and cancer risk, 8451 providing for bounds on the recommended risk values

. 8452303

A detailed discussion of PBTK model structure (CYP rate equation) and parameter 8453 uncertainties is provided in Section 5.3. While th e structure and equations used in the existing 8454 model have been described in multiple peer-review p ublications over the past two decades, there 8455 are discrepancies between dichloromethane kinetics observed in vitro and the model parameters 8456 obtained from in vivo data. However, an alternativ e (dual-binding-site) CYP metabolic equation 8457 appears to resolve these discrepancies. Integratio n of the alternate rate equation into the PBTK 8458 modeling, and then quantitative risk assessment, wi ll likely require several years of further 8459 research, and hence is beyond the scope of the curr ent assessment. Since the GST activity in the 8460

current model is within a factor of three of that m easured in vitro (when both are evaluated on a 8461 per gram of liver basis), the impact of that model uncertainty is also expected to be no more than 8462 a factor of three. Sensitivity to the human PBTK p arameter distributions was evaluated by 8463 rescaling the parameters to the mean values obtaine d by David et al. (2006) for a specific data set 8464 (DiVincenzo and Kaplan, 1981) for which the GST act ivity was close to a numerical average of 8465

those obtained across individual data sets. When t his was done, the upper bound estimates on 8466 GST dosimetry (for low, fixed inhalation or oral ex posures) in the GST-T1 +/+ subpopulation 8467 increased by over an order of magnitude, as did the estimate of the mean activity for an 8468 inhalation exposure, although the estimated mean GS T activity for an oral exposure only 8469 increased about two-fold. So while correspondence of the in vivo GST activity with that 8470 measured in vitro suggests a lower degree of quanti tative uncertainty, it is possible that revision 8471 of the PBTK model could have a larger impact. The ultimate impact will depend on how 8472 revisions effect model predictions for both the ani mal and the human. If the predicted GST 8473 metabolism per unit exposure increases in both mice and humans by a similar factor, there will 8474 be little impact on the risk estimate. But if the GST activity predicted in the mouse is decreased 8475 by a factor of 3, while that in the human is increa sed by a factor of 3, for example, then the net 8476 impact would be an increase of 9-fold in human risk estimates. 8477

The PBTK animal models were utilized deterministic ally; i.e., the single-value parameter 8478

estimates for the rat PBTK model were used for rat dosimetry simulations and the mean 8479 parameter estimates from the Bayesian analysis of M arino et al. (2006) were used for the mouse 8480 dosimetry simulations. To assess the effect of usi ng point estimates of parameter values for 8481 calculation of rodent dosimetry, a sensitivity anal ysis was performed to identify model 8482 parameters most influential on the predictions of d ose metrics used to estimate oral and 8483

inhalation cancer risks. As was described in the R fD and RfC sensitivity analysis calculation, 8484 this procedure used a univariate analysis in which the value of an individual model parameter 8485 was perturbed by an amount ( ∆ ), in the forward and reverse direction (i.e., an i ncrease and 8486 decrease from the nominal value), and the change in the output variable was determined. Results 8487 are for the effects of a perturbation of ±1% from t he nominal value of each parameter on the 8488 output values at the end of a minimum of 10,000 sim ulated hours. This time was chosen to 8489 achieve a stable daily value of the dose metric, gi ven that the simulated bioassay exposures did 8490304

not include weekend exposures. The exposure conditi ons represented the lowest bioassay 8491 exposure, resulting in significant increases in the critical effect. For inhalation exposures in 8492 mice, the blood:air partition coefficient, followed closely by the first-order GST-mediated 8493 metabolism rate (k f C), had the greatest impact on the dose metric for liver cancer (mg 8494 dichloromethane metabolized via GST pathway per lit er liver per day) (Figure 5-17). For 8495 drinking water exposures in mice, the k f C, followed by the CYP-mediated maximum reaction 8496 velocity (V max c), affected the liver cancer dose metric to the gr eatest extent (Figure 5-18). For 8497 mice inhaling dichloromethane, the lung cancer dose metric (mg dichloromethane metabolized 8498 via GST pathways per liter lung per day), like the liver cancer metric, was highly affected by the 8499

k f

C and the blood:air partition coefficient (Figure 5 -19). However, since GST-mediated lung 8500

metabolism is calculated as a constant fraction of the liver metabolism rate (A2 × k f C), the lung 8501 cancer dose metric was most sensitive to the propor tional yield of liver GST-mediated metabolic 8502 activity attributed to the lung. The blood:air par tition coefficient was experimentally determined, 8503 lending high confidence to its value. Values for t he three metabolic parameters were determined 8504 by computational optimization against data sets not directly measuring dichloromethane or its 8505 metabolites in the target/metabolizing tissues. It is uncertain how alternative values for these 8506 three parameters would affect the estimation of ani mal BMDL 10 values and, ultimately, the OSFs 8507

and IURs. 8508

8509

Inhalation exposure: liver GST

-0.75 -0.5 -0.25 0 0.25 0.5 0.75 1 QCC VPR VLC VSC PB VMAXC A2 KFCParameter

Normalized sensitivity coefficient 8510 Figure 5-17. Sensitivity coefficients for long-ter m mass GST-mediated 8511 metabolites per liver volume from a long-term avera ge daily inhalation 8512 concentration of 2000 ppm in mice. KFC = GST-mediated metabolism rate; A2 = 8513 proportion of liver GST metabolism attributed to th e lung; VMAXC = CYP-mediated 8514 maximum rate of metabolism; PB = blood:air partitio n coefficient; VSC = slowly 8515 perfused tissue volume; VLC = liver volume; VPR = V entilation perfusion ratio ; 8516 QCC = cardiac output constant. 8517305 8518

Oral exposure: liver GST

-0.75 -0.5 -0.25 0 0.25 0.5 0.75 1 QCC VPR VLC VSC PB VMA KA A2 KFCParameter

Normalized sensitivity coefficient 8519 Figure 5-18. Sensitivity coefficients for long-ter m mass GST-mediated metabolites 8520 per liver volume from a long-term average daily dri nking water concentration of 8521 500 mg/L in mice. KFC = GST-mediated metabolism rate; A2 = proport ion of liver 8522 GST metabolism attributed to the lung; KA = oral ab sorption rate from gut; VMAXC = 8523 CYP-mediated maximum rate of metabolism; PB = blood :air partition coefficient; VSC 8524 = slowly perfused tissue volume; VLC = liver volume ; VPR = Ventilation perfusion 8525 ratio ; QCC = cardiac output constant. 8526 8527 8528 Inhalation exposure: lung GST

-0.75 -0.5 -0.25 0 0.25 0.5 0.75 1 QCC VPR VLC VSC PB VMAXC A2 KFCParameter

Normalized sensitivity coefficient 8529 Figure 5-19. Sensitivity coefficients for long-ter m mass GST-mediated metabolites 8530 per lung volume from a long-term average daily inha lation concentration of 500 8531 ppm in mice. KFC = GST-mediated metabolism rate; A2 = proporti on of liver GST 8532 metabolism attributed to the lung; VMAXC = CYP-medi ated maximum rate of 8533 metabolism; PB = blood:air partition coefficient; V SC = slowly perfused tissue volume; 8534 VLC = liver volume; VPR = Ventilation perfusion rat io ; QCC = cardiac output 8535

constant. 8536 8537306

The comparison of the OSF derived from the oral exp osure data and from the route–to8538 route extrapolation from the inhalation data provid es a crude measure of the uncertainty in 8539 recommending a human OSF based on the available rod ent bioassay data. The oral cancer slope 8540 factor based on route-to-route extrapolations from liver tumors in mice exposed by inhalation are 8541 about an order of magnitude lower than those based on the liver tumor responses in mice 8542 exposed via drinking water. This difference may be explained, at least partially, by the 8543 heterogeneity of hepatic cell types within the sinu soid, resulting in regio-specific toxicity. Oral 8544 exposure may result in a higher internal exposure o f hepatocytes in the periportal region 8545 (particularly those lining the portal vein, through which all gastrointestinal-absorbed 8546 dichloromethane passes) than in the centrilobular r egion (Syracuse Research Corporation [SRC], 8547 1989). Further, the metabolic capacity of hepatic cells is also regio-specific, with higher CYP 8548 activity found in the centrilobular region compared to the periportal region. Thus, liver perfusion 8549 via the systemic arterial circulation (through whic h inhaled dichloromethane would be 8550 introduced) or portal drainage of the gastrointesti nal tract may influence regio-specific 8551 hepatotoxicity, resulting in the route-of-exposure effects on toxicity. The available PBTK 8552 models do not have the capability to predict regio- specific disposition of dichloromethane in the 8553

liver. 8554 There is uncertainty as to whether the reactivity of the toxic dichloromethane metabolites 8555 is sufficiently high enough to preclude systemic di stribution. Therefore, alternative derivations 8556 of cancer risk values were performed under the assu mption that high reactivity leads to complete 8557

clearance from the tissue in which the active metab olite is formed (scaling factor = 1.0). The 8558 difference in scaling factor (7.0 for allometric sc aling versus 1.0) results in a 7-fold decrease in 8559 estimated cancer toxicity values. Using a whole-bo dy GST metabolism dose metric, the 8560 resulting OSF and IUR for liver and lung cancer wer e approximately five-fold higher than when 8561 tissue-specific dose metrics were used (Table 5-16 and Table 5-22). The mechanistic data 8562 support the notion that reactive metabolites produc ed in the target tissues are not well distributed 8563 and produce deleterious effects in the metabolizing tissues soon after generation. Thus, there is 8564 less uncertainty in the cancer risk values derived by using a tissue-specific GST metabolism dose 8565 metric compared with those derived using a whole-bo dy GST metabolism dose metric. 8566

8567 Sensitive human populations 8568 Possible sensitive populations include persons wit h altered CYP (e.g., obese individuals, 8569 alcoholics, diabetics, and the very young) and GST (e.g., GST-T1 homozygous conjugators) 8570 metabolic capacity. The PBTK model includes an est imate of the variability of CYP metabolism 8571 (sixfold variation), within the general population but does not specifically address what could be 8572 greater variation in these other groups. However, the known polymorphisms for GST-T1 8573 expression were integrated into the human model. T he distributions of human IUR values (from 8574 which the recommended [i.e., mean] values were take n) show that the 99 th percentiles are 8575307

approximately 4-fold and 6-fold higher than means f or liver and lung cancer. For the distribution 8576 of OSFs, the 99 th percentile is approximately threefold higher than mean for liver cancer. 8577 To further characterize the potential sensitivity o f specific subpopulations, internal dose 8578 distributions for oral exposure to 1 mg/kg-day or i nhalation exposure to 1 mg/m 3 were estimated 8579 for 1-year-old children and 70-year-old men and wom en to compare with the broader population 8580

results used to estimate cancer risks above. Since the recommended cancer risk estimate is based 8581

on the GST-T1 +/+ subpopulation, this analysis was also restricted t o that subpopulation, so that 8582 only the factors of age and gender would differ. T he impact of considering other GST-T1 groups 8583

can be seen where risk estimates for the GST-T1 +/– and entire population mix are given above. 8584 Specification of age- and gender-specific parameter s are as described in Appendix B. This 8585 sensitivity analysis is qualitatively similar to th at described previously for the noncancer 8586 assessments of dichloromethane, where the variabili ty in human equivalent administered dose 8587

and HEC values was estimated. 8588 For this analysis, however, consideration of exclus ively GST-T1 +/+ individuals will 8589 clearly narrow any estimate of variability. This a nalysis will also differ from that for noncancer 8590

effects in that the inverse of the former relations hip is being considered (i.e., the variation in a 8591 specific internal dose for a fixed exposure is bein g computed, whereas for the human equivalent 8592 administered dose and HEC the variability in exposu re levels corresponding to a fixed internal 8593 dose are estimated. The results of this analysis f or oral exposures are shown in Figure 5-20 and 8594 Table 5-27 and for inhalation exposures in Figure 5 -21 and Table 5-28. 8595 For the oral exposure analysis, the distribution of internal doses shows little variation 8596 among the different age/gender groups (Figure 5-20, Table 5-27). The cancer analysis is based 8597 on a very low internal dose, where little enzymatic saturation is expected to occur, allowing for 8598 efficient first-pass metabolism, which is independe nt of differences in respiration; differences 8599 will be more significant at the higher doses analyz ed for the noncancer human equivalent applied 8600 dose. Thus, the consideration of only GST metaboli sm and the narrower range of metabolic rate 8601 for that pathway in the +/+ population at low oral exposure rates results in minimal age/gender 8602 sensitivity differences (the 7-year-old female is o nly 5% more sensitive from pharmacokinetic 8603 factors than the general population). 8604 For inhalation, an internal liver GST dose (mean va lue) about 2.5 times higher in the 8605 child than the general population is predicted due to the higher inhalation rate. The results for 8606 the liver GST dose for inhalation, Figure 5-21 and Table 5-28 indicate that the 70-year-old male 8607 and female populations are only slightly shifted fr om the general population, while the 8608 population for the 1-year-old child is a distinct u pper tail of the general distribution. 8609

8610308 0 3 6 9 12 15

0 0.05 0.1 0.15 0.2 0.25 0.3 Internal dose (mg GST metabolites/L liver/day)Frequency density(population fraction/dose unit)

General 1 yo child 70 yo Male 70 yo Female Internal dose distribution, GST-T1 ++ population, oral ingestion

8611 Figure 5-20. Histograms for a liver-specific dose of GST metabolism (mg 8612 GST metabolites per liter liver per day) for the ge neral population ( 0.5- to 8613 80-year-old males and females) and specific age/gender groups, within the 8614 population of GST-T1 +/+ genotypes, given a daily oral dose-rate of 1 mg/kg

8615 day dichloromethane. 8616 8617 8618 Table 5-27 . Statistical characteristics of human internal doses for 1 mg/kg-day oral exposures in specific populations

Internal dose (mg/L liver per day) a Population Mean 95 th percentile 99 th percentile All ages b 7.96 ×10 -2 1.91 ×10 -1 2.89 ×10 -1

1-year-old children 6.60 ×10 -2 1.47 ×10 -1 2.05 ×10 -1

70-year-old men 8.22 ×10 -2 1.97 ×10 -1 2.98 ×10 -1

70-year-old women 8.66 ×10 -2 2.18 ×10 -1 3.37 ×10 -1

a Liver-specific GST-T1 metabolism in GST-T1 +/+ individuals exposed orally to 1 mg/kg-day dichloromethane.

b 0.5- to 80-year-old males and females. 8619309 0 0.05 0.1 0.15 0.2 0.25

0 10 20 30 40

ng GST metabolites/L liver/dayFrequency density (fraction/dose unit)

All ages, M & F 1 yo Child

70 yo Male

70 yo Female Internal dose distribution, GST-T1 ++ population, 1 mg/m3 inhalation

8620 Figure 5-21. Histograms for liver-specific dose of GST metabolism (mg GST 8621 metabolites per liter liver per day) for the genera l population (0.5- to 80-year8622 old males and females) and specific age/gender groups, within the populati on 8623 of GST-T1 +/+ genotypes, given a continuous inhalation exposure to 1 mg/m 3 8624 dichloromethane. 8625 8626

8627

Table 5-28 . Statistical characteristics of human internal doses for 1 mg/m 3 inhalation exposures in specific subpopulations

Internal dose (mg/L liver per day) a Population Mean 95 th percentile 99 th percentile All ages b 5.63 × 10 –6 1.56 × 10 –5 2.60s × 10 –5 1-year-old children 1.41 × 10 –5 3.30 × 10 –5 4.70 × 10 –5 70-year-old men 4.36 × 10 –6 1.11 × 10 –5 1.62 × 10 –5 70-year-old women 3.55 × 10 –6 9.41 × 10 –6 1.45 × 10 –5

a Liver-specific GST-T1 metabolism in GST-T1 +/+ individuals exposed continuously by inhalation to 1 mg/m 3 dichloromethane.

b 0.5- to 80-year-old males and females. 8628

8629

8630

8631310

6. MAJOR CONCLUSIONS IN THE CHARACTERIZATION OF HA ZARD AND 8632

DOSE RESPONSE 8633

8634 8635

6.1. HUMAN HAZARD POTENTIAL 8636 Dichloromethane (CASRN 75-09-2), also known as meth ylene chloride, is a colorless 8637 liquid with a penetrating ether-like odor. It is p roduced by the direct reaction of methane with 8638 chlorine at either high temperatures or low tempera tures under catalytic or photolytic conditions. 8639 The principal uses for dichloromethane have been in paint strippers and removers, as a propellant 8640 in aerosols, in the manufacture of drugs, pharmaceu ticals, film coatings, electronics, and 8641 polyurethane foam, and as a metal-cleaning solvent. 8642 Dichloromethane is rapidly absorbed through both o ral administration and inhalation 8643 exposure with a steady-state saturation occurring w ith inhalation. Results from studies of 8644 animals show that, following absorption, dichlorome thane is rapidly distributed throughout the 8645 body and has been detected in all tissues that have been evaluated. Metabolism of 8646 dichloromethane involves two primary pathways. Dic hloromethane is metabolized to CO and to 8647

a lesser extent CO 2 in a CYP-dependent oxidative pathway (CYP2E1) that is predominant at low 8648 exposure levels. The other major pathway for dichl oromethane metabolism involves the 8649 conjugation of dichloromethane to GSH, catalyzed by GST (GST-T1). This results in the 8650 formation of a GSH conjugate that is eventually met abolized to CO 2 . The conjugation of 8651

dichloromethane to GSH results in the formation of two reactive intermediates that have been 8652 hypothesized to be involved in dichloromethane carc inogenicity, S-(chloromethyl)glutathione 8653 and formaldehyde. Formation of formaldehyde leads to several covalent modifications of 8654 cellular macromolecules, including DNA-protein cros s-links (Casanova et al., 1996) and RNA8655 formaldehyde adducts (Casanova et al., 1997). Evid ence is also available that 8656 S-(chloromethyl)glutathione can result in both DNA SSBs and DNA mutations, presumably 8657 through DNA alkylation (Green, 1997; Graves and Gre en, 1996; Graves et al., 1996, 1994a; 8658 Hashmi et al., 1994). However, DNA reaction product s (e.g., DNA adducts) produced by S8659 (chloromethyl)glutathione have not been found, poss ibly due to potential instability of these 8660 compounds (Watanabe et al., 2004; Hashmi et al., 19 94). 8661 Information on noncancer effects in humans exposed orally to dichloromethane are 8662 restricted to case reports of neurological impairme nt (general CNS depression), liver and kidney 8663 effects (as severe as organ failure), and gastroint estinal irritation in individuals who ingested 8664 amounts ranging from about 25 to 300 mL (Chang et a l., 1999; Hughes and Tracey, 1993). The 8665 animal toxicity database identifies hepatic effects (hepatic vacuolation, nonneoplastic liver foci) 8666 as the critical dose-dependent noncancer endpoint a ssociated with oral exposure to 8667 dichloromethane. The most frequently observed live r effect was hepatocyte vacuolation, seen 8668 with drinking water exposure (90 days) in F344 rats at ≥ 166 mg/kg-day and B6C3F 1 mice at 8669311

586 mg/kg-day (Kirschman et al., 1986) and with gav age exposure (14 days) in CD-1 mice at 8670 333 mg/kg-day (Condie et al., 1983). Hepatocyte de generation or necrosis was observed in 8671 female F344 rats exposed in drinking water for 90 d ays to 1,469 mg/kg-day (Kirschman et al., 8672 1986) and in female F344 rats exposed by gavage for 14 days to 337 mg/kg-day (Berman et al., 8673 1995). In the chronic-duration (104-week) study, l iver effects were described as nonneoplastic 8674 foci in F344 rats exposed to drinking water doses b etween 50 and 250 mg/kg-day (Serota et al., 8675 1986a). In the reproductive oral administration st udies, no significant effect on reproductive 8676 function or parameter was observed in rats up to 22 5 mg/kg-day (General Electric Co., 1976) or 8677 in mice up to 500 mg/kg-day (Raje et al., 1988). T he NOAEL and LOAEL for altered 8678 neurological functions in female F344 rats were 101 and 337 mg/kg-day (as reported by Moser et 8679 al., 1995). 8680 Acute inhalation exposure of humans to dichlorometh ane has been associated with 8681 cardiovascular impairments due to decreased oxygen availability from COHb formation and 8682 neurological impairment from interaction of dichlor omethane with nervous system membranes 8683 (Bos et al., 2006; ACGIH, 2001; ATSDR, 2000; Cherry et al., 1983; Putz et al., 1979; Gamberale 8684 et al., 1975; Winneke, 1974). Relatively little is known about the long-term neurological effects 8685 of chronic exposures, although there are studies th at provide some evidence of an increased 8686 prevalence of neurological symptoms among workers w ith average exposures of 75–100 ppm 8687 (Cherry et al., 1981) and long-term effects on some neurological measures (i.e., possible 8688 detriments in attention and reaction time in comple x tasks) in workers whose past exposures 8689 were in the 100–200 ppm range (Lash et al., 1991). These studies are limited by the relatively 8690 small sample sizes and low power for detecting stat istically significant results for these 8691 endpoints. 8692 Following repeated inhalation to dichloromethane, t he liver is the most sensitive target 8693 for noncancer toxicity in rats and mice. Lifetime exposure was associated with hepatocyte 8694 vacuolation and necrosis in F344 rats exposed to 1, 000 ppm 6 hours/day (Mennear et al., 1988; 8695 NTP, 1986), hepatocyte vacuolation in Sprague-Dawle y rats exposed to 500 ppm 6 hours/day 8696 (Nitschke et al., 1988a; Burek et al., 1984), and h epatocyte degeneration in B6C3F 1 mice 8697 exposed to 2,000 ppm 6 hours/day (lower concentrati ons were not tested in mice) (Mennear et 8698 al., 1988; NTP, 1986). Other effects observed incl ude renal tubular degenerations in F344 rats 8699

and B6C3F 1 mice at 2,000 ppm, testicular atrophy in B6C3F 1 mice at 4,000 ppm, and ovarian 8700 atrophy in B6C3F 1 mice at 2,000 ppm. 8701 Other studies with inhalation exposure to dichlorom ethane revealed no significant effects 8702 on reproductive performance in rats (up to 1,500 pp m) (Nitschke et al., 1988b), although some 8703 evidence of a decrease in fertility index was seen in male mice exposed to 150 and 200 ppm 8704 (Raje et al., 1988), and no adverse effects on feta l development of mice or rats exposed up to 8705 1,250 ppm were seen by Schwetz et al. (1975). Decre ases in fetal BW and changes in behavioral 8706 habituation were observed in Long-Evans rats expose d to 4,500 ppm during the gestational 8707312

period (Bornschein et al., 1980; Hardin and Manson, 1980). Exposure-related noncancer effects 8708 on the lungs consisted of foreign-body pneumonia in rats exposed to 8,400 ppm 6 hours/day for 8709 13 weeks (NTP, 1986), Clara cell vacuolation in mic e exposed to 4,000 ppm 6 hours/day for 8710 13 weeks (Foster et al., 1992), and pulmonary conge stion in guinea pigs exposed to 5,000 ppm 8711 7 hours/day for 6 months (Heppel et al., 1944). Se veral neurological mediated parameters 8712 including decreased activity (Kjellstrand et al., 1 985; Weinstein et al., 1972; Heppel and Neal, 8713 1944), impairment of learning and memory (Alexeef a nd Kilgore, 1983), and changes in 8714 responses to sensory stimuli (Rebert et al., 1989) are reported from acute and short-term 8715 dichloromethane exposure. Evidence of a localized immunosuppressive effect in the lung, 8716 resulting from inhalation dichloromethane exposure, was seen in an acute exposure (3 hours, 8717 100 ppm) study in CD-1 mice (Aranyi et al., 1986). 8718 Numerous in vitro studies have demonstrated mutag enic and genotoxic effects 8719 associated with dichloromethane exposure. For exa mple, bacterial assays, yeast, and fungi 8720 provide evidence that the mutagenic action of dichl oromethane in bacterial systems is enhanced 8721 by metabolic activation (e.g., Dillon et al., 1992; Jongen et al., 1982; Gocke et al., 1981). 8722 Positive results from assays of DNA damage with in vitro mammalian systems provide support 8723 that dichloromethane genotoxicity is linked to meta bolism by GST enzymes (Graves et al., 1996, 8724 1995, 1994b). Consistent evidence for several geno toxic endpoints in target tissues (liver and 8725 lung) in mice following in vivo exposure to dichlor omethane provides supporting evidence that 8726 GST-pathway metabolites are key actors in the mutag enic and carcinogenic mode of action for 8727

dichloromethane. Pretreatment of mice with buthion ine sulphoximine, a GSH depletor, caused a 8728

decrease to levels seen in controls in the amount o f DNA damage detected immediately after in 8729 vivo exposure in liver and lung tissue, indicating GSH involvement in the genotoxic process 8730 (Graves et al., 1995). DNA damage (detected by the comet assay) was also reported in liver and 8731 lung tissues from male CD-1 mice sacrificed 4 hours after administration of a single oral dose of 8732 1,720 mg/kg of dichloromethane (Sasaki et al., 1998 ). In this study, DNA damage in lung and 8733

liver was not detected 3 hours after dose administr ation, and no DNA damage occurred at either 8734 time point in several other tissues in which a carc inogenic response was not seen in chronic 8735 animal cancer bioassays (e.g., stomach, kidney, bon e marrow). The weight of evidence from 8736 these studies suggests that dichloromethane is carc inogenic by a mutagenic mode of action. 8737 Dichloromethane is “likely to be carcinogenic in hu mans” by the inhalation and oral 8738 routes of exposure under the Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a). 8739 Results from several 2-year bioassays provide adequ ate evidence of the carcinogenicity of 8740 dichloromethane in mice and rats exposed by inhalat ion, as well as adequate data to describe 8741 dose-response relationships. Oral exposure to dich loromethane produced statistically significant 8742 increases in hepatocellular adenomas and carcinomas in male B6C3F 1 mice (Serota et al., 1986b; 8743 Hazelton Laboratories, 1983). Inhalation exposure to concentrations of 2,000 or 4,000 ppm 8744 dichloromethane produced increased incidences of lu ng and liver tumors in B6C3F 1 mice 8745313

(Maronpot et al., 1995; Foley et al., 1993; Kari et al., 1993; Mennear et al., 1988; NTP, 1986). 8746 Significantly increased incidences of benign mammar y tumors (adenomas or fibroadenomas) 8747 were observed in male and female F344/N rats expose d by inhalation to 2,000 or 4,000 ppm 8748 (Mennear et al., 1988; NTP, 1986). A statistically significant increased incidence of brain or 8749 CNS tumors has not been observed in any of the anim al cancer bioassays, but a 2-year study 8750 using relatively low exposure levels (0, 50, 200, a nd 500 ppm) in Sprague-Dawley rats observed 8751 a total of six astrocytoma or glioma (mixed glial c ell) tumors in the exposed groups (Nitschke et 8752 al., 1988a). These tumors are exceedingly rare in rats, and there are few examples of statistically 8753 significant trends in animal bioassays (Sills et al ., 1999). 8754

8755

6.2. DOSE-RESPONSE 8756

6.2.1. Oral RfD 8757 The available oral toxicity data for animals identi fy hepatic effects (hepatic vacuolation, 8758 nonneoplastic liver foci) as the most sensitive non cancer endpoint associated with chronic oral 8759 exposure to dichloromethane. The 104-week drinking -water study in F344 rats (Serota et al., 8760 1986a) was selected as the principal study for RfD derivation because the study provided a 8761 sensitive endpoint (nonneoplastic liver foci) and u sed lower doses in comparison to other chronic 8762 oral administration studies. In this study, four d oses (6, 52, 125, and 235 mg/kg-day in males; 6, 8763 58, 136, and 263 mg/kg-day in females) were used. A NOAEL of 6 mg/kg-day in males and 8764 females and a LOAEL of 52 (male) and 58 (female) mg /kg-day for nonneoplastic alterations of 8765

liver foci was identified. 8766

An RfD of 7 × 10 –3 mg/kg-day is recommended for use in humans. The R fD derivation 8767 process involved first fitting all available dichot omous models in BMDS version 2.0 to the 8768

incidence data for male rats. The male data were u sed because a greater sensitivity was seen in 8769 males compared with females in this study. A dose metric of average daily mass of 8770 dichloromethane metabolized via the CYP pathway per unit volume of liver was derived from an 8771 EPA-modified rat PBTK model (see Appendix C). This metric was chosen because there are no 8772 data to support the role of a specific metabolite i n the development of the noncancer liver lesions 8773 seen in oral and inhalation exposure studies and th e CYP-metabolism dose metric was 8774

determined to be most consistent with the data. Th en, the lower 95% confidence limit on the 8775 dose associated with a 10% risk for liver lesions ( BMDL 10 ) was derived, based on the best fitting 8776 model (in terms of the value of the AIC and examina tion of model fit and residuals). Because 8777 the metric is a rate of metabolism, rather than the concentration of putative toxic metabolites, and 8778 the clearance of these metabolites may be slower pe r volume tissue in the human compared with 8779 the rat, this rodent internal dose metric for nonca ncer effects was adjusted by dividing by a 8780 pharmacokinetic scaling factor to obtain a human eq uivalent internal BMDL 10 . This BMDL 10 8781 was then converted to the human equivalent dose by using a human PBTK model (adapted from 8782 David et al., 2006; see Appendix B) that utilizes M onte Carlo sampling techniques to provide a 8783314

distribution of human equivalent doses. The first percentile of the distribution of human 8784 equivalent doses was chosen to include the most sen sitive population, while staying within 8785 bounds of what is considered computationally stable . The first percentile human equivalent 8786 administered dose was used as a point of departure and was divided by a composite UF of 30 (3 8787 [10 0.5 ] to account for uncertainty about interspecies tox icodynamic equivalence, 3 [10 0.5 ] to 8788 account for uncertainty about toxicodynamic variabi lity in humans, and 3 [10 0.5 ] for database 8789 deficiencies) to arrive at an RfD of 7 × 10 –3 mg/kg-day. 8790 Use of the mean value (3.6 × 10 –1 mg/kg-day) of the human equivalent administered do se 8791

distribution instead of the 1 st percentile, with an additional UF of 3 (10 0.5 ) to account for human 8792 toxicokinetic variability, would yield a candidate RfD of 4 × 10 –3 , which is relatively similar to 8793

the recommended RfD of 1 × 10 –3

. 8794

8795

6.2.2. Inhalation RfC 8796 The liver is the most sensitive target for noncance r toxicity in rats and mice, following 8797 repeated inhalation exposure to dichloromethane. N onneoplastic liver lesions (specifically, 8798 hepatic vacuolation) in rats are the critical nonca ncer effect from chronic dichloromethane 8799 inhalation in animals. Inhalation bioassays with S prague-Dawley rats identified the lowest 8800 inhalation LOAEL for nonneoplastic liver lesions in the database: 500 ppm (6 hours/day, 8801 5 days/week for 2 years) (Nitschke et al., 1988a; B urek et al., 1984). Nitscke et al. (1988a) 8802 identified a NOAEL of 200 ppm for hepatocyte vacuol ation in female rats. Because the Nitschke 8803 et al. (1988a) study more adequately covers the ran ge spanning the BMR compared with the 8804 study by Burek et al. (1984), the former study was selected as the principal study for derivation 8805

of a chronic inhalation RfC. 8806 An RfC of 0.2 mg/m 3 is derived based on the observed critical effect i n the principal 8807 study. As was described above for the RfD, the RfC derivation process was based on a dose 8808 metric of average daily mass of dichloromethane met abolized via the CYP pathway per unit 8809

volume of liver. This metric was derived from an E PA-modified rat PBTK model (see Appendix 8810 C). Then, the lower 95% confidence limit on the do se associated with a 10% risk for liver 8811 lesions (BMDL 10 ) was derived, based on the best fitting model in t erms of the value of the AIC 8812

and examination of model fit and residuals. Becaus e the metric is a rate of metabolism, rather 8813 than the concentration of putative toxic metabolite s, and the clearance of these metabolites may 8814 be slower per volume tissue in the human compared w ith the rat, this rodent internal dose metric 8815 for noncancer effects was adjusted by dividing by a pharmacokinetic scaling factor to obtain a 8816 human-equivalent internal BMDL 10 . This BMDL 10 was then converted to the HEC by using a 8817 human PBTK model (adapted from David et al., 2006; see Appendix B) that utilizes Monte 8818 Carlo sampling techniques to provide a distribution of HECs. 8819 The first percentile HEC was used as a point of de parture. This percentile was chosen 8820 because it included the most sensitive population w hile staying within bounds of what is 8821315

considered computationally stable. This point of d eparture was divided by a composite UF of 8822 100 (3 [10 0.5 ] to account for uncertainty about interspecies tox icodynamic equivalence, 3 [10 0.5 ] 8823 to account for uncertainty about toxicodynamic vari ability in humans, and 10 for database 8824 deficiencies) to arrive at an RfC of 0.2 mg/m 3

. 8825 Use of the mean value (47.36 mg/m 3 )of the HEC distribution instead of the 1 st percentile, 8826 with an additional UF of 3 (10 0.5 ) to account for human toxicokinetic variability wo uld yield a 8827

candidate RfC identical to the recommended value of 0.2 mg/m 3 . In addition, two comparison 8828 values derived from occupational studies produced v alues of 3.5 mg/m 3 (Cherry et al., 1983) and 8829 0.55 mg/m 3 (Lash et al., 1991). The animal-derived candidate RfC is preferable to the human8830

derived candidate RfC because of the uncertainties about the exposure durations for the workers 8831 in the Cherry et al. (1983) study and uncertainties regarding the exposures and effect sizes in 8832 Lash et al. (1991) and because the RfC based on the rat data is more health protective. 8833

8834

6.2.3. Uncertainties in Reference Dose and Referen ce Concentration Values 8835 One data uncertainty identified is the potential f or neurodevelopmental effects. Animal 8836 bioassays have not identified gross or microscopic effects on neural tissues from long-term 8837 exposures or single (Schwetz et al., 1975) or multi generational (Nitschke et al., 1988b) 8838 developmental toxicity studies. However, behaviora l changes were observed in pups born to rats 8839 exposed to high levels (4,500 ppm) of dichlorometha ne (Bornschein et al., 1980; Hardin and 8840 Manson, 1980); 4,500 ppm was the only dose used in this study. Thus uncertainty exists as to 8841 the development of neurological effects from lower gestational exposures in animals, or in 8842 humans. In addition, immunotoxicity data revealed an additional area of data uncertainty 8843 specifically with respect to inhalation exposure. Data from Aranyi et al. (1986) demonstrated 8844 evidence of immunosuppression, following a single 1 00 ppm dichloromethane exposure for 3 8845 hours in CD-1 mice. The weight of evidence for non neoplastic effects in humans and animals 8846 suggests that the development of liver lesions is t he most sensitive effect, with a UF applied 8847 because of the lack of neurodevelopmental studies a nd, for the RfC, the uncertainty regardingthe 8848 lack of a low dose developmental study. 8849 The extrapolation of internal dichloromethane dosi metry from nonneoplastic rat 8850 responses to human risk was accomplished by using P BTK models for dichloromethane in rats 8851

and humans. Uncertainties in rat and human dosimet ry used for RfD and RfC derivation can 8852

arise from uncertainties in the PBTK models to accu rately simulate the toxicokinetics of 8853 dichloromethane for animals under bioassay conditio ns and humans experiencing relatively low, 8854 chronic environmental exposures. Further, the dose metric used in the models is the rate of 8855 metabolism to a putative toxic metabolite, rather t han the concentration (average or area under 8856 the concentration curve of the metabolite), so the model specifically fails to account for rodent8857

human differences in clearance or removal of the to xic metabolite. A scaling factor, based on 8858 BW ratios, was used to account for this difference. 8859316

Uncertainties in the human population model were q uantitatively accounted for by 8860 utilizing hierarchical Bayesian calibration methods during model development (David et al., 8861 2006; Marino et al., 2006). The rat model was modi fied and utilized in a deterministic manner. 8862 Data were not available to perform a hierarchical B ayesian calibration in the rat, but 8863 uncertainties in the rat model predictions were ass essed qualitatively. For both oral and 8864 inhalation exposures, the liver volume, followed cl osely by the volume of slowly perfused 8865 tissues, had the greatest impact on the internal do se of mg dichloromethane metabolized via CYP 8866 pathway per liter tissue per day. This was due to the fact that the dose metric is a tissue-specific 8867 concentration, the majority of CYP metabolism is at tributed to the liver, and changes in liver 8868 volume have a greater impact on the total CYP metab olism than either of the individual V max 8869 values. There is high confidence in the values use d for volume of liver and slowly perfused 8870 tissues in the rat, as these are well studied (Brow n et al., 1997). Therefore, the uncertainties 8871

associated with use of the rat PBTK model should no t markedly affect the values of the RfD and 8872

RfC. 8873 An additional uncertainty inherent in this process , however, is the lack of knowledge 8874 concerning the most relevant dose metric (e.g., a s pecific metabolite) within the context of the 8875 development of the noncancer liver effects. This b asic research question represents a data gap, 8876 and this uncertainty is not addressed quantitativel y or qualitatively in the assessment. 8877 The effect of dichloromethane on human populations that are sensitive due to 8878 pharmacokinetic differences was addressed quantitat ively by using a human probabilistic PBTK 8879 model to generate distributions of human exposures likely to occur given a specified internal 8880

BMDL 10 . The model and resulting distributions take into account the known differences in 8881 human physiology and metabolic capability with rega rd to dichloromethane dosimetry. The first 8882 percentile values of the distributions of human equ ivalent doses (Table 5-3) and HECs (Table 5- 8883 7) served as points of departure for candidate RfDs and RfCs, respectively, to protect 8884 toxicokinetically sensitive individuals. No data a re available regarding toxicodynamic 8885 differences within a human population. Therefore, a UF of 3 for possible differences in human 8886 toxicodynamic responses is intended to be protectiv e for sensitive individuals. 8887

8888 6.2.4. Oral Cancer Slope Factor 8889 The recommended oral cancer slope factor for dichlo romethane is 1 × 10 –3 (mg/kg-day) –1

, 8890 which is based on liver tumor responses in male B6C 3F 1 mice exposed to dichloromethane in 8891 drinking water for 2 years (Serota et al., 1986b; H azelton Laboratories, 1983). This value was 8892 derived by using a tissue-specific GST metabolism d ose metric with allometric scaling to 8893 account for uncertainty regarding the reactivity an d clearance of the metabolite(s) involved in the 8894 carcinogenic response. 8895 There was only one adequate oral exposure cancer bi oassay (Serota et al., 1986a, b; 8896 Hazelton Laboratories, 1983) evaluating the carcino genic potential of orally administered 8897317

dichloromethane in F344 rats and B6C3F 1 mice. Significant increases in incidence of liver 8898 adenomas and carcinomas were observed in male (tren d p -value = 0.058) but not female B6C3F 1 8899 mice (Serota et al., 1986b; Hazelton Laboratories, 1983). In F344 rats (Serota et al., 1986a), no 8900

increased incidence of liver tumors was seen in mal e rats, and the pattern in female rats was 8901 characterized by a jagged stepped pattern of increa sing incidence of hepatocellular carcinoma or 8902 neoplastic nodules; a similar pattern, but based on more sparse data, was seen when limited to 8903 hepatocellular carcinomas. Statistically significa nt increases in tumor incidences were observed 8904 in the 50 and 250 mg/kg-day groups (incidence rates of 10 and 14%, respectively) but not in the 8905 125 mg/kg-day group (incidence rate of 3%). Incide nce was also increased (10%) in a group 8906 exposed for 78 weeks followed by a 26-week period o f no exposure. The derivation of the oral 8907 cancer slope factor is based on the male mice data because of their greater sensitivity to liver 8908 cancer compared with female rats. 8909 A modified mouse PBTK model of Marino et al. (2006) was used to approximate the 8910 internal dose of daily dichloromethane (mg) metabol ized via the GST pathway per unit volume 8911 of liver from the daily oral administered doses. T his approach was taken based on evidence that 8912 GST-pathway metabolites produced from dichlorometha ne are primarily responsible for 8913 dichloromethane carcinogenicity in mouse liver. Th e multistage dose-response model (BMDS 8914 version 2.0) was used to fit the mouse liver tumor incidence and PBTK model-derived internal 8915

dose data and derive a mouse internal BMD and BMDL associated with 10% extra risk 8916 (BMDL 10 ). The human BMDL 10 was derived by multiplying the mouse BMDL 10 by allometric 8917 scaling factor (BW human /BW mouse ) 0.25 ≈ 7). Linear extrapolation from the internal human 8918

BMDL 10 (0.1/BMDL 10 ) was used to derive oral risk factors for liver tu mors based on tumor 8919 responses in male mice. The linear low-dose extrap olation approach for agents with a mutagenic 8920

mode of action was selected because GST-metabolism of dichloromethane is expected to occur 8921 at and below exposures producing the mouse BMDL 10 , even though CYP2E1 metabolism is 8922 expected to be unsaturated and to represent the pre dominant metabolic pathway in the liver. 8923 Currently, there are no data from chronic oral canc er bioassays in mice providing support for a 8924 nonlinear dose-response relationship. 8925 Probability distributions of human oral cancer slo pe factors were derived by using a 8926 human PBTK model (adapted from David et al. [2006]; see Appendix B). The cancer reference 8927 values (OSF and IUR) were derived for a sensitive p opulation: a population composed entirely 8928

of carriers of the GST-T1 +/+ homozygous genotype (that is, the group that would be expected to 8929 be most sensitive to the carcinogenic effects of di chloromethane). In addition, cancer values 8930 derived for a population reflecting the estimated f requency of GST-T1 genotypes in the current 8931 U.S. population (20% GST-T1 –/– , 48% GST-T1 +/– , and 32% GST-T1 +/+ ) were presented. All 8932

simulations also included a distribution of CYP act ivity based on data from Lipscomb et al. 8933 (2003). The mean OSF based on liver tumors in mice exposed to dichloromethane in drinking 8934 water, 1 × 10 –3 (mg/kg-day) –1 , based on what is assumed to be the most sensitive of the 8935318

populations, the GST-T1 +/+ group, is the recommended OSF to be used in determ inistic risk 8936 assessments for chronic oral exposures to dichlorom ethane. 8937

An OSF derived from the liver tumor data in the S erota et al. (1986b) study, using 8938 administered dose dosimetry rather than PBTK modeli ng, is approximately one order of 8939 magnitude higher than the current recommended value of 1 × 10 –3 (per mg/kg-day). There is 8940 approximately one to two orders of magnitude differ ence among the values based on different 8941 dose metrics, scaling factors, and populations (Tab le 6-1). 8942

The recommended OSF of 1 × 10 –3 (per mg/kg-day) is based on a tissue-specific GST 8943 internal dose metric with allometric scaling. Alth ough the involvement of the GST pathway in 8944 carcinogenic response has been established, some un certainty remains as to the metabolite(s) 8945

involved and the rate at which those metabolites ar e cleared. The value derived specifically for 8946

the GST-T1 +/+ population is recommended to provide protection fo r the population that is 8947 hypothesized to be most sensitive to the carcinogen ic effect. Application of ADAFs to the 8948 cancer OSF is recommended in combination with appro priate exposure data when assessing 8949 risks associated with early-life exposure (see sect ion 5.4.4 for more details). 8950 89513198952Table 6-1. Comparison of OSFs derived by using various assumptions and metrics, based on liver tumorsin male micePopulationaDose metricSpecies,sexTumorScalingfactorMean OSF(mg/kg-day)–1Source(Table)GST-T1+/+Tissue-specific GST-metabolism ratebMouse, male Liver 7.0 1.4 × 10–3Table 5-13Tissue-specific GST-metabolism rate Mouse, male Liver 1.0 2.0 × 10–4Table 5-13Whole-body GST metabolism rate Mouse, male Liver 7.0 8.1 × 10–4Table 5-13Route-to-route extrapolation, tissue-specific metabolism Mouse, male Liver 7.0 1.0 × 10–4Table 5-14Route-to-route extrapolation, tissue-specific metabolism Mouse, male Liver 1.0 1.5 × 10–5Table 5-14Route-to-route extrapolation, whole-body metabolism Mouse, male Liver 7.0 5.8 × 10–5Table 5-14Mixed Tissue-specific GST-metabolism ratebMouse, male Liver 7.0 8.0 × 10–4Table 5-13Tissue-specific GST-metabolism rate Mouse, male Liver 1.0 1.2 × 10–4Table 5-13Whole-body GST metabolism rate Mouse, male Liver 7.0 4.6 × 10–4Table 5-13Route-to-route extrapolation, tissue-specific metabolism Mouse, male Liver 7.0 5.8 × 10–5Table 5-14Route-to-route extrapolation, tissue-specific metabolism Mouse, male Liver 1.0 8.3 × 10–6Table 5-14Route-to-route extrapolation, whole-body metabolism Mouse, male Liver 7.0 3.3 × 10–5Table 5-14Applied dose (human equivalent dose) Mouse, male Liver1.0×10–2Table 5-151995 IRIS assessment Mouse, male Liver7.5 × 10–3aGST-T1+/+= homozygous, full enzyme activity; Mixed = genotypes based on a population reflecting the estimatedfrequency of genotypes in the currentU.S. population: 20% GST-T1–/–, 48% GST-T1+/–, and 32% GST-T1+/+(Haber et al., 2002).Bolded value is the basis for the recommended OSF of 1×10–3per mg/kg-day.320 8953 6.2.5. Cancer Inhalation Unit Risk 8954

The recommended cancer IUR is 1 × 10

– 8 ( μ g/m 3 ) –1 for the development of liver and lung 8955 cancers, based on data from male B6C3F 1 mice, using a tissue-specific GST metabolism dose 8956 metric. Data for liver and lung tumors in male and female B6C3F 1 mice, following exposure to 8957 airborne dichloromethane, were used to develop IURs for dichloromethane (Mennear et al., 8958 1988; NTP, 1986). This study was selected as the p rincipal study to derive an IUR for 8959 dichloromethane because of the completeness of the data, adequate sample size, and clear dose 8960 response. In the NTP (1986) study, significant inc reases in incidence of liver and lung adenomas 8961

and carcinomas were observed in both sexes of B6C3F 1 mice exposed 6 hours/day, 5 days/week 8962 for 2 years. 8963 The PBTK model of Marino et al. (2006) for dichloro methane in the mouse was used to 8964 calculate long-term daily average internal liver do ses. The selected internal dose metrics for 8965 liver tumors and lung tumors were long-term average daily mass of dichloromethane 8966 metabolized via the GST pathway per unit volume of liver and lung, respectively. This approach 8967 was taken based on evidence that GST-pathway metabo lites produced from dichloromethane are 8968 primarily responsible for dichloromethane carcinoge nicity in mouse liver. The multistage dose8969 response model (BMDS version 2.0) was used to fit t he mouse liver tumor incidence and PBTK 8970

model-derived internal dose data and derive a mouse internal BMD and BMDL associated with 8971 10% extra risk (BMDL 10 ). The human BMDL 10 was derived by multiplying the mouse BMDL 10 8972 by allometric scaling factor (BW human /BW mouse ) 0.25 ≈ 7). A linear extrapolation approach using 8973

the internal human BMDL 10 for liver and lung tumors was used to calculate hu man tumor risk 8974 factors by dividing the BMR of 0.1 by the human BMD L for each tumor type. Currently, there 8975 are no data from chronic inhalation cancer bioassay s in mice or rats providing support for a 8976 nonlinear dose-response relationship. 8977 The human PBTK model (adapted from David et al. [20 06]; see Appendix B) provided 8978

distributions of human internal dose metrics of dai ly mass of dichloromethane metabolized via 8979 the GST pathway per unit volume of liver and lung r esulting from chronic inhalation exposure to 8980

a unit concentration of 1 μ g/m 3 dichloromethane (0.00029 ppm). As with the OSF, t he cancer 8981 IUR was derived for a sensitive population: a popul ation composed entirely of carriers of the 8982 GST-T1 homozygous positive genotype (that is, the g roup that would be expected to be most 8983 sensitive to the carcinogenic effects of dichlorome thane). In addition, cancer values derived for 8984 a population reflecting the estimated frequency of GST-T1 genotypes in the current U.S. 8985 population (20% GST-T1 –/– , 48% GST-T1 +/– , and 32% GST-T1 +/+ ) were also presented. The 8986 distributions of IURs for liver or lung tumors were generated by multiplying the human tumor 8987 risk factor for each tumor type and sex by the dist ribution of internal doses from chronic 8988 exposure to 1 μ g/m 3 dichloromethane. A procedure to combine risks for liver and lung tumors 8989 using different dose metrics for the different tumo rs (i.e., liver-specific and lung-specific 8990321

metabolism for liver and lung tumors, respectively) , was used to derive the recommended IUR of 8991

1 × 10 – 8 ( μ g/m 3 ) –1 based on what is assumed to be the most sensitive o f the populations, the 8992

GST-T1 +/+ group. 8993

The current recommended IUR value of 1 × 10 – 8 ( μ g/m 3 ) – 1 is approximately 1.5 orders 8994 of magnitude lower than the previous IRIS value of 4.7 × 10 – 7 ( μ g/m 3 ) – 1 and similar to the 8995 occupational exposure-based risk value of 4.17 × 1 0 – 8 ( μ g/m 3 ) – 1 promulgated by OSHA (1997) 8996

and derived from an estimated risk of 3.62 × 10 – 3 for a lifetime occupational inhalation exposure 8997 of 25 ppm. The current use of the updated mouse PB TK model, with blood and tissue 8998 equilibrium partition coefficients and metabolic pa rameters updated with MCMC calibration, 8999 resulted in approximately three- and four-fold incr eases in the values of internal liver and lung 9000 dose metrics, respectively, associated with the dic hloromethane exposure concentrations, 9001 compared with estimates from the model used in the previous IRIS assessment. For a unit 9002 inhalation exposure, the mean internal lung GST dose predicted for the entire population 9003 predicted by the MCMC updated human PBTK model is a pproximately thirteen-fold lower 9004 compared with the human PBTK model used in the U.S. EPA (1995) assessment. The mean 9005

internal liver GST dose, however, is approximately the same as (only 16% higher than) that 9006 obtained with the previous PBTK parameters. For un it oral exposures, the mean internal liver 9007

GST

dose predicted by the MCM updated model is about 8 0% of that predicted using the 9008 previous parameters, while the mean whole-body GST dose is predicted to be about 50% of that 9009 predicted using the previous parameters. 9010

An IUR derived from the liver tumor data of the NT P (1986) study using applied 9011 concentration dosimetry rather than PBTK modeling, 3.7 × 10 –7 ( μ g/m 3 ) –1 , is approximately one 9012 order of magnitude higher than the currently recomm ended value of 1 × 10 –8 ( μ g/m 3 ) –1 (Table 9013 6-2). There is approximately one- to two- orders o f magnitude difference among the values 9014 based on different dose metrics, scaling factors, a nd populations. 9015

The recommended IUR value of 1 × 10 –8 ( μ g/m 3 ) –1 is based on a tissue-specific GST9016 internal dose metric with allometric scaling. Alth ough the involvement of the GST pathway in 9017 carcinogenic response has been established, some un certainty remains as to the metabolite(s) 9018

involved and the rate at which those metabolites ar e cleared. The value derived specifically for 9019

the GST-T1 +/+ population is recommended to provide protection fo r the population that is 9020 hypothesized to be most sensitive to the carcinogen ic effect. Application of ADAFs to the 9021 cancer IUR is recommended when assessing risks asso ciated with early-life exposure (see 9022 section 5.4.4 for more details). 902332290249025Table 6-2. Comparison of IURs derived by using various assumptions and metricsPopulationaDose metricSpecies, sex, Tumor typeScalingfactorIURb(μg/m3)–1Source(Table)GST-T1+/+Tissue-specific GST-metabolism ratecMouse, male Liver and lung7.0 1.1 × 10–8Table 5-20GST-T1+/+Tissue-specific GST-metabolism rate Mouse, male Liver 7.0 7.3 × 10–9Table 5-19GST-T1+/+Tissue-specific GST-metabolism rate Mouse, male Lung 7.0 4.8 × 10–9Table 5-19GST-T1+/+Tissue-specific GST-metabolism rate Mouse, male Liver and lung 1.0 1.6 × 10–9Table 5-20GST-T1+/+Tissue-specific GST-metabolism rate Mouse, male Liver 1.0 1.0 × 10–9Table 5-19GST-T1+/+Tissue-specific GST-metabolism rate Mouse, male Lung 1.0 6.8 × 10–10Table 5-19GST-T1+/+Whole-body GST metabolism rate Mouse, male Liverand lung 7.0 1.4 × 10–8Table 5-20GST-T1+/+Whole-body GST metabolism rate Mouse, male Liver7.0 4.6 × 10–9Table 5-19GST-T1+/+Whole-body GST metabolism rate Mouse, male Lung 7.0 1.0 × 10–8Table 5-19Mixed Tissue-specific GST-metabolism rate Mouse, male Liver and lung 7.0 5.5 × 10–9Table 5-20Mixed Tissue-specific GST-metabolism rate Mouse, male Liver 7.0 3.4 × 10–9Table 5-19Mixed Tissue-specific GST-metabolism rate Mouse, male Lung 7.0 2.6 × 10–9Table 5-19Mixed Tissue-specific GST-metabolism rate Mouse, male Liver and lung 1.0 7.9 × 10–10Table 5-20Mixed Tissue-specific GST-metabolism rate Mouse, male Liver 1.0 4.8 × 10–10Table 5-19Mixed Tissue-specific GST-metabolism rate Mouse, male Lung 1.0 3.7 × 10–10Table 5-19Mixed Whole-body GST metabolism rate Mouse, male Liver and lung 7.0 7.9 × 10–9Table 5-20Mixed Whole-body GST metabolism rate Mouse, male Liver 7.0 2.7 × 10–9Table 5-19Mixed Whole-body GST metabolism rate Mouse, male Lung 7.0 6.0 × 10–9Table 5-19Administered concentration (HEC) Mouse, male Liver3.7 × 10–7Table 5-21Administered concentration (HEC)Mouse, male Lung9.2 × 10–7Table 5-211995 IRIS assessmentcMouse, male Liver and lung 12.7 4.7 × 10–7aGST-T1+/+= homozygous, full enzyme activity; Mixed = genotypes based on a population reflecting the estimatedfrequency of genotypes in the current U.S.population: 20% GST-T1–/–, 48% GST-T1+/–, and 32% GST-T1+/+(Haber et al., 2002).bBased on mean value of the derived distributionsBolded value is the basis for the recommended IUR of 1 × 10–8μg/m3per mg/kg-day.9026323

6.2.6. Uncertainties in Cancer Risk Values 9027 The database of animal bioassays identifies the li ver and lung as the most sensitive target 9028 organs for dichloromethane-induced tumor developmen t, and there is high confidence that the 9029 dose-response data for liver and lung cancer in mic e represent the best available data for 9030 derivation of human cancer risks. A dose-response relationship was seen with respect to liver 9031 cancer in mice exposed orally and with respect to l iver and lung cancer in mice exposed by 9032 inhalation. Statistically significant increases in benign mammary gland tumors were observed in 9033 one study of F344 rats exposed by inhalation to 2,0 00 or 4,000 ppm (Mennear et al., 1988; NTP, 9034 1986); evidence for a tumorigenic mammary gland res ponse in Sprague-Dawley rats was limited 9035 to increased numbers of benign mammary tumors per a nimal at levels of 50–500 ppm (Nitschke 9036 et al., 1988a) or 500–3,500 ppm (Burek et al., 1984 ). An oral (gavage) study in female Sprague9037 Dawley rats reported an increased incidence of mali gnant mammary tumors, mainly 9038 adenocarcinomas (8, 6, and 18% in the control, 100, and 500 mg/kg dose groups, respectively), 9039 but the increase was not statistically significant. Data were not provided to allow an analysis that 9040 accounts for differing mortality rates (Maltoni et al., 1988). The toxicokinetic or mechanistic 9041 events that might lead to mammary gland tumor devel opment in rats are unknown, although 9042 CYP2E1 (El-Rayes et al., 2003; Hellmold et al., 199 8) and GST-T1 expression has been detected 9043 in human mammary tissue (Lehmann and Wagner, 2008). Rare CNS tumors were observed in 9044 one study in rats, a study spanning a relatively lo w range of exposures (0–500 ppm) (Nitschke et 9045 al., 1988a). These cancers were not seen in two oth er studies in rats, both involving higher doses 9046 (1,000–4,000 ppm) (NTP, 1986; Burek et al., 1984), or in a similar high-dose study in mice 9047 (NTP, 1986). The relative rarity of the tumors pre cludes the use of the low-dose exposure study 9048 (Nitschke et al., 1988a) in a quantitative dose-res ponse assessment. The available epidemiologic 9049 studies provide some evidence of an association bet ween dichloromethane and brain cancer. The 9050 available epidemiologic studies do not provide an a dequate basis for the evaluation of the role of 9051

dichloromethane in breast cancer because there are currently no cohort studies with adequate 9052 statistical power and no case-control studies with adequate exposure methodology to examine 9053 this relationship. 9054 There is uncertainty as to whether the reactivity of the toxic dichloromethane metabolites 9055 is sufficiently high enough to preclude systemic di stribution. Therefore, alternative derivations 9056 of cancer risk values were performed under the assu mption that high reactivity leads to complete 9057

clearance from the tissue in which the active metab olite is formed (scaling factor = 1.0). The 9058 difference in scaling factor (7.0 for allometric sc aling versus 1.0) results in a 7-fold decrease in 9059 estimated cancer toxicity values. Using a whole-bo dy GST metabolism dose metric, the 9060 resulting OSF and IUR for liver and lung cancer wer e approximately five-fold higher than when 9061 tissue-specific dose metrics were used (Table 5-16 and Table 5-22). The mechanistic data 9062 support the notion that reactive metabolites produc ed in the target tissues are not well distributed 9063 and produce deleterious effects in the metabolizing tissues soon after generation. Thus, there is 9064324

less uncertainty in the cancer risk values derived by using a tissue-specific GST metabolism dose 9065 metric compared with those derived using a whole-bo dy GST metabolism dose metric. 9066 Uncertainty in the ability of the PBTK models to es timate animal and human internal 9067 doses from lifetime bioassay low-level environmenta l exposures may affect the confidence in the 9068 cancer risk extrapolated from animal data. Uncerta inties in the mouse and human model 9069 parameter values were integrated quantitatively int o parameter estimation by utilizing 9070 hierarchical Bayesian methods to calibrate the mode ls at the population level (David et al., 2006; 9071 Marino et al., 2006). The use of Monte Carlo sampl ing to define human model parameter 9072

distributions allowed for derivation of human distr ibutions of dosimetry and cancer risk, 9073 providing for bounds on the recommended risk values . However, the PBTK animal models were 9074 utilized deterministically, and a sensitivity analy sis was performed to identify model parameters 9075 most influential on the predictions of dose metrics used to estimate oral and inhalation cancer 9076 risks. For inhalation exposures in mice, the blood :air partition coefficient, followed closely by 9077 the first-order GST-mediated metabolism rate, had t he greatest impact on the dose metric for 9078 liver cancer (mg dichloromethane metabolized via GS T pathway per liter liver per day). For 9079 drinking water exposures in mice, the first-order G ST-mediated metabolism rate, followed by the 9080 CYP-mediated maximum reaction velocity (V max c) affected the liver cancer dose metric to the 9081 greatest extent. For mice inhaling dichloromethane , the lung cancer dose metric (mg 9082 dichloromethane metabolized via GST pathways per li ter lung per day), like the liver cancer 9083 metric, was highly affected by the first-order GST- mediated metabolism rate and the blood:air 9084 partition coefficient. However, the lung cancer do se metric was most sensitive to the 9085 proportional yield of liver GST-mediated metabolic activity attributed to the lung. The blood:air 9086 partition coefficient was experimentally determined , lending high confidence to its value. In 9087 contrast, values for the three metabolic parameters were determined by computational 9088 optimization against data sets not directly measuri ng dichloromethane or its metabolites in the 9089 target/metabolizing tissues. It is uncertain how a lternative values for these three parameters 9090

would affect the estimation of animal BMDL 10 values and, ultimately, IURs and OSFs. In 9091 addition, specific uncertainty remains concerning t he human PBTK parameter distributions. In 9092 addition, while the structure and equations used in the existing model have been described 9093 extensively in peer-reviewed publications, uncertai nty remains concerning the model structure, 9094 and specifically the potential of an alternative (d ual-binding-site) CYP metabolic rate equation 9095 for dichloromethane Integration of the alternate r ate equation into the PBTK modeling, and then 9096 quantitative risk assessment, will likely require s everal years of further research, and hence is 9097 beyond the scope of the current assessment. 9098325 7. REFERENCES

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Andersen, ME; Clewell, HJ, III; Gargas, ML; et al. (1991) Physiologically based pharmacokinetic modeling with dichloromethane, its metabolite, carb on monoxide, and blood carboxyhemoglobin in rats and humans. Toxicol Appl Pharmacol 108:14–27.

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Angelo, MJ; Pritchard, AB; Hawkins, DR; et al. (198 6a) The pharmacokinetics of dichloromethane. I. Disposition in B6C3F 1 mice following intravenous and oral administration . Food Chem Toxicol 24(9):965–974.

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