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www.epa.gov/iris

1 

2 TOXICOLOGICAL REVIEW

3 

4 OF

5 

6 DICHLOROMETHANE 

7 (METHYLENE CHLORIDE) 

8 

9 (CAS No. 75-09-2) 

10 

11 

12 

13 In Support of Summary Information on the

14 Integrated Risk Information System (IRIS)

15 

16 December 2009 

17 

18 

19 

20 NOTICE 

21 

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

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

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

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

26 of its technical accuracy and science policy implications. 

27 

28 

29 

30 U.S. Environmental Protection Agency 

31 Washington, DC 
32 

33 DISCLAIMER 

34 

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

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

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

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

39 names or commercial products does not constitute endorsement or recommendation for use. 

40 

41 
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42 CONTENTS—TOXICOLOGICAL REVIEW OF DICHLOROMETHANE 

43 (CAS No. 75-09-2) 

44 

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

46 LIST OF TABLES........................................................................................................................ vii 

47 LIST OF FIGURES ..................................................................................................................... xiii 

48 LIST OF ACRONYMS .............................................................................................................. xvii 

49 FOREWORD ............................................................................................................................... xix 

50 AUTHORS, CONTRIBUTORS, AND REVIEWERS ................................................................ xx

51 1. INTRODUCTION ......................................................................................................................1 

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

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

54 3.1. ABSORPTION.................................................................................................................. 5 

55 3.1.1. Oral — Gastrointestinal Tract Absorption ............................................................. 5 

56 3.1.2. Inhalation—Respiratory Tract Absorption............................................................. 5 

57 3.2. DISTRIBUTION ............................................................................................................... 7 

58 3.3. METABOLISM................................................................................................................. 9 

59 3.3.1. The CYP2E1 Pathway.......................................................................................... 11 

60 3.3.2. The GST Pathway................................................................................................. 14 

61 3.4. ELIMINATION............................................................................................................... 19 

62 3.5. PHYSIOLOGICALLY BASED TOXICOKINETIC MODELS .................................... 21 

63 3.5.1. Probabilistic Mouse PBTK Dichloromethane Model (Marino et al., 2006) ........ 26 

64 3.5.2. Probabilistic Human PBTK Dichloromethane Model (David et al., 2006) ......... 31 

65 3.5.3. Evaluation of Rat PBTK Dichloromethane Models............................................. 38 

66 3.5.4. Comparison of Mouse, Rat and Human PBTK Models....................................... 39 

67 HAZARD IDENTIFICATION......................................................................................................43 

68 4.1. STUDIES IN HUMANS ................................................................................................. 43 

69 4.1.1. Introduction—Case Reports, Epidemiologic, and Clinical Studies ..................... 43 

70 4.1.2. Noncancer Studies................................................................................................ 43 

71 4.1.2.1. Case Reports of Acute, High-dose Exposures........................................ 43 

72 4.1.2.2. Controlled Experiments Examining Acute Effects.................................. 44 

73 4.1.2.3. Observational Studies Focusing on Clinical Chemistries, Clinical 

74 Examinations, and Symptoms................................................................. 45 

75 4.1.2.4. Observational Studies Using Workplace Medical Program Data ......... 51 

76 4.1.2.5. Studies of Ischemic Heart Disease Mortality Risk ................................. 55 

77 4.1.2.6. Studies of Suicide Risk............................................................................ 56 

78 4.1.2.7 Studies of Infectious Disease Risk .......................................................... 57 

79 4.1.2.8. Studies of Reproductive Outcomes ......................................................... 57 

80 4.1.2.9. Summary of Noncancer Studies.............................................................. 60 

81 4.1.3. Cancer Studies...................................................................................................... 62 

82 4.1.3.1. Identification and Selection of Studies for Evaluation of Cancer 

83 Risk ......................................................................................................... 62 

84 4.1.3.2. Description of the Selected Studies......................................................... 63 

85 4.1.3.3. Cellulose Triacetate Film Base Production Cohorts.............................. 63 

86 4.1.3.4. Cellulose Triacetate Fiber Production Cohorts..................................... 70 

87 4.1.3.5. Solvent-Exposed Workers—Hill Air Force Base, Utah.......................... 75 

88 4.1.3.6. Case-Control Studies of Specific Cancers and Dichloromethane.......... 76 

89 4.1.3.7. Summary of Cancer Studies by Type of Cancer ..................................... 83 
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90 4.2. SUBCHRONIC AND CHRONIC STUDIES AND CANCER BIOASSAYS IN 

91 ANIMALS—ORAL AND INHALATION .................................................................... 92 

92 4.2.1. Oral Exposure: Overview of Noncancer and Cancer Effects............................... 92 

93 4.2.1.1. Toxicity Studies of Subchronic Oral Exposures: Hepatic Effects .......... 93 

94 4.2.1.2. Toxicity Studies of Chronic Oral Exposures: Hepatic Effects and 

95 Carcinogenicity....................................................................................... 96 

96 4.2.2. Inhalation Exposure: Overview of Noncancer and Cancer Effects.................... 102 

97 4.2.2.1. Toxicity Studies of Subchronic Inhalation Exposures: General, 

98 Renal, and Hepatic Effects ................................................................... 103 

99 4.2.2.2. Toxicity Studies from Chronic Inhalation Exposures........................... 107 

100 4.3. REPRODUCTIVE/DEVELOPMENTAL STUDIES—ORAL AND 

101 INHALATION.............................................................................................................. 124 

102 4.3.1. Reproductive Toxicity Studies ........................................................................... 127 

103 4.3.1.1. Oral (Gavage) Studies.......................................................................... 127 

104 4.3.1.2. Inhalation Studies................................................................................. 128 

105 4.3.2. Developmental Toxicity Studies ........................................................................ 129 

106 4.3.2.1. Oral (Gavage) Studies and Culture Studies......................................... 129 

107 4.3.2.2. Inhalation Studies................................................................................. 130 

108 4.4. OTHER DURATION- OR ENDPOINT-SPECIFIC STUDIES ................................... 132 

109 4.4.1. Short-term (2-Week) Studies of General and Hepatic Effects in Animals ........ 132 

110 4.4.2. Immunotoxicity Studies in Animals................................................................... 133 

111 4.4.3. Neurotoxicology Studies in Animals.................................................................. 135 

112 4.4.3.1. Neurotoxicology Studies—Oral Exposures.......................................... 140 

113 4.4.3.2. Neurotoxicology Studies—Inhalational Exposure................................ 141 

114 4.5. MECHANISTIC DATA AND OTHER STUDIES IN SUPPORT OF THE 

115 MODE OF ACTION..................................................................................................... 148 

116 4.5.1. Genotoxicity Studies .......................................................................................... 148 

117 4.5.1.1. In Vitro Genotoxicity Assays ................................................................ 148 

118 Bacterial, Yeast, and Fungi mutagenicity assays............................................... 148 

119 4.5.1.2. In Vivo Genotoxicity Assays ................................................................. 158 

120 4.5.2. Mechanistic Studies of Liver Effects ................................................................. 167 

121 4.5.3. Mechanistic Studies of Lung Effects.................................................................. 170 

122 4.5.4. Mechanistic Studies of Neurological Effects..................................................... 175 

123 4.6. SYNTHESIS OF MAJOR NONCANCER EFFECTS ................................................. 177 

124 4.6.1. Oral Exposures ................................................................................................... 177 

125 4.6.1.1. Summary of Human Data ..................................................................... 177 

126 4.6.1.2. Summary of Animal Data...................................................................... 178 

127 4.6.2. Inhalation Exposures .......................................................................................... 182 

128 4.6.2.1. Summary of Human Data ..................................................................... 182 

129 4.6.2.2. Summary of Animal Studies.................................................................. 183 

130 4.6.3. Mode of Action Information .............................................................................. 190 

131 4.6.3.1. Mode of Action for Nonneoplastic Liver Effects................................... 190 

132 4.6.3.2. Mode of Action for Nonneoplastic Lung Effects................................... 191 

133 4.6.3.3. Mode of Action for Neurological Effects.............................................. 191 

134 4.6.3.4. Mode of Action for Reproductive and Developmental Effects.............. 192 

135 4.6.3.5. Mode of Action for Immunotoxicity ...................................................... 193 

136 4.7. EVALUATION OF CARCINOGENICITY ................................................................. 194 

137 4.7.1. Summary of Overall Weight of Evidence .......................................................... 194 

138 4.7.2. Synthesis of Human, Animal, and Other Supporting Evidence ......................... 195 
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139 4.7.3. Mode of Action Information .............................................................................. 204 

140 4.7.3.1. Hypothesized Mode of Action ............................................................... 204 

141 4.7.3.2. General Conclusions About the Mode of Action for Tumors in 

142 Rodents and Possible Relevance to Humans........................................ 211 

143 4.8. SUSCEPTIBLE POPULATIONS AND LIFE STAGES ............................................. 213 

144 4.8.1. Possible Childhood Susceptibility...................................................................... 213 

145 4.8.2. Possible Gender Differences .............................................................................. 215 

146 4.8.3. Other Susceptible Populations............................................................................ 215 

147 5. DOSE-RESPONSE ASSESSMENTS....................................................................................217 

148 5.1. ORAL REFERENCE DOSE (RfD) .............................................................................. 217 

149 5.1.1. Choice of Principal Study and Critical Effect—with Rationale and 

150 Justification........................................................................................................ 217 

151 5.1.2. Derivation Process for Noncancer Reference Values ........................................ 220 

152 5.1.3. Evaluation of Dose Metrics for Use in Noncancer Reference Value 

153 Derivations......................................................................................................... 223 

154 5.1.4. Methods of Analysis—Including Models (PBTK, BMD, etc.).......................... 224 

155 5.1.5. RfD Derivation—Including Application of Uncertainty Factors (UFs)............. 229 

156 5.1.6. Previous RfD Assessment ................................................................................... 230 

157 5.1.7. RfD Comparison Information............................................................................. 230 

158 5.2. INHALATION REFERENCE CONCENTRATION (RfC)......................................... 234 

159 5.2.1. Choice of Principal Study and Critical Effect—with Rationale and 

160 Justification........................................................................................................ 234 

161 5.2.2. Derivation Process for Reference Concentration Values ................................... 238 

162 5.2.3. Methods of Analysis—Including Models (PBTK, BMD, etc.).......................... 238 

163 5.2.4. RfC Derivation—Including Application of Uncertainty Factors (UFs)............. 243 

164 5.2.5. Previous RfC Assessment................................................................................... 245 

165 5.2.6. RfC Comparison Information............................................................................. 245 

166 5.3. UNCERTAINTIES IN THE ORAL REFERENCE DOSE AND INHALATION 

167 REFERENCE CONCENTRATION............................................................................. 250 

168 5.4. CANCER ASSESSMENT ............................................................................................ 262 

169 5.4.1. Cancer Oral Slope Factor ................................................................................... 263 

170 5.4.1.1. Choice of Study/Data—with Rationale and Justification ..................... 263 

171 5.4.1.2. Derivation Process for Cancer Oral Slope Factor .............................. 263 

172 5.4.1.3. Dose-Response Data............................................................................. 266 

173 5.4.1.4. Dose Conversion and Extrapolation Methods: Cancer Oral Slope 

174 Factor ................................................................................................... 266

175 5.4.1.5. Oral Cancer Slope Factor.................................................................... 273 

176 5.4.1.6. Alternative Derivation Based on Route-to-Route Extrapolation.......... 273 

177 5.4.1.8. Previous IRIS Assessment: Cancer Oral Slope Factor ........................ 276 

178 5.4.1.9. Comparison of Cancer Oral Slope Factors Using Different 

179 Methodologies....................................................................................... 276 

180 5.4.2. Cancer Inhalation Unit Risk ............................................................................... 278 

181 5.4.2.1. Choice of Study/Data—with Rationale and Justification ..................... 278 

182 5.4.2.2. Derivation Process for Cancer Inhalation Unit Risk ........................... 279 

183 5.4.2.3. Dose-Response Data............................................................................. 279 

184 5.4.2.4. Dose Conversion and Extrapolation Methods: Cancer Inhalation 

185 Unit Risk ............................................................................................... 280 

186 5.4.2.5. Cancer Inhalation Unit Risk................................................................. 287 

187 5.4.2.8. Previous IRIS Assessment: Cancer Inhalation Unit Risk..................... 291 
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188 5.4.2.8. Comparison of Cancer Inhalation Unit Risk Using Different 

189 Methodologies....................................................................................... 292 

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

191 Assessment......................................................................................................... 294 

192 5.4.4. Application of ADAFs ....................................................................................... 296 

193 5.4.4.1. Application of ADAFs in Oral Exposure Scenarios............................. 297 

194 5.4.4.2. Application of ADAFs in Inhalation Exposure Scenarios.................... 297 

195 5.4.5. Uncertainties in Cancer Risk Values.................................................................. 298 

196 6. MAJOR CONCLUSIONS IN THE CHARACTERIZATION OF HAZARD AND 

197 DOSE RESPONSE ................................................................................................................310 

198 6.1. HUMAN HAZARD POTENTIAL ............................................................................... 310 

199 6.2. DOSE-RESPONSE ....................................................................................................... 313 

200 6.2.1. Oral RfD ............................................................................................................. 313 

201 6.2.2. Inhalation RfC .................................................................................................... 314 

202 6.2.3. Uncertainties in Reference Dose and Reference Concentration Values ............ 315 

203 6.2.4. Oral Cancer Slope Factor ................................................................................... 316 

204 6.2.5. Cancer Inhalation Unit Risk ............................................................................... 320 

205 6.2.6. Uncertainties in Cancer Toxicity Values............................................................ 323 

206 7. REFERENCES .......................................................................................................................325 

207 

208 

209 APPENDIX A: EXTERNAL REVIEW PANEL COMMENTS .............................................. A-1 

210 

211 APPENDIX B: HUMAN PBTK DICHLOROMETHANE MODEL ........................................B-1 

212 B-1. Human Model Description...............................................................................................B-1 

213 B-2. Revisions to Parameter Distributions of David et al. (2006) ...........................................B-3 

214 B-3. CYP2E1 and GST-T1 ......................................................................................................B-4 

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

216 B-4.1. Age................................................................................................................................B-8 

217 B-4.2. Gender...........................................................................................................................B-9 

218 B-4.3. Body Weight ...............................................................................................................B-10 

219 B-4.4. Alveolar Ventilation....................................................................................................B-12 

220 B-4.5. Cardiac Output............................................................................................................B-13 

221 B-4.6. Fat Fraction .................................................................................................................B-14 

222 B-4.7. Liver Fraction..............................................................................................................B-14 

223 B-4.8. Tissue volume normalization......................................................................................B-16 

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

225 

226 APPENDIX C: RAT DICHLOROMETHANE PBTK MODELS ............................................C-1 

227 C-1. PBTK Models Selected for Comparison Against Dichloromethane Kinetic Data 

228 Sets .................................................................................................................................C-1 

229 C-2. Model Performance..........................................................................................................C-4 

230 C-2.1. Evaluation of Model Structure for Description of Carboxyhemoglobin Levels...........C-4 

231 C-2.2. Evaluation of Model Structure for Prediction of Uptake, Blood and Liver 

232 Concentrations, and Expiration of Dichloromethane .....................................................C-5 

233 C-2.3. Evaluation of Model Structure on Relative Flux of CYP and GST Metabolism of 

234 Dichloromethane ............................................................................................................C-9 

235 C-2.4. Evaluation of Model Predictions of Oral Absorption of Dichloromethane................C-11 

236 C-3. Model Option Summary ................................................................................................C-15 
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237 

238 APPENDIX D: SUMMARY OF BENCHMARK DOSE (BMD) MODELING OF 

239 NONCANCER ENDPOINTS .......................................................................... D-1 

240 D-1. Oral Reference Dose: BMD Modeling of Nonneoplastic Liver Lesion Incidence 

241 Data For Rats Exposed to Dichloromethane In Drinking Water For 2 Years 

242 (Serota et al., 1986a) ..................................................................................................... D-1 

243 D-2. Inhalation Reference Concentration: BMD Modeling of Nonneoplastic Liver 

244 Lesion Incidence Data For Rats Exposed to Dichloromethane by Inhalation For 2 

245 Years (Nitschke et al., 1988a) ....................................................................................... D-5 

246 

247 APPENDIX E: SUMMARY OF BENCHMARK DOSE (BMD) MODELING OF CANCER 

248 ENDPOINTS.....................................................................................................E-1 

249 E-1. Oral Cancer Slope Factors: BMD Modeling of Liver Tumor Incidence Data for 

250 Mice Exposed to Dichloromethane in Drinking Water for 2 Years (Serota et al., 

251 1986b; Hazelton Laboratories, 1983).............................................................................E-1 

252 E-1.1. Modeling results for the internal liver metabolism metric............................................E-3 

253 E-1.2. Modeling results for the whole body metabolism metric .............................................E-5 

254 E-2. Cancer Inhalation Unit Risk: BMD Modeling of Liver and Lung Tumor 

255 Incidence Data for Male Mice Exposed to Dichloromethane via Inhalation for 2 

256 Years (Mennear et al., 1988; NTP, 1986). ...................................................................E-7 

257 E-2.1. Modeling results for the internal liver metabolism metric, liver tumors ......................E-9 

258 E-2.2. Modeling results for the internal lung metabolism metric, lung tumors.....................E-11 

259 E-2.3. Modeling results for the whole body metabolism metric, liver tumors......................E-13 

260 E-2.4. Modeling results for the whole body metabolism metric, lung tumors ......................E-15 

261 

262 APPENDIX F. COMPARATIVE CANCER INHALATION UNIT RISK BASED ON FEMALE 

263 MICE DATA .....................................................................................................F-1 

264 

265 APPENDIX G: COMPARATIVE CANCER INHALATION UNIT RISK BASED ON 

266 BENIGN MAMMARY GLAND TUMORS IN RATS................................... G-1 

267 

268 APPENDIX H: SOURCE CODE AND COMMAND FILES FOR DICHLOROMETHANE 

269 PBTK MODELS .............................................................................................. H-1 
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270 

271 LIST OF TABLES 

272 

273 Table 2-1. Physical properties and chemical identity of dichloromethane.................................... 3 

274 Table 3-1. Distribution of radioactivity in tissues 48 hours after inhalation exposure of mature 

275 male Sprague-Dawley rats (n = 3) for 6 hours ............................................................ 7 

276 Table 3-2. Brain and perirenal fat dichloromethane and blood CO concentrations in male Wistar 

277 rats exposed by inhalation to dichloromethane at constant exposure concentrations 

278 compared with intermittently high exposure concentrations....................................... 8 

279 Table 3-3. Mean prevalences of the GST-T1 null (–/–) genotype in human ethnic groups ........ 15 

280 Table 3-4. GST-T1 enzyme activities toward dichloromethane in human, rat, mouse, and 

281 hamster tissues (liver, kidney, and erythrocytes)....................................................... 17 

282 Table 3-5. Values for parameter distributions in a B6C3F1 mouse probabilistic PBTK model for 

283 dichloromethane compared with associated values for point parameters in earlier 

284 deterministic B6C3F1 mouse PBTK models for dichloromethane............................ 29 

285 Table 3-6. Internal daily doses for B6C3F1 mice exposed to dichloromethane for 2 years 

286 (6 hours/day, 5 days/week) calculated with different PBTK models ........................ 30 

287 Table 3-7. Results of calibrating metabolic parameters in a human probabilistic PBTK model for 

288 dichloromethane with individual kinetic data for 42 exposed volunteers and MCMC 

289 analysis ...................................................................................................................... 32 

290 Table 3-8. Parameter distributions used in human Monte Carlo analysis for dichloromethane by 

291 David et al. (2006) ..................................................................................................... 34 

292 Table 3-9. Parameter distributions for the human PBTK model for dichloromethane used by 

293 EPA............................................................................................................................ 36 

294 Table 3-10. Parameter values for the rat PBTK model for dichloromethane used by EPA ........ 39 

295 Table 3-11. Parameters in the mouse, rat, and human PBTK model for dichloromethane used by 

296 the EPA...................................................................................................................... 41 

297 Table 4-1. Percentage of male General Electric plastic polymer workers reporting neurologic 

298 symptoms or displaying abnormal values in measures of neurological function, 

299 hepatic function, and cardiac function....................................................................... 54 

300 Table 4-2. Ischemic heart disease mortality risk in four cohorts of dichloromethane-exposed 

301 workers ...................................................................................................................... 56 

302 Table 4-3. Suicide risk in two cohorts of dichloromethane-exposed workers............................. 57 

303 Table 4-4. Mortality risk in Eastman Kodak cellulose triacetate film base production workers, 

304 Rochester, New York................................................................................................. 66 

305 Table 4-5. Mortality risk by cumulative exposure in Eastman Kodak cellulose triacetate film 

306 base production workers, Rochester, New York ....................................................... 67 

307 Table 4-6. Mortality risk in Imperial Chemical Industries cellulose triacetate film base 

308 production workers, Brantham, United Kingdom: 1,473 men employed 1946–1988, 

309 followed through 1994............................................................................................... 70 

310 Table 4-7. Mortality risk in Hoechst Celanese Corporation cellulose triacetate fiber production 

311 workers, Rock Hill, South Carolina: 1,271 men and women employed 1954–1977, 

312 followed through 1990............................................................................................... 72 

313 Table 4-8. Cancer mortality risk in Hoechst Celanese Corporation cellulose triacetate fiber 

314 production workers, Cumberland, Maryland: 2,909 men and women employed 

315 1970–1981, followed through 1989........................................................................... 74 

316 Table 4-9. Summary of cohort studies of cancer risk and dichloromethane exposure................ 85 

317 Table 4-10. Summary of case-control studies of cancer risk and dichloromethane exposure..... 87 
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318 Table 4-11. Incidences of histopathologic changes in livers of male and female F344 rats 

319 exposed to dichloromethane in drinking water for 90 days....................................... 94 

320 Table 4-12. Incidences of histopathologic changes in livers of male and female B6C3F1 mice 

321 exposed to dichloromethane in drinking water for 90 days....................................... 96 

322 Table 4-13. Studies of chronic oral dichloromethane exposures (up to 2 years)......................... 97 

323 Table 4-14. Incidences of nonneoplastic liver changes and liver tumors in male and female F344 

324 rats exposed to dichloromethane in drinking water for 2 years................................. 98 

325 Table 4-15. Incidences for focal hyperplasia and tumors in the liver of male B6C3F1 mice 

326 exposed to dichloromethane in drinking water for 2 years...................................... 100 

327 Table 4-16. Studies of chronic inhalation dichloromethane exposures ..................................... 108 

328 Table 4-17. Incidences of nonneoplastic histologic changes in male and female F344/N rats 

329 exposed to dichloromethane by inhalation (6 hours/day, 5 days/week) for 2 years 110 

330 Table 4-18. Incidences of selected neoplastic lesions in male and female F344/N rats exposed to 

331 dichloromethane by inhalation (6 hours/day, 5 days/week) for 2 years .................. 112 

332 Table 4-19. Incidences of nonneoplastic histologic changes in B6C3F1 mice exposed to 

333 dichloromethane by inhalation (6 hours/day, 5 days/week) for 2 years .................. 114 

334 Table 4-20. Incidences of neoplastic lesions in male and female B6C3F1 mice exposed to 

335 dichloromethane by inhalation (6 hours/day, 5 days/week) for 2 years .................. 115 

336 Table 4-21. Incidences of selected neoplastic histologic changes in male and female Sprague337 Dawley rats exposed to dichloromethane by inhalation (6 hours/day, 5 days/week) 

338 for 2 years ................................................................................................................ 119 

339 Table 4-22. Incidences of selected nonneoplastic histologic changes in male and female 

340 Sprague-Dawley rats exposed to dichloromethane by inhalation (6 hours/day, 

341 5 days/week) for 2 years .......................................................................................... 121 

342 Table 4-23. Incidences of selected neoplastic histologic changes in male and female Sprague343 Dawley rats exposed to dichloromethane by inhalation (6 hours/day, 5 days/week) 

344 for 2 years ................................................................................................................ 123 

345 Table 4-24. Summary of studies of reproductive and developmental effects of dichloromethane 

346 exposure in animals ................................................................................................. 125 

347 Table 4-25. Reproductive outcomes in F344 rats exposed to dichloromethane by inhalation for 

348 14 weeks prior to mating and from GDs 0–21 ........................................................ 128 

349 Table 4-26. Studies of neurobehavioral changes from dichloromethane, by route of exposure and 

350 type of effect ............................................................................................................ 137 

351 Table 4-27. Studies of neurophysiological changes as measured by evoked potentials resulting 

352 from dichloromethane, by route of exposure........................................................... 138 

353 Table 4-28. Studies of neurochemical changes from dichloromethane, by route of exposure. 139 

354 Table 4-29. Results from in vitro genotoxicity assays of dichloromethane with bacteria, yeast, or 

355 fungi......................................................................................................................... 149 

356 Table 4-30. Results from in vitro genotoxicity assays of dichloromethane with mammalian 

357 systems, by type of test............................................................................................ 155 

358 Table 4-31. Results from in vivo genotoxicity assays of dichloromethane in insects............... 158 

359 Table 4-32. Results from in vivo genotoxicity assays of dichloromethane in mice.................. 160 

360 Table 4-33. Results from in vivo genotoxicity assays of dichloromethane in rats and hamsters

361 ................................................................................................................................. 164 

362 Table 4-34. Comparison of in vivo dichloromethane genotoxicity assays targeted to lung or liver 

363 cells, by species ....................................................................................................... 165 

364 Table 4-35. NOAELs and LOAELs in selected animal studies involving oral exposure to 

365 dichloromethane for short-term, subchronic, or chronic durations ......................... 180 
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366 Table 4-36. NOAELs and LOAELs in animal studies involving inhalation exposure to 

367 dichloromethane for subchronic or chronic durations, hepatic, pulmonary, and 

368 neurologic effects..................................................................................................... 185 

369 Table 4-37. NOAELs and LOAELs in selected animal studies involving inhalation exposure to 

370 dichloromethane, reproductive and developmental effects ..................................... 188 

371 Table 4-38. Incidence of liver tumors in male B6C3F1 mice exposed to dichloromethane in a 2-

year oral exposure (drinking water) studya372 .............................................................. 197 

373 Table 4-39. Incidences of liver tumors in male and female F344 rats exposed to dichloromethane 

374 in drinking water for 2 years.................................................................................... 198 

375 Table 4-40. Incidences of selected neoplastic lesions in B6C3F1 mice exposed to 

376 dichloromethane by inhalation (6 hours/day, 5 days/week) for 2 years .................. 200 

377 Table 4-41. Incidences of selected neoplastic lesions in F344/N rats exposed to dichloromethane 

378 by inhalation (6 hours/day, 5 days/week) for 2 years.............................................. 201 

379 Table 4-42. Incidences of mammary gland tumors in two studies of male and female Sprague380 Dawley rats exposed to dichloromethane by inhalation (6 hours/day, 5 days/week) 

381 for 2 years ................................................................................................................ 202 

382 Table 4-43. Comparison of internal dose metrics in inhalation and oral exposure scenarios, in 

383 male mice and rats ................................................................................................... 204 

384 Table 5-1. Incidence data for nonneoplastic liver lesions and internal liver doses, based on 

385 various metrics, in male and female F344 rats exposed to dichloromethane in 

386 drinking water for 2 years (Serota et al., 1986a) ..................................................... 225 

387 Table 5-2. BMD modeling results for incidence of noncancer liver lesions in male and female 

388 F344 rats exposed to dichloromethane in drinking water for 2 years, based liver389 specific CYP metabolism dose metric (mg dichloromethane metabolism via CYP 

390 pathway per liter liver tissue per day)...................................................................... 227 

391 Table 5-3. RfD for dichloromethane based on PBTK model-derived probability distributions of 

392 human drinking water exposures extrapolated from nonneoplastic liver lesion 

393 incidence data for male rats exposed via drinking water for 2 years, based on liver394 specific CYP metabolism dose metric (mg dichloromethane metabolized via CYP 

395 pathway per liter liver tissue per day)...................................................................... 228 

396 Table 5-4. Potential points of departure with applied UFs and resulting RfDs......................... 232 

397 Table 5-5. Incidence data for nonneoplastic liver lesions (hepatic vacuolation) and internal liver 

398 doses, based on various metrics, in female Sprague-Dawley rats exposed to 

399 dichloromethane via inhalation for 2 years (Nitschke et al., 1988a) ....................... 239 

400 Table 5-6. BMD modeling results for incidence of noncancer liver lesions in female Sprague401 Dawley rats exposed to dichloromethane by inhalation for 2 years, based on liver 

402 specific CYP metabolism metric (mg dichloromethane metabolized via CYP 

403 pathway per liter liver tissue per day)...................................................................... 241 

404 Table 5-7. Inhalation RfC for dichloromethane based on PBTK model-derived probability 

405 distributions of human inhalation exposure extrapolated from nonneoplastic liver 

406 lesion data for female rats exposed via inhalation for 2 years, based on liver-specific 

407 CYP metabolism dose metric (mg dichloromethane metabolized via CYP pathway 

408 per liter liver tissue per day) .................................................................................... 242 

409 Table 5-8. Potential points of departure with applied UFs and resulting RfCs ......................... 248 

410 Table 5-9. Statistical characteristics of human equivalent applied doses in specific populations 

of the GST-T1 -/- 411 group............................................................................................ 260 

Table 5-10. Statistical characteristics of HECs in specific populations of the GST-T1 -/- 412 group

413 ................................................................................................................................. 262 
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414 Table 5-11. Incidence data for liver tumors and internal liver doses, based on GST metabolism 

415 dose metrics, in male B6C3F1 mice exposed to dichloromethane in drinking water 

416 for 2 years ................................................................................................................ 267 

417 Table 5-12. BMD modeling results and tumor risk factors for internal dose metric associated 

418 with 10% extra risk for liver tumors in male B6C3F1 mice exposed to 

419 dichloromethane in drinking water for 2 years, based on liver-specific GST 

420 metabolism and whole body GST metabolism dose metrics................................... 269 

421 Table 5-13. Cancer OSFs for dichloromethane based on PBTK model-derived internal liver 

422 doses in B2C3F1 mice exposed via drinking water for 2 years, based on liver423 specific GST meabolism and whole body metabolism dose metrics, by population 

424 genotype ................................................................................................................. 272 

425 Table 5-14. Alternative route-to-route cancer OSFs for dichloromethane extrapolated from male 

426 B6C3F1 mouse inhalation liver tumor incidence data using a tissue-specific GST 

427 metabolism dose metric, by population genotype ................................................... 274 

428 Table 5-15. Cancer OSF based on a human BMDL10 using administered dose for liver tumors in 

429 male B6C3F1 mice exposed to dichloromethane in drinking water for 2 years ...... 275 

430 Table 5-16. Comparison of OSFs derived using various assumptions and metrics, based on 

431 tumors in male mice................................................................................................. 277 

432 Table 5-17. Incidence data for liver and lung tumors and internal doses, based on GST 

433 metabolism dose metrics, in male and female B6C3F1 mice exposed to 

434 dichloromethane via inhalation for 2 years ............................................................. 280 

435 Table 5-18. BMD modeling results and tumor risk factors associated with 10% extra risk for 

436 liver and lung tumors in male and female B6C3F1 mice exposed by inhalation to 

437 dichloromethane for 2 years, based on liver-specific GST metabolism and whole 

438 body GST metabolism dose metrics........................................................................ 283 

439 Table 5-19. IURs for dichloromethane based on PBTK model-derived internal liver and lung 

440 doses in B6C3F1 male mice exposed via inhalation for 2 years, based on liver441 specific GST metabolism and whole body metabolism dose metrics, by population 

442 genotype................................................................................................................... 286 

443 Table 5-20. Upper bound estimates of combined human IURs for liver and lung tumors 

resulting from lifetime exposure to 1 μg/m3 444 dichloromethane, based on liver-specific 

445 GST metabolism and whole body metabolism dose metrics, by population genotype

446 ................................................................................................................................. 289 

447 Table 5-21. Inhalation units risks based on human BMDL10 values using administered 

448 concentration for liver and lung tumors in B6C3F1 mice exposed by inhalation to 

449 dichloromethane for 2 years .................................................................................... 291 

450 Table 5-22. Comparison of IURs derived by using various assumptions and metrics.............. 293 

451 Table 5-23. Comparison of key B6C3F1 mouse parameters differing between prior and current 

452 PBTK model application ......................................................................................... 295 

453 Table 5-24. Application of ADAFs to dichloromethane cancer risk following a lifetime (70-

454 year) oral exposure .................................................................................................. 297 

455 Table 5-25. Application of ADAFs to dichloromethane cancer risk following a lifetime (70-

456 year) inhalation exposure......................................................................................... 298 

457 Table 5-26. Summary of uncertainty in the derivation of cancer risk values for dichloromethane

458 ................................................................................................................................. 299 

459 Table 5-27. Statistical characteristics of human internal doses for 1 mg/kg-day oral exposures in 

460 specific populations ................................................................................................. 308 

Table 5-28. Statistical characteristics of human internal doses for 1 mg/m3

461 inhalation exposures 

462 in specific subpopulations........................................................................................ 309 
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463 Table 6-1. Comparison of OSFs derived by using various assumptions and metrics, based on 

464 liver tumors in male mice ........................................................................................ 319 

465 Table 6-2. Comparison of IURs derived by using various assumptions and metrics................ 322 

466 

467 Table B-1. Parameter distributions used in human Monte Carlo analysis for 

468 dichloromethane by David et al. (2006) ............................................................................ B-2 

469 

470 Table B-2. Parameters for BW distributions as functions of age and gender ......................... B-10 

471 

472 Table B-3. Parameter distributions for the human PBTK model used by the EPA................. B-18 

473 

474 Table C-1. Parameter values used in rat PBTK models............................................................. C-3 

475 

476 Table C-2. Effect of PBTK model configuration on predicted dichloromethane metabolite 

477 production in the liver of rats from inhalation of 200 or 1,000 ppm dichloromethane 

478 for 4 hours (Andersen et al., 1991) or 2,000 or 4,000 ppm dichloromethane for 

479 6 hours/day, 5 days/week, for 2 years (NTP, 1986)......................................................... C-10 

480 

481 Table C-3. Observations and predictions of total expired dichloromethane resulting from 

482 gavage doses in rats ......................................................................................................... C-15 

483 

484 Table D-1. Incidence data for nonneoplastic liver lesions and internal liver doses, based 

485 on various metrics, in male and female F344 rats exposed to dichloromethane in 

486 drinking water for 2 years (Serota et al., 1986a) ............................................................... D-1 

487 

488 Table D-2. BMD modeling results for incidence of noncancer liver lesions in male and female 

489 F344 rats exposed to dichloromethane in drinking water for 2 years, based liver-specific 

490 CYP metabolism dose metric (mg dichloromethane metabolism via CYP pathway per liter 

491 liver tissue per day)............................................................................................................ D-2 

492 

493 Table D-3. Incidence data for nonneoplastic liver lesions (hepatic vacuolation) and internal liver 

494 doses, based on various metrics, in female Sprague-Dawley rats exposed to 

495 dichloromethane via inhalation for 2 years (Nitschke et al., 1988a) ................................. D-5 

496 

497 Table D-4. BMD modeling results for incidence of noncancer liver lesions in female 

498 Sprague-Dawley rats exposed to dichloromethane by inhalation for 2 years, based 

499 on liver specific CYP metabolism metric (mg dichloromethane metabolized via 

500 CYP pathway per liter liver tissue per day)....................................................................... D-6 

501 

502 Table E-1. Incidence data for liver tumors and internal liver doses, based on GST 

503 metabolism dose metrics, in male B6C3F1 mice exposed to dichloromethane in 

504 drinking water for 2 years...................................................................................................E-1 

505 

506 Table E-2. BMD modeling results and tumor risk factors for internal dose metric 

507 associated with 10% extra risk for liver tumors in male B6C3F1 mice exposed to 

508 dichloromethane in drinking water for 2 years, based on liver-specific GST 

509 metabolism and whole body GST metabolism dose metrics..............................................E-2

510 
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511 Table E-3. Incidence data for liver and lung tumors and internal doses, based on GST 

512 metabolism dose metrics, in male B6C3F1 mice exposed to dichloromethane via 

513 inhalation for 2 years ..........................................................................................................E-7 

514 

515 Table E-4. BMD modeling results and tumor risk factors associated with 10% extra risk 

516 for liver and lung tumors in male B6C3F1 mice exposed by inhalation to 

517 dichloromethane for 2 years, based on liver-specific GST metabolism and whole 

518 body GST metabolism dose metrics ...................................................................................E-8 

519 

520 Table F-1. Incidence data for liver and lung tumors and internal doses, based on GST 

521 metabolism dose metrics, in female B6C3F1 mice exposed to dichloromethane via 

522 inhalation for 2 years ..........................................................................................................F-1

523 

524 Table F-2. BMD modeling results and tumor risk factors associated with 10% extra risk for liver 

525 and lung tumors in female B6C3F1 mice exposed by inhalation to dichloromethane for 2 

526 years, based on liver-specific GST metabolism and whole body GST metabolism dose 

527 metrics.................................................................................................................................F-3 

528 

529 Table F-3. IURs for dichloromethane based on PBTK model-derived internal liver and lung 

530 doses in B6C3F1 female mice exposed via inhalation for 2 years, based on liver-specific 

531 GST metabolism and whole body metabolism dose metrics, by population genotype .....F-4 

532 

533 Table F-4. Upper bound estimates of combined human IURs for liver and lung tumors resulting from 

lifetime exposure to 1 μg/m3 534 dichloromethane, based on liver-specific GST metabolism and whole 

535 body metabolism dose metrics, by population genotype, using female mouse data for derivation of 

536 risk factors............................................................................................................................F-5 

537 

538 Table G-1. Incidence data for mammary gland tumors and internal doses, based on different dose 

539 metrics, in male and female F344 rats exposed to dichloromethane via inhalation for 

540 2 years ................................................................................................................................ G-1 

541 

542 Table G-2. BMD modeling results associated with 10% extra risk for mammary gland tumors in 

543 F344 rats exposed by inhalation to dichloromethane for 2 years, based on AUC for 

544 dichloromethane in blood .................................................................................................. G-3 

545 

546 Table G-3. IURs for dichloromethane, based on benign mammary tumors and PBTK model547 derived internal doses in F344N rats exposed via inhalation for 2 years, based on AUC for 

548 dichloromethane in blood dose metric............................................................................... G-4 

549 

550 
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551 LIST OF FIGURES 

552 

553 Figure 3-1. Proposed pathways for dichloromethane metabolism............................................... 10 

554 Figure 3-2. Schematics of PBTK models (1986–2006) used in the development of estimates for 

555 dichloromethane internal dosimetry. ......................................................................... 22 

556 Figure 3-3. Schematic of mouse PBTK model used by Marino et al. (2006).............................. 27 

557 Figure 3-4. Schematic of human PBTK, used by David et al. (2006). ........................................ 31 

558 Figure 3-5. Schematic of rat PBTK model used in current assessment....................................... 38 

559 Figure 5-1. Exposure response array for oral exposure to dichloromethane ............................. 219 

560 Figure 5-2. Process for deriving noncancer oral RfDs and inhalation RfCs using rodent and 

561 human PBTK models............................................................................................... 221 

562 Figure 5-3. PBTK model-derived internal doses (mg dichloromethane metabolized via the CYP 

563 pathway per liter liver per day) in rats and humans and their associated external 

564 exposures (mg/kg-day), used for the derivation of RfDs. ....................................... 226 

565 Figure 5-4. Comparison of RfDs derived from selected point of departures for endpoints 

566 presented in Table 5-4. ............................................................................................ 233 

567 Figure 5-5. Exposure response array for chronic (animal) or occupational (human) inhalation 

568 exposure to dichloromethane (log Y axis)............................................................... 235 

569 Figure 5-6. Exposure response array for subacute to subchronic inhalation exposure to 

570 dichloromethane (log Y axis) .................................................................................. 236 

571 Figure 5-7. PBTK model-derived internal doses (mg dichloromethane metabolized via the CYP 

572 pathway/L liver/day) in rats and humans versus external exposures (ppm)............ 240 

573 Figure 5-8. Comparison of RfCs derived from selected point of departures for endpoints 

574 presented in Table 5-8. ............................................................................................ 249 

575 Figure 5-9. Comparison dichloromethane oxidation rate data with alternate kinetic models. .. 253 

576 Figure 5-10. Sensitivity coefficients for long-term mass CYP- and GST-mediated metabolites 

577 per liver volume from a daily drinking water concentration of 10 mg/L in rats. .... 257 

578 Figure 5-11. Sensitivity coefficients for long-term mass CYP- and GST-mediated metabolites 

579 per liver volume from a long-term average daily inhalation concentration of 500 ppm 

580 in rats. (KA is not included since it has no impact on inhalation dosimetry.)........ 257 

581 Figure 5-12. Frequency density of human equivalent applied doses in specific populations in 

582 comparison to a general population (0.5- to 80-year-old males and females) estimate 

583 for an internal dose of 15.1 mg dichloromethane metabolized by CYP per liter liver 

per day; all groups were restricted to the GST-T1 -/- 584 population)............................ 260 

585 Figure 5-13. Frequency density of HECs in specific populations in comparison to a general 

586 population (0.5- to 80-year-old males and females) estimate for an internal dose of 

587 128.1 mg dichloromethane metabolized by CYP per liter liver per day ; all groups 

restricted to the GST-T1 -/- 588 population). .................................................................. 262 

589 Figure 5-14. Process for deriving cancer OSFs and IURs by using rodent and human PBTK 

590 models...................................................................................................................... 265 

591 Figure 5-15. PBTK model-derived internal doses (mg dichloromethane metabolized via the GST 

592 pathway per liter liver per day) in mice and humans and their associated external 

593 exposures (mg/kg-day) used for the derivation of cancer OSFs based on liver tumors 

594 in mice. .................................................................................................................... 268 

595 Figure 5-16. PBTK model-derived internal doses (mg dichloromethane metabolized via the GST 

596 pathways per liter tissue per day) for liver (A) and lung (B) in mice and humans, and 

597 their associated external exposures (ppm), used for the derivation of cancer 

598 inhalation unit risks.................................................................................................. 282 
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599 Figure 5-20. Histograms for a liver-specific dose of GST metabolism (mg GST metabolites per 

600 liter liver per day) for the general population (0.5- to 80-year-old males and females) 

and specific age/gender groups, within the population of GST-T1+/+ 601 genotypes, given 

602 a daily oral dose-rate of 1 mg/kg-day dichloromethane. ......................................... 308 

603 Figure 5-21. Histograms for liver-specific dose of GST metabolism (mg GST metabolites per 

604 liter liver per day) for the general population (0.5- to 80-year-old males and females) 

and specific age/gender groups, within the population of GST-T1+/+ 605 genotypes, given 

a continuous inhalation exposure to 1 mg/m3

606 dichloromethane.............................. 309 

607 

608 Figure B-1. Schematic of the David et al. (2006) PBTK model for dichloromethane in the 

609 human .................................................................................................................................B-1 

610 

611 Figure B-2. Total CYP2E1 activity (Vmax) normalized to the average total activity in 14-18 

612 year-old individuals (Vmax(14-18)) plotted against normalized body weight (BW) for 

613 individuals ranging from six months to 18 years of age. Data source: Johnsrud et al. 

614 (2003) ................................................................................................................................B-6 

615 

616 Figure B-3. Body-weight scaled CYP2E1 activity (Vmaxc) normalized to the average scaled 

617 activity in 14-18 year-old individuals (Vmaxc (14-18)) plotted against age individuals 

618 ranging from six months to 18 years of age. Data source: Johnsrud et al. (2003)..........B-7 

619 

620 Figure B-4. U.S. age distribution, 6 months-80 years (values from U.S. Census Bureau).........B-9 

621 

622 Figure B-5. U.S. age-specific gender distribution (values from U.S. Census Bureau) ..............B-9 

623 

624 Figure B-6. Function fits to age-dependent data for BW mean and standard deviations 

625 for males and females in the United States (values from Portier et al., 2007).................B-11 

626 

627 Figure B-7. Example body weight (BW) histogram from Monte Carlo simulation for 0.5-

628 to 80-year-old males and females in the United States (simulated n = 10,000)...............B-12 

629 

630 Figure B-8. Mean value respiration rates for males and females as a function of age (values from 

631 Clewell et al., 2004) .........................................................................................................B-13 

632 

633 Figure B-9. Geometric standard deviations (GSDs) for respiration rates for males and 

634 females as a function of age (values from Arcus-Arth and Blaisdell, 2007) ...................B-13 

635 

636 Figure B-10. Fraction body fat (VFC) over various age ranges in males and females (data 

637 from Clewell et al., 2004).................................................................................................B-15 

638 

639 Figure B-11. Fraction liver (VLC) as a function of age (data from Clewell et al., 2004) ........B-16 

640 

641 Figure C-1. Schematic of the Andersen et al. (1991) PBTK model for dichloromethane 

642 in the rat ..............................................................................................................................C-2 

643 

644 Figure C-2. Observations of exhaled dichloromethane (DCM) and carbon monoxide (CO) after a 

645 bolus oral dose of 200 mg/kg in rats (data of Angelo et al., 1986b). ................................C-5 

646 
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647 Figure C-3. Observations and predictions (models A-D) of Gargas et al. (1986) data for 

648 respiratory uptake by 3 rats of 100, 500, 1000, or 3000 ppm dichloromethane in a 9-

649 L closed chamber. ............................................................................................................C-6 

650 

651 Figure C-4. Observations (points) and predictions (curves, Models A-D) of Angelo et al. 

652 (1986b) data for dichloromethane in blood following 10 and 50 mg/kg iv injection 

653 in rats. Model predictions with doses at 56% of the nominal doses, i.e., 5.6 and 28 

654 mg/kg, are shown for comparison as dashed lines for model D.........................................C-7 

655 

656 Figure C-5. Observations and predictions (Models A-D) of Andersen et al. (1987) data for 

657 dichloromethane in rat blood from inhalation of 200 and 1000 ppm dichloromethane for 4 

658 hours....................................................................................................................................C-8 

659 

660 Figure C-6. Observations and predictions (Models A-D) of Angelo et al. (1986b) data for 

661 percent of dichloromethane dose expired as dichloromethane following 10 and 

662 50 mg/kg iv injection in rats. Model predictions with doses at 56% of the nominal 

663 doses, i.e., 5.6 and 28 mg/kg, are shown for comparison as dashed lines for model 

664 D..........................................................................................................................................C-9 

665 

666 Figure C-7. Observations and Model D predictions of Angelo et al. (1986b) data for (A) percent 

667 dose expired as dichloromethane, (B) blood dichloromethane, (C) percent expired as 

668 carbon monoxide, and (D) liver dichloromethane in rats given a single dichloromethane 

669 gavage doses of 50 and 200 mg/kg, using a numerically fitted GI absorption rate (KA) of 

670 1.8. ..................................................................................................................................C-12 

671 

672 Figure C-8. Model predictions with of blood carboxy-hemoglobin (COHb, percent of 

673 total Hb) from a single gavage dose of 526 mg/kg DCM in rats, compared to the 

674 data of Pankow et al. (1991). Model simulations performed with Model D (heavy 

675 read line, KA = 1.8/h) or with an alternate value of KA = 0.47/h, fit to these data. ........C-13 

676 

677 Figure D-1. Predicted (logistic model) and observed incidence of noncancer liver lesions in male 

678 F344 rats exposed to dichloromethane in drinking water for 2 years 

679 (Serota et al., 1986a) ......................................................................................................... D-3 

680 

681 Figure D-2. Predicted (log-probit model) and observed incidence of noncancer liver lesions in 

682 female Sprague-Dawley rats inhaling dichloromethane for 2 years 

683 (Nitschke 1988a) ............................................................................................................. D-7 

684 

685 Figure E-1. Predicted and observed incidence of animals with hepatocellular carcinoma or 

686 adenoma in male B6C3F1 mice exposed to dichloromethane in drinking water 

687 for 2 years, using liver-specific metabolism dose metric (Serota et al., 1986b; Hazelton 

688 Laboratories, 1983).............................................................................................................E-3 

689 

690 Figure E-2. Predicted and observed incidence of animals with hepatocellular carcinoma or 

691 adenoma in male B6C3F1 mice exposed to dichloromethane in drinking water 

692 for 2 years, using whole-body metabolism dose metric (Serota et al., 1986b; Hazelton 

693 Laboratories, 1983).............................................................................................................E-5 

694 
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695 Figure E-3. Predicted and observed incidence of animals with hepatocellular carcinoma or 

696 adenoma in male B6C3F1 mice exposed by inhalation to dichloromethane for 2 years, 

697 using liver-specific metabolism dose metric (Mennear et al., 1988; NTP, 1986)..............E-9 

698 

699 Figure E-4. Predicted and observed incidence of animals with carcinoma or adenoma in 

700 the lung of male B6C3F1 mice exposed by inhalation to dichloromethane for 2 

701 years, using liver-specific metabolism dose metric (Mennear et al., 1988; NTP, 

702 1986).................................................................................................................................E-11 

703 

704 Figure E-5. Predicted and observed incidence of animals with hepatocellular carcinoma or 

705 adenoma in male B6C3F1 mice exposed by inhalation to dichloromethane for 2 years, 

706 using whole-body metabolism dose metric (Mennear et al., 1988; NTP, 1986)..............E-13 

707 

708 Figure E-6. Predicted and observed incidence of animals with carcinoma or adenoma in 

709 the lung of male B6C3F1 mice exposed by inhalation to dichloromethane for 2 

710 years, using whole-body metabolism dose metric (Mennear et al., 1988; NTP, 

711 1986).................................................................................................................................E-15 

712 

713 Figure G-1. PBTK model-derived internal doses (daily average AUC for dichloromethane in 

714 blood) in rats and humans, and their associated external exposures (ppm) used for the 

715 derivation of cancer IURs, based on mammary tumors in rats. ........................................ G-2

716 
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717 LIST OF ACRONYMS 

718 

719 

720 ACGIH American Conference of Governmental Industrial Hygienists 

721 ADAF age-dependent adjustment factor 

722 AEGL acute exposure guideline level 

723 AIC Akaike’s Information Criterion 

724 ALT alanine aminotransferase 

725 AP alkaline phosphatase 

726 AST aspartate aminotransferase 

727 ATSDR Agency for Toxic Substances and Disease Registry 

728 AUC area under the curve of a concentration versus time plot 

729 BAER brainstem-auditory evoked response 

730 BMD benchmark dose 

731 BMDL10 95% lower bound on the BMD 

732 BMR benchmark response level 

733 BW body weight

734 CAEP cortical-auditory-evoked potential 

735 CASRN Chemical Abstracts Service Registry Number 

736 CHO Chinese hamster ovary 

737 CI confidence interval 

738 CMR Chemical Marketing Reporter 

739 CNS central nervous system 

740 COHb carboxyhemoglobin 

741 CV coefficient of variation 

742 CYP cytochrome P450 

743 DNA deoxyribonucleic acid 

744 EPA U.S. Environmental Protection Agency

745 FEP flash-evoked potential

746 FOB functional observational battery 

747 GD gestation day 

748 GSH reduced glutathione 

749 GST glutathione S-transferase 

750 HEC human equivalent concentration 

751 HPRT hypoxanthine-guanine phosphoribosyl transferase 

752 IARC International Agency for Research on Cancer 

ICD-9 International Classification of Diseases 9th 753 ed. 

754 IgM immunoglobulin M 

755 IRIS Integrated Risk Information System 

756 IUR inhalation unit risk 

757 LOAEL lowest-observed-adverse-effect level 

758 LOH loss of heterozygosity 

759 MCHC mean corpuscular hemoglobin concentration 

760 MCMC Markov Chain Monte Carlo 

761 Mg milligrams

762 mRNA messenger ribonucleic acid 

763 NADPH nicotinamide adenine dinucleotide phosphate 

764 NIOSH National Institute of Occupational Safety and Health 

765 NLM National Library of Medicine 
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766 NOAEL no-observed-adverse-effect level 

767 NRC National Research Council 

768 NTP National Toxicology Program 

769 OR odds ratio 

770 OSF oral slope factor 

771 OSHA Occupational Safety and Health Administration 

772 PBTK physiologically based toxicokinetic 

773 PND postnatal day 

774 QCC cardiac output 

775 RfC reference concentration 

776 RfD reference dose 

777 SD standard deviation 

778 SEM standard error of the mean

779 SEP somatosensory-evoked potential 

780 SMR standardized mortality ratio 

781 SRC Syracuse Research Corporation 

782 SSB single-strand break 

783 TWA time-weighted average 

784 UF uncertainty factor 

785 VPR ventilation:perfusion ratio 
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786 FOREWORD 

787 

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

789 for the hazard and dose-response assessment in IRIS pertaining to exposure to dichloromethane. 

790 It is not intended to be a comprehensive treatise on the chemical or toxicological nature of 

791 dichloromethane. 

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

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

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

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

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

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

798 assessment process. 

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

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

801 hotline.iris@epa.gov (email address). 

802 

803 
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804 AUTHORS, CONTRIBUTORS, AND REVIEWERS 

805 

806 

807 CHEMICAL MANAGERS

808 

809 Glinda S. Cooper, Ph.D. 

810 Ambuja S. Bale, Ph.D., DABT 

811 Office of Research and Development, IRIS Program 

812 U.S. Environmental Protection Agency 

813 Washington, DC 

814 

815 AUTHORS

816 

817 Glinda S. Cooper, Ph.D. 

818 Ambuja S. Bale, Ph.D., DABT 

819 Andrew Rooney, Ph.D. 

820 Paul Schlosser, Ph.D. 

821 Allan Marcus, Ph.D. 

822 Gene (Ching-Hung) Hsu, Ph.D., DABT 

823 National Center for Environmental Assessment 

824 Office of Research and Development 

825 U.S. Environmental Protection Agency 

826 Washington, DC 

827 

828 John C. Lipscomb, Ph.D., DABT 

829 National Center for Environmental Assessment 

830 U.S. Environmental Protection Agency 

831 Cincinnati, OH 

832 

833 Peter McClure, Ph.D., DABT 

834 Michael Lumpkin, Ph.D. 

835 Fernando Llados, Ph.D. 

836 Mark Osier, Ph.D., DABT 

837 Daniel Plewak, B.S. 

838 Syracuse Research Corporation 

839 Syracuse, NY 

840 

841 Elizabeth Dupree Ellis, Ph.D. 

842 Oak Ridge Institute for Science and Education 

843 Center for Epidemiologic Research 

844 Oak Ridge, TN 

845 

846 REVIEWERS

847 This document has been peer reviewed by EPA scientists. 
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848 INTERNAL EPA REVIEWERS

849 

850 Ghazi Dannan, Ph.D 

851 Karen Hogan, M.S. 

852 Jennifer Jinot, Ph.D 

853 Paul White, Ph.D 

854 Samantha Jones, Ph.D. 

855 Jamie Strong, Ph.D. 

856 National Center for Environmental Assessment 

857 Office of Research and Development 

858 U.S. Environmental Protection Agency 

859 

860 David Herr, Ph.D 

861 National Health and Environmental Effect Research Laboratory 

862 Office of Research and Development 

863 U.S. Environmental Protection Agency 

864 

865 

866 
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867 1. INTRODUCTION 

868 

869 

870 This document presents background information and justification for the Integrated Risk 

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

872 dichloromethane. IRIS Summaries may include oral reference dose (RfD) and inhalation 

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

874 carcinogenicity assessment. 

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

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

877 mode of action. The RfD (expressed in units of mg/kg-day) is defined as an estimate (with 

878 uncertainty spanning perhaps an order of magnitude) of a daily exposure to the human 

879 population (including sensitive subgroups) that is likely to be without an appreciable risk of 

deleterious effects during a lifetime. The inhalation RfC (expressed in units of mg/m3

880 ) is 

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

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

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

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

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

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

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

888 derived for chronic exposure duration. 

889 The carcinogenicity assessment provides information on the carcinogenic hazard 

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

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

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

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

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

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

plausible upper bound on the estimate of risk per μg/m3896 air breathed. 

897 Development of these hazard identification and dose-response assessments for 

898 dichloromethane has followed the general guidelines for risk assessment as set forth by the 

899 National Research Council (1983). EPA Guidelines and Risk Assessment Forum Technical 

900 Panel Reports that may have been used in the development of this assessment include the 

901 following: Guidelines for the Health Risk Assessment of Chemical Mixtures (U.S. EPA, 1986a), 

902 Guidelines for Mutagenicity Risk Assessment (U.S. EPA, 1986b), Recommendations for and 

903 Documentation of Biological Values for Use in Risk Assessment (U.S. EPA, 1988a), Guidelines 

904 for Developmental Toxicity Risk Assessment (U.S. EPA, 1991), Interim Policy for Particle Size 
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905 and Limit Concentration Issues in Inhalation Toxicity Studies (U.S. EPA, 1994a), Methods for 

906 Derivation of Inhalation Reference Concentrations and Application of Inhalation Dosimetry

907 (U.S. EPA, 1994b), Use of the Benchmark Dose Approach in Health Risk Assessment (U.S. EPA, 

908 1995), Guidelines for Reproductive Toxicity Risk Assessment (U.S. EPA, 1996), Guidelines for 

909 Neurotoxicity Risk Assessment (U.S. EPA, 1998a), Science Policy Council Handbook: Risk 

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

911 EPA, 2000b), Supplementary Guidance for Conducting Health Risk Assessment of Chemical 

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

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

914 Supplemental Guidance for Assessing Susceptibility from Early-Life Exposure to Carcinogens 

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

916 Framework for Assessing Health Risk of Environmental Exposures to Children (U.S. EPA, 

917 2006b). 

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

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

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

921 in the development of this document. The relevant literature was reviewed through April 2009. 
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922 2. CHEMICAL AND PHYSICAL INFORMATION 

923 

924 

925 Dichloromethane is a colorless liquid with a penetrating ether-like odor (Lewis, 

1997).1926 Selected chemical and physical properties of dichloromethane are listed in Table 2-1. 

927 

Table 2-1. Physical properties and chemical identity 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 CH2Cl2 O’Neil et al. (2001) 

Boiling point 40°C Lide (2000) 

Melting point −95.1°C Lide (2000) 

Vapor pressure 1.15 × 102

 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 × 104 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 Kow = 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 × 105 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 × 106 Pa Holbrook (2003) 

Viscosity 0.430 cP at 20°C Lewis (1997) 

Henry’s constant 3.25 × 10–3 atm m3

/mol at 25°C Leighton and Calo (1981) 

OH reaction rate constant 1.42 × 10–13 cm3

/molecule sec at 25°C Atkinson (1989) 

Chemical structure 

Cl C Cl

H

H

928 

929 

930 Dichloromethane is produced by two methods of manufacturing (International Agency 

931 for Research on Cancer [IARC], 1999). The older method involves the direct reaction of 

932 methane with chlorine either at high temperatures or at lower temperatures under catalytic or 

933 photolytic conditions (Holbrook, 2003). The more common method used today involves an 

934 initial reaction of hydrochloric acid with methanol to yield methyl chloride. Excess methyl 

935 chloride is then reacted in the gas phase thermally with chlorine to produce dichloromethane 

936 (Holbrook, 2003). This process can also be carried out catalytically or photolytically. 
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937 Dichloromethane became an important industrial chemical in the U.S. during 

938 World War II (Hardie, 1964). Dichloromethane has been used in paint strippers and removers, 

939 as a propellant in aerosols, in the manufacture of drugs, pharmaceuticals, film coatings, 

940 electronics, and polyurethane foam, and as a metal-cleaning solvent. Dichloromethane can also 

941 be used in the decaffeination process of coffee and tea (ATSDR, 2000). The U.S. production 

942 was 3.8 million pounds in 1941 and 8.3 million pounds in 1944 (Searles and McPhail, 1949). 

943 Dichloromethane production rose sharply in the decades following the war due to the increased 

944 demand for this substance for use mainly in paint strippers (Hardie, 1964; Searles and McPhail, 

945 1949). U.S. production in 1947, 1955, 1960, and 1962 was approximately 19, 74, 113, and 

946 144 million pounds, respectively (Hardie, 1964; Searles and McPhail, 1949). As other solvent 

947 uses and its use in aerosol propellants became important, demand for this substance increased 

948 further (Anthony, 1979). Dichloromethane production continued to rise dramatically through the 

949 1970s; production capacities were 520 million pounds in 1973 and 830 million pounds in 1979 

950 (Chemical Marketing Reporter [CMR], 1979, 1973). 

951 After 1980, production of dichloromethane began to decline. Production capacities fell 

952 from 722 million pounds in 1982 to 465 million pounds in 1997 (CMR, 1997, 1982). The total 

953 U.S. production capacity for dichloromethane in 2000 was 535 million pounds (CMR, 2000). 

954 The demand for dichloromethane decreased from 600 million pounds in 1979 to 200 million 

955 pounds in 1999 (CMR, 2000, 1979). The decline in production of and demand for 

956 dichloromethane over the past 2 decades has been attributed to increased regulation, the use of 

957 alternative chemicals in aerosol spray cans, and concern over dichloromethane carcinogenicity 

958 (Holbrook, 2003; ATSDR, 2000). 

959 Dichloromethane in the environment will partition mainly to air (National Library of 

960 Medicine [NLM], 2003). In air, dichloromethane exists as a vapor. Some of the 

961 dichloromethane released to soil or water is expected to volatilize to air. In soil, 

962 dichloromethane is expected to be highly mobile and may migrate to groundwater. The potential 

963 for dichloromethane to bioconcentrate in aquatic or marine organisms is low. Dichloromethane 

964 may biodegrade in soil or water under both aerobic and anaerobic conditions. 

 

1

 To avoid confusion, “dichloromethane” is used throughout this summary even if a specific paper used the term 

“methylene chloride.” 
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965 3. TOXICOKINETICS 

966 

967 

968 3.1. ABSORPTION 

969 3.1.1. Oral — Gastrointestinal Tract Absorption 

970 There are currently no data available on absorption of dichloromethane following oral 

971 intake in humans. However, after oral administration in animals, dichloromethane is rapidly and 

972 nearly completely absorbed in the gastrointestinal tract (Angelo et al., 1986a, b; McKenna and 

973 Zempel, 1981). Angelo et al. (1986b) reported that, following administration of single 

974 radiolabeled oral doses (10, 50, or 200 mg/kg) to mature male F344 rats, 97% of the label was 

975 detected in the exhaled air within 24 hours, indicating nearly complete absorption. At several 

976 time points within 40 minutes of dose administration, less than 2% of the dose was found in the 

977 lower part of the gastrointestinal tract, indicating that the majority of dichloromethane absorption 

978 occurs in the upper gastrointestinal tract (Angelo et al., 1986b). Similar results were reported in 

979 mature male B6C3F1 mice exposed to up to 50 mg/kg (Angelo et al., 1986a). In mature male 

980 Sprague-Dawley rats administered a single dose (1 or 50 mg/kg) of radiolabeled 

981 dichloromethane, less than 1% of the label was found in feces collected for 48 hours after dose 

982 administration (McKenna and Zempel, 1981). Absorption of dichloromethane generally follows 

983 first-order kinetics (Angelo et al., 1986a), and no evidence for a dichloromethane-specific carrier 

984 has been presented. The vehicle appears to affect the rate, but not the extent, of gastrointestinal 

985 absorption, with an aqueous vehicle resulting in a more rapid absorption of dichloromethane than 

986 an oil-based vehicle (Angelo et al., 1986a). 

987 

988 3.1.2. Inhalation—Respiratory Tract Absorption 

989 Several studies in humans have demonstrated the absorption of dichloromethane

990 following inhalation exposure. In a study by Astrand et al. (1975), 14 male volunteers (ages 19–

29) were exposed to about 870 mg/m3

 (250 ppm) or 1,740 mg/m3991 (500 ppm) for 30 minutes 

992 while resting or exercising on a bicycle ergometer. There was a pause of about 20 minutes 

993 without exposure between rest and exercise periods. Uptake of dichloromethane was estimated 

994 at about 55% while resting and about 40, 30, and 35% at respective workloads of 50, 100, and 

995 150 watts. Blood levels of dichloromethane correlated directly with exposure concentrations, 

996 and did not appear to increase when a workload was applied (Astrand et al., 1975). Similar 

997 reports of rapid uptake and a direct correlation between dichloromethane exposure level and 

998 blood levels in humans have been presented by other groups (DiVincenzo and Kaplan, 1981; 

999 DiVincenzo et al., 1971). 

1000 With extended (1–2 hours or greater) exposure, uptake tends to reach a steady-state level, 

1001 at which point blood dichloromethane levels remain more or less constant (DiVincenzo and 

1002 Kaplan, 1981; DiVincenzo et al., 1972; Riley et al., 1966). DiVincenzo et al. (1972) reported 
6 DRAFT - DO NOT CITE OR QUOTE 

1003 that in humans exposed to 100 or 200 ppm of dichloromethane for 2 hours (without physical 

1004 exercise), dichloromethane was rapidly absorbed, reaching an approximate steady state, as 

1005 assessed by levels of unchanged dichloromethane in the expired air, within the first 15–

1006 30 minutes of exposure. A later study by the same group (DiVincenzo and Kaplan, 1981) 

1007 similarly reported a rapid absorption of dichloromethane in volunteers exposed to 50–200 ppm 

1008 for 7.5 hours on each of 5 consecutive days. A steady-state level, as assessed by levels of 

1009 unchanged dichloromethane in the expired air, was reached quickly (1–2 hours), with exhaled 

1010 dichloromethane levels increasing with increasing exposure level. A similar pattern was seen 

1011 with blood dichloromethane levels. Estimated pulmonary uptake was 69–75% and did not vary 

1012 appreciably with exposure concentration. In another experiment in which one of the 

1013 investigators was seated during exposure to 100 ppm dichloromethane for 2 hours, 

1014 concentrations of dichloromethane in expired air reached an apparent plateau of about 70 ppm 

1015 within the first hour of exposure (Riley et al., 1966). 

1016 Body fat may influence absorption of dichloromethane, as evidenced by data from an 

1017 experiment involving 12 men ages 21–35, divided into two groups (n = 6 per group) based on 

1018 percent body fat (Engström and Bjurström, 1977). The mean percent body fat in the leaner 

1019 group was 7.8% (standard error of the mean [SEM] 1.9), range 2.3–13.6%, compared with 

1020 25.1% (SEM 2.8), range 18.3–36.2%, in the more overweight group. Total uptake of 

dichloromethane during a light exercise period (50 watts21021 ) for 1 hour with an exposure level of 

1022 750 ppm was positively correlated with percent body fat (r = 0.81), and the estimated amount of 

1023 dichloromethane in fat storage was also correlated with percent body fat (r = 0.84). 

1024 A pattern of absorption similar to that seen in humans has been seen in animals. Initially, 

1025 dichloromethane is readily absorbed following inhalation exposure, as evidenced by rapid 

1026 appearance of dichloromethane in blood, tissues, and expired air (Withey and Karpinski, 1985; 

1027 Stott and McKenna, 1984; Anders and Sunram, 1982; Carlsson and Hultengren, 1975; Roth et 

1028 al., 1975). For example, absorption of inhaled 500 ppm dichloromethane in anesthetized, mature 

1029 male F344 rats reached an apparent plateau within 10–20 minutes and was relatively constant for 

1030 up to 2 hours (Stott and McKenna, 1984). In these experiments, absorption was calculated from 

1031 measurements of exposure (nose only) and effluent concentrations and ventilation flow rate in 

1032 intact animals; double tracheostomized rats were used to measure absorption in the isolated 

1033 upper respiratory tract and the lower respiratory tract. At a ventilation rate of 53 mL/minute, 

1034 absorption expressed as mean percentage of dichloromethane available for absorption was 44% 

1035 (standard deviation [SD] 10) in intact rats, 13.2% (SD 3.6) in the upper respiratory tract, and 

1036 37% (SD 4.1) in the lower respiratory tract. 

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 exercise effort that was exerted by the individuals in the study). 
7 DRAFT - DO NOT CITE OR QUOTE 

1039 3.2. DISTRIBUTION 

1040 Results from studies of animals show that, following absorption, dichloromethane is 

1041 rapidly distributed throughout the body and has been detected in all tissues that have been 

evaluated. Twenty minutes after a single intravenous dose of 10 mg [14 1042 C]-dichloromethane/kg to 

1043 mature male B6C3F1 mice (Angelo et al., 1986a), total label was greatest in the liver 

1044 (6.72 μg-equivalents/g tissue), with lower levels reported in the lung (1.82 μg-equivalents/g 

1045 tissue), kidney (1.84 μg-equivalents/g tissue), and the remainder of the carcass 

1046 (1.90 μg-equivalents/g tissue). By 4 hours post administration, levels in the liver had fallen to 

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

1048 were 0.23 μg-equivalents/g tissue. The levels in the kidney rose sharply in the first hour 

1049 postexposure but then fell and remained steady at ~1.60 μg-equivalents/g tissue for the 

1050 remaining 3 hours of the study (Angelo et al., 1986a). McKenna et al. (1982) exposed groups of 

mature male Sprague-Dawley rats to 50, 500, or 1,500 ppm [14 1051 C]-labeled dichloromethane for 

1052 6 hours and examined tissues at 48 hours for presence of radiolabel; results are shown in 

1053 Table 3-1. The greatest concentration of label was found in the liver, followed by the kidney and 

1054 lung. 

1055 

Table 3-1. Distribution of radioactivity in tissues 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 

1057 Source: McKenna et al. (1982). 

1058 

1059 

1060 As noted in the preceding section, adipose tissue may affect the uptake of 

1061 dichloromethane, and there is also evidence of a relation between adiposity and dichloromethane 

1062 storage. In the study by Engström and Bjurström (1977) involving 12 men ages 21–35 exposed 

1063 to 750 ppm dichloromethane during a 1 hour light exercise (50 watts) period, dichloromethane 

1064 was measured in body fat biopsy specimens at 1, 2, 3, and 4 hours postexposure. All specimens 

1065 were taken from the buttocks. The concentration of dichloromethane (per gram tissue) was 
8 DRAFT - DO NOT CITE OR QUOTE 

1066 negatively correlated with percent body fat, but the total estimated amount of dichloromethane in 

1067 fat tissue 4 hours postexposure was higher in subjects with a higher amount of fat (r = 0.84). 

1068 Carlsson and Hultengren (1975) exposed groups of 10 mature male Sprague-Dawley rats 

to [14C]-dichloromethane for 1 hour at a mean concentration of 1,935 mg/m31069 (557 ppm) and SD 

of 90 mg/m31070 (26 ppm). The initial levels were highest in the white adipose tissue (approximately 

1071 80 μg dichloromethane per gram tissue) compared with approximately 35, 20, and 5 μg1072 equivalent dichloromethane/g tissue in the liver, kidney and adrenal glands, and brain, 

1073 respectively. These initial levels in the adipose quickly fell to less than 10 μg-equivalent 

1074 dichloromethane/g tissue; more moderate declines were seen in the other tissues. 

1075 With acute 6-hour exposure scenarios, peak exposure concentrations may have a greater 

1076 influence on dichloromethane levels in the brain and perirenal fat than time-weighted average 

1077 (TWA) concentrations during the exposure period (Savolainen et al., 1981). In rats exposed over 

1078 a 6-hour period for 5 days/week to a TWA of 1,000 ppm dichloromethane consisting of two 

1079 1-hour peak concentrations (2,800 ppm) interspersed with exposure to 100 ppm, levels of 

1080 dichloromethane in the brain and perirenal fat were significantly higher than corresponding 

1081 levels in rats exposed to constant levels of 1,000 ppm. This difference was not seen with blood 

1082 carbon monoxide (CO) levels (Table 3-2). With constant exposure concentrations of 500 or 

1083 1,000 ppm, perirenal fat levels of dichloromethane approximately doubled following 2 weeks of 

1084 exposure compared with 1 week of exposure, indicating that some storage of dichloromethane in 

1085 fat tissue can occur with repeated exposure scenarios (Table 3-2). In contrast, brain levels of 

1086 dichloromethane in rats exposed for 1 week were higher than brain levels in rats exposed for 2 

1087 weeks. One possible explanation of these observations is that there is an induction of enzymes 

1088 involved in dichloromethane metabolism in liver and other tissues with repeated exposure and 

1089 dichloromethane in fat is poorly metabolized. 

1090 

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

concentrations in male Wistar rats exposed by inhalation to 

dichloromethane at constant exposure concentrations compared with 

intermittently high exposure concentrations

Exposure weeks 

Exposure level 1 2 1 2 1 2 a

(TWA, ppm) Brain (nmol/g) Perirenal fat (nmol/g) Blood CO (nmol/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 

aGroups of 5 rats were exposed to 0, 50, or 1,000 ppm 6 hours/day or 100 ppm interspersed with two 1-hour 

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). 

1091 
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1092 Placental transfer 

1093 Dichloromethane is capable of crossing the placental barrier and entering the fetal 

1094 circulation. Anders and Sunram (1982) reported that when pregnant Sprague-Dawley rats (n = 

1095 3) were exposed to 500 ppm dichloromethane for 1 hour on gestational day (GD) 21, mean 

1096 maternal blood levels were 176 nmol/mL (SEM 50), while fetal levels were 115 nmol/mL (SEM 

1097 40); interestingly, the levels of CO, a metabolite of dichloromethane, were similar in both the 

1098 maternal blood (167 nmol/mL, SEM 12) and fetal blood (160 nmol/mL, SEM 31). Withey and 

1099 Karpinski (1985) also reported higher maternal compared with fetal dichloromethane levels 

1100 based on a study of five pregnant Sprague-Dawley rats exposed to 107–2,961 ppm of 

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

1102 found in the fetal circulation. 

1103 

1104 Blood-brain barrier transfer 

1105 Dichloromethane is thought to readily transfer across the blood-brain barrier, as 

1106 evidenced by the detection of radioactivity in brain tissue 48 hours after exposures of rats to 

1107 radiolabeled dichloromethane at concentrations of 50, 500, or 1,500 ppm for 6 hours (McKenna 

1108 et al., 1982) (see Table 3-1), and the historical demonstrations that dichloromethane has transient 

1109 sedative and anesthetic properties in humans (for review of these reports, see Mattsson et al. 

1110 [1990] and Winneke [1974]). Dichloromethane is no longer used as an anesthetic gas because 

1111 the margin between anesthetic and lethal doses is narrow (Winneke, 1974). 

1112 

1113 3.3. METABOLISM 

1114 Metabolism of dichloromethane involves two primary pathways, outlined in Figure 3-1 

1115 (Agency for Toxic Substances and Disease Registry [ATSDR], 2000; Guengerich, 1997; Hashmi 

1116 et al., 1994; Gargas et al., 1986). Dichloromethane is metabolized to CO in a cytochrome 

1117 P450 (CYP)-dependent oxidative pathway that is predominant at low exposure levels. The CYP1118 related pathway results in the addition of oxygen, followed by spontaneous rearrangement to 

formyl chloride, and then to CO; each spontaneous rearrangement releases H+

 and Cl–1119 ions. At 

1120 higher exposure levels, the CYP pathway becomes saturated and a second pathway begins to 

1121 predominate. Glutathione S-transferase (GST)-catalyzed addition of glutathione (GSH) is the 

1122 initial step in this pathway. The replacement of one of the chlorine atoms with the S-glutathione 

group results in formation of S-(chloromethyl)glutathione and the release of H+ and Cl–1123 ions. 

1124 Hydration of S-(chloromethyl)glutathione results in an S-glutathionyl methanol molecule, which 

1125 can spontaneously form formaldehyde or rearrange to form an S-glutathione formaldehyde 

1126 molecule, and then further rearrange to formate. Both formaldehyde and formate can then be 

1127 further metabolized to CO2. 

1128 
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GSTT1

C H

Cl

Cl

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

CO2

CO2 C

H

O

G - S

C

H

O

OH

Formic acid CO2

(minor pathway)

1129 

1130 

1131 Figure 3-1. Proposed pathways for dichloromethane metabolism. 

1132 

1133 Adapted from: ATSDR (2000); Guengerich (1997); Hashmi et al. (1994); Gargas et al. (1986). 

1134 

1135 

1136 As described in the following discussion of the two pathways, a metabolic balance 

1137 appears to exist between them, with the CYP pathway tending to be relatively more active at 

1138 lower doses and the GST pathway metabolizing the majority of a dichloromethane dose at higher 

1139 exposure levels once the CYP pathway has become saturated. Exposure to other agents may 

1140 shift this balance. For example, pretreatment with compounds that deplete GSH (e.g., buthionine 

1141 sulfoximine, diethylmaleate, phorone) resulted in an increase in blood carboxyhemoglobin 

1142 (COHb) levels, following a single injection of dichloromethane, relative to animals that did not 

1143 receive GSH depletion, indicating a shift to the CYP pathway (Oh et al., 2002). Similarly, co1144 exposure to agents that compete for CYP2E1 results in a shift toward the GST pathway and away 

1145 from CO production (Lehnebach et al., 1995; Pankow and Jagielki, 1993; Pankow et al., 

1146 1991a, b; Glatzel et al., 1987; Roth et al., 1975).

1147 
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1148 3.3.1. The CYP2E1 Pathway 

1149 There is considerable evidence of the importance of the CYP2E1 metabolic pathway in 

1150 studies in animals (Oh et al., 2002; Wirkner et al., 1997; Kim and Kim, 1996; Lehnebach et al., 

1151 1995; Pankow et al., 1991a, b; Pankow and Hoffmann, 1989; Pankow, 1988; Glatzel et al., 1987; 

1152 Angelo et al., 1986a, b; Landry et al., 1983; Anders and Sunram, 1982; McKenna et al., 1982; 

1153 McKenna and Zempel, 1981; Rodkey and Collison, 1977; Carlsson and Hultengren, 1975; Roth 

1154 et al., 1975; Fodor et al., 1973) and humans (Takeshita et al., 2000; DiVincenzo and Kaplan, 

1155 1981; Astrand et al., 1975). These studies demonstrate that exposure to dichloromethane, 

1156 regardless of exposure route, results in the formation of CO, as assessed by direct measurements 

1157 of elevated levels of CO in expired air and increased levels of COHb in the blood. 

1158 The first step in the CYP2E1 pathway is the formation of formyl chloride (Figure 1). 

1159 Watanabe and Guengerich (2006) conducted a series of studies to investigate the downstream 

1160 metabolites of formyl chloride, and reported only marginal (3% maximum at pH 9) formation of 

1161 S-formyl GSH from formyl chloride in the presence of GSH. Therefore, most (>97%) of the 

1162 formyl chloride is metabolized further to carbon monoxide. Furthermore, CO formation from 

1163 formyl chloride was independent of GSH presence in the assay. 

1164 Results from numerous studies in rats in which CYP2E1 metabolism was blocked or 

1165 induced indicate that the generation of CO occurs as a result of metabolism of dichloromethane 

1166 by the CYP2E1 pathway (Figure 3-1). Co-exposure of rats to a high dose of ethanol 

1167 (174 mmol/kg), which is metabolized by CYP2E1, and dichloromethane (1.6, 6.2, 15.6 mmol/kg) 

1168 resulted in no increase in blood COHb, indicating that the metabolic pathway for CO formation 

1169 had been either blocked or saturated (Glatzel et al., 1987). Similar results have been seen with 

1170 coadministration of other known CYP substrates, including diethyldithiocarbamate (Lehnebach 

1171 et al., 1995), methanol (Pankow and Jagielki, 1993), benzene, toluene, and three xylene isomers 

1172 (Pankow et al., 1991b). Pretreatment of animals with CYP inducers (e.g., benzene, toluene, 

1173 xylenes, methanol, isoniazid), particularly those that induce CYP2E1, resulted in an increased 

1174 level of CO formation, as assessed by COHb formation or measurement in expired air, following 

1175 single exposures to dichloromethane (Kim and Kim, 1996; Pankow and Jagielki, 1993; Pankow 

1176 et al., 1991b; Pankow and Hoffmann, 1989; Pankow, 1988). Pretreatment with disulfuram, a 

1177 CYP2E1 blocker, resulted in a complete lack of formation of COHb following dichloromethane 

1178 exposure, indicating that CYP2E1 is the isozyme responsible for metabolism of dichloromethane 

1179 (Kim and Kim, 1996). 

1180 Evidence in hamster and rat studies suggests that the CYP2E1 pathway becomes 

1181 saturated at high dichloromethane exposure levels; comparable data from studies in mice were 

1182 not found. In hamsters, mean COHb percentages were elevated to a similar degree (about 28–

1183 30%, compared with <1% in controls) in three groups exposed by inhalation to 500, 1,500, or 

1184 3,500 ppm dichloromethane for 6 hours (Burek et al., 1984). After 21 months of exposure by 

1185 this protocol, mean COHb percentages in the three exposure groups remained similarly elevated, 
12 DRAFT - DO NOT CITE OR QUOTE 

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

1187 lack of accumulation of dichloromethane and CYP2E1 metabolites with chronic exposure. 

1188 McKenna et al. (1982) found that blood COHb levels in rats increased when inhalation exposure 

1189 concentration was increased from 50 to 500 ppm but that similar levels of COHb were reported 

1190 following exposure to 1,500 ppm as following exposure to 500 ppm; the peak blood COHb 

1191 percentages were approximately 10%. In rats exposed to 0, 50, 200, or 500 ppm for 6 hours/day, 

1192 5 days/week for 2 years, mean COHb percentages were 2.2, 6.5, 12.5, and 13.7%, respectively, 

1193 suggesting that saturation of the CYP2E1 pathway is approached at 200 ppm (Nitschke et al., 

1194 1988a). In male F344 rats exposed for 4 hours to dichloromethane concentrations of about 150, 

1195 300, 600, 1,000, and 2,000 ppm, mean COHb percentages (estimated from a figure) were about 

1196 4% at 150 ppm and about 8% at each of the four higher exposure concentrations (Gargas et al., 

1197 1986). McKenna and Zempel (1981) reported that increasing the oral dose of labeled 

1198 dichloromethane from 1 mg/kg to 50 mg/kg in rats resulted in a lower fraction of the total dose 

1199 being metabolized to CO. Single injections of 3 and 6 mmol/kg of dichloromethane in rats 

1200 resulted in nearly identical levels of blood COHb (Oh et al., 2002). 

1201 In human subjects exposed to dichloromethane in the workplace, saturation of CYP 

1202 metabolism appears to be approached in the 400–500 ppm range (Ott et al., 1983e). Blood 

1203 samples were drawn during working hours from 136 fiber production workers who were exposed 

1204 to dichloromethane, acetone, and methanol. TWA exposure concentrations for the workers were 

1205 determined by personal monitoring techniques, and percent COHb levels in the blood samples 

1206 were determined. Estimated TWA concentrations in the exposed workers showed a bimodal 

1207 distribution. The lower mode of exposure concentrations showed the highest frequency in the 

1208 150–200 ppm range, while the higher mode showed the highest frequency in the range of 450–

1209 500 ppm. Plots of percent COHb against TWA exposure concentrations showed that saturation 

1210 begins to be apparent in the 400–500 ppm range of exposure concentrations. 

1211 The liver is the tissue most enriched in CYP2E1 catalytic activity, but CYP2E1 protein 

1212 and messenger ribonucleic acid (mRNA) have been detected in other human tissues, including 

1213 the lung, brain, kidney, pancreas, bladder, small intestine, and blood lymphocytes (Nishimura et 

1214 al., 2003). As such, the liver is expected to be the main site of CYP metabolism of 

1215 dichloromethane, but other tissues are also expected to metabolize dichloromethane via this 

1216 pathway. Of particular relevance given the neurologic effects seen with dichloromethane are the 

1217 distribution and inducibility of CYP2E1 in different areas of the brain (Miksys and Tyndale, 

1218 2004). Individuals with decreased CYP2E1 activity may experience decreased generation of CO 

1219 and an increased level of GST-related metabolites following exposure to dichloromethane. As a 

1220 result, these individuals may be more susceptible to the chronic effects of dichloromethane from 

1221 GST-related metabolites than individuals with higher levels of CYP2E1 activity. Conversely, 

1222 individuals with higher CYP2E1 activity may experience relatively increased generation of CO 
13 DRAFT - DO NOT CITE OR QUOTE 

1223 at a given dichloromethane exposure level and, therefore, may be more susceptible to the acute 

1224 toxicity of dichloromethane (from CO). 

1225 Results from studies examining human interindividual variation in CYP2E1 activities 

1226 (e.g., catalytic activities, protein levels, or mRNA levels) indicate that individuals may vary in 

1227 their ability to metabolize dichloromethane through the CYP2E1 pathway. In a study of liver 

1228 samples from 30 Japanese and 30 Caucasian individuals, two- to threefold variation was found in 

1229 the levels of CYP2E1 protein, whereas catalytic activity toward substrates associated with 

1230 CYP2E1 (e.g., 7-ethoxycoumarin) displayed a wider range of values, approximately 25-fold; no 

1231 clear gender-specific or ethnic differences were found in hepatic levels of CYP2E1 protein or 

1232 enzymatic activities associated with CYP2E1 (Shimada et al., 1994). In a study of 

1233 interindividual variation in 70 healthy human subjects (40 men and 30 women) given an oral 

1234 dose of chlorzoxazone, a therapeutic agent whose metabolism and blood clearance has been 

1235 related to CYP2E1 levels, a three- to fourfold range in plasma half-life and clearance values was 

1236 observed, with no clear or dramatic age- or gender-specific differences (Kim et al., 1995). A six1237 to sevenfold range in chlorzoxazone hydroxylation activity was reported for a group of 

1238 69 healthy, smoking and nonsmoking male and female volunteers with mixed ethnic 

1239 backgrounds; the range was markedly increased when a group of 72 alcoholic inpatients was 

1240 included (Lucas et al., 1999). In studies of human liver microsomes, four- to sixfold ranges in 

1241 CYP2E1-dependent oxidation of trichloroethylene have been reported (Lipscomb et al., 2003, 

1242 1997). CYP2E1 protein levels in 50 specimens of human lymphocytes from healthy individuals 

1243 showed an approximate fivefold range (Bernauer et al., 2000), and a 3.7-fold range in liver 

1244 CYP2E1 mRNA levels was reported for a group of 24 patients with chronic hepatitis (Haufroid 

1245 et al., 2003). More recently, a threefold range was reported for maximal rates of hepatic 

1246 CYP2E1-catalyzed metabolism of dichloromethane, which were estimated with a modified 

1247 physiologically based toxicokinetic (PBTK) model originally developed by Andersen et al. 

1248 (1987) and kinetic data (e.g., dichloromethane breath and blood concentrations) for 13 volunteers 

1249 (10 males and 3 females) exposed to one or more concentrations of dichloromethane by 

1250 inhalation for 7.5 hours (Sweeney et al., 2004). In summary, most studies indicate a three- to 

1251 sevenfold variability in CYP2E1 activity, as assessed by various types of measurements, among 

1252 “healthy” volunteers. However, various clinical factors (i.e., obesity, alcoholism, use of specific 

1253 medications) or co-exposures (i.e., to various solvents) (Lucas et al., 1999) may result in greater 

1254 variation, and thus the potential for saturation at lower exposures, within the general population. 

1255 Several genetic polymorphisms for the human CYP2E1 gene have been described, but 

1256 clear and consistent correlations with interindividual variation in CYP2E1 protein levels or 

1257 associated enzyme activities have not been identified (Ingelman-Sundberg, 2004; Lucas et al., 

1258 2001; Kim et al., 1995; Shimada et al., 1994). The most frequently studied CYP2E1 

1259 polymorphisms, RsaI/PstI, are located in the 5 ́-flanking region of the gene, and mutations are 

1260 thought to lead to increased CYP2E1 protein expression via transcription (Lucas et al., 2001). 
14 DRAFT - DO NOT CITE OR QUOTE 

1261 Available data indicate that the frequency of this polymorphism, as well as other CYP2E1 

1262 polymorphisms, varies among ethnic groups. For example, Stephens et al. (1994) examined 

1263 blood samples from 126 African-Americans, 449 European Americans, and 120 Taiwanese 

1264 subjects and found frequencies for a rare RsaI allele (C2) of 0.01 in African-Americans, 0.04 in 

1265 European Americans, and 0.28 in Taiwanese subjects. In a study of 102 Mexicans, the reported 

1266 mutation frequency at the RsaI C2 allele was 0.30 (Mendoza-Cantú et al., 2004). 

1267 

1268 3.3.2. The GST Pathway 

1269 The other major pathway for dichloromethane metabolism involves the conjugation of 

1270 dichloromethane to GSH, catalyzed by GST. This results in the formation of a GSH conjugate 

1271 that is eventually metabolized to CO2 (Figure 3-1). The conjugation of dichloromethane to GSH 

1272 results in formation of two reactive intermediates that have been proposed to be involved in 

1273 dichloromethane toxicity, S-(chloromethyl)glutathione and formaldehyde. In studies with rat, 

1274 mouse, and human liver cytosol preparations in the presence of GSH, examination of metabolites 

with 13 1275 C-NMR indicated that S-(chloromethyl)glutathione was an intermediate in the pathway to 

1276 formaldehyde (Hashmi et al., 1994). Formaldehyde formation from dichloromethane has been 

1277 noted in human (Bruhn et al., 1998; Hallier et al., 1994; Hashmi et al., 1994), rat, and mouse 

1278 (Casanova et al., 1997; Hashmi et al., 1994) cells in vitro. Formation of free hydrogen ion is also 

1279 hypothesized, although no direct evidence supporting this has been presented. 

1280 The GST pathway has approximately a 10-fold lower affinity for dichloromethane than 

1281 the CYP pathway (Reitz et al., 1989; Andersen et al., 1987). At lower exposure concentrations, 

1282 the CYP pathway is expected to predominate, but, as exposure concentrations increase, the GST 

1283 pathway is expected to gain in relative importance as a dispositional pathway for absorbed 

1284 dichloromethane. Based on in vitro studies with liver preparations, the estimated Michaelis1285 Menten kinetic constants (Kms) in GST assays with dichloromethane were about 137 mM in a 

1286 B6C3F1 mouse preparation and about 44 mM in two human preparations (Reitz et al., 1989). In 

1287 contrast, estimated Kms in CYP assays were about 1.8, 1.4, and 2.0 mM in B6C3F1 mouse, F344 

1288 rat, and Syrian golden hamster preparations, respectively. In four human liver preparations, 

1289 estimated CYP Kms were about 2.6, 2.0, 0.9, and 2.8 mM (Reitz et al., 1989). 

1290 Early investigations indicated that in humans GSTs of the α-, μ-, and π-classes were not 

1291 responsible for the metabolism of dichloromethane (Bogaards et al., 1993). Tissue samples that 

1292 metabolized substrates specific to those GST classes did not conjugate dichloromethane to GSH. 

1293 Later investigations identified the recently-characterized GST theta class (Meyer et al., 1991), 

1294 specifically GST-theta1-1 (GST-T1), as the GST isoenzyme responsible for the metabolism of 

1295 dichloromethane (Mainwaring et al., 1996; Blocki et al., 1994). In the absence of the GST-T1 

1296 gene, no deoxyribonucleic acid (DNA)-protein cross-links were formed by human liver cells 

1297 exposed to dichloromethane (Casanova et al., 1997), and formaldehyde production was not 

1298 detected in human erythrocytes (Hallier et al., 1994). In a mouse model with a disrupted GST-
15 DRAFT - DO NOT CITE OR QUOTE 

1299 T1 gene, GST activity with dichloromethane in liver and kidney cytosol samples was 

1300 substantially lower compared with wild-type GST mice (Fujimoto et al., 2007). 

1301 A polymorphism of the GST-T1 gene has been demonstrated in humans. People with 

1302 two functional copies of the gene (+/+) readily conjugate GSH to dichloromethane. Individuals 

1303 having only one working copy of the gene (+/–) display relatively decreased conjugation ability. 

1304 Individuals with no functional copy of the gene (–/–) do not express active GST-T1 protein and 

1305 do not metabolize dichloromethane via a GST-related pathway (Thier et al., 1998). Results from 

1306 studies of GST-T1 genotypes in human blood samples indicate that average prevalences of the 

1307 GST-T1 null (–/–) genotype are higher in Asian ethnic groups (47–64%) than in other groups, 

1308 including Caucasians (19–20%), African-Americans (22%), and mixed groups (19%) (Raimondi 

1309 et al., 2006; Garte et al., 2001; Nelson et al., 1995) (see Table 3-3). Although information on the 

1310 age distribution of study subjects was not generally reported in these analyses, there is little 

1311 reason to expect effect modification by age since this is not a gene linked to early mortality. 

1312 Based on data collected by Nelson et al. (1995) and U.S. 2000 census data (and assuming Hardy1313 Weinberg equilibrium), Haber et al. (2002) calculated U.S. average distributions of GST-T1 

1314 genotypes as follows: 32% +/+; 48% +/–; and 20% –/–. 

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 reported 

Mexican American 9.7% (n = 73) Not reported Not reported 

Other Not reported Not reported 19.4 % (n = 1,485)

aNelson et al. (1995) examined prevalence of the null GST-T1 genotype from analysis of blood samples from 

subjects of various ethnicities as noted above. 

bGarte 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 polymorphisms in 

genes for bioactivating enzymes. 

cRaimondi et al. (2001) collected GST-T1 genotype data from 35 case-control studies of cancer and GST-T1 

genotype; data in this table are for control subjects. The “other” group in this study is defined as Latino, AfricanAmerican, and mixed ethnicities. 

1316 

1317 

1318 Results from a study of the distribution of activity levels for in vitro conjugation of 

1319 dichloromethane with GSH in 22 human liver samples are roughly reflective of these estimates 

1320 of the distribution of this polymorphism (Bogaards et al., 1993). No activity was found in 3/22 

1321 of the liver samples. Eleven of the samples showed low activity levels (0.21–0.41 nmol 
16 DRAFT - DO NOT CITE OR QUOTE 

1322 product/minute/mg protein), and eight samples showed high activity levels ranging from 0.82 to 

1323 1.23 nmol/minute/mg protein. In another study of seven human subjects, lysates of erythrocytes 

1324 showed high activities for producing formaldehyde from dichloromethane (presumably via GST1325 T1) in three subjects (15.4, 17.7, and 17.8 nmol product/minute/mg hemoglobin) and lower 

1326 activity in the other four subjects (4.3, 6.0, 7.2, and 7.6 nmol product/minute/mg hemoglobin) 

1327 (Hallier et al., 1994). 

1328 Comparisons of mice, rats, humans, and hamsters for the ability to metabolize 

1329 dichloromethane via the GST pathway in liver and lung tissues indicate that mice appear to be 

1330 the most active at metabolizing dichloromethane (Sherratt et al., 2002; Thier et al., 1998; 

1331 Casanova et al., 1997, 1996; Hashmi et al., 1994; Reitz et al., 1989). Reitz et al. (1989) reported 

1332 mean (± SD) GST enzymatic activity levels with dichloromethane as substrate (in units of nmol 

1333 product formed/minute/mg protein) in liver cytosol preparations to be: 25.9 ± 4.2 units in 

1334 B6C3F1 mice (n = 15 determinations per preparation); 7.05 ± 1.7 units in F344 rats (n = 6); and 

1335 1.27 ± 0.21 units (n = 6) in Syrian golden hamsters. Mean GST activity levels in liver 

1336 preparations from four human subjects (accident victims screened for human immunodeficiency 

1337 virus and hepatitis B and C and obtained through a transplant center) were 2.62 ± 0.44 units (n = 

1338 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 

1339 al., 1989). The finding that one of the four individuals was unable to conjugate dichloromethane 

1340 with GST was reflective of the estimated frequency of the GST-T1 null genotype in the U.S. 

1341 population (see Table 3-3). Mean GST activity levels in lung cytosol preparations showed a 

1342 similar rank order among species: 7.3 ± 1.4 units in mice (n = 4), 1.0 ± 0.1 units in rats (n = 4), 

1343 0.0 ± 0.2 units in hamsters (n = 4), and 0.37 ± 0.25 units in a pooled lung preparation from the 

1344 same four human subjects (n = 2). Reitz et al. (1989) noted that relative abilities of these animal 

1345 species to metabolize dichloromethane via the GST pathway correlated with their cancer 

1346 sensitivities in long-term inhalation bioassays: (1) B6C3F1 mice showed statistically significant 

1347 increased incidence of liver and lung tumors in a 2-year cancer bioassay (National Toxicology 

1348 Program [NTP], 1986); (2) rats showed much less evidence of increased incidence of liver 

1349 tumors, and no increased risk of lung tumors at equivalent exposure concentrations but showed 

1350 increased incidence of nonmalignant mammary tumors (NTP, 1986; Burek et al., 1984); and 

1351 (3) Syrian golden hamsters did not show tumorigenic responses at any site (Burek et al., 1984). 

1352 Thier et al. (1998) conducted a study evaluating the activity of GST-T1 after treatment of 

1353 dichloromethane in the cytosol of liver and kidney homogenates from hamsters (pooled male and 

1354 females), rats (pooled male and female), male mice, and female mice and for humans classified 

1355 as nonconjugators, low conjugators, or high conjugators of GST to dichloromethane. Little 

1356 information is provided about the human samples other than that 13 kidney cancer patients were 

1357 the source of the kidney samples; normal tissue identified by pathological exam was used. Blood 

1358 samples from 10 of these patients were collected and enzyme activities measured in erythrocytes 

1359 from 9 of these samples were reported. Results of conjugation of dichloromethane to GSH from 
17 DRAFT - DO NOT CITE OR QUOTE 

1360 these studies are presented in Table 3-4. As can be seen from the table, activity levels (expressed 

1361 as nmol/minute per mg of cytosolic protein) of humans varied considerably, with nonconjugators 

(presumed to be GST-T1–/– 1362 ) having no detectable activity, low conjugators (presumed to be 

GST-T1+/–) having moderate activity, and high conjugators (presumed to be GST-T1+/+ 1363 ) having 

1364 approximately twice the activity seen in low conjugators. In the liver, the activity of rat GST 

1365 conjugation was over twofold that seen in human high conjugators, while levels in mice were 

1366 >11-fold (males) or 18-fold (females) greater than those of human high conjugators. In the 

1367 kidney, the activity of high-conjugator humans was approximately 1.8-fold that of rats and was 

1368 comparable to the activity of both male and female mice. The data in Table 3-4 show the 

1369 following order for GST-T1 activities with dichloromethane as substrate: in liver preparations, 

1370 mouse >> rat > human high conjugators > human low conjugators > hamster > human 

1371 nonconjugators and, in kidney preparations, female mouse ≈ male mouse ≈ human high 

1372 conjugators > rat ≈ human low > hamster > human nonconjugators. In addition, the data indicate 

1373 that activity levels in liver, kidney, and erythrocytes of human subjects are in correspondence 

1374 with the nonconjugator, low conjugator, and high conjugator designations. 

1375 

Table 3-4. GST-T1 enzyme activities toward dichloromethane 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 measured 

Mouse, female 29.7 ± 6.31 (5) 3.88 ± 0.90 (5) Not measured 

Hamster 0.27 ± 0.20 (6) 0.25 ± 0.21 (6) Not measured 

aMean ± SD with number of samples noted in parentheses. 

Source: Adapted from Thier et al. (1998). 

1376 

1377 

1378 Sherratt et al. (2002) reported that, on a per mg basis, native recombinant mouse GST-T1 

1379 (purified after expression in Escherichia coli) was approximately twofold more active toward 

1380 dichloromethane than native recombinant human enzyme, as well as being approximately 

1381 fivefold more efficient (as assessed by the ratio of kcat/Km). 

1382 The distribution of GST-T1 in human tissues has been examined with antibodies raised 

1383 against recombinant human GST-T1 (Sherratt et al., 2002, 1997). Immunoblotting of sodium 

1384 dodecyl sulfate polyacrylamide gel electrophoresis gels loaded with tissue extracts from a 

1385 73 year-old man who had died with brochopneumonia and atherosclerosis indicated the 
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1386 following order of expression of GST-T1: liver ≈ kidney > prostate ≈ small intestine > 

1387 cerebrum ≈ pancreas ≈ skeletal muscle > lung ≈ spleen ≈ heart ≈ testis (Sherratt et al., 1997). 

1388 It was estimated that the levels of cross-reacting materials in the cerebrum, pancreas, or skeletal 

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

1390 and testis were less than 5% of the levels in the liver. Comparison of the amounts of cross1391 reacting material in soluble liver extracts from a B6C3F1 mouse and five human subjects (i.e., 

1392 normal liver tissue samples from biopsies of secondary liver tumors) found that levels of GST1393 T1 protein were higher in the mouse extracts than in any of the human liver extracts (Sherratt et 

1394 al., 2002). Densitometer analysis indicated that the GST-T1 level in the mouse liver extract was 

1395 about fivefold higher than those in human liver extracts displaying the highest level. Cross1396 reacting material was not detectable in liver extracts from one of the five human subjects, 

1397 indicating that this individual may have been GST-T1 null (Sherratt et al., 2002). 

1398 Results from in situ hybridization with oligonucleotide anti-sense probes for GST-T1 

1399 mRNA levels and immunohistochemical studies with antibodies to GST-T1 have indicated that 

1400 there may be subtle differences between mice and humans in the intracellular localization of 

1401 GST-T1 in the liver. Mainwaring et al. (1996) reported that staining for GST-T1 mRNA was 

1402 higher in liver slices from B6C3F1 mice than in liver slices from F344 rats and that staining in 

1403 human liver samples was very low. Although the number of mouse and rat liver samples 

1404 examined in this study was not indicated in the available report, it was reported that slices from 

1405 five human liver samples were examined. No information was provided regarding the clinical 

1406 history of the sources of the human samples. In mouse liver, staining for GST-T1 mRNA was 

1407 enhanced in the limiting plate hepatocytes, in nuclei, in bile-duct epithelial cells, and in lesser 

1408 amounts in the centrilobular cells in general. In rat liver, a similar pattern was observed, except 

1409 no enhanced staining was observed in the limiting plate hepatocytes or in nuclei. Staining for 

1410 GST-T1 mRNA in the human liver samples showed an even distribution throughout the liver 

1411 lobule, and no mention of a specific nuclear localization was made (Mainwaring et al., 1996). 

1412 Quondamatteo et al. (1998), using antibodies to GST-T1, subsequently reported a similar 

1413 localization of GST-T1 protein in nuclei of cells in mouse liver slices. In another study using 

1414 antibodies raised against recombinant human GST-T1 or a peptide derived from the deduced 

1415 mouse GST-T1 primary sequence, Sherratt et al. (2002) reported that nuclear staining was 

1416 observed in all cells in mouse liver slices (from five individual B6C3F1 mice) showing the 

1417 presence of mouse GST-T1; staining in the cytoplasm was only detected in cells with very high 

1418 levels of GST-T1. In liver slices obtained from two human subjects (males, ages 60 and 

1419 61 years, with a secondary liver tumor and what was described as a “cavernous hemangioma” 

1420 without malignancy, respectively), the most intense nuclear staining was associated with bile 

1421 duct epithelial cells, but there was heterogeneity of staining within hepatocytes; some cells 

1422 showed nuclear staining, but others only exhibited cytoplasmic staining (Sherratt et al., 2002). 
19 DRAFT - DO NOT CITE OR QUOTE 

1423 In summary, the relative amount of dichloromethane metabolized via the GST pathway 

1424 increases with increasing exposure concentrations. As the high affinity CYP pathway becomes 

1425 saturated (either from high exposure levels of genetic or other factors that decrease CYP2E1 

1426 activity), the GST pathway increases in relative importance as a dispositional pathway for 

1427 dichloromethane. Two reactive metabolites (S-(chloromethyl)glutathione and formaldehyde) 

1428 resulting from this pathway have been identified. GST-T1 is the GST isozyme that catalyzes 

1429 conjugation of dichloromethane with GST. Interindividual variation in the ability to metabolize 

1430 dichloromethane via GST-T1 is associated with genetic polymorphisms in humans. Estimated 

1431 U.S. population prevalence of nonconjugators (–/– at the GST-T1 locus) is about 20%, but higher 

1432 prevalences (47–64%) have been reported for Asians (Raimondi et al., 2006; Haber et al., 2002; 

1433 Garte et al., 2001; Nelson et al., 1995). The prevalences for low (+/– at the GST-T1 locus) and 

1434 high (+/+) conjugators have been estimated at 48 and 32%, respectively (Haber et al., 2002). 

1435 The liver and kidney are the most enriched tissues in GST-T1, but evidence is available for the 

1436 presence of GST-T1 in other tissues at lower levels, including the brain and lung. In humans, 

1437 GST-T1 expression in the brain is lower than that seen in the liver or kidney but higher than in 

1438 the lung. Comparisons of mice, rats, humans, and hamsters for the ability to metabolize 

1439 dichloromethane via the GST pathway in liver (based on measurement of tissue-specific enzyme 

1440 activity) indicate the following rank order: mice > rats > or ≈ humans > hamsters. This relative 

1441 ranking corresponds to the rank order of the strength of the association between inhalation 

1442 exposure to dichloromethane and liver tumors in long-term cancer bioassays with mice, rats, and 

1443 hamsters. In mouse liver tissue, GST-T1 appears to be localized in the nuclei of hepatocytes and 

1444 bile-duct epithelium, but rat liver does not show preferential nuclear localization of GST-T1. In 

1445 human liver tissue, some hepatocytes show nuclear localization of GST-T1 and others show 

1446 localization in cytoplasm, as well as in bile duct epithelial cells. The apparent species 

1447 differences in intracellular localization of GST-T1 may play a role in species differences in 

1448 susceptibility to dichloromethane carcinogenicity if nuclear production of 

1449 S-(chloromethyl)glutathione is more likely to lead to DNA alkylation than cytoplasmic 

1450 production. 

1451 

1452 3.4. ELIMINATION 

1453 Dichloromethane is eliminated mainly through exhalation either of the parent compound 

1454 or as the two primary metabolites CO2 and CO (Angelo et al., 1986a, b; McKenna et al., 1982; 

1455 DiVincenzo and Kaplan, 1981; DiVincenzo et al., 1972, 1971). In human studies, 

1456 dichloromethane is rapidly eliminated from the body following the cessation of exposure, with 

1457 much of the parent compound completely removed from the bloodstream and expired air by 

1458 5 hours postexposure in experiments using exposure levels of 90, 100, or 210 ppm (DiVincenzo 

1459 et al., 1972, 1971; Riley et al., 1966). Studies in rats have similarly demonstrated that 

1460 elimination from the blood is rapid, with elimination half-times in F344 rats on the order of 4–
20 DRAFT - DO NOT CITE OR QUOTE 

1461 6 minutes following intravenous doses in the range of 10–50 mg/kg (Angelo et al., 1986a). In a 

1462 study using Sprague-Dawley rats, Carlsson and Hultengren (1975) demonstrated variability in 

1463 elimination rates between different types of tissues, with the most rapid elimination seen in the 

1464 adipose and brain tissue, while elimination from liver, kidneys, and adrenals proceeded more 

1465 slowly. 

1466 In a study using human volunteers, DiVincenzo and Kaplan (1981) reported a dose1467 related increase in CO in the expired breath after inhalation exposure to 50–200 ppm of 

1468 dichloromethane, with a net elimination as CO on the order of 25–35% of the absorbed dose. 

1469 Similar results have been reported in animal studies. Following gavage administration of 50 or 

200 mg/kg-day doses of [14 1470 C]-labeled dichloromethane in water to groups of six mature male 

1471 F344 rats for up to 14 days, >90% of the label was recovered in the expired air within 24 hours 

1472 of dose administration (Angelo et al., 1986b). Following administration of the first of 14 daily 

1473 50 mg/kg-day doses, radioactivity in parent compound, CO2, and CO in the 24-hour expired 

1474 breath accounted for 66, 17, and 16% of the administered radioactivity, respectively; similar 

1475 patterns were reported for 24-hour periods following administration of the seventh and 

1476 fourteenth 50 mg/kg-day dose. Following administration of the first 200 mg/kg-day dose, 

1477 radioactivity in parent compound, CO2, and CO in the 24-hour expired breath accounted for 77, 

1478 9, and 6%, respectively, of the administered radioactivity (Angelo et al., 1986b). In mature, male 

Sprague-Dawley rats given a smaller dose (1 mg/kg) of [14 1479 C]-labeled dichloromethane, 

1480 radioactivity in parent compound, CO2, and CO in 48-hour expired breath accounted for 12, 35, 

1481 and 31%, respectively; these data indicate that, at lower dose levels, a greater percentage of the 

1482 administered dose was metabolized by the CYP pathway and eliminated in the expired breath, 

1483 compared with higher dose levels (McKenna and Zempel, 1981). Similar patterns of 

1484 radioactivity distribution in parent compound, CO2, and CO in expired breath were found in 

1485 mature male B6C3F1 mice following gavage administration of 50 mg/kg-day (in water), or 

500 or 1,000 mg/kg-day (in corn oil), [14 1486 C]-labeled dichloromethane (Angelo et al., 1986a). For 

1487 example, radioactivity in parent compound, CO2, and CO in 24-hour expired breath accounted 

1488 for 61, 18, and 11% of the administered radioactivity, following administration of a single 

1489 50 mg/kg dose to a group of six mice (Angelo et al., 1986a). Exhalation rates were similarly 

1490 high following inhalation exposure of mature male Sprague-Dawley rats (>90%) (McKenna et 

1491 al., 1982) or following intravenous administration of dichloromethane to mature male F344 rats 

1492 (Angelo et al., 1986b). 

1493 Elimination of dichloromethane in the urine of exposed humans is generally small, with 

1494 total urinary dichloromethane levels on the order of 20–25 μg or 65–100 μg in 24 hours 

1495 following a 2-hour inhalation exposure to 100 or 200 ppm, respectively (DiVincenzo et al., 

1496 1972). However, a direct correlation between urinary dichloromethane and dichloromethane 

1497 exposure levels was found in volunteers, despite the comparatively small urinary elimination 

1498 (Sakai et al., 2002). Following administration of a labeled dose in animals, regardless of 
21 DRAFT - DO NOT CITE OR QUOTE 

1499 exposure route, generally <5–8% of the label is found in the urine and <2% in the feces 

1500 (McKenna et al., 1982; McKenna and Zempel, 1981; DiVincenzo et al., 1972, 1971). 

1501 

1502 3.5. PHYSIOLOGICALLY BASED TOXICOKINETIC MODELS 

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

1504 from 1986 to 2006. These models are mathematical representations of the body and its 

1505 absorption, distribution, metabolism, and elimination of dichloromethane and select metabolites, 

1506 based on the structure of the Ramsey and Andersen (1984) model for styrene. The models’ 

1507 equations are designed to mimic actual biological behavior of dichloromethane, incorporating in 

1508 vitro and in vivo data to define physiological and metabolic equation parameters. As such, the 

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

1510 dichloromethane and metabolite internal dosimeters (i.e., instantaneous blood and tissue 

1511 concentration, area under the curve [AUC] of concentration versus time plots, rate of metabolite 

1512 formation), allowing for the extrapolation of toxicity data across species, route of exposure, and 

1513 high to low exposure levels. The development of dichloromethane PBTK models has resulted in 

1514 either increased biological detail and functionality or refinement of model parameters with newly 

1515 available data. The former type of development provides more options for toxicity data 

1516 extrapolation, while the latter serves to increase confidence in model predictions and decrease 

1517 uncertainty in risk assessments for which the models were, or will be, applied. This section of 

1518 the document describes each of the models reported in the scientific literature and/or used by the 

1519 regulatory community (i.e., Occupational Safety and Health Administration [OSHA], EPA) and 

1520 their contribution to the advancement of predictive dosimetry and data extrapolation for 

1521 dichloromethane. In some instances, model development was accomplished by the addition of 

1522 new biological compartments (e.g., tissue systems). Diagrams of the compartmental structure of 

1523 the models are shown in Figure 3-2. Significant statistical advances in parameter estimation also 

1524 have been incorporated in model development. For this reason, some animal and human PBTK 

1525 models may be described as deterministic (Sweeney et al., 2004; Casanova et al., 1996; Reitz et 

1526 al., 1988a, b; U.S. EPA, 1988b, 1987a, b; Andersen et al., 1987; Gargas et al., 1986) in which 

1527 point estimates for each model parameter are used, resulting in point estimates for dosimetry. 

1528 Others may be described as probabilistic (Jonsson and Johanson, 2001; El-Masri et al., 1999; 

1529 OSHA, 1997), in which probability distributions for each parameter were defined, resulting in 

1530 probability distributions for dosimetry. The latter approach, particularly utilizing a Bayesian 

1531 hierarchical statistical model structure (described below) (David et al., 2006; Marino et al., 2006) 

1532 to estimate parameter values, allows for the introduction of intra- and interspecies variability into 

1533 model predictions and quantitative assessment of model uncertainty. Both deterministic (U.S. 

1534 EPA, 1988b, 1987a, b) and probabilistic (OSHA, 1997) applications have been used to develop 

1535 regulatory values. As discussed below, subsequent applications of the developed models for 

1536 cancer risk assessment have resulted in significantly different estimates of human cancer risk. 
22 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 

1580 Figure 3-2. Schematics of PBTK models (1986–2006) used in the 

1581 development of estimates for dichloromethane internal dosimetry. 

1582 

1583 Key references: Model A—Gargas et al. (1986); B—Andersen et al. (1987); C—Andersen et al. 

1584 (1991); D—Casanova et al. (1996); E—Sweeney et al. (2004); F—OSHA (1997); G—Jonsson 

1585 and Johanson (2001). Models C–G all build on the structure in model B. Models E and G have 

1586 been applied in humans; all others have been applied in humans and rodents (mice and/or rats). 

1587 CYP = CYP pathway metabolites; GST = GST pathway metabolites.

Gas 

Exchange Lung

Fat

Richly 

Perfused

Slowly 

Perfused

Liver

GST CYP

Blood

Alveolar 

Air

Endogenous 

Production

Gas 

Exchange

Fat

Richly 

Perfused

Slowly 

Perfused

Liver

GST CYP

CO Sub 

Model

GST CYP 3-2A 3-2B

Gas 

Exchange Lung

Fat

Richly 

Perfused

Slowly 

Perfused

Liver

GST CYP

GST CYP

Formaldehyde

DNA-protein cross-links

Blood

Alveolar 

Air

Endogenous 

Production

Gas 

Exchange Lung

Fat

Richly 

Perfused

Slowly 

Perfused

Liver

GST CYP

GST CYP

CO Sub 

Model

3-2C 3-2D

Gas 

Exchange Lung

Richly 

Perfused

Perineral

Fat

Subcutaneous 

Fat

Liver

GST CYP

GST CYP

Working 

Muscle

Resting 

Muscle

Gas 

Exchange Lung

Fat

Richly 

Perfused

Slowly 

Perfused

Liver

GST CYP

GST CYP

Bone 

Marrow

3-2F

3--2G

Blood

Alveolar 

Air

Endogenous 

Production

Gas 

Exchange Lung

Fat

Richly 

Perfused

Slowly 

Perfused

Liver

GST CYP

GST CYP

CO Sub 

Model

CYP

3-2E
23 DRAFT - DO NOT CITE OR QUOTE 

1588 The deterministic rat model of Gargas et al. (1986), based on previous work by Ramsey 

1589 and Andersen (1984) examining inhalation pharmacokinetics of styrene in rats, was the first 

1590 PBTK model for dichloromethane. It was comprised of four compartments (fat, liver, richly 

1591 perfused tissues, and slowly perfused tissues [Figure 3-2A]) and described flows and partitioning 

1592 of parent material and metabolites through the compartments with differential equations. 

1593 Metabolism, which was restricted to the liver compartment, was described as two competing 

1594 pathways: the GST pathway, described with a linear first-order kinetic model, and the CYP 

1595 pathway, described with a saturable Michaelis-Menten kinetic model. Rate constants for the 

1596 CYP and GST pathways in rats were determined by optimization of the model with in vivo gas 

1597 uptake data. COHb production was modeled both endogenously and from CYP-mediated 

1598 metabolism of dichloromethane. This model demonstrated the dose-dependent flux through the 

1599 competing CYP and GST metabolic pathways and the effect of CYP inhibition on COHb 

1600 generation. 

1601 Andersen et al. (1987) extended the rat model of Gargas et al. (1986) to include a lung 

1602 compartment, including CYP and GST metabolism pathways within the lung, in rats, mice, 

1603 hamsters, and humans (Figure 3-2B). Physiological flow rates were allometrically scaled among 

1604 species by 3⁄4 power of body weight (BW). Rate constants for the CYP and GST pathways in 

1605 rodents were determined by optimization of the model with in vivo gas uptake data. CYP rate 

1606 constants for humans were derived from data on dichloromethane uptake in human subjects 

1607 (number of subjects not reported). Human GST rate constants were derived by allometric 

1608 scaling of the animal GST rate constants. Model predictions compared favorably with kinetic 

1609 data for human subjects exposed by inhalation to dichloromethane (Andersen et al., 1987). 

1610 Using the mouse cancer bioassay data from NTP (1986), Andersen et al. (1987) compared the 

1611 linear body surface area-derived or the PBTK model-derived human liver and lung dose 

1612 surrogates associated with tumor development (mg dichloromethane metabolized via GST 

1613 pathway/volume tissue/day). They reported that PBTK model-extrapolated human liver and lung 

1614 internal doses were for inhalation exposure 167- and 144-fold lower and for drinking water 

1615 exposure 45- and 213-fold lower, respectively, than body surface area scaled internal doses. The 

1616 study authors suggested that the lower model-predicted human internal dose surrogates were due 

1617 to the need to saturate the CYP pathway before appreciable tumorigenic metabolite levels could 

1618 be attained, which is not captured by extrapolation based on body surface area. 

1619 U.S. EPA (1988b, 1987a, b) slightly modified the Andersen et al. (1987) model for mice 

1620 by using different alveolar ventilation and cardiac flow rates and used the mouse and human 

1621 models to derive human cancer risks from animal tumor incidence data. The flow rate 

1622 parameters in the Andersen et al. (1987) model were based on a human breathing rate of 

12.5 m3

/day (reflecting a resting rate), compared with the EPA value of 20 m31623 /day (reflecting 

average daily activity level) and a mouse breathing rate of 0.084 m31624 /day (based on allometric 

1625 scaling of bioassay-specific BWs), compared with the rate commonly used by EPA, 
24 DRAFT - DO NOT CITE OR QUOTE 

0.043 m31626 /day (U.S. EPA, 1987a). The internal dose metric used in the applications of the model 

1627 to cancer risk assessment was reflective of the amount of dichloromethane metabolized by the 

1628 GST pathway. In addition to using the mouse and human PBTK models to account for species 

1629 differences in dosimetry, a body surface area correction factor of 12.7 was applied to low-dose 

1630 slopes of estimated dose-response relationships for liver and lung tumors in mice to account for 

1631 presumed higher human responsiveness, relative to mice, to dichloromethane-induced cancer 

1632 (U.S. EPA, 1987a). The factor of 12.7 is the cube root of the ratio of human to mouse reference 

1633 BWs; this BW scaling factor was applied to adjust for interspecies toxicodynamic variability 

1634 (i.e., presumed differences in the lifetime impact in mice and humans of a given daily amount of 

1635 dichloromethane metabolically activated per liter of tissue) (Rhomberg, 1995). A human cancer 

inhalation unit risk (IUR) of 4.7 × 10–7 per (μg/m3

1636 ), based on this analysis, was placed on IRIS in 

1637 September 1990. 

1638 The Andersen et al. (1987) models were also modified by addition of submodel structures 

1639 for estimation of new dosimeters of interest. Andersen et al. (1991) added the capability to 

1640 specifically describe the kinetics of dichloromethane, CO, and COHb in rats and humans with 

1641 the addition of the Coburn-Forster-Kane equation to describe CO and COHb kinetics 

1642 (Figure 3-2C). However, equations were not added for metabolism of dichloromethane to CO in 

1643 the lung. Casanova et al. (1996) extended the Andersen et al. (1987) mouse model to include a 

1644 submodel that predicted the formation of formaldehyde and DNA-protein cross-links in the liver 

1645 (Figure 3-2D). 

1646 Further refinements of the Andersen et al. (1987) models allowed for incorporation of 

1647 new data. New in vitro measurements of metabolic rate constants in human and animal tissues 

1648 were incorporated into the Andersen et al. (1987) models by Reitz and coworkers (Reitz, 1991; 

1649 Reitz et al.1988a, b). Sweeney et al. (2004) modified the Andersen et al. (1987) human PBTK 

1650 model, adding extrahepatic CYP metabolism in richly perfused tissues (Figure 3-2E) to obtain a 

1651 better fit of the model to kinetics data for humans. Data for 13 volunteers (10 men and 

1652 3 women) who were exposed to one or more concentrations of dichloromethane for 7.5 hours 

1653 included dichloromethane concentrations in breath and blood, COHb concentrations in blood, 

1654 and CO concentrations in exhaled breath. Individual CYP Vmaxc (maximal velocity) values were 

1655 obtained by optimizing model predictions to match time-course data simultaneously for 

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

individual values of CYP Vmaxc ranged from 7.4 to 23.6 mg/hour/kg0.7 1657 , indicating an 

1658 approximate threefold range in maximal CYP metabolic activity. 

1659 The significance of metabolic variability for the kinetics of dichloromethane in animals 

1660 and humans was explored by several investigators using PBTK models. Dankovic and Bailer 

1661 (1994) used the updated human model presented by Reitz et al. (1988a, b) to explore the 

1662 consequences of interindividual variability in vitro kinetic constants for the CYP and GST 

1663 pathways (based on data for four human subjects) and reported that predicted GST-metabolized 
25 DRAFT - DO NOT CITE OR QUOTE 

1664 doses to the lung and liver could range from about zero to up to fivefold greater than those 

1665 predicted with the values of these rate constants used in the Reitz et al. (1988a, b) model. 

1666 El-Masri et al. (1999) replaced parameter estimates in the mouse and human PBTK models 

1667 presented by Casanova et al. (1996) with probability distributions, including published 

1668 information on the distribution of GST-T1 polymorphism in human populations, and used Monte 

1669 Carlo simulations to estimate distributions of cancer potency of dichloromethane in mice, 

1670 distributions of the amount of DNA-protein cross-links formed in the liver of humans, and 

1671 distributions of human cancer risks at given exposure levels of dichloromethane. The analysis 

1672 showed that, at exposure levels of 1, 10, 100, and 1,000 ppm dichloromethane, average and 

1673 median cancer risk estimates were 23–30% higher when GST-T1 polymorphism was not 

1674 included in the model. 

1675 Given the demonstrated influence of population variability in dichloromethane 

1676 metabolism on PBTK model-derived cancer risk estimates (El-Masri et al., 1999; Dankovic and 

1677 Bailer, 1994), PBTK model development has included a more formal statistical treatment of data 

1678 for physiological and metabolic variability. Bayesian statistical approaches have been applied to 

1679 develop probabilistic PBTK models for dichloromethane. Probabilistic models account for 

1680 variability between individuals in model parameters by replacing point estimates for the model 

1681 parameters with probability distributions. Calibration or fitting of probabilistic PBTK models to 

1682 experimental toxicokinetic data is facilitated by a Bayesian technique called Markov Chain 

1683 Monte Carlo (MCMC) simulation, which quantitatively addresses both variability and 

1684 uncertainty in PBTK modeling (Jonsson and Johanson, 2003). 

1685 OSHA (1997) used MCMC simulation to fit probabilistic versions of the Reitz et al. 

1686 (1988a, b) and Andersen et al. (1991, 1987) mouse and human models, which included 

1687 probability distributions for all model parameters. GST- and CYP-mediated metabolism 

1688 occurred in the liver and lung compartments (see Figure 3-2F). The model parameters were 

1689 modified to focus on occupational exposure scenarios; that is, a parameter distribution for work 

1690 intensity (using data from Astrand [1989]) was added, which adjusted physiological flow rates as 

1691 a function of work intensity as measured in watts. In addition, updated measurements of 

1692 blood:air and tissue:air partition coefficients (Clewell et al., 1993) were used to describe 

1693 distributions for these parameters. The Clewell et al. (1993) blood:air partition coefficient of 23 

1694 is higher than the value of 8.29 reported by Andersen et al. (1987) and used by EPA (U.S. EPA, 

1695 1988b, 1987a, b). The newer Clewell et al. (1993) value for mice is the preferred value, since it 

1696 is much closer to the values for rats (19.4) and hamsters (22.5) rather than humans (9.7), as 

1697 reported by Andersen et al. (1987). Distributions of metabolic, physiological, and partitioning 

1698 parameters in the mouse and human models were updated by using Bayesian methods with data 

1699 for mice and humans in published studies of mouse and human physiology and dichloromethane 

1700 kinetic behavior. 
26 DRAFT - DO NOT CITE OR QUOTE 

1701 Jonsson et al. (2001) used additional human kinetics data to expand the PBTK model of 

1702 Reitz et al. (1988a, b) and added new model compartments (Figure 3-2G). These investigators 

1703 used MCMC simulation to develop a probabilistic model from the Reitz et al. (1988a, b) human 

1704 model by using published in vitro measurements of liver Vmax for the CYP pathway (Reitz et al., 

1705 1989) and kinetic data for five human subjects exposed by inhalation to dichloromethane 

1706 (Astrand et al., 1975). A working muscle compartment was added to the basic Andersen et al. 

1707 (1987) and Reitz et al. (1988a, b) structure (see Figure 3-2G). Jonsson and Johanson (2001) 

1708 refined and extended this probabilistic model by including an additional fat compartment (to 

1709 provide a better description of the experimental data for the time course of dichloromethane in 

1710 subcutaneous fat), incorporating (with MCMC simulation) kinetic data for dichloromethane in an 

1711 additional 21 human subjects and including three GST-T1 genotypes/phenotypes 

1712 (nonconjugators –/–, low conjugators +/–, high conjugators +/+). Monte Carlo simulations were 

1713 then used with the refined probabilistic model to predict human liver cancer risk estimates at 

1714 several dichloromethane exposure levels using an algorithm similar to the one used by El-Masri 

et al. (1999), using DNA-protein cross-links as the internal dose metric. The mean, 50th, 90th 1715 , 

and 95th 1716 percentile human cancer risk values from Jonsson et al. (2001) and El-Masri et al. 

1717 (1999) were very similar, within onefold of one another for simulated exposure levels up to 

1718 100 ppm. 

1719 The most statistically rigorous and data-intensive PBTK model development was 

1720 performed by Marino et al. (2006) for mice and David et al. (2006) for humans. Development of 

1721 these models used multiple mouse and human data sets in a Bayesian hierarchical statistical 

1722 structure to quantitatively capture population variability and reduce uncertainty in model 

1723 dosimetry and the resulting risk values. EPA used these models in the derivation of reference 

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

1725 more detail below. 

1726 

1727 3.5.1. Probabilistic Mouse PBTK Dichloromethane Model (Marino et al., 2006) 

1728 Marino et al. (2006) used MCMC analysis to develop a probabilistic PBTK model for 

1729 dichloromethane in mice, using the Andersen et al. (1987) model structure as a starting point 

1730 (Figure 3-3). Metabolic kinetic parameters (Vmaxc, Km, kfC, A1, and A2) (Table 3-5) were 

1731 calibrated with this Bayesian methodology by using several experimental data sets. Distribution 

1732 parameters (i.e., means and coefficients of variation [CVs]) for other physiological parameters 

1733 (i.e., BW, fractional flow rates, and fractional tissue volumes) and partition coefficients were 

1734 taken from the general literature as noted by Clewell et al. (1993). Marino et al. (2006) noted 

1735 that using distributions for these latter parameters from the general literature (based on a large 

1736 number of animals) was better than updating them based on the relatively smaller number of 

1737 animals in the available dichloromethane kinetic studies. Clewell et al. (1993) determined 

1738 blood:air and tissue:air partition coefficients (means and CVs) with tissues from groups of male 
27 DRAFT - DO NOT CITE OR QUOTE 

1739 and female B6C3F1 mice. These partition coefficients were derived by using a vial equilibration 

1740 method similar to that used by prior investigators (Andersen et al., 1987; Gargas et al., 1986). 

1741 Tissue:air partition coefficients were approximately two to three times lower than previously 

1742 utilized values with the exception of the liver coefficient, which was similar to previous values 

1743 (Table 3-5). The blood:air partition coefficient (23) from Clewell et al. (1993) is higher than the 

1744 previously reported value of 8.3 (Gargas et al., 1986). The higher value is more in line with 

1745 values measured in rats (19.4) and hamsters (22.5) and, thus, is more reasonable than the older 

1746 value of 8.3. Table 3-5 shows mean and CVs for physiological parameters and partition 

1747 coefficients in the Marino et al. (2006) mouse model as well as values used in earlier 

1748 deterministic PBTK mouse models for dichloromethane. 

1749 

1750 

1751 

1752 

1753 

1754 

1755 

1756 

1757 

1758 

1759 

1760 

1761 

1762 

1763 

1764 

1765 

1766 

1767 Figure 3-3. Schematic of mouse PBTK model used by Marino et al. (2006). 

1768 

1769 

1770 The Bayesian calibration of the cardiac output constant (QCC), ventilation:perfusion ratio 

1771 (VPR), and metabolic parameters was divided into three sequential steps: using kinetic data from 

1772 closed chamber studies with mice treated with an inhibitor of CYP2E1 (trans-1,2-

1773 dichloroethylene) in order to minimize the oxidative pathway and enable a more precise estimate 

1774 of parameters for the GST pathway, followed by kinetic data for mice given intravenous 

1775 injections of dichloromethane to estimate metabolism parameters in the absence of pulmonary 

1776 absorption processes and, finally, kinetic data for naïve mice exposed to dichloromethane in 

1777 closed chambers (Marino et al., 2006). The initial prior distributions were based on mean values 

Blood 

Alveolar 

air 

Endogenous 

production 

Gas 

exchange 

Lung 

Fat 

Richly 

perfused 

Slowly 

perfused 

Liver 

GST CYP

GST CYP 

CO sub 

model 
28 DRAFT - DO NOT CITE OR QUOTE 

1778 used by Andersen et al. (1987) for the metabolic parameters and by OSHA (1997) for the 

parameters for VPR, ratio of lung Vmax to liver Vmax (A1), and ratio of lung GST 1st 1779 order kinetic 

1780 constant (lung KF) to liver KF (A2). Posterior distributions from the first Bayesian analysis were 

1781 used as prior distributions for the second step, and posterior distributions from the second step 

1782 were used as prior distributions for the final updating. Final results from the Bayesian 

1783 calibration of the mouse probabilistic model are shown in Table 3-5. 
29

Table 3-5. Values for parameter distributions in a B6C3F1 mouse probabilistic PBTK model for 

dichloromethane compared with associated values for point parameters in earlier deterministic B6

PBTK models for dichloromethane 

Marino et al. (2006)a

Parameter Prior mean Prior CV

Final posterior 

 mean 

Final posterior 

CV 

U.S. 

(198

1987

Fractional flow rates (fraction of QCC)b

QFC Fat QLC Liver QRC Rapidly perfused tissues QSC Slowly perfused tissues 

0.05 0.24 0.52 0.19 

0.60 0.96 0.50 0.40 

0.0.0.0.

Fractional tissue volumes (fraction of BW)b

VFC Fat VLC Liver VLuC Lung VRC Rapidly perfused tissues VSC Slowly perfused tissues 

0.04 0.04 

0.0115 

0.05 0.78 

0.30 0.06 0.27 0.30 0.30 

0.0.

0.0

0.0.

Partition coefficientsc

PB Blood:air PF Fat:blood PL Liver:blood PLu Lung:blood PR Rapidly perfused:blood PS Slowly perfused:blood 

23 

5.1 1.6 

0.46 0.52 0.44 

0.15 0.30 0.20 0.27 0.20 0.20 

These parameters were taken from an 

extensive literature database derived from 

a large number of animals; therefore, 

further Bayesian updating does not 

inform on the true mean and variance for 

these values. 

8.141.1.1.0.

Flow rates QCC Cardiac output (L/hr/kg0.74) 

VPR ventilation:perfusion ratio 

28.0 1.52 

0.58 0.75 

24.2 1.45 

0.19 0.20 

14

1

Metabolism parameters Vmaxc Maximum CYP metabolic rate (mg/hr/kg0.7) 

Km CYP affinity (mg/L) 

kfC First-order GST metabolic rate constant (kg0.3/hr) 

A1 Ratio of lung Vmaxc to liver Vmaxc 

A2 Ratio of lung kfC to liver kfC 

11.1 

0.396 

1.46 

0.462 0.322 

2 2 2 

0.55 0.55 

9.27 

0.574 

1.41 

0.207 0.196 

0.21 0.42 0.28 0.36 0.37 

11

0.3

1.

0.40.1

1784 

a 1785 MCMC analysis was used to update prior distributions (means and CVs) for flow rate and metabolic parameters in a sequential process w

1786 data from mouse studies, as explained further in the text. Final values for posterior distributions are given in this table. 

b 1787 Source: Andersen et al. (1987, 1991). c 1788 Source: Clewell et al. (1993). dBased on a mouse breathing rate of 0.043 m3 1789 /day. eBased on a mouse breathing rate of 0.084 m3 1790 /day. 
30

1791 Marino et al. (2006) used the Bayesian-calibrated mouse model to calculate internal dose 

1792 metrics associated with exposure conditions in the NTP (1986) B6C3F1 mouse cancer inhalation 

1793 bioassay. The internal dose metric selected was milligrams (mg) dichloromethane metabolized 

1794 by the GST pathway per liter tissue per day. This is the same dose metric used in earlier 

1795 applications of PBTK models to derive human cancer IUR estimates based on cancer responses 

1796 in mice (OSHA, 1997; Andersen et al., 1987; U.S. EPA, 1987a, b). Its use is consistent with 

1797 evidence that dichloromethane metabolism via GST-T1 results in the formation of a reactive 

1798 metabolite that damages DNA and results in the formation of tumors (see section 4.7). The 

1799 model was used to calculate values for this internal dose metric in the lung and liver of mice in 

1800 the NTP (1986) study, using the mean values of the final distributions for the parameters in the 

1801 model. Resultant values were three- to four-fold higher than values calculated with the Andersen 

1802 et al. (1987) and U.S. EPA (1987a, b) versions of the model (Table 3-6). Marino et al. (2006) 

1803 noted that the difference could be primarily attributed to the changes in the partition coefficients 

1804 based on Clewell et al. (1993) as well as to the Bayesian updating of the metabolic parameters 

1805 (see Table 3-5). 

1806 

Table 3-6. Internal daily doses for B6C3F1 mice exposed to 

dichloromethane for 2 years (6 hours/day, 5 days/week) calculated with 

different PBTK models 

PBTK model 

Target organ 

NTP (1986) 

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

Liverb Control 

2,000 ppm 

4,000 ppm 

0 

2,359.99 

4,869.85 

0 

 727.8 

1,670 

0 

 851 

1,811 

Lungb 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/m3; 4,000 ppm = 13,894 mg/m3. 

b

Internal dose expressed as mg dichloromethane metabolized by the GST pathway per liter tissue per day.

1807 

1808 

1809 Marino et al. (2006) noted that inclusion of extrahepatic CYP metabolism in the slowly 

1810 perfused tissue compartment in the mouse model had little impact on the formation of GST 

1811 metabolites in the liver and lung, especially at exposure levels used in the mouse NTP (1986) 

1812 bioassay. To support this contention, the Andersen et al. (1987) model was modified to include 

1813 10% of the liver rate of oxidative metabolism in the slowly perfused tissue compartment (as 

1814 suggested by Sweeney et al. [2004]), and the modified model was used to calculate the formation 

1815 of GST metabolites. If extrahepatic metabolism was included in the slowly perfused tissue 

1816 compartment, there was a 5–6% reduction in the formation of GST metabolites in the lung and 

1817 liver at an exposure level of 50 ppm. At 2,000 or 4,000 ppm, however, there was only a 0.77 or 

1818 0.37% reduction, respectively. Marino et al. (2006) did not discuss the impact of including 
31

1819 extrahepatic metabolism in the rapidly perfused tissue compartment; the same group of 

1820 investigators developed a human PBTK model that included CYP metabolism in the richly 

1821 perfused compartment (David et al., 2006). 

1822 

1823 3.5.2. Probabilistic Human PBTK Dichloromethane Model (David et al., 2006) 

1824 The basic model structure used by David et al. (2006) was that of Andersen et al. (1987) 

1825 with the addition of the CO submodel of Andersen et al. (1991), refinements from the Marino et 

1826 al. (2006) mouse model, and an inclusion of CYP metabolism in richly perfused tissue (Figure 3-

1827 4). David et al. (2006) used Bayesian analysis to develop and calibrate metabolic parameters in a 

1828 human probabilistic PBTK model for dichloromethane, using kinetic data from several studies of 

1829 volunteers exposed to dichloromethane (n = 13 from DiVincenzo and Kaplan [1981]; n = 12 

1830 from Engström and Bjurström [1977]; n = 14 from Astrand et al. [1975]; n = 3 from Stewart et 

1831 al. [1972a], and group means for metabolism parameters from Andersen et al. [1991]). Exhaled 

1832 dichloromethane and CO and blood levels of dichloromethane and COHb were available in the 

1833 studies by Andersen et al. (1991) and DiVincenzo and Kaplan (1981). The other three studies 

1834 included two or three of these measures. The only available data for levels of dichloromethane 

1835 in fat came from the study of Engström and Bjurström (1977) (described in section 3.2 within 

1836 adipose tissue). 

1837 

1838 

1839 

1840 

1841 

1842 

1843 

1844 

1845 

1846 

1847 

1848 

1849 Figure 3-4. Schematic of human PBTK, used by David et al. (2006). 

1850 

1851 

1852 Values (means and SDs or CVs) for the model parameter distributions were selected from 

1853 multiple sources considered to provide the most current scientific evidence for each parameter 

1854 (David et al., 2006). Mean values for cardiac output (QCC), VPR, and all fractional tissue 

1855 volumes and blood flow rates were based on mean values used by EPA (U.S. EPA, 2000d) in a 

1856 PBTK model for vinyl chloride, as were values for CVs for all physiological parameters, except 

1857 CVs for VPR and fractional lung volume, which were set to those used by OSHA (1997). Means 

Blood

Alveolar 

Air

Endogenous 

Production

Gas 

Exchange Lung

Fat

Richly 

Perfused

Slowly

Perfused

Liver

GST CYP

GST CYP

CO Sub 

Model

CYP
32

1858 for the CO submodel parameters were set equal to those in Andersen et al. (1991), except for 

1859 those for the endogenous rate of CO production (REnCOC) and the background amount of CO 

1860 (ABCOC), which were based on data collected by DiVincenzo and Kaplan (1981). Means for 

1861 partition coefficients, the ratio of lung Vmax to liver Vmax, and the ratio of lung KF to liver KF

1862 (A2) were those used by Andersen et al. (1987), whereas prior means for Vmaxc and Km were 

1863 those used by Andersen et al. (1991). The prior mean for the metabolic parameter for CYP 

1864 metabolism in the rapidly perfused tissue was set at 0.03, slightly lower than the value suggested 

1865 by Sweeney et al. (2004). Prior CVs for the metabolic parameters were set at 200%. 

1866 MCMC analysis was used to calibrate metabolic parameters in the human model in a 

1867 two-step approach: (1) posterior distributions were estimated separately by using data from each 

1868 of the five studies with kinetic data for humans exposed to dichloromethane (with durations 

1869 ranging from 1 to 8 hours and concentrations ranging from 50 to 1,000 ppm); and (2) posterior 

1870 distributions were estimated with combined data from the 42 individual subjects from the four 

1871 studies with individual subject data (DiVincenzo and Kaplan, 1981; Engström and Bjurström, 

1872 1977; Astrand et al., 1975; Stewart et al., 1972a). Results from the Bayesian calibration with the 

1873 combined kinetic data for individual subjects are shown in Table 3-7. This analysis resulted in a 

1874 narrowing of the distribution for the CYP2E1 metabolism parameters Vmax and Km, from a fairly 

1875 broad prior distribution with a CV of 200% for both parameters to 13.1 and 33.6%, respectively, 

1876 for Vmax and Km. 

1877 

Table 3-7. Results of calibrating metabolic parameters in a human 

probabilistic PBTK model for dichloromethane with individual kinetic data 

for 42 exposed volunteers and MCMC analysis 

Prior distributions Posterior distributions 

Parameter 

Mean 

(arithmetic) CV 

Mean 

(arithmetic) CV 

Vmaxc —-maximal CYP metabolic rate (mg/hr/kg0.7) 6.25 2 9.42 0.131 

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

kfC—first-order GST metabolic rate (kg0.3/hr) 2 2 0.852 0.711 

A1—ratio of lung Vmaxc to liver Vmaxc 0.00143 2 0.000993 0.399 

A2—ratio of lung kfC to liver kfC 0.0473 2 0.0102 0.728 

FracR—fraction of Vmaxc in rapidly perfused tissues 0.03 2 0.0193 0.786 

Source: David et al. (2006). 

1878 

1879 

1880 The parameter statistics shown in Table 3-7 (values reported by David et al., 2006) are 

1881 summary statistics of the converged parameter chains obtained in that analysis, leaving out any 

1882 evaluation of correlation or covariance among the updated parameters. As such, these statistics 

1883 implicitly include both the inter-individual variability that would have been elucidated by the 
33

1884 Bayesian analysis (variation between mean values for each individual for which data were 

1885 available) and uncertainty in those values. 

1886 David et al. (2006) further refined the human probabilistic model to reflect 

1887 polymorphisms in the GST pathway: homozygous positive (+/+) GST-T1 individuals, 

1888 heterozygous (+/–) GST-T1, and homozygous negative (–/–) GST-T1 individuals with no GST 

1889 activity. Distributions of GST activities for these genotypes in a group of 208 healthy male and 

1890 female subjects from Sweden were scaled to obtain distributions of kfC for each genotype that, 

1891 when weighted by estimated frequencies of the genotypes in the U.S. population, would result in 

1892 an overall population mean equal to the kfC mean for the posterior distribution shown in 

Table 3-7 (0.852 kg0.3/hour). The resultant mean kfC values were 0.676 kg0.3 1893 /hour (SD 0.123) 

for heterozygous individuals and 1.31 kg0.3 1894 /hour (SD 0.167) for homozygous positive 

1895 individuals. The final parameter distributions used by David et al. (2006) are summarized in 

1896 Table 3-8. 

1897 As described in Appendix B, EPA undertook an evaluation of the David et al. (2006) 

1898 model and parameterization, focusing on the adequacy of the characterization of parameter 

1899 distributions in the full human population. EPA’s conclusion is that the reported distributions for 

1900 physiological parameters in particular, but also key metabolic parameters, only represented a 

1901 narrow set of adults (with the exception of BW). The EPA therefore chose to use supplemental 

1902 data sources to define these distributions in a way that should fully characterize the variability in 

1903 the human population for individuals between six months and eighty years of age. The EPA 

1904 incorporated additional data concerning the variability in CYP2E1 activity among humans, based 

1905 on Lipscomb et al. (2003). The Lipscomb et al. (2003) study was based on in vitro analysis of 

1906 liver samples from 75 human tissue donors (activity towards trichloroethylene and measurements 

1907 of protein content) to estimate a distribution of activity in the population. These data support a 

1908 wider distribution in CYP2E1 activity than had been used in the David et al. (2006) model, with 

1909 approximately a sixfold range between the upper and lower bounds in Lipscomb et al. (2003) and 

1910 a twofold range in David et al. (2006). Thus the EPA replaced the David et al. distribution 

1911 parameters by using the same GM = 9.34, but GSD = 1.73. Further, since even the data available 

1912 to Lipscomb et al. (2003) were limited, and the log-normal distribution is naturally bounded to 

1913 be greater than zero, the EPA chose to use a non-truncated distribution for this parameter. (Since 

1914 the distribution form for CYP2E1 was set by David et al. (2006) to be log-normal and the U.S. 

1915 EPA chose to retain that form, even without specific bounds the distribution only includes values 

1916 greater than zero.) Finally, the scaling of CYP2E1 for individuals under the age of 18 was 

1917 adjusted based on the data of Johsrud et al. (2003); the EPA’s analysis of these data indicate 

1918 CYP2E1 activity in children is better predicted when assumed to scale with body weight (BW) 

1919 raised to the 0.88 power, as compared to the more general power of 0.74, used by David et al. 

CYP2E1 activity for individuals over the age of 18 is still assumed to scale as BW0.74 1920 .

1921 
34

Table 3-8. Parameter distributions used in human Monte Carlo analysis for 

dichloromethane by David et al. (2006)

Distribution 

Parameter 

Mean 

(arithmetic) SD Source 

BW Body weight (kg) 70.0 21.0 Humansa

QCC Cardiac output (L/hr/kg0.74) 16.5 1.49 Humansa

VPR Ventilation:perfusion ratio 1.45 0.203 Humansa

QFC Fat 0.05 0.0150 Humansa

QLC Liver 0.26 0.0910 Humansa

QRC Rapidly perfused tissues 0.50 0.10 Humansa

QSC Slow perfused tissues 0.19 0.0285 Humansa

Tissue volumes (fraction BW) 

VFC Fat 0.19 0.0570 Humansa

VLC Liver 0.026 0.00130 Humansa

VLuC Lung 0.0115 0.00161 Humansa

VRC Rapidly perfused tissues 0.064 0.00640 Humansa

VSC Slowly perfused tissues (muscle) 0.63 0.189 Humansa

Partition coefficients 

PB Blood:air 9.7 0.970 Humansb

PF Fat:blood 12.4 3.72 Ratsb

PL Liver:blood 1.46 0.292 Ratsb

PLu Lung:arterial blood 1.46 0.292 Ratsb

PR Rapidly perfused tissue:blood 1.46 0.292 Ratsb

PS Slowly perfused tissue (muscle:blood) 0.82 0.164 Ratsb

Metabolism parameters 

Vmaxc Maximum metabolism rate (mg/hr/kg0.7) 9.42 1.23 Calibrationc

Km Affinity (mg/L) 0.433 0.146 Calibrationc

A1 Ratio of lung VMax to liver Vmax 0.000993 0.000396 Calibrationc

A2 Ratio of lung KF to liver KF 0.0102 0.00739 Calibrationc

FracR Fractional CYP2E1 capacity in rapidly perfused tissue 0.0193 0.0152 Calibrationc

First order metabolism rate (/hr/kg0.3) 

Homozygous (–/–) 0 0 Calibrationc

Heterozygous (+/–) 0.676 0.123 Calibrationc

kfC 

Homozygous (+/+) 1.31 0.167 Calibrationc

aUS EPA, 2000d. Human PBTK model used for vinyl chloride. 

bAndersen et al. (1987). Blood:air partition measured using human samples; other partition coefficients based on 

estimates from tissue measures in rats. 

cBayesian calibration based on five data sets (see text for description); posterior distributions presented in this table.

Source: David et al. (2006). 

1922 

1923 In addition, while the BW distribution in the David et al. (2006) PBTK model used 

1924 ranges from 7 to 130 kg, thus covering 6-month-old children to obese adults, there are age1925 dependent changes and gender-dependent differences in ventilation rates and body fat that are 

1926 not explicitly included. To more accurately reflect the distribution of physiological parameters 

1927 in the entire population, EPA replaced the unstructured distributions of David et al. (2006) with 

1928 distributions based on available information that specifically account for population variability in 

1929 age, gender, and age- and gender-specific distributions or functions for BW, QCC, alveolar 
35

1930 ventilation, body fat (fraction), and liver fraction (see Appendix B for more details of the 

1931 evaluation of each of these parameters). 

1932 The resulting set of parameter distribution characteristics, including those used as defined 

1933 by David et al. (2006) are described in Table 3-9. Using this revised set of distributions, 

1934 including the (revised) CYP and (published) GST activity distributions, and other distributions 

1935 used as defined by David et al. (2006), the model as applied should reflect the full variability in 

1936 the (U.S.) human population. 

1937 
36

1938 

Table 3-9. Parameter distributions for the human PBTK model for dichloromethane used by EPA 

Distribution 

Parameter Shape 

(Geometric) 

mean a SD/GSD a

Lower 

bound 

Upper bound S

BW Body weight (kg) Normal f (age, gender) 1st %tile 99th %tile B-

Flow rates 

QAlvC Alveolar ventilation (L/hour/kg0.75) Normal f (age, gender) f (age) 5th %tile 95th %tile B-4.4; me

SD: Arcus-

vprv Variability in ventilation:perfusion ratio Log-normal 1.00 0.203 0.69 1.42 VPR/VPR

QCC Cardiac output (L/hour/kg0.75) QCCmean = f(QAlvC) QCC = QCCmean/vprv B-4.5; Cle

Fractional flow rates (fraction of QCC) 

QFC Fat Normal 0.05 0.0150 0.0050 0.0950 

QLC Liver Normal 0.26 0.0910 0.010 0.533 

QRC Rapidly perfused tissues Normal 0.50 0.10 0.20 0.80 

QSC Slow perfused tissues Normal 0.19 0.0285 0.105 0.276 

David et al. 

these di

Tissue volumes (fraction BW) 

VFC Fat Normal f (age, gender) 0.3·mean 0.1·mean 1.9·mean

VLC Liver Normal f (age) 0.05·mean 0.85·mean 1.15·mean

VLuC Lung Normal 0.0115 0.00161 0.00667 0.0163 

VRC Rapidly perfused tissues Normal 0.064 0.00640 0.0448 0.0832 

VSC Slowly perfused tissues Normal 0.63 0.189 0.431 0.829 

Fat mean: B-4

mean: B-4

otherwise, 

sampling 

=

Vi 

Partition coefficients 

PB Blood:air Log-normal 9.7 1.1 7.16 13.0 

PF Fat:blood Log-normal 11.9 1.34 4.92 28.7 

PL, PLu, 

& PR 

Liver:blood, lung:arterial blood, and 

rapidly perfused tissue:blood Log-normal 1.43 1.22 0.790 2.59 

PS Slowly perfused tissue (muscle):blood Log-normal 0.80 1.22 0.444 1.46 

Geometric me

converted fro

values 

Metabolism parameters (based on Monte Carte calibration from five human data sets) 

Vmaxc Maximum metabolism rate (mg/hr/kgXvmax) Lognormal 9.34 1.73 (none) (none)

B-3 mea

GSD & boun

Xvmax 

Xvmax 

Km Affinity (mg/L) Log-normal 0.41 1.39 0.154 1.10 Geometric me
37

Table 3-9. Parameter distributions for the human PBTK model for dichloromethane used by EPA 

Distribution 

Parameter Shape 

(Geometric) 

mean a SD/GSD a

Lower 

bound 

Upper bound S

A1 Ratio of lung VMax to liver VMax Log-normal 0.00092 1.47 0.000291 0.00292 A2 Ratio of lung KF to liver KF Log-normal 0.0083 1.92 0.00116 0.0580 

FracR Fractional MFO capacity in rapidly perfused tissue Log-normal 0.0152 2.0 0.00190 0.122 

converted fro

values 

First order metabolism rate ([hour/kg0.3]–1) 

Homozygous (–/–) Normal 0 0 0 0 

kfC Heterozygous (+/–) Normal 0.676 0.123 0.00 1.291

Homozygous (+/+) Normal 1.31 0.167 0.00 2.145

D

aArithmetic mean and SD listed for normal distributions; geometric mean and geometric SD (GSD) listed for log-normal distributions.

1939 
38

1940 3.5.3. Evaluation of Rat PBTK Dichloromethane Models 

1941 Several deterministic PBTK rat models have been reported in the scientific literature 

1942 (Sweeney et al., 2004; Andersen et al., 1991, 1987; Reitz , 1991; Reitz et al.,1988a, b; U.S. EPA 

1943 1988b, 1987a, b; Gargas et al., 1986). Unlike the mouse (Marino et al., 2006) and human (David 

1944 et al., 2006), no hierarchical population model for dichloromethane in the rat exists in which 

1945 parameter uncertainty is quantitatively integrated into model calibration. Rat data are not 

1946 available that would allow for Bayesian calibration of individual metabolic parameters for the 

1947 CYP or GST pathways. Thus, EPA assessed modified versions of deterministic rat PBTK 

1948 models to select the most appropriate model for use in extrapolating internal dosimetry from rats 

1949 to humans, for example in the determination of RfDs and RfCs based on effects seen in the rat. 

1950 This work is described in detail in Appendix C and is based on evaluation of blood levels of 

1951 dichloromethane and the percent saturation of hemoglobin as COHb (%COHb) and expired 

1952 dichloromethane following intravenous injection (Angelo et al., 1986b), closed chamber gas 

1953 uptake (Gargas et al., 1986), and dichloromethane and %COHb blood levels from a 4-hour 

1954 inhalation exposure (Andersen et. al., 1991, 1987). Based on this work, the basic model 

1955 structure of Andersen et al. (1991) was chosen, with the inclusion of lung dichloromethane 

1956 metabolism via CYP (4% of liver metabolite production) and GST (14% of liver metabolite 

1957 production) pathways (estimated from Reitz et al., 1989) (Figure 3-5) with metabolic parameters 

1958 recalibrated against data of Andersen et al. (1991), based on prediction agreement of the various 

1959 parameters with the available rat data sets. Table 3-10 presents the parameter distribution data 

1960 for this model. 

1961 

1962 Figure 3-5. Schematic of rat PBTK model used in current assessment. 

Blood

Alveolar 

Air

Endogenous 

Production

Gas 

Exchange

Lung

Fat

Richly 

Perfused

Slowly 

Perfused

Liver

GST CYP

GST CYP

CO Sub 

Model
39

Table 3-10. Parameter values for the rat PBTK model for 

dichloromethane used by EPA 

Parameter Mean 

Flow rates 

QCC (L/hour/kg0.74) 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 BW0.99) 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/kg0.7) 3.93

Affinity (mg/L) 0.524

Ratio of lung VMax to liver VMax 0.04

Ratio of lung KF to liver KF 0.14

1st order metabolism rate (liver KF) ([hour/kg0.3]

–1) 2.46

Oral absorption constant, ka (1/hr) 1.80 

1963 

1964 3.5.4. Comparison of Mouse, Rat and Human PBTK Models 

1965 The comparison of various parameters across species (Table 3-11) primarily shows the 

1966 modest inter-species differences that are known to occur in physiological parameters, also 

1967 including the approximately 2-fold differences in partition coefficients which occur because of 

1968 differences in rodent versus human blood lipid content. The 2.5-fold lower Vmaxc (CYP activity) 

1969 in rats versus mice is also typical. The most striking difference is the variation in A1 and A2. 

1970 Those values, however, reflect the in vitro differences originally quantified by Lorenz et al. 

1971 (1984) and used in the dichloromethane PBPK modeling of Anderson et al. (1987). Thus these 

1972 differences are based on independent measurements of tissue-specific metabolic capacity, and 
40

1973 while the specific values for mouse and human were refined through Bayesian analysis, the 

1974 ultimate (posterior) values used are within a reasonable range of the in vitro measurements and 

1975 so do not appear to be artifactual. (Since in vivo kinetics often indicate some differences from 

1976 what would be predicted without adjustment from in vitro, it is not surprising that such 

1977 differences occur here.) These differences do explain why lung-specific metrics in particular 

1978 lead to lower internal dose and hence risk predictions in humans compared to whole-body 

1979 metrics. 
41

Table 3-11. Parameters in the mouse, rat, and human PBTK model for dichloromethane used by the E

Mouse a Rat b Human c

Parameter Mean Value Mean CV/GSD (Shape, bounds) Fractional flow rates (fraction of cardiac output) b

QFC Fat QLC Liver QRC Rapidly perfused tissues QSC Slowly perfused tissues 

0.05 0.24 0.52 0.19 

0.09 0.20 0.56 0.15 

0.05 0.26 0.50 0.19 

0.3 (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) 

Fractional tissue volumes (fraction of body weight) bVFC Fat VLC Liver VLuC Lung VRC Rapidly perfused tissues VSC Slowly perfused tissues 

0.04 0.04 

0.0115 

0.05 0.78 

0.07 0.04 

0.0115 

0.05 0.75 

f (age, gender) 

f (age) 

0.0115 

0.064 

0.63 

0.3 (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) 

Partition coefficientsc

PB Blood/air PF Fat/blood PL Liver/blood PLu Lung/blood PR Rapidly perfused/blood PS Slowly perfused/blood 

23.0 

5.1 1.6 

0.46 0.52 0.44 

19.4 6.19 0.73 0.46 0.73 0.41 

9.7 

11.9 1.43 1.43 1.43 0.80 

1.1 (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) 

Flow rates QCC Cardiac output (L/hr/kg0.74) 

VPR ventilation/perfusion ratio 

QAlvC

24.2 1.45 

QCC/VPR 

14.99 

0.94 

QCC/VPR

QCCmean = f(QAlvC)

(variable) 

f (age, gender)

QCC = QCCmean/vprv 

(varies) (LN, 0.69-1.42) 

f (age) (N, 5th-95th %) 

Metabolism parameters Vmaxc Maximum CYP metabolic rate (mg/hr/kgXvmax) 

Xvmax CYP allometric scaling power Km CYP affinity (mg/L) 

kfC First-order GST metabolic rate constant (kg0.3/hr) 

A1 Ratio of lung Vmaxc to liver Vmaxc 

A2 Ratio of lung kfC to liver kfC 

9.27 

0.7 

0.574 

1.41 

0.207 0.196 

3.93 

0.7 

0.524 

2.46 0.04 0.14

9.34 

0.88 for age <18; 

0.7 for age 

0.41 

0 (-/-)e

0.676 (+/-)e

1.31 (+/+)e

0.00092 

0.0083

1.73 (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) 

a Based on Marino et al. (2006) (source for all mouse parameters) b Based on Andersen et al. (1991), with the addition of lung metabolism of dichloromethane via the CYP (4% of liver metabolite production
42

metabolite production) pathways. Physiological parameters and partition coefficients are from Andersen et al. (1991). The values for dichl

in the lung (as a fractional yield of liver metabolism for each pathway) were estimated from the in vitro ratios of enzyme activity (nmol/minliver cytosolic (GST) and microsomal (CYP) tissue fractions (Reitz et al., 1989). Metabolic parameters were re-optimized against the inhalaal. (1991) using a heteroscedasticity parameter value of 2, which uses relative error for the model fitting algorithm. See Appendix C for fur

c Based on David et al. (2006), with changes as noted. Additional sources include Clewell et al. (2004), Arcus-Arth and Blaisdell (2007), aSee identified sections for details. Distribution values (mean and a measure of dispersion) are provided with the CV (mean/SD) presented foand the GSD (italicized) presented for log-normal (LN) distributions. Distributions were truncated, bounds are (upper-lower bound)/mean. e Values for the homozygous (–/–), heterozygous (+/–), and homozygous (+/+) GST-T1 genotypes, respectively. 
43

1980 HAZARD IDENTIFICATION 

1981 

1982 

1983 4.1. STUDIES IN HUMANS 

1984 4.1.1. Introduction—Case Reports, Epidemiologic, and Clinical Studies 

1985 There has been considerable interest in the influence of occupational exposure to 

1986 dichloromethane in relation to a variety of conditions. The recognition that dichloromethane can 

1987 be metabolized and bound to hemoglobin to form COHb, resulting in a reduction in the oxygen 

1988 carrying capacity of the blood (Stewart et al., 1972b), prompted investigations into risk of 

1989 ischemic heart disease and other cardiovascular effects. Reports of neurological effects from 

1990 acute, high-exposure situations contributed to concern about neurological effects of chronic 

1991 exposure to lower levels of dichloromethane. A general interest in potential cancer risk became 

1992 more focused on lung and liver cancer because of the observation of these specific tumors in the 

1993 NTP (1986) experiments in mice. Details of the studies pertaining to the experimental and 

1994 epidemiologic studies of noncancer outcomes (e.g., cardiac, neurologic, hepatic, reproductive) 

1995 are presented in section 4.1.2, and studies of cancer risk are presented in section 4.1.3. 

1996 

1997 4.1.2. Noncancer Studies 

1998 4.1.2.1. Case Reports of Acute, High-dose Exposures 

1999 Numerous case reports have been published that describe health effects resulting from 

2000 acute exposure to dichloromethane. Most of the reports describe health effects resulting from 

2001 inhalation of dichloromethane or dermal contact, but a few involve ingestion. The COHb levels 

2002 in some of these cases were relatively low (7.5–13%), so the initial toxic effects of acute 

2003 dichloromethane exposure appear to be due to its anesthetic properties as opposed to metabolic 

2004 conversion of dichloromethane to CO. 

2005 Bakinson and Jones (1985) reported on a series of 33 cases of acute inhalation exposures 

2006 to dichloromethane that occurred in the workplace over the period 1961–1980. Thirteen had lost 

2007 consciousness, and one of the workers died. Nineteen cases reported general neurological 

2008 effects, 13 reported gastrointestinal symptoms, 4 reported respiratory symptoms, and 1 reported 

2009 hepatic symptoms. Of the 19 with general neurological symptoms, all reported headache, and 

2010 dizziness was reported by 11 workers. Five workers reported one of the following symptoms: 

2011 drunkenness, confusion, lack of coordination, or paresthesia. 

2012 Rioux and Myers (1988) summarized the health effects reported for 26 cases of 

2013 dichloromethane poisoning published in the literature between 1936 and 1986. Three cases 

2014 resulted from abuse-related exposures, 2 from chronic exposures, and 21 from acute exposures. 

2015 The most common effects involved the central nervous system (CNS) (unconsciousness, 

2016 drowsiness, headache, and behavioral symptoms), pulmonary edema and dyspnea, and 

2017 dermatologic symptoms. Even severe symptoms could be reversed, but four deaths occurred. 
44

2018 More than 10 other case reports of fatalities or poisonings have been published since the 

2019 summaries by Rioux and Myers (1988) and Bakinson and Jones (1985), and many of these 

2020 incidents involve inadequately ventilated occupational settings (Jacubovich et al., 2005; Raphael 

2021 et al., 2002; Fechner et al., 2001; Zarrabeitia et al., 2001; Goulle et al., 1999; Mahmud and 

2022 Kales, 1999; Kim et al., 1996; Tay et al., 1995; Manno et al., 1992; Leikin et al., 1990; 

2023 Shusterman et al., 1990). CNS depression and resulting narcosis, respiratory failure, and heart 

2024 failure are common features of these reports. In a survey of workers in furniture stripping shops, 

2025 10 of the 21 workers stated that they sometimes experienced dizziness, nausea, or headache 

2026 during furniture stripping operations (Hall and Rumack, 1990). 

2027 Chang et al. (1999) reported details of six patients who had ingested dichloromethane 

2028 (four in a suicide attempt and two from accidental ingestion during a state of intoxication). The 

2029 estimated amounts ingested were 350 mL or less. COHb levels, which were measured in only 

2030 two of the cases, were 8.4 and 35% (with the latter being seen in a fatal case). As in exposures 

2031 resulting from inhalation, the most common symptoms involved CNS depression, ranging from 

2032 somnolence and weakness to deep coma. Tachypnea (n = 6) and corrosive gastrointestinal tract 

2033 injury (n = 3) were also reported. Hepatic and renal failure and pancreatitis were found in the 

2034 two most severe cases. 

2035 

2036 4.1.2.2. Controlled Experiments Examining Acute Effects 

2037 Several controlled experiments were conducted in the 1970s, examining 

2038 neurophysiological effects and levels of COHb resulting from short-term (1–4 hours) exposures 

2039 to dichloromethane at levels up to 1,000 ppm or longer-term exposures at levels up to 500 ppm. 

2040 The 8-hour threshold limit value before 1975 was 500 ppm (National Institute of Occupational 

2041 Safety and Health [NIOSH], 1986). These studies are described below. With the exception of 

2042 Putz et al. (1979), there is no description in the published reports of the informed consent and 

2043 other human subjects research ethics procedures undertaken in these studies, but there is no 

2044 evidence that the conduct of the research was fundamentally unethical or significantly deficient 

2045 relative to the ethical standards prevailing at the time the research was conducted. 

2046 In 1972, Stewart et al. (1972a, b) reported results from four experiments that were 

2047 initiated because of the chance observation of an elevation in COHb saturation levels in an 

2048 individual (one of the investigators) the morning after he had spent 2 hours working with varnish 

2049 remover. Participants were medical students and faculty (including at least one of the 

2050 coauthors). A total of 11 healthy nonsmoking volunteers were placed in an exposure chamber 

2051 with mean concentrations of dichloromethane ranging from 213 to 986 ppm for 1 or 2 hours. 

2052 These experiments indicated that dichloromethane exposure at these levels resulted in COHb 

2053 saturation levels that exceeded and were more prolonged than those seen with threshold limit 

2054 value exposures to CO. The exposures also resulted in symptoms of CNS depression indicated 

2055 by visual evoked response changes and reports of light-headedness. Although return of COHb 
45

2056 levels to background levels could take >24 hours, all of the other symptoms were reversible 

2057 within a few hours after exposure ceased. 

2058 Winneke (1974) measured auditory vigilance, visual flicker fusion frequency, and 

2059 14 psychomotor tasks in a total of 38 women exposed to dichloromethane levels of 300–800 ppm 

2060 for 4 hours in an exposure chamber. A comparison group (nine females, nine males) exposed to 

2061 100 ppm CO for 5 hours was also included. Exposure to 800 ppm dichloromethane resulted in a 

2062 statistically significant decrease in the performance of 10 of the 14 psychomotor tasks. In tests 

2063 of auditory vigilance and visual flicker fusion, depressed response was seen at 300 ppm and was 

2064 further depressed at 800 ppm. These effects were not seen with CO exposure. 

2065 Forster et al. (1974) exposed four healthy young men to dichloromethane levels ranging 

2066 from 0 to 500 ppm for 7.5 hours/day for a total of 26 days over a 6-week period to investigate 

2067 alterations in hemoglobin affinity for oxygen and altered pulmonary function. While no changes 

2068 were observed in pulmonary function, hemoglobin affinity for oxygen was increased with no 

2069 indication of adaptation to restore this affinity for oxygen to normal. 

2070 Putz et al. (1979) examined the behavioral effects seen after exposure to dichloromethane 

2071 and to CO. Twelve healthy volunteers (six men and six women) each acted as his/her own 

2072 control in separate 4-hour exposures to 70 ppm CO and 200 ppm dichloromethane. These levels 

2073 were chosen so that the COHb level would reach 5% from both the CO and dichloromethane 

2074 exposures. The experiments were conducted in a double-blind manner so that neither the 

2075 investigators nor the participant knew the exposure condition under study at any particular time. 

2076 Informed consent was obtained, and the study was reviewed by the NIOSH Human Subject 

2077 Review Board. The performance tests were dual tasks (an eye-hand coordination task in 

2078 conjunction with a tracking task), with five measures of performance assessed at six time points 

2079 over the 4-hour test period and an auditory vigilance task. Two levels of difficulty were assessed 

2080 for each task to allow assessment of whether the exposure effect was similar in low and high 

2081 difficulty tasks. The tests of eye-hand coordination, tracking tasks, and auditory vigilance 

2082 revealed significant impairment with both exposures under the more difficult task conditions. 

2083 Effects were similar or stronger in magnitude for dichloromethane compared with CO. 

2084 

2085 4.1.2.3. Observational Studies Focusing on Clinical Chemistries, Clinical Examinations, and 

2086 Symptoms 

2087 Studies in currently exposed workers 

2088 Ott et al. (1983a, c, d) evaluated several parameters of hepatic, hematopoietic, and cardiac 

2089 function in workers exposure to dichloromethane in a triacetate fiber production plant in Rock 

2090 Hill, South Carolina. Two hundred sixty-six Rock Hill workers and a comparison group of 

2091 251 workers in an acetate fiber production plant in Narrows, Virginia, were included in the 

2092 examination of urinary and blood measures. These groups included men and women, blacks and 

2093 whites, and smokers and nonsmokers. The median 8-hour TWA exposure for dichloromethane 
46

2094 ranged from 60 to 475 ppm in Rock Hill. Acetone at levels up to over 1,000 ppm was present in 

2095 both plants, but dichloromethane and acetone exposures were inversely related. 

2096 There were differences in blood collection procedures between the two plants and in the 

2097 age, sex, race, and smoking history distribution of the study groups. The demographic and 

2098 smoking differences were accounted for in the analysis by stratification. Statistically significant 

2099 differences were seen between the workers in the two plants for COHb, serum alanine 

2100 aminotransferase (ALT), total bilirubin, and mean corpuscular hemoglobin concentration 

2101 (MCHC) (although the direction and magnitude of these differences were not reported and the 

2102 authors stated that the difference in serum ALT could be due to the differences in blood 

2103 collection procedures, which involved a sitting versus recumbent position of the subjects at the 

2104 exposed and nonexposed plants, respectively) (Ott et al., 1983c). Within the Rock Hill plant, 

2105 analyses were also conducted to examine associations between dichloromethane exposure and 

2106 the clinical parameters within specific race-sex groups by using multiple regression to control for 

2107 smoking status, age, and time of venipuncture. Positive associations were seen with COHb in all 

2108 race-sex groups (increases of 0.7–2.1% per 100 ppm increase in dichloromethane) and with total 

2109 bilirubin (increases of 0.05–0.08 mg/dL per 100 ppm increase in dichloromethane) in all groups 

2110 except nonwhite men (which was a much smaller group, n = 20, than the other groups). Red cell 

2111 count, hematocrit, hemoglobin, and aspartate aminotransferase (AST) were also positively 

2112 associated with dichloromethane exposure in white females. The increase in total bilirubin level 

2113 was not supported by parallel changes in other measures of liver function or red blood cell 

2114 turnover, suggesting that this measure was not reflecting liver damage or hemolysis. 

2115 The increased red cell count, hemoglobin, and hematocrit in women exposed to high 

2116 levels of dichloromethane (up to 475 ppm, 8-hour TWA) may indicate a compensatory 

2117 hematopoietic effect. The fact that these changes were not significant among men may be due to 

2118 higher baseline hemoglobin, which was observed when comparisons were made between 

2119 nonsmoking men and women. No such difference in the baseline values was observed among 

2120 the smoking men and women, suggesting that the compensatory advantage may be lost among 

2121 smokers. 

2122 Ott et al. (1983e) present results from a further investigation of changes in COHb, 

2123 alveolar CO, and oxygen half-saturation pressure in relation to dichloromethane exposure. 

2124 Blood samples were collected before and after shift from 136 Rock Hill and 132 Narrows 

2125 workers. For the Rock Hill workers, personal monitoring for dichloromethane exposure was 

2126 done during the shift. The TWA for dichloromethane ranged from 0–900 ppm, with a bimodal 

2127 distribution (peaks around 150 and 500 ppm) resulting from the layout of the plant. The blood 

2128 samples were used to determine blood COHb and alveolar CO levels, and the partial oxygen 

2129 pressure (P50) (that is, the pressure required to keep 50% of the blood oxygen-carrying capacity 

2130 saturated with oxygen at pH 7.4 and 37°C). Separate analyses were conducted for smokers and 

2131 nonsmokers to account for the smoking-related effects on COHb. Linear relationships were seen 
47

2132 between dichloromethane exposure and the before-shift COHb and alveolar CO levels, reflecting 

2133 residual CO metabolism from the previous day’s exposure. There were significant quadratic 

2134 relationships between dichloromethane exposure and the postshift COHb and alveolar CO levels, 

2135 indicating a partial saturation of the enzyme system metabolizing dichloromethane. The 

2136 P50 group means were lower among the exposed compared with the referents, among smokers 

2137 compared with nonsmokers, and among men compared with women. Given the relationship 

2138 between COHb and P50, an expected decrease in P50 during the shift was observed among the 

2139 exposed. 

2140 Continuous 24-hour cardiac monitoring was also evaluated in a smaller sample of 

2141 24 dichloromethane-exposed workers from the triacetate fiber production plant in Rock Hill, 

2142 South Carolina, and 26 workers from the comparison plant in Narrows, Virginia. This study (Ott 

2143 et al., 1983d) was limited to white men ages 35 or more years. Special efforts were made to 

2144 recruit men with a history of heart disease, because this group was postulated to be most likely to 

2145 demonstrate positive findings. The estimated TWA dichloromethane exposure ranged from 

2146 60 to 475 ppm in the exposed group. The evaluation examined ventricular and supraventricular 

2147 ectopic activity and S-T segment depression in the exposed and nonexposed groups. 

2148 Comparisons were also made between cardiac performance during work hours and nonwork 

2149 hours to discern possible short-term effects of recent exposure. Comparing the findings for the 

2150 24 exposed and 26 referent volunteers indicated no difference in ventricular or supraventricular 

2151 ectopic activity or S-T-segment depression. There was no difference comparing work and 

2152 nonwork hours among exposed. 

2153 Soden et al. (1996) studied all active male workers exposed to dichloromethane at a 

2154 Hoechst Celanese triacetate film production plant in Belgium. The production process was the 

2155 same as the process at the Hoechst Celanese Rock Hill plant, except the Belgium plant was 

2156 newer with better engineering controls to significantly reduce overall levels of the 

2157 dichloromethane, acetone, and methanol used in the process. The objectives of the study were to 

2158 determine the impact of varying levels of dichloromethane exposure on COHb levels, whether 

2159 successive days of dichloromethane exposure affected the COHb levels, and what impact 

2160 smoking had on COHb levels in conjunction with dichloromethane exposure. Workers were 

2161 monitored semiannually for COHb at the end of the work shift and were personally monitored 

2162 for exposure to the three solvents. Smoking status was defined based on a health assessment 

2163 questionnaire, with smokers smoking at least one cigarette per day. Among nonsmokers, a dose 

2164 response was found among COHb levels and average dichloromethane exposure levels in the 

2165 range of 7–90 ppm. The maximum COHb was 4.00% at an average exposure of 90 ppm 

2166 (correlation coefficient = 0.58, p < 0.05). Smokers’ COHb levels were elevated when compared 

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

2168 correlation between dichloromethane air levels and COHb levels was weaker and not statistically 

2169 significant (correlation coefficient 0.20). The maximum COHb level for smokers was 6.35% at 
48

2170 an average dichloromethane air level of 99 ppm. The authors concluded that dichloromethane 

2171 exposures up to the levels observed did not produce COHb levels that are likely to cause cardiac 

2172 symptoms. 

2173 Cherry et al. (1983, 1981) reported the results of health evaluations of two studies of 

2174 triacetate film production workers. Cherry et al. (1981) recruited 46 of the 76 male workers at a 

2175 triacetate film factory, where workers were exposed to dichloromethane and methanol in a ratio 

2176 of 9:1 at air levels of dichloromethane ranging from 75 to 100 ppm. A small comparison group 

2177 (n = 12) of workers at this factory who worked a similar shift pattern (rapidly rotating shifts) but 

2178 who were not exposed to dichloromethane was also included. The men were asked whether they 

2179 had ever experienced cardiac symptoms (pain in the arms, chest pain sitting or lying, or chest 

2180 pain when walking or hurrying) and were asked about the presence, in the past 12 months, of 

2181 neurological disorders (frequent headaches, dizziness, loss of balance, difficulty remembering 

2182 things, numbness and tingling in the hands or feet), affective symptoms (irritability, depression, 

2183 tiredness), and stomachache (as an indicator of symptom overreporting). No difference in 

2184 response was found in history of stomachache (reported by 15% of exposed workers compared 

2185 with 17% nonexposed workers). Six of the exposed and none of the unexposed men responded 

2186 positively to the cardiac symptoms. The exposed group reported an excess of neurological 

2187 symptoms; the number (and proportion) reporting zero, one, two, and three or more symptoms 

2188 were 26 (0.56), 8 (0.17), 9 (0.20), and 3 (0.07), respectively, in exposed workers compared with 

2189 11 (0.92), 1 (0.12), 0 (0.00), and 0 (0.00), respectively, in controls (p < 0.02 for chi-square test of 

2190 linear trend). With respect to affective symptoms, the number (and proportion) reporting zero, 

2191 one, two, and three symptoms were 28 (0.61), 6 (0.13), 7 (0.15), and 5 (0.11), respectively, 

2192 among the exposed workers, and 9 (0.75), 2 (0.17), 1 (0.08), and 0 (0.0), respectively, among the 

2193 unexposed workers. The authors concluded that there was no difference between exposed and 

2194 nonexposed in reporting of affective symptoms based on a chi-square test of linear trend. There 

2195 was no discussion of the statistical power of this test or of tests of the proportion reporting a 

2196 specified number of symptoms (which may be a more appropriate test given the sample size), but 

2197 it is clear that the statistical power of this test was very low. For example, taking the simple 

2198 case of the comparison of the proportion reporting two or more symptoms and using the 

2199 approximate estimates from this study (25 and 10% in the exposed and unexposed, respectively), 

2200 approximately 75 exposed and 300 unexposed workers would be needed for a power of 0.80 

2201 (i.e., an 80% chance of rejecting the null hypothesis when the null hypothesis was false); the 

2202 actual power with the sample size of 46 and 12 is less that 0.10. 

2203 Based on these results, a follow-up study was conducted, which included a larger referent 

2204 group. This study included the symptom list described in the previous paragraph, a standardized 

2205 clinical exam (including an electrocardiograph), and neurological and psychological tests of 

2206 nerve conduction, motor speed and accuracy, intelligence, reading, and memory (Cherry et al., 

2207 1981). Twenty-nine of the original 46 exposed workers participated in the follow-up. The men 
49

2208 who did not participate in the follow-up were similar in age and symptoms to the men who did. 

2209 The new referent group was recruited from another plant with the same owner and a very similar 

2210 process but without dichloromethane exposure. One control, age-matched within 3 years, was 

2211 selected for each exposed worker. No differences between the groups were found in the clinical 

2212 exam, electrocardiogram, or nerve conduction tests. A statistically significant (p < 0.05) deficit 

2213 among the exposed workers was found for coarse motor speed. On two tests of overall 

2214 intelligence, the exposed group did significantly better than the referent, but, on a reading ability 

2215 test designed to assess premorbid educational level, scores for the exposed group were slightly 

2216 lower than for the referent group. (Only one of these three differences, the trail making 

2217 intelligence test, was statistically significant.) With respect to the report of neurological 

2218 symptoms in the past year, the number (and proportion) reporting zero, one, two, and three 

2219 symptoms were 17 (0.59), 4 (0.14), 6 (0.21), and 2 (0.07), respectively, among the exposed 

2220 workers, and 21 (0.72), 6 (0.21), 0 (0.0), and 2 (0.07), respectively, among the unexposed 

2221 workers, with a test of linear trend that was not statistically significant. The authors interpret the 

2222 results as indicating that the differences in neurological symptoms seen in the initial study were 

2223 due to chance and that, taken as a whole, the exposed workers had no detrimental effect 

2224 attributable to dichloromethane exposure. Again, the limitations of the statistical power of the 

2225 analysis and alternative interpretations that might have resulted from approaches taken to 

2226 improve the power were not discussed. These approaches include combining the unexposed 

2227 groups from the two analyses, using the full sample of the exposed group instead of the subset of 

2228 29 who completed the clinical exam, or using a different test (i.e., of a proportion rather than a 

2229 linear trend), 

2230 Cherry et al. (1983) compared dichloromethane-exposed workers at an acetate film 

2231 factory to nonexposed workers (from the same plant but from areas without solvent contact or 

2232 from another film production factory in which solvents were not used). The 56 exposed and 

2233 36 unexposed workers were matched to within 3 years of age. Both factories were on rapid 

2234 rotating shifts. Exposure to dichloromethane ranged from 28–173 ppm, using individual air 

2235 sampling pumps. Blood samples were taken to monitor dichloromethane levels at the beginning 

2236 and end of the shift. Study participants were asked to rate sleepiness, physical and mental 

2237 tiredness, and general health on visual analog scales with the extreme responses at either end. 

2238 Participants were also given a digit symbol substitution test and a test of simple reaction time. 

2239 No differences were seen between exposed and unexposed groups at the beginning of the shift on 

2240 the four visual analogue scales, but the exposed deteriorated more on each of the scales than did 

2241 the controls. This difference in deterioration was statistically significant (p ≤ 0.05) during the 

2242 morning shift but was not statistically significant during the afternoon or night shifts. A 

2243 significant correlation was shown between change in mood over the course of the shift and level 

2244 of dichloromethane in the blood. No difference was seen between the exposed and referents on 

2245 the tests of reaction time or digit substitution. However, among the exposed, deterioration in the 
50

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

2247 levels. 

2248 Anundi et al. (1993) studied 12 men who worked in a graffiti-removing company. Each 

2249 worker filled out a questionnaire about previous occupational and non occupational exposure to 

2250 solvents and use of protective equipment. Half-day breathing zone samples were taken for each 

2251 of the 12 workers, and 15-minute samples were also taken for 10 workers. On the day the air 

2252 sampling was done, a structured interview pertaining to recent diseases or symptoms related to 

2253 allergies, asthma, diseases of the skin, respiratory organs, gastrointestinal tract, urinary organs, 

2254 neurological trauma and disease, and neuropsychiatric symptoms was conducted by a physician, 

2255 and blood and urine samples were collected. The results were compared with those of 233 men 

2256 from the area population. The 12 men (mean age 23 years) had worked between 3 months and 

2257 4.5 years cleaning graffiti from underground stations. No respiratory protection was used, and 

2258 the leather gloves were frequently soaked with solvent. While mixed solvent was used to do the 

2259 cleaning, dichloromethane was the predominant component, as confirmed by the air samples. 

The geometric mean of the TWA calculated from the half-day samples was 127 mg/m3

2260 (range 

18–1,188 mg/m3

2261 ), with half of the samples exceeding the Swedish permissible exposure limit of 

120 mg/m3. The geometric mean of the 15-minute samples was 400 mg/m32262 (range 6–

5,315 mg/m32263 ), with most samples exceeding the Swedish short-time exposure limit of 

300 mg/m32264 . Two workers had clinical laboratory data outside the normal range (urinary α1- or 

2265 β2-microglobulin, serum ALT, γ-glutamyl transpeptidase), which could indicate possible kidney 

2266 and liver damage. The authors stated that in both cases factors other than the solvent exposure 

2267 (i.e., urinary tract medical condition preceding employment, history of renal stones) could have 

2268 influenced these laboratory results. The prevalence of irritation of the eyes and upper respiratory 

2269 tract (blocked nose and nasal catarrh) was much higher in the graffiti-cleaning workers compared 

2270 with the referent group (e.g., >70% of the workers compared with 18% of the comparison group 

2271 reported a blocked nose; ~50% of workers and 15% of the comparison group reported eye 

2272 irritation), but there were no or much smaller differences in abnormal tiredness, headache, 

2273 nausea, or irritative cough. No acute effects on the CNS were noted. 

2274 

2275 Studies in retired workers 

2276 Lash et al. (1991) examined the hypothesis that long-term exposure to dichloromethane 

2277 produces lasting CNS effects as measured by long-term impairment on memory and attention 

2278 centers. Retired aircraft maintenance workers employed in at least 1 of 14 targeted jobs with 

2279 dichloromethane exposure for 6 or more years between 1970 and 1984 were compared to a like 

2280 group of workers without dichloromethane exposure. The unexposed workers were also retired 

2281 aircraft mechanics at the same base and held one of 10 jobs in the jet shop where little solvent 

2282 was used. The exposed group made up of painters and mechanics in the overhaul department 

2283 was chosen to maximize the exposure contrast yet minimize differences in potential confounders 
51

2284 between exposed and nonexposed. Exposures were typically within state and federal guidelines 

2285 for dichloromethane exposure. From 1974 to 1986, when 155 measurements for 

2286 dichloromethane exposure were made, mean breathing zone TWAs ranged from 82 to 236 ppm 

2287 and averaged 225 ppm for painters and 100 ppm for mechanics. 

2288 Data collection occurred in three phases, with an initial questionnaire to all retired 

2289 members of the airline mechanics union to identify eligible workers, followed by a telephone 

2290 survey to collect medical, demographic, and general employment criteria. Subjects who 

2291 qualified were then recruited to participate in the medical evaluation. Sixty percent of the 

2292 1,758 retirees responded to the questionnaire and 259 of these retirees met the eligibility criteria. 

2293 Ninety-one men qualified for the medical evaluation based on the telephone survey; 25 retirees 

2294 exposed to solvents, and 21 unexposed retirees participated in the evaluation. All were men 

2295 between the ages of 55 and 75, without a history of alcoholism or any neurological disorder. The 

2296 25 exposed participants worked an average of 11.6 years in dichloromethane-exposed jobs 

2297 during the target period and 23.8 years in the industry. 

2298 The medical evaluation included a questionnaire about the occurrence of 33 different 

2299 symptoms in the past year, physiological measurement of odor and color vision senses, auditory 

2300 response potential, hand grip strength, and measures of reaction time (simple, choice, and 

2301 complex), short-term visual memory and visual retention, attention, and spatial ability. The only 

2302 large differences (i.e., effect size, or mean difference between groups divided by the SD of the 

2303 outcome measure, of 0.4 or greater) between the two groups were a higher score on verbal 

2304 memory tasks (effect size approximately 0.45, p = 0.11) and lower score on attention tasks 

2305 (effect size approximately –0.55, p = 0.08) and complex reaction time (effect size approximately 

2306 –0.40, p = 0.18) in the exposed compared with the control group. (Although not noted by the 

2307 authors, the power to detect a statistically significant difference between the groups, given this 

2308 sample size, was low [i.e., approximately 0.30 for an effect size of 0.40, using a two-tailed alpha 

2309 of 0.05]) (Cohen, 1987). The authors investigated the possibility of response bias, given the low 

2310 initial response to the mailed questionnaire recruiting retirees and the small number of workers 

2311 from the entire pool of eligible participants who actually participated in the medical evaluation. 

2312 Attempts were made to contact 30% of the questionnaire nonrespondents, with 46% contacted 

2313 and 31% completing the telephone interview. The only difference found between those who 

2314 responded to the mailed questionnaire and those who did not was a higher percentage of 

2315 diagnosed heart disease among the nonrespondents who were 2.5 years older and had been 

2316 retired 1.7 more years than the respondents. Those who were eligible but did not participate in 

2317 the medical evaluation were similar to the exam participants on all characteristics included in the 

2318 interview. The only difference was a higher prevalence of gout among the unexposed who did 

2319 not participate compared to the unexposed who did participate. 

2320 

2321 4.1.2.4. Observational Studies Using Workplace Medical Program Data 
52

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

2323 dichloromethane on six health outcomes identified in the literature or based on biological 

2324 plausibility. Participants in the study were male workers at least 19 years old at two 

2325 General Electric plastic polymer plants where dichloromethane was one of the chemicals used. 

2326 Four dichloromethane exposure categories were established based on full-shift personal air 

2327 monitoring data (8-hour TWA) collected in 1979–1985, job titles, and industrial hygienists’ 

2328 knowledge of plant operations. The mean 8-hour TWA and number of workers in each of the 

2329 four exposure groups were 49.0 ppm for the 19 workers in the highest, 10.9 ppm for the 

2330 49 workers in the intermediate, 3.3 ppm for the 56 workers in the low, and <1.0 ppm for the 

2331 772 workers in the minimal/no exposure group. 

2332 Data from 1984 annual medical exams and 1985 absence data from payroll records were 

2333 evaluated for possible health effects resulting from occupational exposure to dichloromethane. 

2334 A high percentage of workers participated in the annual medical exams, with only 5 of the 

2335 896 eligible for inclusion in the study refusing the exam completely in 1984. Six hypotheses 

2336 were specifically tested regarding dichloromethane exposure in relation to different health 

2337 outcomes: absence due to illness, hepatotoxicity (manifested by nausea, weakness and fatigue, 

2338 palpable liver, abdominal tenderness, jaundice, hepatomegaly, abnormal serum γ-glutamyl 

2339 transferase, ALT, AST, or bilirubin), diabetes mellitus (manifested by weight loss, weakness and 

2340 fatigue, polydypsia, polyuria, impaired vision, excessive weight loss, elevated fasting blood 

2341 sugar, and abnormal urinary glucose or urinary acetone), CNS toxicity (manifested by headache, 

2342 lightheadedness, dizziness and vertigo, ataxia, weakness and fatigue, and abnormalities detected 

2343 in the central motor, central sensory, cranial nerve, gait, neurocoordination, or Bibinski reflex 

2344 examinations), cardiovascular abnormalities (manifested by fatigue, dyspnea, chest pain with 

2345 exertion, palpitations, or abnormalities detected in the point maximum impulse exam, blood 

2346 pressure measurements, or electrocardiogram), and neoplastic breast changes (154 women were 

2347 included in this portion of the study—manifested by painful breast, breast swelling, lump, nipple 

2348 discharge, or abnormalities detected in the breast examination). 

2349 Workers were placed in exposure categories based on their current jobs. In addition, 

2350 exposure to high noise levels occurred in both plants, and workers in each plant had exposure to 

2351 another chemical, either phenol or phosgene. The authors noted that workers tended to move 

2352 from entry-level jobs with high dichloromethane exposure to supervisory jobs with lower 

2353 dichloromethane exposure, based on the seniority system in place at both plants. Thus, current 

2354 exposure levels reported did not necessarily reflect cumulative exposure. Because of the way the 

2355 seniority system moved workers through jobs and the fact that workers were assigned to 

2356 dichloromethane exposure categories based on their current job, age was inversely related to 

2357 exposure and was controlled in the analysis of some of the continuous variables using analysis of 

2358 covariance. Age adjustment was not employed in the analysis of dichotomous variables. The 
53

mean age was 35.3, 39.7, 37.1 and 29.5 years in the minimal/no, low, medium32359 , and high 

2360 exposure groups, respectively. The small number of workers in the exposed groups limited the 

2361 ability to evaluate the effects of dichloromethane exposure on health outcomes related to age, 

2362 since age had to be adjusted in these analyses. The racial distribution did not differ among the 

2363 exposure groups. 

2364 The authors indicated that all the hypotheses were accepted with the exception of CNS 

2365 symptoms. However, it should be noted that the small size and younger age distribution in the 

2366 high exposure group and the lack of adjustment for age in most of the analyses make it difficult 

2367 to interpret the statistical testing that was performed. Data pertaining to neurological, hepatic, 

2368 and cardiac function are shown in Table 4-1. Among the six neurological symptoms evaluated, a 

2369 statistically significant positive exposure-effect relationship between dizziness/vertigo and 

2370 dichloromethane exposure was identified. This trend was driven most strongly by the low 

2371 frequency of this reported symptom in the minimal/no exposure group (1.2%), but there was no 

2372 linear trend across the higher levels of exposure (7.5, 2.1, and 5.3% in the low, medium, and high 

2373 exposure groups, respectively). 

2374 

 

3

 The “medium” exposure group is also referred to as the “intermediate” exposure group in Kolodner et al. (1990). 
54

2375 

Table 4-1. Percentage of male General Electric plastic polymer workers 

reporting neurologic symptoms or displaying abnormal values in measures 

of neurological function, hepatic function, and cardiac function 

Exposure Groupa

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/syncopeb0.1 0.0 2.1 0.0 

 Seizuresb 0.4 0.0 2.1 0.0 

 Paresis/paralysisb 0.7 0.0 0.0 0.0 

 Parasthesisb 4.0 7.5 14.6 0.0 

 Head trauma/concussionb 0.8 1.9 0.0 0.0 

 Peripheral motor examb,c 0.5 0.0 0.0 0.0 

 Peripheral sensory examb,c 1.1 2.4 5.1 0.0 

 Rhomberg examb,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 

Cardiacd

 

Palpitations: percent abnormal 1.2 9.1 2.1 0.0 

Electrocardiogram 

borderline/abnormal 18.5 16.7 19.1 8.3 

bradycardia/tachycardia abnormalitiesb 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 

aMean 8-hour TWA exposure was <1.0, 3.3, 0.9, and 49.0 ppm in the minimal/no, low, medium, and high groups, 

respectively; mean age 35.3, 39.7, 37.1, 29.5 years in the minimal/no, low, medium, and high groups, respectively. 

bThe authors considered these to be screening variables rather than hypothesis-testing variables.

c

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

d

For all cardiac outcomes except bradycardia/tachycardia, n = 728, 54, 47, and 12 in the minimal/no, low, medium, 

and high groups, respectively. For bradycardia/tachycardia, n = 727 in the minimal/no group. 

Source: Kolodner et al. (1990). 

2376 

2377 

2378 Soden (1993) compared health-monitoring data from dichloromethane-exposed workers 

2379 in the Rock Hill triacetate fiber production plant to workers from another plant making polyester 

2380 fibers owned by the same company in the same geographic area. Exposed and control workers 

2381 were chosen from among workers who had worked at least 10 years in their respective areas and 

2382 who participated in the company’s health-monitoring program between 1984 and 1986 and were 
55

2383 still employed on December 31, 1986. Controls were matched by race, age, and gender to each 

2384 Rock Hill worker for a sample size of 150 and 260 in the exposed and control groups, 

2385 respectively. (The aim of the study had been 1:2 matching.) The 8-hour TWAs among the Rock 

2386 Hill workers were those reported by Lanes et al. (1990), namely 475 ppm for dichloromethane, 

2387 900 ppm for acetone, and 100 ppm for methanol. None of these exposures occurred at the 

2388 polyester plant. There was a 90% participation rate in the health-monitoring program. Six 

2389 questions in the health history portion of the health-monitoring program concerned cardiac and 

2390 neurological symptoms (chest discomfort with exercise; racing, skipping, or irregular heartbeat; 

2391 recurring severe headaches; numbness/tingling in hands or feet; loss of memory; dizziness). Part 

2392 of this program included blood samples used for standard clinical hepatic and hematologic 

2393 parameters: serum ALT, AST, total bilirubin, and hematocrit. The clinical measures were 

2394 available for 90 (60%) of the exposed and 120 (46%) of the control group because some 

2395 participants declined this part of the health-monitoring program because similar tests had been 

2396 part of recent personal medical care. 

2397 There was little difference in the frequency of reported symptoms between exposed 

2398 workers and controls: chest discomfort reported by 2.0% of exposed and 4.0% of the controls, 

2399 irregular heartbeat reported by 5.5% of exposed and 6.0% of the controls, recurring severe 

2400 headaches reported by 3.5% of exposed and 5.5% of the controls, numbness/tingling in hands 

2401 and feet reported by 6.4% of exposed and 8.1% of the controls, loss of memory reported by 1.3% 

2402 of exposed and 0.4% of controls, and dizziness reported by 2.7% of exposed and 4.8% of 

2403 controls (Soden, 1993). The levels of the blood values were similar in the exposed and control 

2404 groups, except for a 3.1 IU/L decrease in serum AST activity (p = 0.06). The authors concluded 

2405 that this difference was not clinically significant but did not discuss the potential bias introduced 

2406 by the selective participation in this part of the study. 

2407 

2408 4.1.2.5. Studies of Ischemic Heart Disease Mortality Risk 

2409 Several studies have examined the relation between dichloromethane exposure and risk 

2410 of cardiovascular-related mortality. The methodological details of these studies are described in 

2411 section 4.1.3.2.). No evidence of increased risk of ischemic heart disease mortality was seen in 

2412 two triacetate film production cohort studies (Hearne and Pifer, 1999; Tomenson et al., 1997) or 

2413 in two triacetate fiber production cohort studies (Gibbs et al., 1996; Lanes et al., 1993). 

2414 Information on this outcome was not included in the dichloromethane analysis of civilian Air 

2415 Force base workers (Blair et al., 1998). The standardized mortality ratios (SMRs) for ischemic 

2416 heart disease mortality were <1.0 in all of the cohorts and dose groups examined (Table 4-2). 

2417 The “healthy worker effect” may have contributed to these observations. There are no case2418 control studies of ischemic heart disease and dichloromethane exposure. 2419 
56

Table 4-2. Ischemic heart disease mortality risk in four cohorts of 

dichloromethane-exposed workers 

 Obsa

 Expb SMR 95% CIc

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 

aObs = number of observed deaths. 

bExp = number of expected deaths. 

cCI = confidence interval. 

2420 

2421 

2422 4.1.2.6. Studies of Suicide Risk 

2423 Suicide risk is not an outcome that was a primary hypothesis or motivation of the cohort 

2424 studies but may be relevant given the potential neuropsychological effects of dichloromethane, 

2425 as evidenced from studies of acute and chronic exposure scenarios described previously. In a 

2426 triacetate film production cohort in Rochester, New York, Hearne and Pifer (1999) reported 

2427 14 observed deaths from suicide compared with 7.8 expected, for an SMR of 1.8 (95% 

2428 confidence interval [CI] 0.98–3.0) (Table 4-3). This cohort (“Cohort 1”) consisted of 1,311 men 

2429 who were first employed between 1946 and 1970 and were followed through 1994. Similar 

2430 results were seen in a different, but somewhat overlapping, cohort in this study (“Cohort 2”) of 

2431 1,013 men employed between 1964 and 1970 and followed through 1994 (see section 4.1.3.3.1). 

2432 There was also evidence of increasing suicide risk with dichloromethane exposure, particularly 

2433 in the highest exposure group, in the study of triacetate fiber production workers in Maryland 

2434 (Gibbs, 1992). The triacetate fiber production cohort study in Rock Hill, South Carolina, has 

2435 published what appears to be erroneous information about suicide risk. In the 1993 paper (Lanes 

2436 et al., 1993), 4 observed and 5.21 expected cases were reported (SMR 0.77), but the SMR that 

2437 was reported with these data was 1.19 (95% CI 0.39, 2.8). This ratio would correspond to 

2438 6 observed and around 5.2 expected cases. Information on suicide was not included in the other 

2439 film and fiber cohort studies (Tomenson et al., 1997) or in the analysis of civilian Air Force base 

2440 workers (Blair et al., 1998). There are no case-control studies of suicide risk and 

2441 dichloromethane exposure. 

2442 
57

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

 

 Obsa

 Expb 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 productionc 

Gibbs (1992) 50–100 ppm 8 6.4 1.3 0.54–2.5 

350–700 ppm 8 4.4 1.8 0.78–3.6 

aObs = number of observed deaths. 

bExp = number of expected deaths. 

cOne additional study provided data on suicide risk, but some kind of error seems to be present: 4 observed and 

5.21 expected cases were reported in Lanes et al. (1993), which would be an SMR of 0.77, but the SMR that was 

reported with these data was 1.19 (95% CI 0.39, 2.8). This ratio would correspond to 6 observed and around 

5.2 expected cases. 

2443 

2444 

2445 4.1.2.7 Studies of Infectious Disease Risk 

2446 There is limited information pertaining to infectious disease risk in relation to 

2447 dichloromethane exposure. Only one of the cohort studies (Hearne and Pifer, 1999) reported 

2448 data for the broad category of infectious and parasitic disease mortality. In Cohort 1 of this 

2449 analysis, there were no observed deaths in this category (5.6 expected), and in Cohort 2 there 

2450 were 3 observed and 4.7 expected, for an SMR of 0.64. The detailed report by Gibbs (1992) of 

2451 the cellulose triacetate fiber production cohorts in Maryland (Gibbs et al., 1996) also contained 

2452 information on the facility in South Carolina that was the site of the report by Lanes et al. (1993, 

2453 1990). Slightly elevated risks of mortality due to influenza and pneumonia were seen among the 

2454 male workers in the high exposure group in Maryland (7 observed, 5.62 expected, SMR 1.25) 

2455 and in South Carolina (3 observed, 1.33 expected, SMR 2.26). Among females, there were few 

2456 observed or expected cases (in Maryland, 1 observed, 0.23 expected, SMR 4.36; in South 

2457 Carolina, 0 observed and 0.74 expected). 

2458 

2459 4.1.2.8. Studies of Reproductive Outcomes 

2460 Pregnancy outcomes in women exposed to dichloromethane have been investigated in 

2461 two studies. Taskinen et al. (1986) studied spontaneous abortions among women employed in 

2462 eight pharmaceutical factories between 1973 and 1980. Data on pregnancy outcomes were 

2463 collected from a national hospital and clinic discharge registry in Finland from 1973 to 1981 by 

2464 matching the worker rosters to the registry. Exposure to dichloromethane was one of eight 

2465 solvents or classes of solvents included in the study. The study consisted of two parts. The first 

2466 investigated the rate of spontaneous abortions (number of spontaneous abortions divided by the 

2467 sum of spontaneous abortions and births) during, before, or after employment in the 

2468 pharmaceutical industry. One hundred and forty-two spontaneous abortions and 1,179 births 

2469 were identified among the female workers at the eight plants. Employment hire and termination 
58

2470 dates were obtained from plant records. The spontaneous abortion rate was 10.9% during 

2471 employment compared with 10.6% before and after employment. These results compared to a 

2472 rate of 8.5% in the general population in the geographic area where the factories were located. 

2473 The rate of spontaneous abortions among workers declined over the period of the study, with a 3-

2474 year moving average of 15% at the beginning declining to 9.5% at the end of the study. Over the 

2475 same period, the industrial hygiene allegedly improved in the plants. Ten congenital 

2476 malformations of different types were identified among the women (five among those who were 

2477 employed in the pharmaceutical industry during the pregnancy and five among those whose 

2478 pregnancies occurred before or after this employment). 

2479 The second part of the study by Taskinen et al. (1986) was a case-control study of the risk 

2480 of spontaneous abortions in relation to workplace exposures during pregnancy. The source 

2481 population consisted of women who were employed in one of the eight Finnish pharmaceutical 

2482 factories during at least 1 week of the first trimester of pregnancy during the study period. Cases 

2483 (n = 44) were selected from this population based on hospital or clinic records indicating a 

2484 spontaneous abortion, and 130 controls (women who had given birth) were age-matched 

2485 (3:1 matching, age within 2.5 years) to each case. Occupational exposure data were obtained by 

2486 questionnaires completed by the plant physician or the nursing staff, blinded to the case status of 

2487 the study member, in consultation with labor protection chiefs and department foremen. The 

2488 questionnaire requested information about job history and job tasks, exposure to eight specific 

2489 solvents or classes of solvents (aliphatic solvents, alicyclic solvents, toluene, xylene, benzene, 

2490 chloroform, dichloromethane, and other solvents), antineoplastic agents, carcinogens, hormones, 

2491 antibiotics, heavy lifting, known chronic diseases, acute diseases during pregnancy, smoking 

2492 status, and previous pregnancies. Exposure frequency to each solvent was based on the 

2493 cumulative weighted sum of the number of days/week the woman was exposed to the solvent. 

2494 While overall response to the questionnaire was 93%, less than half the questionnaires contained 

2495 information about smoking or previous pregnancies, precluding inclusion of these variables in 

2496 the analysis. The distribution of broad categories of occupations (i.e., pharmaceutical workers 

2497 and packers, laboratory assistants) was similar in both groups. However, exposure to each of the 

2498 solvents was higher in the cases compared with controls, and the results for dichloromethane 

2499 were relatively strong. For dichloromethane, the prevalence of exposure was 28.9 and 14.3% in 

2500 cases and controls, respectively, resulting in an odds ratio (OR) of 2.3 (95% CI, 1.0, 5.7). There 

2501 was also evidence of an increasing risk with higher exposure frequency, with an odds ratio (OR) 

2502 of 2.0 (95% CI 0.6, 6.6) with exposures of less than once a week and 2.8 (95% CI 0.8, 9.5) with 

2503 exposures of once a week or more. An association was also seen with exposure to four or more 

2504 solvents (OR 3.4, 95% CI 1.0, 12.5), and weaker associations were seen with other specific 

2505 solvents (e.g., chloroform, toluene). 

2506 Bell et al. (1991) investigated the relation between birth weight and maternal exposure to 

2507 airborne dichloromethane as a result of living around the triacetate film facility in Rochester, 
59

2508 New York. For this population-based cross-sectional study, birth certificates were obtained for 

2509 all births in 1976–1987 in Monroe County, where the triacetate film facility is located. Multiple 

2510 births and births of infants weighing <750 grams were excluded. Data abstracted from the 

2511 certificate included date of birth, census tract of residence, age, race, educational level of the 

2512 mother and father, sex, gestational age, multiple births, month of the pregnancy that prenatal care 

2513 began, total previous births, total previous live births, and conditions present during the 

2514 pregnancy. An air dispersion modeling system for 250 air emissions, including 

2515 dichloromethane, predicting average annual ground level concentrations in the surrounding 

2516 community, was used to assign dichloromethane exposure levels to each birth mother. One of 

2517 four levels of exposure was assigned to each census tract based on the isopleth of exposure in 

2518 which more than half of the census tract population resided. Because of the few births among 

2519 nonwhites that occurred in areas of higher exposure, the study was restricted to whites 

2520 (n = 91,302). The number of births that occurred in each of the four exposure levels was 

n = 1,085 in the high-exposure group (50 μg/m3

2521 [0.014 ppm]), n = 1,795 in the moderateexposure group (25 μg/m3

 [0.007 ppm]), n = 6,044 in the low-exposure group (10 μg/m32522 

2523 [0.003 ppm]), and n = 82,076 in the no-exposure group. At the levels of dichloromethane 

2524 exposure in this population no significant adverse effect on birth weight was found. There was 

2525 an 18.7 g decrease in birthweight (95% CI –51.6, 14.2) in the high- compared with the no2526 exposure group, adjusting for maternal age, maternal education, parity, previous pregnancy loss, 

2527 late start of prenatal care, sex of the child, and pregnancy complications. No significant 

2528 association was found between any combination of exposure levels and birth weight. There was 

2529 no association between exposure group and risk of a low birthweight infant (i.e., <2,500 g, 

2530 OR 1.0 [95% CI 0.81, 1.2] in the high- compared with the no-exposure group). The authors 

2531 point out a number of problems with assignment of dichloromethane exposure. It is possible that 

2532 the dichloromethane exposure was overestimated using the model. Comparisons to ambient air 

2533 sampling levels collected six times/year resulted in the dichloromethane exposure derived from 

2534 the model being twice as high as the ambient air samples. There was also inaccuracy in the 

2535 assignment of dichloromethane exposure level to each birth because the exposure assignment 

2536 was made using the predominant value of the isopleth for a census tract. 

2537 Two studies have investigated the occurrence of oligospermia among men occupationally 

2538 exposed to dichloromethane exposure. Kelly (1988) studied 34 men employed in an automotive 

2539 plant as bonders, finishers, and press operators. These men were self-referred to a health center 

2540 for a variety of complaints, including neurological symptoms, musculoskeletal symptoms, and 

2541 shortness of breath. Twenty-six of the men were bonders and eight were finishers or press 

2542 operators. The job as bonder consisted of dipping hands into an open bucket of dichloromethane 

2543 and splashing it onto plastic automobile parts. The dichloromethane exposure for bonders 

2544 averaged 68 ppm with a range from 3.3 to 154.4 ppm. Eight men, all of whom were bonders, 

2545 reported symptoms of testicular and epidydimal tenderness, with confirmation on medical exam. 
60

2546 They ranged in age from 20 to 47 years old and had been bonders for up to 2.9 years. The COHb 

2547 levels for the eight workers with genital symptoms ranged from 1.2 to 17.3%, with an average of 

2548 6.9% anywhere from 4 to 90 hours postexposure. The COHb levels for the two men who 

2549 smoked were among the highest, namely 7.3 and 17.3%. Four of the eight workers agreed to 

provide semen samples; their sperm counts were 2–26 × 106/cm32550 . The authors stated that men 

with sperm counts as low as 25 × 106/cm32551 may still be fertile, but none of these men had had any 

2552 children since working with dichloromethane despite not using contraceptives. There was one 

2553 miscarriage. All four men reported dipping their hands into open buckets of dichloromethane 

2554 without any protective equipment, and two men reported feeling dizzy, giddy, and high at work. 

2555 Based on the results of the Kelly (1988) case report, Wells et al. (1989) planned to do a 

2556 study of oligospermia among 20 exposed and 20 unexposed to dichloromethane. The exposed 

2557 workers were unvasectomized men who had worked for the 3 months prior to recruitment in 

2558 furniture stripping shops. Eleven men were recruited from among 14 eligible workers at six 

2559 different shops where dichloromethane was utilized. Names of acquaintances of the exposed 

2560 were solicited as potential referents. Only one exposed man provided any names. Therefore, the 

2561 study was redirected as a case report on the 11 exposed men. The mean TWA dichloromethane 

2562 exposure was 122 ppm (range 15–366 ppm) with a mean COHb of 5.8% (range 2.2–13.5%). 

2563 The mean COHb for smokers, 10.2% (range 8.1–13.5), was higher than for nonsmokers, 3.9% 

2564 (range 2.2–5.9), and the nonsmoker levels were higher than the 2% level considered to be the 

upper limit of normal in nonsmoking populations. The mean sperm count was 54 × 106/cm32565 

(range 23–128 × 106/cm3) compared to a population value of 47 × 106/cm32566 for the same 

2567 geographic based on samples analyzed at the same laboratory. Using the standard definition for 

oligospermia of 20 × 106/cm32568 , none of the 11 workers had oligospermia. 

2569 

2570 4.1.2.9. Summary of Noncancer Studies 

2571 The clinical and workplace studies of noncancer health effects of dichloromethane 

2572 exposure have examined markers of disease and specific clinical endpoints relating to cardiac, 

2573 neurological disease, hepatic function, and reproductive health. 

2574 

2575 Cardiac effects 

2576 The effect of dichloromethane on the formation of COHb (Stewart et al., 1972b) raised 

2577 concerns about potential risk of cardiovascular damage. To date, there is little evidence of 

2578 cardiac damage related to dichloromethane exposure in the cohort studies of dichloromethane2579 exposed workers that examined ischemic heart disease mortality risk (Hearne and Pifer, 1999; 

2580 Tomenson et al., 1997; Gibbs et al., 1996; Lanes et al., 1993) or in two small cardiac monitoring 

2581 studies (Ott et al., 1983d; Cherry et al., 1981). However, limitations in these studies should be 

2582 noted, including the healthy worker effect and the absence of data pertaining to workers who 

2583 died before the establishment of the analytic cohort (Gibbs et al., 1996; Gibbs, 1992). 
61

2584 Neurological effects 

2585 The acute effects of dichloromethane exposure on neurological function seen in 

2586 numerous case reports have also been established in experimental studies in humans (Putz et al., 

2587 1979; Winneke, 1974; Stewart et al., 1972a, b). Relatively less is known about the long-term 

2588 effects of chronic exposures in humans. Some data from studies of workers suggest that the 

2589 effects of dichloromethane are relatively short-lived. For example, in the study by Cherry et al. 

2590 (1983) of 56 exposed and 36 unexposed workers, alterations in mood or in digit substitution test 

2591 results were seen during the course of a work shift but were not seen at the beginning of a shift. 

2592 No difference in four neurological symptoms was seen in an analysis of exposed workers 

2593 (average exposure 475 ppm, ≥10-year duration) and an unexposed comparison group by Soden 

2594 (1993). Other data suggest an increase in prevalence of neurological symptoms among workers 

2595 (Cherry et al., 1981) and possible detriments in attention and reaction time in complex tasks 

2596 among retired workers (Lash et al., 1991). These latter two studies are limited by the small 

2597 sample size. Thus Cherry et al. (1981) and Lash et al. (1991) have low power for detecting 

2598 statistically significant results and consequently should not be interpreted as definitive analyses 

2599 showing no effects. Rather, these analyses provide some evidence of an increased prevalence of 

2600 neurological symptoms among workers with average exposures of 75–100 ppm (Cherry et al., 

2601 1981) and long-term effects on specific neurological measures (i.e., attention and reaction time) 

2602 in workers whose past exposures, at least for part of their work history, were in the 100–200 ppm 

2603 range (Lash et al., 1991). The increased risk of suicide (approximately a twofold increased risk) 

2604 seen in two of the worker cohort studies (Hearne and Pifer, 1999; Gibbs, 1992) is an additional 

2605 indication of potential neurological consequences of dichloromethane exposure. Adequate 

2606 studies addressing these specific issues are not available. Thus, given the suggestions from the 

2607 currently available studies, the statement that there are no long-term neurological effects of 

2608 chronic exposures to dichloromethane cannot be made with confidence. 

2609 

2610 Hepatic effects 

2611 Three studies provide data pertaining to markers of hepatic damage (i.e., serum enzymes 

2612 and bilirubin levels) (Soden, 1993; Kolodner et al., 1990; Ott et al., 1983c). Two of these studies 

2613 were based in the Rock Hill, South Carolina, cellulose triacetate fiber plant (Soden, 1993; Ott et 

2614 al., 1983c), with the most recent of the studies focusing on workers with more than 10 years 

2615 duration in a high exposure area (average exposure estimated as 475 ppm). There is some 

2616 evidence of increasing levels of serum bilirubin with increasing dichloromethane exposure in Ott 

2617 et al. (1983c) and Kolodner et al. (1990), but there are no consistent patterns with respect to the 

2618 other hepatic enzymes examined (serum γ-glutamyl transferase, serum AST, serum ALT). These 

2619 studies do not provide clear evidence of hepatic damage in dichloromethane-exposed workers, to 

2620 the extent that this damage could be detected by these serologic measures; however, these data 
62

2621 are limited and thus the absence, presence, or extent of hepatic damage is not known with 

2622 certainty. 

2623 

2624 Immune effects 

2625 Only limited, and somewhat indirect, evidence pertaining to immune-related effects of 

2626 dichloromethane in humans is available. No risk was seen in the broad category of infectious 

2627 and parasite-related mortality reported by Hearne and Pifer (1999), but there was some evidence 

2628 of an increased risk for influenza and pneumonia-related mortality at two cellulose triacetate 

2629 fiber production work sites in Maryland and South Carolina (Gibbs, 1992). In the Maryland 

2630 facility, an increased risk of cervical cancer was seen among the 938 female workers, with an 

2631 SMR of 3.0 (95% CI, 0.96, 6.9) in the 50–100 ppm group and 5.4 (95% CI 0.13, 30.1) in the 

2632 350–700 ppm group (Gibbs et al., 1996). Cervical cancer is viral mediated (human papilloma 

2633 virus), and immunosuppression is a risk factor for development of this disease, as seen by the 

2634 increased risk in immunocompromised patients and people taking immunosuppressant 

2635 medications (Leitao et al., 2008; Ognenovski et al., 2004). 

2636 

2637 Reproductive effects 

2638 Studies pertaining to various reproductive effects and dichloromethane exposure from 

2639 workplace settings (Wells et al., 1989; Kelly, 1988; Taskinen et al., 1986) or environmental 

2640 settings (Bell et al., 1991) have examined possible associations with spontaneous abortion 

2641 (Taskinen et al., 1986), low birth weight (Bell et al., 1991), and oligospermia (Wells et al., 1989; 

2642 Kelly, 1988). Of these, the data pertaining to spontaneous abortion provide the strongest 

2643 evidence of an adverse effect of dichloromethane exposure, particularly with respect to the case2644 control study in which the strongest association was seen specifically with the higher frequency 

2645 category of dichloromethane exposure. It is a small study (44 cases, 130 controls), however, 

2646 with limited quantitative exposure assessment and multiple exposures (although the association 

2647 seen with dichloromethane was among the highest seen among the solvents) and so cannot be 

2648 considered to firmly establish the role of dichloromethane in induction of miscarriage. However, 

2649 the high exposure scenario, including the potential for substantial dermal exposure in the study 

2650 of Kelly (1988), also suggests the potential for adverse male reproductive effects. 

2651 

2652 4.1.3. Cancer Studies 

2653 4.1.3.1. Identification and Selection of Studies for Evaluation of Cancer Risk 

2654 Twelve epidemiologic studies of cancer risk were identified and included in this 

2655 evaluation: four cohorts for which the primary solvent exposure was to dichloromethane (two in 

2656 film production settings and two in cellulose triacetate fiber production), one large cohort of 

2657 civilian employees at a military base with exposures to a variety of solvents but that included an 

2658 assessment specifically of dichloromethane exposure, and seven case-control studies of specific 
63

2659 cancers with data on dichloromethane exposure. One additional study (Ott et al., 1985), a cohort 

2660 of 1,919 men employed at Dow Chemical facilities, was identified but was not included in the 

2661 summary. The analysis was based on exposure to a combined group of chlorinated methanes 

2662 (e.g., carbon tetrachloride, chloroform, methyl chloride, and dichloromethane), and it was not 

2663 possible from the data presented to assess the individual effects of dichloromethane. 

2664 

2665 4.1.3.2. Description of the Selected Studies 

2666 In this section, the study setting, methods (including exposure assessment techniques), 

2667 results pertaining to incidence or mortality from specific cancers, and a brief summary of 

2668 primary strengths and limitations are summarized for each of the 12 selected studies. When two 

2669 papers of the same cohort were available, the results from the longer period of follow-up are 

2670 emphasized in the summary. Information from earlier reports is used when these reports contain 

2671 more details regarding working conditions, study design, and exposure assessment. The 

2672 description of individual studies is followed by a summary of the evidence available from these 

2673 studies relating to specific types of cancer. 

2674 

2675 4.1.3.3. Cellulose Triacetate Film Base Production Cohorts 

2676 4.1.3.3.1. Cellulose triacetate film base production—Rochester, New York (Eastman Kodak). 

2677 Friedlander et al. (1978) reported a cohort mortality study of workers in an Eastman Kodak 

2678 facility in Rochester, New York. This study was expanded and extended several times during 

2679 the next 20 years (Hearne and Pifer, 1999; Hearne et al., 1990, 1987). The latest analysis 

2680 provided data on two overlapping cohorts. The first cohort (Cohort 1) consisted of 1,311 male 

2681 workers employed in the roll coating division (n = 1,070) or the dope and distilling departments 

2682 (n = 241) of the Eastman Kodak facility in Rochester, New York. Men who began working in 

2683 these areas after 1945 and were employed in these areas for at least 1 year (including seasonal or 

2684 part-time work that equaled 1 full-time year equivalent) from 1946 to 1970 were included. 

2685 Follow-up time was calculated from the end of the first year of employment in the study area 

2686 through December 31, 1994. The mean duration of work in Cohort 1 was 17 years. The total 

2687 number of person-years of follow-up was 46,112, and the mean duration of follow-up was 

2688 35.2 years (range 25–49 years). The second cohort (Cohort 2) included 1,013 male workers in 

2689 the roll coating division who were employed for at least 1 year in this division between 1964 and 

2690 1970. Follow-up time was calculated from January 1, 1964, or the date of first employment in 

2691 the roll coating division for those who were employed there before 1964 and those who began in 

2692 1964 or later, respectively. Follow-up continued through December 31, 1994. The mean 

2693 duration of work in Cohort 2 was 24 years. Total follow-up time was 26,251 person-years, and 

2694 the mean duration of follow-up was 25.9 years (range 25–31 years). Cohort 2 was the focus of 

2695 previous analyses by Friedlander et al. (1978) and Hearne et al. (1990, 1987). 
64

2696 For both cohorts, causes of death were based on the underlying causes of death recorded 

2697 on the death certificates, which were routinely obtained by the company for the processing of life 

2698 insurance claims. The expected number of deaths was calculated using appropriate age-, sex-, 

2699 calendar time-, and cause-specific death rates for men in New York State (excluding New York 

2700 City). In addition, another referent group was also used in the analysis of the second cohort. 

2701 This other referent was based on the age-, sex-, calendar time-, and cause-specific death rates of 

2702 other hourly male workers employed at the Eastman Kodak plant in Rochester, New York. (An 

2703 internal referent group was also described for Cohort 1, but data for that analysis were not 

2704 presented.) 

2705 Dichloromethane was first used in the film production process at the Eastman Kodak 

2706 facility around 1944 (Hearne et al., 1987). Cellulose triacetate was dissolved in dichloromethane 

2707 and then cast into a thin film onto revolving wheels. The film was then cured by circulating hot 

2708 air in the coating machines, and the solvent was recovered and redistilled. 1,2-Dichloropropane 

2709 and 1,2-dichloroethane were also used as solvents from the 1930s to the 1960s, but 

2710 dichloromethane was predominant (ratio 17:2:1 for dichloromethane:1,2-dichloropropane:1,2-

2711 dichloroethane in general workplace air measurements) (Hearne et al., 1987). 

2712 The exposure assessment in the Rochester, New York, Eastman Kodak cohort studies 

2713 was based on employment records (start and stop dates for specific jobs in the relevant areas of 

2714 the company) in combination with air monitoring data used to estimate the exposure level for a 

2715 given job, location, and time period (Hearne et al., 1987). Air monitoring began in the 1940s, 

2716 but few data are available before 1959. In the most recent update (Hearne and Pifer, 1999), more 

2717 than 1,500 area and 2,500 breathing zone air samples were used in the exposure assessment 

2718 process. Reductions in exposures in the dope department and the distilling department began 

2719 after 1965. The highest exposure jobs were operator and maintenance workers (dope 

2720 department) and filter washing and waste operator (distilling department), with estimated 8-hour 

2721 TWA exposures of 100–520 ppm between 1946 and 1985. There was little change in estimated 

2722 exposures for jobs in the roll coating division from the 1940s through 1985, but some reduction 

2723 was seen from 1986 to 1994. The mean 8-hour TWA exposures were 39 ppm for Cohort 1 and 

2724 26 ppm for Cohort 2. 

2725 These data were used to estimate a cumulative exposure index (i.e., the summation across 

2726 all jobs held by an individual of the product of the average dichloromethane concentration as 

2727 ppm and the duration of employment in that job). The authors refer to this as a “career exposure 

2728 index.” Additional adjustment in these estimates was made for respiratory protection, but the 

2729 details of this adjustment were not described. For Cohort 1, the cumulative exposure categories 

2730 used in exposure-effect analyses were <150, 150–349, 350–799, and ≥800 ppm. For Cohort 2, 

2731 the cumulative exposure categories were <400, 401–799, 800–1,199, and ≥1,200 ppm. The cut 

2732 points were chosen to produce an approximately equal number of expected total deaths in these 

2733 categories. Two different methods to calculate expected number of deaths within each exposure 
65

2734 category were used for each cohort analysis. For Cohort 1, an internal comparison was made 

2735 based on the distribution of person-years within each exposure category, and an external 

2736 comparison was made applying New York State mortality rates. For Cohort 2, the internal 

2737 comparison using the distribution of person-years within each exposure category was also used, 

2738 but the external comparison was based on mortality rates in other hourly workers at the 

2739 Rochester, New York, Eastman Kodak work site. 

2740 There was no increased risk of mortality for all sites of cancer or for lung cancer in either 

2741 cohort analysis (Table 4-4). Data pertaining to smoking history, obtained from a survey of 

2742 workers in the New York film production facility, indicate that smoking rates were similar in the 

2743 exposed group, the internal comparison group, and the general population; therefore, it is 

2744 unlikely that differences in smoking could be masking an effect of dichloromethane. 

2745 The only specific sites for which there were increased SMRs in both cohorts were brain 

2746 and CNS cancer, Hodgkin’s lymphoma, and leukemia. Pancreatic cancer mortality risk was 

2747 increased in Cohort 2 but not in Cohort 1. None of these associations were statistically 

2748 significant, and the Hodgkin’s lymphoma observations were based on a total of only four cases 

2749 in both cohorts combined and so were very imprecise. Within Cohort 2, there was little 

2750 difference in results for most sites, using the different referent groups, but the point estimates for

2751 the SMRs for brain and CNS cancer, Hodgkin’s lymphoma, and leukemia were somewhat higher 

2752 using the New York State referent group compared with the internal Eastman Kodak referent 

2753 group. An attenuation of the dichloromethane association seen in the analyses using the internal 

2754 Kodak referent group would be expected if the risk of specific cancers was increased in this 

2755 comparison group, possibly because of other workplace exposures. 
66

Table 4-4. Mortality risk in Eastman Kodak cellulose triacetate film base production workers, RNew York 

Cohort 1: 

1,311 men employed 1946–1970, followed 

through 1994 

Cohort 2: 

1,013 men employed 1964–1970, followed 

New York referent group New York referent group Koda

Cancer type Obsb Expb SMR 95% CI Obs Exp SMR 95% CI Exp Cancer, all sites 93 105.8 0.88 0.71–1.08 91 102.0 0.89 0.72–1.10 94.7  Livera 1 2.4 0.42 0.01–2.36 1 2.4 0.42 0.01–2.33 1.8 

 Pancreas 5 5.5 0.92 0.30–2.14 8 5.3 1.51 0.65–2.98 5.1  Lunga 27 36.0 0.75 0.49–1.09 28 34.2 0.82 0.55–1.19 31.3 

 Braina 6 2.8 2.16 0.79–4.69 4 2.1 1.88 0.51–4.81 2.7 

 Lymphatic system 5 6.6 0.75 0.24–1.78 6 5.7 1.06 0.39–2.30 5.7 

 Non-Hodgkin’s 2 4.1 0.49 0.06–1.76 3 3.5 0.85 0.17–2.50 3.6  Hodgkin’s 2 1.1 1.82 0.20–6.57 2 0.6 3.13 0.35–11.30 0.9  Multiple myeloma 1 1.5 0.68 0.01–3.79 1 1.5 0.65 0.01–3.62 1.3 

 Leukemia 8 3.9 2.04 0.88–4.03 6 3.5 1.73 0.63–3.76 4.4 

aLiver includes liver and biliary duct; lung includes lung, trachea, and bronchus; brain includes brain and CNS. bObs = number observed deaths, Exp = number of expected deaths. 

Source: Hearne and Pifer (1999). 

2756 
67

2757 The authors presented the exposure-effect analysis, based on the estimated cumulative 

2758 dichloromethane exposure groups, for all sites of cancer, pancreatic cancer, lung cancer, brain 

2759 cancer, and leukemia (Table 4-5). 

2760 

Table 4-5. Mortality risk by cumulative exposure in Eastman Kodak 

cellulose triacetate film base production workers, Rochester, New York 

Cohort, cancer, 

referent group SMRs (number of observed deaths) 

Cohort 1a

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 

Lungb

 internal

 New York 

0.78 (5) 

0.52 

1.07 (6) 

0.90 

1.25 (9) 

0.86 

0.90 (7) 

0.77 

Brainb

 

 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 2c

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 

Lungb

 Internal

 New York 

 

0.95 (6) 

0.80 

 

1.15 (12) 

1.00 

0.94 (6) 

0.89 

0.82 (4) 

0.79 

Brainb

 

 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 

aCohort 1: 1,311 men employed 1946–1970 in the roll coating division, dope department, or distilling department, 

followed through 1994; mean exposure (cumulative exposure years) 66, 244, 543, and 1,782 ppm-years in the four 

dose groups, respectively. 

bLung includes lung, trachea, and bronchus; brain includes brain and CNS. 

cCohort 2: 1,013 men employed 1964–1970 in the roll coating division, followed through 1994; mean exposure 

(cumulative exposure years) 168, 581, 981, and 1,670 ppm-years in the four dose groups, respectively. 

Source: Hearne and Pifer (1999). 

2761 

2762 
68

2763 There is no evidence of an exposure-effect for all site cancer mortality or lung cancer 

2764 mortality risk. The relatively sparse number of deaths for the other specific cancer types makes 

2765 it difficult to interpret the data. The patterns for pancreatic cancer differ between the two 

2766 cohorts, with increased risk at the higher dose in Cohort 1 and a U-shaped curve seen in 

2767 Cohort 2. For brain cancer mortality, a higher SMR was seen in the groups with cumulative 

2768 exposure levels of 800 ppm-years or greater compared with lower exposure groups. For 

2769 leukemia, in both cohorts, an increased mortality risk is seen in the highest exposure group 

2770 (mean approximately 1,700 ppm-years). 

2771 A strength of the Eastman Kodak cohort studies was the sampling data for 

2772 dichloromethane that allowed an assessment of each worker’s exposure, using the monitoring 

2773 data and the worker’s job history, making exposure-effect analyses possible. Follow-up of the 

2774 vital status of the cohort was >99% (Hearne and Pifer, 1999). There was also some information 

2775 on smoking history, too, for workers in the plant, based on a survey conducted in 1986 (Hearne 

2776 et al., 1987). A difficulty in interpreting the data, however, is that there was some overlap 

2777 between the cohorts: 707 of the men were included in both Cohort 1 and Cohort 2. Data are not 

2778 presented in a way that would allow the reader to eliminate duplicate cases and person-years so 

2779 that cases are only counted once when examining both cohorts. A strength of the Cohort 

2780 1 sampling strategy, compared with that of Cohort 2, is that Cohort 1 is limited to workers who 

2781 began work at the plant after 1945. These workers would not have had workplace exposure to 

2782 methanol and acetone, which were used at the plant in the film production process prior to that 

2783 time. Also, follow-up began with the beginning of employment in the relevant area. In contrast, 

2784 Cohort 2 was limited to workers who were employed from 1964 to 1970, so exposed workers 

2785 who left or died before 1964 were not included. The relatively small number of cases with 

2786 specific low incidence cancers (e.g., brain, leukemia) is also a limitation of the analyses of both 

2787 of the cohorts in this study. In addition, the exposure levels in both cohorts (mean 8-hour TWA 

2788 39 and 26 ppm in Cohorts 1 and 2, respectively) is relatively low compared with values seen in 

2789 other workplaces, including the cellulose triacetate fiber production cohorts described in Ott et 

2790 al. (1983a) and Gibbs et al. (1996). Also, the outcome assessment is based on mortality 

2791 (underlying cause from death certificates) rather than incidence data, and, because the Kodak 

2792 studies were limited to men, there is no information on risk of breast cancer or other female 

2793 reproductive cancers. 

2794 

2795 4.1.3.3.2. Cellulose triacetate film base production—Brantham, United Kingdom (Imperial 

2796 Chemical Industries). Tomenson et al. (1997) reported the results of a retrospective cohort 

2797 mortality study of 1,473 men who worked at a film-base production facility in Brantham, 

2798 England, anytime between 1946 and 1988 in jobs that were considered to have dichloromethane 

2799 exposure. The start of the follow-up period was not specified by the authors but is likely to have 

2800 been 1946 or the date of first employment at the plant. Follow-up of the cohort continued 
69

2801 through December 31, 1994, and vital status was based on national records (United Kingdom 

2802 National Health Service Central Register and the Department of Social Security). Causes of 

2803 death were based on the underlying causes of death recorded on the death certificates. The 

2804 expected number of deaths was calculated using age-, sex-, calendar time-, and cause-specific 

2805 death rates for England and Wales. In addition, a comparison using mortality rates for the local 

2806 areas (Tendring and Samford) for 1968–1978 and analyses limited to workers who had been 

2807 employed for at least 3 months were also made, but the results of these analyses were not 

2808 presented. The mean duration of work in the cohort was 9 years, the total number of person2809 years was 39,759, and the mean duration of follow-up was 27.0 years (7–49 years). 

2810 This facility produced cellulose diacetate film from 1950 to 1988, with other types of 

2811 films also manufactured beginning in the 1960s. Dichloromethane was the solvent used in this 

2812 process, and exposure occurred in the production of the triacetate film base and the casting of the 

2813 film into rolls. The exposure assessment was based on more than 2,700 personal or air 

2814 monitoring samples collected since 1975. An exposure matrix was constructed, assigning jobs to 

2815 1 of 20 work groups with similar exposure potential for each of four different time periods 

2816 (before 1960, 1960–1969, 1970–1979, and 1980–1988). For the 1980–1988 period, exposure 

2817 estimates for specific jobs were based on about 330 personal monitoring samples. For the earlier 

2818 time periods, information about work tasks and location was used in combination with the 

2819 information about the number of, use of, speed of, and problems with casting machines at 

2820 different times from their initial introduction in 1950. The highest exposures were estimated to 

2821 be in the casting machine operators and cleaners. Lifetime cumulative exposure to 

2822 dichloromethane was calculated as the product of the mean level of exposure for the assigned 

2823 work group and the duration of employment in that job summed across all jobs. Three categories 

2824 of cumulative exposure were used for the analysis of ever exposed workers: <400, 400–700, and 

2825 800+ ppm-years. Approximately 30% of the workers in the cohort were classified as 

2826 “unassigned” for the calculation of exposure group because sufficient information needed to 

2827 determine exposures (i.e., the location and tasks assigned to laborers and maintenance workers) 

2828 was not available. The mean 8-hour TWA exposure was estimated as 19 ppm for the cohort. 

2829 There was no increased risk of mortality for all sites of cancer (Table 4-6), and the SMRs 

2830 for most of the specific cancer sites examined (stomach, colon, rectum, liver, pancreas, lung, and 

2831 prostate) were less than 1.0. The only specific sites for which there was an increased SMR (i.e., 

2832 1.1 or higher) were brain and CNS cancer and leukemia, and these estimates were based on few 

2833 (less than five) observed cases (Table 4-6). Tomenson et al. (1997) present the exposure-effect 

2834 analysis, based on the estimated cumulative dichloromethane exposure groups, for all sites of 

2835 cancer, pancreatic cancer, and lung cancer, and there is no evidence of an increasing effect with 

2836 increasing exposure level in these analyses. A formal exposure-effect analysis for brain cancer 

2837 or leukemia was not presented. However, the authors described two of the brain cancer cases as 

2838 having “minimal” exposure to dichloromethane (and thus presumably would have been in the 
70

2839 <400 ppm-year cumulative exposure group). One case was estimated as having 572 ppm-years 

2840 cumulative exposure, and the other case was an electrician classified in the unassigned exposure 

2841 group. He had worked for 21 years at an exposure level “that was unlikely to have exceeded 

2842 15 ppm 8 hour TWA.” 

2843 

Table 4-6. Mortality risk in Imperial Chemical Industries cellulose 

triacetate film base production workers, Brantham, United Kingdom: 

1,473 men employed 1946–1988, followed through 1994

Cancer type Observed Expecteda 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 

aExpected, calculated from observed and SMR data reported by the authors by using the following formula: 

expected = 100 × observed ÷ SMR; SMRs and CIs were not calculated for categories with zero observed cases. 

Source: Tomenson et al. (1997). 

2844 

2845 

2846 A strength of this study was the monitoring data available that allowed assignment of 

2847 cumulative exposure categories for use in exposure-effect analyses. However, 30% (439) of 

2848 exposed workers had insufficient work histories to determine lifetime cumulative exposure. Air 

2849 measurements were not available until 1975, and personal measures were not available until 

2850 1980. In addition, the duration of exposure was relatively low (mean, 9 years), the mean 

2851 exposure level was relatively low (mean 8-hour TWA, 19 ppm), and there were very few deaths 

2852 from specific types of cancer, which limit the statistical power of the study to examine 

2853 associations among dichloromethane and specific cancers. Other limitations, as were also noted 

2854 in the Kodak cohort studies, include the use of mortality rather than incidence to define risk, the 

2855 reliance solely on underlying causes of death from death certificates to classify specific cancer 

2856 types and the lack of information on breast cancer risk. 

2857 

2858 4.1.3.4. Cellulose Triacetate Fiber Production Cohorts 

2859 4.1.3.4.1. Cellulose triacetate fiber production—Rock Hill, South Carolina (Hoechst Celanese 

2860 Corporation). Two cohorts of cellulose triacetate fiber workers have been studied in Rock Hill, 

2861 South Carolina (Lanes et al., 1993, 1990; Ott et al., 1983a, b), and Cumberland, Maryland (Gibbs 

2862 et al., 1996; Gibbs, 1992). Workers were exposed to dichloromethane, methanol, and acetone in 

2863 both facilities. 

2864 Ott et al. (1983a, b) conducted a retrospective cohort mortality study of 1,271 acetate 

2865 fiber production workers (551 men and 720 women) employed at least 3 months from 1954 to 
71

2866 1977 at Dow Chemical Company, Rock Hill, South Carolina. This analysis focused on ischemic 

2867 heart disease mortality risk, and there was no presentation of cancer risk. The Rock Hill cohort 

2868 study was updated twice, through September 30, 1986 (Lanes et al., 1990), and December 31, 

2869 1990 (Lanes et al., 1993), and analyses of cancer mortality risks were included in these later 

2870 reports. Causes of death information was obtained from death certificates, with coding based on 

2871 the underlying and contributing causes (Ott et al., 1983a). The referent used in the updates was 

2872 the general population of York County, South Carolina, and analyses were adjusted for age, race, 

2873 gender, and calendar period. Because the results of the mortality risk analyses were similar for 

2874 both updates, those from the 1993 paper are presented here. The mean duration of work in the 

2875 cohort was not reported, but 56% worked for fewer than 5 years (calculated from Tables 3 and 

2876 4 of Ott et al., 1983b). The mean duration of follow-up was 23.6 years in the analysis through 

2877 1986 (Lanes et al., 1990) but was not reported in the later paper (Lanes et al., 1993). The 

2878 1993 report added approximately 4.25 years of follow-up, which would result in an estimate of 

2879 approximately 28 years follow-up for this report. 

2880 The Rock Hill, South Carolina, plant began producing cellulose triacetate fiber in 1954. 

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

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

2883 measurements taken in 1977–1978 were assumed to be representative of operations since 

2884 dichloromethane use began in 1954, based on review of processing operations. The median 

2885 8-hour TWA exposures were estimated as 140, 280, and 475 ppm in the low, moderate, and high 

2886 categories of exposure (Ott et al., 1983a). Employment records provided information on jobs 

2887 held and dates employed, and this was used in conjunction with the exposure estimates for 

2888 specific jobs and work areas to classify individual exposures. However, detailed work history 

2889 information was only available for 475 (37%) of the workers (Lanes et al., 1990), and it is not 

2890 clear how the exposure assessment was applied to workers with missing job history data. 

2891 Methanol was also used in the cellulose triacetate fiber production process, and methanol 

2892 exposure was estimated as one-tenth that of dichloromethane. Acetone exposure was used in the 

2893 production of acetate (cellulose diacetate) fiber at an adjacent part of the plant. The exposure to 

2894 acetone was inversely related to that of dichloromethane, with estimated median 8-hour TWAs 

2895 of 1,080 ppm acetone in the low dichloromethane exposure group and 110 ppm acetone in the 

2896 moderate and high dichloromethane groups in the Rock Hill plant (Ott et al., 1983a). 

2897 In the latest follow-up (Lanes et al., 1993), there was no increase in mortality risk from 

2898 cancer (all sites) or from cancer of the lung or pancreas (Table 4-7). The SMR for liver and bile 

2899 duct cancer, based on four observed cases, was 2.98 (95% CI 0.81, 7.63). This was lower than 

2900 the SMR of 5.75 (95% CI 1.82, 13.8) that was reported in the 1990 analysis, based on these same 

2901 four cases but on a shorter follow-up period (and thus lower number of expected cases). Three 

2902 of these cases were bile duct cancers. This was the first cohort study that included women, and 

2903 this study provided data on breast cancer risk. There were 3 observed breast cancer deaths 
72

2904 compared with 5.59 expected, yielding an SMR of 0.54 (95% CI 0.11, 1.57). No data were 

2905 provided pertaining to reproductive risk factors (e.g., pregnancy history) for breast cancer among 

2906 the women in this cohort, so it is difficult to assess whether these potential confounders are likely 

2907 to have been distributed differently in the cohort compared with the referent group. Information 

2908 about brain cancer, Hodgkin’s lymphoma, and leukemia (Table 4-7) was not included in this 

2909 report but was included in the report by Gibbs (1992) (see Table 11 of that report). 

2910 

Table 4-7. Mortality risk in Hoechst Celanese Corporation cellulose 

triacetate fiber production workers, Rock Hill, South Carolina: 1,271 men 

and women employed 1954–1977, followed through 1990

Cancer type Observed Expected SMRb 95% CIb, 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 CNSa

 1 1.5 0.67 0.2–3.71 

 Hodgkin’s lymphomaa 0 0.24 – – 

 Leukemiaa 1 1.11 0.90 0.02–5.0 

 Breast cancer (women) 3 5.59 0.54 0.11–1.57 

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

b

SMRs and CIs were not calculated for categories with zero observed cases. 

cCIs were calculated from Breslow and Day (1987, Table 2.10). 

Source: Lanes et al. (1993). 

2911 

2912 

2913 There are a number of limitations of this study, including the small size of the cohort, 

2914 small number of observed cancer deaths, availability of detailed work history information for 

2915 only 37% of the workers, and use of mortality rather than incidence data. The exposure levels at 

2916 this plant were high, but the duration of exposure for most of the cohort was relatively short 

2917 (<5 years). It is the first cohort study, however, that included women and provided information 

2918 on breast cancer risk. 

2919 

2920 4.1.3.4.2. Cellulose triacetate fiber production—Cumberland, Maryland (Hoechst Celanese 

2921 Corporation). Gibbs et al. (1996) studied a cohort of 2,909 cellulose triacetate fiber production 

2922 workers (1,931 men and 978 women) at a Hoechst Celanese plant in Cumberland, Maryland. 

2923 This retrospective cohort mortality study included all workers who were employed on or after 

2924 January 1, 1970, and who worked at least 3 months. This study also included a very small 

2925 comparison group (256 men, 46 women) that was described as a “0” or “no” exposure group of 

2926 workers at the plant who worked in jobs that were not considered to have had dichloromethane 

2927 exposure, for a total of 2,187 men and 1,024 women in the exposed and nonexposed groups. 
73

2928 The plant closed in 1981, and mortality was followed through 1989. Since 1955, 

2929 employees of this plant were exposed to dichloromethane, methanol, acetone, and finishing oils 

2930 used as lubricants. Before 1955, acetone was the only exposure. Industrial hygiene monitoring 

2931 focusing on dichloromethane, acetone, and methanol began in the late 1960s. Exposure 

2932 groupings (low, 50–100 ppm, and high, 350–700 ppm) were assigned by area in which 

2933 employees worked. The extrusion and spinning workers and jet wipers were among the high 

2934 exposure group (300–1,250 ppm 8-hour TWA). The SMR analysis that was reported used 

2935 Allegany County, Maryland, as the comparison group. Cause of death information was obtained 

2936 from death certificates, but the authors did not state whether they used underlying or underlying 

2937 and contributing cause of death information. The mean duration of work in the cohort was not 

2938 reported. The total follow-up period included 49,828 person-years (16,292 in the high exposure 

2939 group and 33,536 in the low exposure group), and the mean duration of follow-up was 17.2 years 

2940 (range 8–20 years). These data were found in Hearne and Pifer (1999, Table 7). 

2941 There was little evidence of an increase in mortality risk from cancer (all sites) or from 

2942 cancer of the liver and bile duct, pancreas, or brain in men or in women (Table 4-8). An 

2943 increasing risk with increasing exposure level was seen for prostate cancer mortality in men. 

2944 The p-value for the trend was not given, but the authors describe it as a “nonstatistically 

2945 significant dose-response relationship.” A statistically significant SMR for prostate cancer death 

2946 was seen in the 350–700 ppm group when latency (at least 20 years since first exposure) was 

2947 included in the analysis (SMR = 2.08, p < 0.05). Cervical cancer mortality risk was increased, 

2948 but the small number of cases in the high exposure group did not allow a precise assessment of 

2949 the pattern with respect to exposure level. There was no increased risk of breast cancer. 

2950 
74

2951 

Table 4-8. Cancer mortality risk in Hoechst Celanese Corporation cellulose 

triacetate fiber production workers, Cumberland, Maryland: 2,909 men and 

women employed 1970–1981, followed through 1989 

 Men (n = 1,931) Women (n = 978) 

Cancer type, 

exposure levela

 Obsb Expb SMR 95% CIc Obsb Expb SMR 95% CIc

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 

Braina 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’sa 

50–100 ppm 1 0.4 2.5 0.06–13.9 0 0.23 

350–700 ppm 0 0.41 0 0.02 

Leukemiaa 

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 

Breasta 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 

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

bObs = number of observed deaths, Exp = number of expected deaths. Referent group = Allegany County, 

Maryland. SMRs and CIs were not calculated for categories with zero observed cases. 

cCIs were calculated from Breslow and Day (1987, Table 2.10). 

Sources: Gibbs et al. (1996); Gibbs (1992). 

2952 

2953 

2954 A primary limitation of this study is that workers who were exposed before 1970 but 

2955 were not working at the plant in 1970 were not included in the cohort. The authors had 

2956 attempted to create a cohort of all workers who were employed on or after January 1, 1954, but 

2957 problems with the completeness of the personnel file made it impossible to use this study design. 
75

2958 From what the author (Gibbs, 1992) was able to determine, the records of workers who had died, 

2959 left the company, or retired before the mid to late 1960s (when a new personnel system was 

2960 developed) were not available. Additional limitations include the small size of the cohort, small 

2961 number of observed cancer deaths, and use of mortality (death certificate) data. This is 

2962 particularly problematic for cancers with relatively high survival rates (such as prostate cancer 

2963 and cervical cancer), since incidence rates are not estimated well by mortality rates in this 

2964 situation. 

2965 

2966 4.1.3.5. Solvent-Exposed Workers—Hill Air Force Base, Utah 

2967 Spirtas et al. (1991) and Blair et al. (1998) evaluated exposure to dichloromethane in 

2968 relation to mortality risk in successive retrospective cohort studies of 14,457 civilian workers 

2969 employed at Hill Air Force Base in Utah for at least 1 year from 1952–1956. The analysis was 

2970 limited to the workers that were white or who had missing data on race, resulting in a sample 

2971 size of 14,066 (10,461 men, 3,605 women). Spirtas et al. (1991) examined mortality through 

2972 1982 (3,832 deaths), and Blair et al. (1998) updated mortality through 1990 (4,195 deaths). The 

2973 underlying and contributing causes of death information from death certificates was used to 

2974 classify cause-specific mortality. SMRs were calculated by using mortality rates from the Utah 

2975 population, and an internally standardized life table method was used to adjust for age at entry 

2976 into the cohort and competing causes of death. In the Blair et al. (1998) analysis, adjusted 

2977 relative risks (rate ratios) were estimated from a Poisson regression analysis with unexposed 

2978 workers as the referent. The mean duration of work was not reported. In the analysis through 

2979 1982 (Spirtas et al., 1991), there were 22,770 person-years of follow-up in men and 

2980 3,091 person-years of follow-up in women who were classified as exposed to dichloromethane. 

2981 The total number of workers classified as exposed to dichloromethane was 1,222 (Stewart et al., 

2982 1991), which would yield an estimated mean of approximately 21 years follow-up through 1982. 

2983 The total number of person-years included in the later report (Blair et al., 1998), with the 

2984 addition of 8 more years of follow-up, was not reported but would be expected to increase the 

2985 mean follow-up time to approximately 29 years. 

2986 Two industrial hygienists developed the exposure assessment based on walkthrough 

2987 surveys, interviews with management and labor representatives, review of historical records, job 

2988 descriptions, monitoring data and other information pertaining to chemicals used, and 

2989 organization of the work site (Blair et al., 1998; Spirtas et al., 1991). Each worker was assigned 

2990 exposure by using information on the worker’s job history, which included job titles, department 

2991 codes, and dates of employment. The most detailed exposure assessment was done for 

2992 trichloroethylene, the primary focus of the study. Dichloromethane, one of 25 other exposures 

2993 analyzed, was classified as a dichotomous exposure (ever exposed, never exposed). 

2994 Blair et al. (1998) presented the mortality risk for three specific cancers in relation to 

2995 15 of the 25 chemicals classified as dichotomized exposures. The rate ratios for non-Hodgkin’s 
76

2996 lymphoma and multiple myeloma in relation to dichloromethane in men were 3.0 (95% CI 0.9, 

2997 10.0) and 3.4 (95% CI 0.9, 13.2), respectively. These rate ratios, (particularly those for multiple 

2998 myeloma), were considerably higher than the rate ratios for any of the other chemicals examined, 

2999 in which the next highest observed rate ratio was 1.8 for Freon. No cases of either of these 

3000 cancers were observed in women with dichloromethane exposure, but the rate ratio for breast 

3001 cancer in these women was 3.0 (95% CI 1.0–8.8). Associations of similar magnitude (rate ratios 

3002 of 3.0–4.0) were also seen among breast cancer and some other exposures (Freon, solder flux, 

3003 isopropyl alcohol, and trichloroethane). 

3004 This is the largest of the cohort studies that were identified that included women and 

3005 specifically reported data pertaining to breast cancer risk. The major limitation of this study is 

3006 that the exposure assessment for dichloromethane was based on a dichotomized classification. In 

3007 addition, exposure to many different types of solvents was common; thus, it is difficult to 

3008 completely separate the effects of individual exposures. Some aspects of reproductive history, 

3009 such as age at first pregnancy, are known risk factors for breast cancer. Reproductive history 

3010 was not included in this analysis, but Blair et al. (1998) note that it is unlikely that these factors

3011 would confound the results of a few specific chemicals, since the referent group was an internal 

3012 group within the cohort (and thus would be expected to be similar in terms of socioeconomic 

3013 status) and there was no association overall between solvent exposures and breast cancer 

3014 mortality. 

3015 

3016 4.1.3.6. Case-Control Studies of Specific Cancers and Dichloromethane 

3017 Seven site-specific cancer case-control studies included dichloromethane as an exposure 

3018 of interest. These studies involve six cancer sites: brain and CNS (Cocco et al., 1999; Heineman 

3019 et al., 1994), breast (Cantor et al., 1995), kidney (Dosemeci et al., 1999), pancreas (Kernan et al., 

3020 1999), rectum (Dumas et al., 2000), and childhood leukemia (Infante-Rivard et al., 2005). A 

3021 synopsis of cohort studies in humans is provided in Table 4-9. 

3022 

3023 4.1.3.6.1. Case-control studies of brain cancer. Heineman et al. (1994) studied the association 

between astrocytic brain cancer (International Classification of Diseases 9th 3024 ed. [ICD-9] codes 

3025 191, 192, 225, and 239.7) and occupational exposure to chlorinated aliphatic hydrocarbons. 

3026 Cases were identified by using death certificates from southern Louisiana, northern New Jersey, 

3027 and the Philadelphia area. This analysis was limited to white males who died between 1978 and 

3028 1981. Controls were randomly selected from the death certificates of white males who died of 

3029 causes other than brain tumors, cerebrovascular disease, epilepsy, suicide, and homicide. The 

3030 controls were frequency matched to cases by age, year of death, and study area.

3031 Next of kin were successfully located for interview for 654 cases and 612 controls, which 

3032 represents 88 and 83% of the identified cases and controls, respectively. Interviews were 

3033 completed for 483 cases (74%) and 386 controls (63%). There were 300 cases of astrocytic 
77

3034 brain cancer (including astrocytoma, glioblastoma, mixed glioma with astrocytic cells). The 

3035 ascertainment of type of cancer was based on review of hospital records, which included 

3036 pathology reports for 229 cases and computerized tomography reports for 71 cases. After the 

3037 exclusion of 66 controls with a possible association between cause of death and occupational 

3038 exposure to chlorinated aliphatic hydrocarbons (some types of cancer, cirrhosis of the liver), the 

3039 final analytic sample consisted of 300 cases and 320 controls. 

3040 In the next-of-kin interviews, the work history included information about each job held 

3041 since the case (or control) was 15 years old (job title, description of tasks, name and location of 

3042 company, kinds of products, employment dates, and hours worked per week). Occupation and 

3043 industry were coded based on four-digit Standard Industrial Classification and Standard 

3044 Occupational Classification (Department of Commerce) codes. The investigators developed 

3045 matrices linked to jobs with likely exposure to dichloromethane, five other chlorinated aliphatic 

3046 hydrocarbons (carbon tetrachloride, chloroform, methyl chloroform, tetrachloroethylene, and 

3047 trichloroethylene), and organic solvents (Gomez et al., 1994). This assessment was done blinded 

3048 to case-control status. Exposure was defined as the probability of exposure to a substance (the 

3049 highest probability score for that substance among all jobs), duration of employment in the 

3050 exposed occupation and industry, specific exposure intensity categories, average intensity score 

3051 (the three-level semiquantitative exposure concentration assigned to each job multiplied by 

3052 duration of employment in the job, summed across all jobs), and cumulative exposure score 

3053 (weighted sum of years in all exposed jobs with weights based on the square of exposure 

3054 intensity [1, 2, 3] assigned to each job). Secular trends in the use of specific chemicals were 

3055 considered in the assignment of exposure potential. Exposures were lagged 10 or 20 years to 

3056 account for latency. Thus, this exposure assessment procedure was quite detailed. 

3057 Adjusting for age and study area, the OR for the association between any exposure to 

3058 dichloromethane and risk of astrocytic brain cancer was 1.3 (95% CI 0.9, 1.8). There was a 

3059 statistically significant trend (p < 0.05) with increasing probability of exposure to

3060 dichloromethane with an OR = 1.0 (95% CI, 0.7, 1.6) for low probability, OR = 1.6 (95% CI 0.8, 

3061 3.0) for medium probability, and OR = 2.4 (95% CI 1.0, 5.9) for high probability compared with 

3062 the referent group of unexposed men. An increased risk with higher duration of exposure was 

3063 also observed, with OR = 1.7 (95% CI 0.9, 3.6) for 21 or more years of work in exposed jobs for 

3064 all exposed workers and OR = 6.1 (95% CI 1.1, 43.8) for the combination of 21 years or more of 

3065 work in a high probability of exposure job. Similar results were seen in additional analyses, 

3066 controlling for age, study area, employment in electronics occupations and industries, and 

3067 exposure to carbon tetrachloride, tetrachloroethylene, and trichloroethylene. There was also 

3068 evidence of an association between astrocytic brain cancer risk and dichloromethane exposure, 

3069 based on the average intensity score, with an OR = 1.1 (95% CI 0.7, 1.7) for the low-medium 

3070 intensity group and an OR = 2.2 (95% CI 1.1, 4.1) for the high intensity group, and trend p-value 

3071 <0.05. The combination of high intensity and high duration (21 or more years) was strongly 
78

3072 associated with risk (OR = 6.1 [95% CI 1.5, 28.3]), and a weaker association (OR = 1.4 [95% CI 

3073 0.6, 3.2]) was seen for high intensity and shorter duration (2–20 years). The association between 

3074 cumulative exposure score (low, medium, and high) and astrocytic brain cancer risk was 

3075 nonlinear (ORs of 0.9, 1.9, and 1.2 in the low, medium, and high exposure categories, 

3076 respectively). 

3077 The strengths of this case-control study include a large sample size, detailed work 

3078 histories (including information not just about usual or most recent industry and occupation but 

3079 also about tasks and products for all jobs held since age 15), and comprehensive exposure 

3080 assessment and analysis along several different dimensions of exposure. The major limitations 

3081 were the lack of direct exposure information and potential inaccuracy of the descriptions of work 

3082 histories that were obtained from next-of-kin interviews. Heineman et al. (1994) acknowledge 

3083 these limitations in the report, and, in response to a letter by Norman and Boggs (1996) 

3084 criticizing the methodology and interpretation of the study, Heineman et al. (1996) noted that, 

3085 while the lack of direct exposure information must be interpreted cautiously, it does not 

3086 invalidate the results. Differential recall bias between cases and controls was unlikely because 

3087 work histories came from next-of-kin for both groups, the industrial hygienists made their 

3088 judgments blinded to disease status, and the strong associations that were seen with the exposure 

3089 measures for dichloromethane were not seen with the other solvents included in the analysis. 

3090 The relatively strong and statistically significant associations between dichloromethane and 

3091 astrocytic brain tumors were seen along multiple measures of exposure, suggesting that the 

3092 results were unlikely to be spurious. Nondifferential misclassification would, on average, 

3093 attenuate true associations and would be unlikely to result in the types of exposure-response 

3094 relationships that were observed in this study. 

3095 Norman and Boggs (1996) described an apparent inconsistency in the estimated trends in 

3096 dichloromethane and carbon tetrachloride exposure based on the methodology used in this case3097 control study (described in more detail in Gomez et al. [1994]). In response, Gomez (1996) 

3098 noted that the apparent inconsistency was actually due to an error in the labeling of the lines on 

3099 one of the figures in the report rather than an inconsistency with the estimated trends. Another 

3100 point raised by Norman and Boggs (1996) was that the Heineman et al. (1994) findings were 

3101 surprising in light of the lack of brain carcinogenesis in animals. In response, Heineman et al. 

3102 (1996) pointed out that carcinogens commonly cause different cancers in animals and humans. It 

3103 can also be noted that brain tumors are exceedingly rare in animal bioassays (Sills et al., 1999). 

3104 Norman and Boggs (1996) also suggested that the results of the Heineman et al. (1994) study be 

3105 given no weight when compared with the results of the cohort studies. The authors responded by 

3106 pointing out that the cohort studies had low statistical power and large CIs around their point 

3107 estimates but were not inconsistent with an association between dichloromethane and brain 

3108 cancer (Heineman et al., 1996). This point is strengthened further by the more recent results 

3109 from the Rochester, New York, Eastman Kodak cohort (Hearne and Pifer, 1999), described 
79

3110 previously, since an increased SMR for brain and CNS cancers was seen in the longer follow-up 

3111 period of this cohort. 

3112 In another case-control study of brain cancer and dichloromethane exposure, Cocco et al. 

3113 (1999) identified 12,980 female cases of cancer of the brain and CNS through the underlying 

3114 cause of death listing (ICD codes 191 and 192) on death certificates from 24 states from 1984 to 

3115 1992. (This collection of death certificates is a data set created by the National Center for Health 

3116 Statistics, NIOSH, and the National Cancer Institute to facilitate research on occupational 

3117 exposures and mortality risk.) The cases included 161 women with meningioma (ICD-9 codes 

3118 192.1, 192.3). Four women who died of nonmalignant diseases, excluding neurological 

3119 disorders, were chosen as controls for each case. The controls were frequency matched to the 

3120 cases by state, race, and 5-year age group. Occupation data were based on the occupation fields 

3121 in the death certificates. This job was coded based on the three-digit industry and three-digit 

3122 occupation (Department of Census) codes. The investigators developed job exposure matrices 

3123 that were applied to these industry/occupation codes. The job exposure matrices included 

3124 probability and intensity scores for 11 occupational hazards, one of which was dichloromethane, 

3125 but also included other solvents, electromagnetic fields, chlorinated aliphatic hydrocarbons, 

3126 benzene, lead, nitrosamines, insecticides, herbicides, and public contact. The investigators used 

3127 logistic regression models to estimate ORs, adjusting for each workplace exposure, marital 

3128 status, three levels of socioeconomic status (based on occupation), and age at death. For each 

3129 chemical, four levels of intensity and probability were defined (unexposed, low, medium, and 

3130 high). 

3131 A weak association between dichloromethane exposure and brain/CNS cancer was seen 

3132 (OR 1.2 [95% CI 1.1, 1.3]) (Cocco et al., 1999). There was no exposure-related trend in the 

3133 association between probability or intensity of exposure and brain cancer. A similar but more 

3134 imprecise association was seen with meningioma cancer (OR 1.2 [95% CI 0.7, 2.2]). There were 

3135 too few cases of meningioma to stratify by exposure probability and intensity. 

3136 The major limitations of this study are the use of mortality rather than incidence data and 

3137 the reliance on occupation data from death certificates. The death certificate occupation data are 

3138 based on “usual” occupation, which may be more prone to misclassification in studies of women 

3139 because of gender-related differences in work patterns (i.e., shorter duration jobs for women 

3140 compared with men). A relatively broad job exposure matrix was applied to the job information, 

3141 and typically more generic job exposure matrices result in less sensitive assessment with limited 

3142 ability to detect exposure-response trends (Teschke et al., 2002). Nondifferential 

3143 misclassification of outcome and exposure would generally result in attenuated effect estimates. 

3144 

3145 4.1.3.6.2. Case-control studies of breast cancer. Cantor et al. (1995) conducted a case-control 

3146 study of occupational exposures and breast cancer, using the 24 state (1984–1989) death 

3147 certificate data described in the previous section. Cases were women with breast cancer coded as 
80

3148 the underlying cause of death (ICD-9 code 174). Four female controls per case were selected 

3149 from all noncancer deaths, frequency matched by age (5-year age groups) and ethnicity (black, 

3150 white). The occupation listed on the death certificate was coded based on the three-digit industry 

3151 and three-digit occupation (Department of Census) codes, and this was used with a job exposure 

3152 matrix developed by the investigators to assess 31 workplace exposures, one of which was 

3153 dichloromethane. Four exposure probability and three exposure level scores were assigned. 

3154 ORs for probability and level were calculated for each ethnic group, adjusting for age at death 

3155 and a measure of socioeconomic status (based on occupation). After excluding subjects whose 

3156 death certificate occupations were listed as homemaker, there were 29,397 white cases and 

3157 4,112 black cases (total 33,509) and 102,955 white controls and 14,839 black controls (total 

3158 117,794). 

3159 There was little evidence of an association between exposure probability and breast 

3160 cancer mortality using the probability exposure metric. The ORs were 1.05 (95% CI 0.97, 1.1) 

3161 and 0.76 (95% CI 0.3, 2.0) in probability level 3 and level 4, respectively, for white women and 

3162 1.13 (95% CI 0.9, 1.4) in probability level 3 for black women. (There were too few black 

3163 women in exposure probability level 4 for analysis.) Weak associations were seen with exposure 

3164 level. In white women, an OR of 1.17 (95% CI 1.1, 1.3) was seen with the highest exposure 

3165 level, and in black women the OR in this exposure group was 1.46 (95% CI 1.2, 1.7). In the 

3166 analysis that jointly considered exposure level and probability ratings but excluded the lowest 

3167 probability of exposure, the OR for the highest category of exposure level was 1.28 in whites 

3168 (p < 0.05) and 1.21 in blacks. 

3169 As with the Cocco et al. (1999) case-control study that used a similar methodology, the 

3170 limitations of this study include the use of an outcome defined by mortality rather than incidence, 

3171 use of usual occupation information as recorded in death certificates, and use of a very broad job 

3172 exposure matrix to classify 31 different exposures. Although information on pregnancy and 

3173 lactation history (known risk factors for breast cancer) was not available, the authors did adjust 

3174 for socioeconomic status by using the occupation data, which may have corrected for some of the 

3175 potential confounding due to reproductive history. 

3176 

3177 4.1.3.6.3. Case-control studies of pancreatic cancer. Kernan et al. (1999) conducted a case3178 control study of 63,097 pancreatic cancer cases, using the 24-state (1984–1993) death certificate 

3179 data. The diagnosis of pancreatic cancer was based on underlying cause of death (ICD-9 code 

3180 157). Four controls who had died during the same time period of causes other than cancer were 

3181 selected for each case, frequency-matched by state, race, gender, and 5-year age group 

3182 (n = 252,386). Usual occupation and industry, based on the occupation data in the death 

3183 certificate, were coded by using the three-digit (Department of Census) codes. A job-exposure 

3184 matrix was used with the industry and occupation codes to evaluate exposure intensity and 

3185 probability (each categorized as high, medium, or low) for formaldehyde, dichloromethane, 
81

3186 10 other solvents, and a combined “organic solvents” measure. Race- and gender- specific 

3187 analyses were conducted by using logistic regression to estimate ORs and 95% CIs, adjusting for 

3188 age, marital status (ever, never married), residential area (metropolitan, nonmetropolitan), and 

3189 region (east, south central, south, and west). 

3190 The point estimates for the ORs in the low, medium, and high intensity categories in the 

3191 four race-gender groups ranged from 0.8 to 1.3, with no exposure-effect trend seen in any group. 

3192 The only statistically significant OR was for high exposure intensity in white females (OR 1.3 

3193 [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 

3194 low intensity in this group. An elevated OR was seen with high exposure probability in black 

3195 males (OR 2.2 [95% CI 1.0, 4.8]) but not in white females (OR 1.0 [95% CI 0.8, 1.4]) or white 

3196 males (OR 1.0 [05% CI 0.8, 1.3]), and the ORs were 0.9 for medium exposure probability in 

3197 these three groups. There were relatively few black females in this study, resulting in imprecise 

3198 estimates (OR 2.0 [95% CI 0.8, 5.4] for medium exposure and OR 1.5 [95% CI 0.6, 3.6] for high 

3199 exposure). 

3200 The limitations of this study, as with the other case-control studies that used the 24-state 

3201 death certificate data set, include the reliance on cause of death data from death certificates rather

3202 than medical-record validated incidence data and the use of death certificate occupation data. 

3203 The job exposure matrix used with the occupation data was more focused than those used in 

3204 Cocco et al. (1999) and Cantor et al. (1995). Although the analysis adjusted for some 

3205 sociodemographic characteristics, it did not include measures of smoking history or diabetes, 

3206 which are known risk factors for pancreatic cancer (Lowenfels and Maisonneuve, 2005). 

3207 

3208 4.1.3.6.4. Case-control studies of renal cancer. Dosemeci et al. (1999) reported data from a 

3209 population-based case-control study of the association between occupation exposures and renal 

3210 cancer risk. The investigators identified newly diagnosed patients with histologically confirmed 

3211 renal cell carcinoma from the Minnesota Cancer Surveillance System from July 1, 1988, to 

3212 December 31, 1990. The study was limited to white cases, and age and gender-stratified controls 

3213 were ascertained by using random digit dialing (for subjects ages 20–64) and from Medicare 

3214 records (for subjects 65–85 years). Of the 796 cases and 796 controls initially identified, 

3215 438 cases (273 men, 165 women) and 687 controls (462 men, 225 women) with complete 

3216 personal interviews were included in the occupational analysis. 

3217 Data were obtained through in-person interviews that included demographic variables, 

3218 residential history, diet, smoking habits, medical history, and drug use. The occupational history 

3219 included information about the most recent and usual industry and occupation (coded using the 

3220 standard industrial and occupation codes, Department of Commerce), job activities, hire and 

3221 termination dates, and full- and part-time status. A job exposure matrix developed by the 

3222 National Cancer Institute was used with the coded job data to estimate exposure status to 

3223 dichloromethane and eight other chlorinated aliphatic hydrocarbons. 
82

3224 ORs were adjusted for age, smoking, hypertension and use of drugs for hypertension, and 

3225 body mass index. No association between renal cell carcinoma and exposure to dichloromethane 

3226 was observed in men (OR 0.85 [95% CI 0.6, 1.2]), women (OR 0.95 [95% CI 0.4, 2.2]), or both 

3227 sexes combined (OR 0.87 [95% CI 0.6, 1.2]). 

3228 A strength of this study includes the use of incident cases of renal cancer from a defined 

3229 population area, with confirmation of the diagnosis using histology reports. The occupation 

3230 history was based on usual and most recent job, in combination with a relatively focused job 

3231 exposure matrix. In contrast to the type of exposure assessment that can be conducted in cohort 

3232 studies within a specific workplace, however, exposure measurements, based on personal or 

3233 workplace measurements, were not used, and a full lifetime job history was not obtained. 

3234 

3235 4.1.3.6.5. Case-control studies of rectal cancer. Dumas et al. (2000) reported data from a case3236 control study of occupational exposures and rectal cancer conducted in Montreal, Quebec, 

3237 Canada. The investigators identified 304 newly diagnosed cases of primary rectal cancer, 

3238 confirmed on the basis of histology reports, between 1979 and 1985; 257 of these participated in 

3239 the study interview. One control group (n = 1,295) consisted of patients with other forms of 

3240 cancer (excluding lung cancer and other intestinal cancers) recruited through the same study 

3241 procedures and time period as the rectal cancer cases. A population-based control group 

3242 (n = 533), frequency matched by age strata, was drawn by using electoral lists and random digit 

3243 dialing. The occupational assessment consisted of a detailed description of each job held during 

3244 the working lifetime, including the company, products, nature of work at site, job activities, and 

3245 any additional information from the interviews that could furnish clues about exposure. The 

3246 percentage of proxy respondents was 15.2% for cases, 19.7% for other cancer controls, and 

3247 12.6% for the population controls. 

3248 A team of industrial hygienists and chemists blinded to subjects’ disease status translated 

3249 jobs into potential exposure to 294 substances with three dimensions (degree of confidence that 

3250 exposure occurred, frequency of exposure, and concentration of exposure). Each of these 

3251 exposure dimensions was categorized into none, any, or substantial exposure. Logistic 

3252 regression models adjusted for age, education, proxy versus subject responder status, cigarette 

3253 smoking, beer drinking, and body mass index. Using the cancer control group, the OR for any 

3254 exposure to dichloromethane was 1.2 (95% CI 0.5, 2.8) and the OR for substantial exposure 

3255 (confident that exposure occurred with 5 or more years of exposure at medium or high frequency 

3256 and concentration) was 3.8 (95% CI 1.1, 12.2). The results using the population-based control 

3257 group for this exposure were not presented. 

3258 The strengths of this study were the large number of incident cases, specific information 

3259 about job duties for all jobs held, and a definitive diagnosis of rectal cancer. However, the use of 

3260 the general population (rather than a known cohort of exposed workers) reduced the likelihood 

3261 that subjects were exposed to dichloromethane, resulting in relatively low statistical power for 
83

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

3263 was used to evaluate 294 chemicals. 

3264 4.1.3.6.6. Case-control studies of childhood leukemia. Infante-Rivard et al. (2005) examined 

3265 the association between maternal occupational exposures, before and during pregnancy, and risk 

3266 of childhood acute lymphoblastic leukemia (ICD-9 code 204.0) by using data from a population3267 based case-control study in Quebec, Canada. Incident cases diagnosed from 1980–2000 were 

3268 identified from the cancer hospitals in the province, and diagnosis was confirmed based on 

3269 clinical records from an oncologist or hematologist. Between 1980 and 1993, cases ages 0–

3270 9 years at diagnosis were included, and from 1994 to 2000 the age range was expanded to 

3271 14 years. The number of eligible cases identified was 848, and, of these, 790 parents (93%) 

3272 participated in the study. Population-based controls, individually matched to the sex and age at 

3273 diagnosis of the cases, were identified from government registries of all children in the province 

3274 (1980–1993) and the universal health insurance files (1994–2000). The parents of 790 (86%) of 

3275 the 916 eligible controls who were identified participated in the study. 

3276 Data were collected by using a structured telephone interview. Some information (i.e., 

3277 job title, dates, type of industry, industry name and address) was obtained for all jobs held since 

3278 age 18, and additional information (e.g., materials and machines used, typical activities) was 

3279 obtained for jobs held by the mother from 2 years before the pregnancy through the birth of the 

3280 child. Specialized exposure modules were also used to collect information about specific jobs 

3281 (e.g., nurse, waitress, hair dresser, textile dry cleaner). All of this information was reviewed by 

3282 chemists and industrial hygienists, blinded to case-control status, to classify exposure to over 

3283 300 chemicals, although the primary focus of the study was on solvents (21 individual 

3284 substances, including dichloromethane, and six mixtures). The exposure assessment included 

3285 ratings of confidence (possible, probable, and definite), frequency of exposure during a normal 

3286 workweek (<5, 5–30, or >30% of the time), and level of concentration (low = slightly above 

3287 background, high = highest possible exposure in the study population, and medium for in3288 between levels). 

3289 A weak association was seen between any dichloromethane exposure during the 2 years 

3290 before pregnancy up to the birth and risk of leukemia in the child (OR 1.34 [95% CI 0.54, 3.34]), 

3291 and results were similar when limited to exposures during pregnancy. Stronger associations 

3292 were seen with probable or definite exposure (OR 3.22 [95% CI 0.88, 11.7]) compared with 

3293 possible or no exposure. The estimates for categories based on concentration and frequency 

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

3295 

3296 4.1.3.7. Summary of Cancer Studies by Type of Cancer 

3297 The cohort and case-control studies with data relevant to the issue of dichloromethane 

3298 exposure and cancer risk are summarized in Tables 4-9 and 4-10, respectively. The strongest of 

3299 the cohort studies, in terms of design, are two of the triacetate film base production cohorts 
84

3300 (Cohort 1 in New York and the United Kingdom cohort, reported in Hearne and Pifer [1999] and 

3301 Tomenson et al. [1997], respectively). These are the cohorts with the most extensive exposure 

3302 assessment information. The start of eligibility for cohort entrance corresponds with the 

3303 beginning of the time when the exposure potential at the work site began, and the follow-up 

3304 period is relatively long (mean >25 years). Although Cohort 2 of the New York film base 

3305 production study has similar exposure data and follow-up, this cohort was limited to workers 

3306 employed between 1964 and 1970 and therefore would have missed anyone leaving (possibly 

3307 because of illness or death) before this time. In addition, because of the overlap between 

3308 Cohort 1 and Cohort 2, including both cohorts in an evaluation would be double counting 

3309 experiences of some individuals. Several limitations of the triacetate film base production 

3310 cohorts should be noted, however. One of these limitations concerns the generalizability of the 

3311 results, given the relatively low exposure level (mean 8-hour TWA <40 ppm) compared to the 

3312 other cellulose triacetate fiber production cohorts (Gibbs et al., 1996; Ott et al., 1983a). 

3313 Exposures in small, poorly ventilated work areas are also often much higher than those seen in 

3314 these film base production cohorts (Estill and Spencer, 1996; Anundi et al., 1993). Other 

3315 limitations include the limited power to detect a risk of low-incidence cancers (including brain 

3316 and leukemia), the lack of women and thus lack of data pertaining to breast cancer, and the use 

3317 of mortality rather than incidence data. Although the exposure levels in the cohorts involved in 

3318 cellulose triacetate fiber production were much higher than those of the film production cohorts, 

3319 the duration of exposure was relatively short in the South Carolina cohort (Lanes et al., 1993), 

3320 and the majority of workers were missing job history data. In the Maryland triacetate fiber 

3321 production plant, duration of exposure was not reported and the length of follow-up was 

3322 relatively short (mean, 17 years) (Gibbs et al., 1996). Also, the cohort began in 1970, even 

3323 though production began in 1955, and the missing personnel records made it impossible to 

3324 recreate an inception cohort. The exposure assessment in the study of civilian Air Force base 

3325 workers (Blair et al., 1998) allowed for only a dichotomized classification of exposure, and there 

3326 was considerable exposure to other solvents among these workers. This Air Force base study 

3327 was the largest of the cohort studies that included women and presented data pertaining to breast 

3328 cancer. 
85

Table 4-9. Summary of cohort studies of cancer risk and dichloromethane exposure 

Cohort 

Total n, exposure levela and 

duration, length of follow-up Inclusion criteriab

Exposure assessment; 

Outcome assessment 

Hearne and Pifer (1999) Cellulose triacetate film base production; New York  Cohort 1

n = 1,311 men Mean, 39 ppm mean duration, 17 yr mean follow-up, 35 yr 

Began working after 1945; worked at least 1 yr 

Work history (job records) and personal/air monitoring; death certificate (underlying causes) 

Elevated brain canand leukno risk focancer (s

 Cohort 2 

n = 1,013 men mean, 26 ppm mean duration, 24 yr mean follow-up, 26 yr 

Employed at least 1 yr between 1964 and 1970 

(potential exposure began 1946) 

Work history (job records) and personal/air monitoring; death certificate (underlying causes) 

Elevated brain canleukemia(SMRs bfor liver 4) 

Tomenson et al. (1997) Cellulose triacetate film base production; United Kingdom 

n = 1,473 men mean, 19 ppm mean duration, 9 yr mean follow-up, 27 yr 

Employed anytime between 1946 and 1988

Work history (job records) and personal/air monitoring; death certificate (underlying causes) 

Elevated brain canelevation liver, lunTable 4-

Lanes et al. (1993) Cellulose triacetate fiber production; South Carolina 

n = 551 men and 720 women (total n = 1,271); median 140, 280, and 475 ppm in low, moderate, and high, respectively; 56% <5 yr work duration; mean follow-up, ~28 yr 

Worked at least 3 months in the preparation or extrusion areas from 1954 to 1977 

Job history data and personal/air monitoring of specific areas (but job history data available for 37%); death certificate (underlying and contributing causes) 

Elevated cancer (Sin earlier risk for lcancer (s

Gibbs et al. (1996) Cellulose triacetate fiber production; Maryland 

n = 1,931 men and 978 women (total n = 2,909); 50–100 ppm in low and 350–700 ppm in high exposure; duration not reported; mean follow-up 

17 yr 

Employed on or after January 1, 1970, for at least 3 months (potential exposure began 1955) 

Work history (job records) and personal/air monitoring; death certificate (fields used not stated) 

Elevated cancer (mcervical 

3.0), and (high expin men; wno risk focancer (s
86

Table 4-9. Summary of cohort studies of cancer risk and dichloromethane exposure 

Cohort 

Total n, exposure levela and 

duration, length of follow-up Inclusion criteriab

Exposure assessment; 

Outcome assessment 

Blair et al. (1998) Air Force Base, Utah 

n = 10,461 men and 3,605 women (total n = 14,066)d 

dichotomized (yes, no) exposure duration not reported mean follow-up ~29 yr 

Employed at least 1 yr from 1952 to 1956 (potential exposure began 1939) 

Work history (job records) and industrial hygiene assessment based on work site review (dichotomized exposure); (underlying and contributing causes)

Elevated Hodgkinmultiple and brea3.0) (see 

a8-hour TWA.

bIf dichloromethane was used at the plant before the first date of entrance into the cohort, the year that potential exposure began is noted

cResults are described as elevated if SMR was around 1.5 or higher. There is limited statistical power for these cause-specific analyses i

statistical significance of individual estimates is not presented in this table. RR = relative risk. 

dIncludes whites and unknown race. 

3329 
87

3330 

Table 4-10. Summary of case-control studies of cancer risk and dichloromethane exposure 

Cancer type, 

reference 

Location 

n cases, n controls (source), demographic 

group Exposure assessment 

Brain  Heineman et al. (1994) 

Louisiana, New Jersey, Philadelphia 300 cases, 320 controls (death certificates); cancer confirmed by hospital records; white men 

Job exposure matrix applied to detailed information on all jobs held (at least 1 year) since age 15, as obtained from next-of-kin interviews; probability, duration, intensity, and cumulative exposure scores; six solvents evaluated 

OR 1.3 for any ewith increased p

<0.05, OR 2.4 foincreased duratiostrongest effects plus high duratiointensity and higcombinations; no cumulative expo4.1.3.6.1) 

Brain  Cocco et al. (1999) 

24 states, U.S. 12,980 cases, 51,920 controls (death certificates); women 

Job exposure matrix applied to death certificate occupation; probability, and intensity scores; 11 exposures evaluated 

Weak association trend with proba(see section 4.1.3

Breast  Cantor et al. (1995) 

24 states, U.S. 33,509 cases, 117,794 controls (death certificates); black and white women 

Job exposure matrix applied to death certificate job data, probability, and exposure level; 31 substances evaluated 

Little evidence oexposure probabwith exposure leblacks (see sectio

Pancreas  Kernan et al. (1999) 

24 states, U.S. 63,037 cases, 252,386 controls (death certificates); black and white men and women 

Job exposure matrix applied to death certificate occupation, probability, and intensity scores; 11 chlorinated solvents and formaldehyde evaluated 

Little evidence ointensity or prob4.1.3.6.3) 

Kidney  Dosemeci et al. (1999) 

Minnesota 438 incident cases (Minnesota cancer registry), 687 controls (random digit dialing and Medicare records); cancer confirmed by histology; men and women 

Job exposure matrices applied to most recent and usual job, as ascertained from interviews; nine solvents evaluated 

No evidence of iwith dichlorome0.95 in women) 

Rectum  Dumas et al. (2000) 

Montreal, Canada 257 incident cases, 1,295 other cancer controls from 19 hospitals; 533 populationbased controls (electoral rolls and random digit dialing), cancer confirmed by histology; men 

Job exposure matrix applied to detailed 

information on all jobs held, as ascertained from interviews; 294 substances evaluated 

Little evidence oexposure (OR 1.small, “substanti3.8) (using cancepopulation contrexposure) (see se
88

Table 4-10. Summary of case-control studies of cancer risk and dichloromethane exposure 

Cancer type, 

reference 

Location 

n cases, n controls (source), demographic 

group Exposure assessment 

Childhood leukemia (acute lymphoblastic leukemia)  Infante-Rivard et al. 

(2005) 

Quebec, Canada 790 incident cases (hospitals—all provinces), 790 population-based controls (government population registries); cancer based on oncologist or hematologist diagnosis ages 0–14,a both sexes 

Systematic review of detailed information on all jobs held by the mother from 2 years before pregnancy through birth of the child; 21 individual substances and six mixtures evaluated (mostly solvents); confidence, frequency, and concentration of exposure rated 

Little evidence oexposure (OR 1.associations (OR possible/no expodefinite and with frequency and co4.1.3.6.6) 

aFrom 1980 to 1993, study was limited to diagnoses of ages 0–9, but this was expanded between 1994 and 2000 to ages 0–14. 

3331 
89

3332 Case-control studies offer the potential for increased statistical power for assessing 

3333 associations with relatively rare cancers, such as brain cancer and leukemia. Case-control 

3334 studies are often designed to examine incidence rather than mortality, which is of particular 

3335 importance in etiologic research for diseases with relatively high survival rates and diseases in 

3336 which survival may be strongly related to factors that are difficult to adjust for without detailed 

3337 data collection (e.g., access to health care). There is a considerable range in the detail and 

3338 quality of the exposure assessment used in case-control studies, however. Case-control studies 

3339 rarely include specific measurements taken at specific work sites of individual study participants. 

3340 Although it is more difficult to determine absolute exposure levels without these individual 

3341 measurements, the exposure assessment methodology used in case-control studies can result in 

3342 useful between-group comparisons of risk if the intra-group variability is less than the inter3343 group variability in potential exposure levels. Among the case-control studies with data 

3344 pertaining to cancer risk and dichloromethane exposure, the two studies with the strongest 

3345 designs are the study of brain cancer by Heineman et al. (1994) and the study of childhood 

3346 leukemia by Infante-Rivard et al. (2005). These are the studies that obtained detailed 

3347 information about all jobs held (rather than just the usual or most recent job), focused on a 

3348 relatively small number of exposures, and used medical record data to confirm the diagnosis. 

3349 Heineman et al. (1994) obtained the work history from interviews with next-of-kin, however, 

3350 which is most likely to have resulted in nondifferential misclassification of exposure, and thus 

3351 attenuation in the observed associations. The use of death certificate data to classify disease and 

3352 occupational exposures in the three studies using the large 24 state death certificate database 

3353 (brain cancer: Cocco et al. [1999]; breast cancer: Cantor et al. [1995]; pancreatic cancer: Kernan 

3354 et al. [1999]) is also likely to have resulted in nondifferential misclassification of both outcome 

3355 and exposure (and thus attenuated associations). 

3356 Considering the issues described above with respect to the strengths and limitations of the 

3357 available epidemiologic studies, a summary of the epidemiologic evidence relating to 

3358 dichloromethane exposure and specific types of cancer can be made, as described below. The 

3359 available epidemiologic data suggest an association between dichloromethane and brain cancer 

3360 and liver cancer, but not lung cancer. 

3361 

3362 4.1.3.7.1. Brain and CNS cancer. An increased risk of brain and CNS cancers was seen in the 

3363 strongest cohort studies; SMRs were 2.16 in Cohort 1 in New York (Hearne and Pifer, 1999) and 

3364 1.45 in the United Kingdom cohort (Tomenson et al., 1997). These estimates are based on a 

3365 small number of observations (six cases in New York and four in the United Kingdom) and so 

3366 are relatively imprecise. It is only in the latest follow-up of the New York film base production 

3367 cohort that an elevated SMR was observed, further suggesting that the statistical power of the 

3368 other cohort studies for examining risk of this disease may be quite low. Two case-control 

3369 studies of dichloromethane exposure and brain cancer have been conducted (Cocco et al., 1999; 
90

3370 Heineman et al., 1994). The Heineman et al. (1994) study, which is the stronger study in terms 

3371 of exposure assessment strategy and confirmation of diagnosis, reported relatively strong trends 

3372 with increasing probability, duration, and intensity measures of exposure, but a nonlinear trend 

3373 was seen with the cumulative exposure metric. This difference could reflect a more valid 

3374 measure of relevant exposures in the brain from the intensity measure, as suggested by the study 

3375 in rats reported by Savolainen et al. (1981) in which dichloromethane levels in the brain were 

3376 much higher with a higher intensity exposure scenario compared with a constant exposure period 

3377 with an equivalent TWA (see section 3.2). The available epidemiologic studies provide some 

3378 evidence of an association between dichloromethane and brain cancer, and this area of research 

3379 represents a data gap in the understanding of the carcinogenic potential of dichloromethane. 

3380 

3381 4.1.3.7.2. Liver and biliary duct cancer. Liver and biliary duct cancer are relatively uncommon 

3382 (age-adjusted incidence 6.2 per 100,000 person-years) (SEER website, seer.cancer.gov, accessed 

3383 April 2006), so it is difficult to study in most occupational cohorts of limited size. The cohort 

3384 study with the higher exposures, the Rock Hill, South Carolina, triacetate fiber production plant, 

3385 suggested an increased risk of liver cancer (Lanes et al., 1993, 1990). The SMR for liver and 

3386 bile duct cancer was 2.98 (95% CI 0.81, 7.63) in the latest update of this cohort. This 

3387 observation was based on four cases; two of these cases were biliary duct cancers. As the 

3388 follow-up period has increased, the strength of this association has decreased, although it is 

3389 relatively strong (albeit with wide CIs). The decrease in the SMR with increasing follow-up 

3390 reflects the increase in number of expected cases, because the four observed cases were seen 

3391 earlier in the follow-up period. No other cohort study has reported an increased risk of liver 

3392 cancer mortality, although it should be noted that there is no other inception cohort study of a 

3393 population with exposure levels similar to those of the Rock Hill plant, and no data from a case3394 control study of liver cancer are available pertaining to dichloromethane exposure. The available 

3395 epidemiologic studies, with biological plausibility inferred from the results from studies in mice 

3396 and female rats (see section 4.2) (NTP, 1986; Serota et al., 1986a, b; Nitschke 1988a), provide 

3397 some evidence of an association between dichloromethane and liver and biliary duct cancer, 

3398 although it should be noted that this evidence is based on very limited epidemiologic data. 

3399 

3400 4.1.3.7.3. Lung cancer. In the stronger cohort studies (Cohort 1 in the New York Eastman 

3401 Kodak Company triacetate film production study reported by Hearne and Pifer [1999] and the 

3402 United Kingdom triacetate film production study reported by Tomenson et al. [1997]), the SMRs 

3403 for lung cancer were well below 1.0. The New York study had also obtained data on smoking 

3404 history that indicated it was unlikely that differences in smoking could be masking an effect of 

3405 dichloromethane (Hearne et al., 1987). Lung cancer is a common cancer (age-adjusted incidence 

3406 61 per 100,000 person-years) (SEER website, seer.cancer.gov, accessed April 2006), so the 

3407 expected rates, even in small cohorts, are based on relatively robust estimates. The only group in 
91

3408 any study that had an increased risk for lung cancer was the high-exposure women in the 

3409 triacetate fiber production cohort in Maryland (Gibbs et al., 1996). However, this was based on 

3410 only two cases and was a highly imprecise estimate (SMR 2.3 [95% CI 0.28, 8.3]). No case3411 control study of dichloromethane exposure and lung cancer risk is available. The available 

3412 epidemiologic studies do not provide evidence for an association between dichloromethane and 

3413 lung cancer, although it should be noted that the studies with the best designs are limited to 

3414 relatively low exposure levels. 

3415 

3416 4.1.3.7.4. Pancreatic cancer. An early study (Hearne et al., 1990) of Cohort 2 of the New York 

3417 triacetate film production cohort had reported 8 observed and 4.2 expected pancreatic cancer 

3418 deaths, for a twofold increased SMR (p = 0.13). This association was reduced in the subsequent 

3419 follow-up (SMR 1.5 [95% CI 0.7, 3.0]) (Hearne and Pifer, 1999) but was not seen in the more 

3420 methodologically sound Cohort 1 (SMR 0.92) or in any of the other cohorts. A meta-analysis of 

3421 the cohort studies (using the data of Hearne et al. [1990]) reported a summary association of 

3422 1.42 (95% CI 0.80, 2.53) (Ojajärvi et al., 2001). This summary measure would be further 

3423 reduced with the updated data for Cohort 2 and the addition of Cohort 1 from Hearne and Pifer 

3424 (1999). The only case-control study of pancreatic cancer mortality risk and dichloromethane 

3425 exposure (based on death certificate data) did not report consistent patterns with respect to 

3426 intensity or exposure among the race-sex groups studied. The available epidemiologic studies do 

3427 not provide evidence for an association between dichloromethane and pancreatic cancer. 

3428 

3429 4.1.3.7.5. Leukemia and lymphoma. Each of the individual hematopoietic cancers is relatively 

3430 uncommon, with age-adjusted incidence rates of 5 per 100,000 person-years or less (SEER 

3431 website, seer.cancer.gov, accessed April 2006). The relatively inconsistent (point estimates 

3432 ranging from 0.50 or less to 2.0 or higher) and imprecise measures of association between 

3433 dichloromethane exposure and non-Hodgkin’s lymphoma, Hodgkin’s lymphoma, myeloma, and 

3434 leukemia are thus expected, given the relatively small size of the available cohort studies. Only 

3435 one case-control study of any of these diseases and dichloromethane is available, and this is a 

3436 study of childhood leukemia (acute lymphoblastic leukemia) in relation to maternal occupational 

3437 history (Infante-Rivard et al., 2005). This is a large, population-based study of confirmed 

3438 incident cases of leukemia, with a detailed exposure assessment pertaining to the period before 

3439 and during pregnancy. A threefold increased risk was seen with probable or definite exposure 

3440 (OR 3.22 [95% CI 0.88, 11.7]) compared with possible or no exposure. The available 

3441 epidemiologic studies do not provide an adequate basis for the evaluation of the role of 

3442 dichloromethane in any of the specific hematopoietic cancers because of the small size of the 

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

3444 
92

3445 4.1.3.7.6. Breast cancer. Only one large cohort study included women and reported data 

3446 pertaining to breast cancer risk (Blair et al., 1998), and this is a cohort with a limited exposure 

3447 assessment (dichotomized) and multiple exposures. A relatively strong association was seen 

3448 between dichloromethane exposure and breast cancer mortality in this study (rate ratio 3.0 [95% 

3449 CI 1.0, 8.8]). Similar associations were seen with several other chemicals, and the potential 

3450 effect of confounding and misclassification of these exposures may have biased the estimate in 

3451 either direction. The only case-control study of breast cancer risk and dichloromethane exposure 

3452 used the 24-state death certificate data to classify exposure and disease. The available 

3453 epidemiologic studies do not provide an adequate basis for the evaluation of the role of 

3454 dichloromethane in breast cancer because there are currently no cohort studies with adequate 

3455 statistical power and no case-control studies with adequate exposure methodology to examine 

3456 this relationship. 

3457 

3458 4.2. SUBCHRONIC AND CHRONIC STUDIES AND CANCER BIOASSAYS IN 

3459 ANIMALS—ORAL AND INHALATION 

3460 4.2.1. Oral Exposure: Overview of Noncancer and Cancer Effects 

3461 Results from studies of animals exposed by the oral route for short-term, subchronic, and 

3462 chronic durations identify the liver and the nervous system as the most sensitive targets for 

3463 noncancer toxicity from repeated oral exposure to dichloromethane. In a 90-day exposure study, 

3464 nonneoplastic histopathologic changes in the liver were observed in F344 rats exposed to 

3465 drinking water doses of ≥166 mg/kg-day (males) or ≥209 mg/kg-day (females) (Kirschman et al., 

3466 1986). Similar changes were seen in F344 rats in a 2-year exposure of ≥50 mg/kg-day (Serota et 

3467 al., 1986a). 

3468 The 2-year oral exposure study in F344 rats did not produce evidence of increasing 

3469 incidence of liver tumors across all of the dose groups in males or females (Serota et al., 1986a). 

3470 In females, however, a jagged stepped pattern of increasing incidence was observed. In a 

3471 parallel study in B6C3F1 mice (Serota et al., 1986b; Hazelton Laboratories, 1983), a clearer trend 

3472 with respect to hepatic cancer was seen in males but not females. 

3473 None of the chronic oral exposure studies included a systematic measurement of potential 

3474 neurological effects. One 14-day study focusing on neurobehavioral changes is available, 

3475 however. Changes in autonomic, neuromuscular, and sensorimotor functions were observed in 

3476 F344 rats exposed for 14 days to gavage doses ≥337 mg/kg-day (Moser et al., 1995) (see section 

3477 4.4.3 for more details). 

3478 No effects on reproductive parameters were observed in Charles River CD rats exposed 

3479 for 90 days to gavage doses as high as 225 mg/kg-day (General Electric Co., 1976) or in 

3480 pregnant F344 rats exposed to gavage doses of up to 450 mg/kg-day on gestation days (GDs) 6–

3481 19 (Narotsky and Kavlock, 1995). However, no oral exposure studies examining developmental 

3482 neurobehavioral effects have been conducted (see section 4.3 for more details). 
93

3483 4.2.1.1. Toxicity Studies of Subchronic Oral Exposures: Hepatic Effects 

3484 Kirschman et al. (1986) examined the toxicity of dichloromethane in a 90-day drinking 

3485 water study in F344 rats (20/sex/dose level). The nominal concentration of dichloromethane in 

3486 the water was 0.15, 0.45, or 1.5%. Based on BW and water consumption data, average intakes 

3487 were reported to be 0, 166, 420, or 1,200 mg/kg-day for males and 0, 209, 607, or 1,469 mg/kg3488 day for females. Clinical chemistry tests (hematological and chemical variables in samples of 

3489 blood and urine) and tissue histopathology were evaluated in groups of five rats/sex/dose level 

3490 after 1 month of treatment. These endpoints were also evaluated in the remaining rats sacrificed 

3491 after 90 days of exposure. 

3492 Exposure to dichloromethane did not affect mortality or cause adverse clinical signs of 

3493 toxicity. Gross necropsy was also unremarkable. Reported changes in mean values for clinical 

3494 chemistry variables, compared with controls, included elevated serum ALT activities for all 

3495 treated males at 1 month and for the high-dose females at 3 months, elevated serum AST activity 

3496 in high-dose females at 3 months, elevated serum lactate dehydrogenase activities in mid- and 

3497 high-dose females at 3 months, and decreases in serum concentrations of fasting glucose, 

3498 cholesterol, and triglycerides in all exposed groups of both sexes at 1 and 3 months. Actual 

3499 values for clinical chemistry variables, however, were not presented in the report. 

3500 No histopathologic alterations were seen in tissues after 1 month of treatment (a detailed 

3501 description of tissues examined was not presented). In rats exposed for 3 months, exposure3502 related histopathologic changes were restricted to the liver. Elevated, statistically significant, 

3503 incidences of hepatocytic vacuolation were observed in all exposed male groups and in the mid3504 and high-dose female groups (see Table 4-11). The most frequently observed vacuolation was 

3505 described as generalized and occurring throughout the lobule, and Oil Red-O-staining indicated 

3506 most were lipid-containing vacuoles. The incidences of generalized vacuolation scored as mild 

3507 or moderate were higher in all of the female dose groups compared with the controls. The 

3508 authors stated that the no-observed-adverse-effect level (NOAEL) based on this study is less than 

3509 200 mg/kg-day and the lowest-observed-adverse-effect level (LOAEL) for males was 

3510 166 mg/kg-day. The authors did not explicitly provide a LOAEL for females. The results 

3511 indicate that 166 mg/kg-day and 209 mg/kg-day were the LOAELs for liver effects in male and 

3512 female rats, respectively. 

3513 
94

3514 

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

female F344 rats exposed to dichloromethane in drinking water for 90 days 

Lesion, by sex Controls Low dose Mid dose High dose

Males—n per groupa 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) 

 

10b

 (67) 9b (60) 7b (47)

Generalized vacuolation severity: 

 minimal 

 mild 

 moderate 

 0 (0) 

 0 

 0 

 0 

 5b (33) 

 4 

 0 

 1 

 8b (53)

 7 

 1 

 0 

 6b (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 groupa 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) 13b (87) 15b (100) 15b

 (100)

Generalized vacuolation severity: 

 minimal 

 mild 

 moderate 

 5 (33) 

 5 

 0 

 0 

 13b (87) 

 8 

 4 

 1 

 15b (100) 

 6 

 5 

 4 

 15b (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 

 11b

 (28) 

 2 

 4 

 3 

 2 

Hepatocyte degeneration 0 (0) 0 (0) 0 (0) 12b (80) 

Focal granuloma 0 (0) 0 (0) 4 (27)c 6b (40) 

a

20 per group; 5 sacrificed at 1 month; these endpoints for the remaining 15 per group. 

b

Statistical significance testing not reported by authors; Fisher’s exact test for comparison with control p--

value <0.05 (two-sided). 

c

Statistical significance testing not reported by authors; Fisher’s exact test for comparison with control p -

value <0.10 (two-sided). Authors stated LOAEL = 166 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 

3516 Kirschman et al. (1986) conducted a similar 90-day study in B6C3F1 mice (20/sex/dose 

3517 level). The estimated average intakes were 0, 226, 587, or 1,911 mg/kg-day for males and 231, 

3518 586, or 2,030 mg/kg-day for females. Six mice (two controls, two low dose, and two mid dose) 

3519 died during the study from unknown causes. Administration of dichloromethane did not cause 

3520 adverse clinical signs of toxicity or affect food consumption, ophthalmology, or serum ALT 

3521 activity. Gross necropsy examinations also were unremarkable. 

3522 Histopathologic evaluation of tissues from mice killed after 1 month of treatment did not 

3523 reveal any compound-related effects. Evaluation at 3 months showed subtle generalized or 
95

3524 centrilobular changes in the liver (characterized as increased vacuolation with fat deposition), 

3525 which was evident in all exposed groups and most prominent in mid- and high-dose female 

3526 groups (Table 4-12). The most frequently detected change was characterized as a generalized 

3527 vacuolation. Some evidence was found for an increase in severity of the generalized vacuolation 

3528 with increasing exposure level, but the incidence of this lesion in the control mice was 

3529 substantial, especially in females (Table 4-12). Incidences for centrilobular vacuolation were 

3530 significantly increased only for the mid-dose female group. No other changes were found. 

3531 Using the results from this study to select doses for a chronic study, Kirschman et al. 

3532 (1986) expressed the opinion that the mid-dose level (587 mg/kg-day) was the LOAEL in this 

3533 study. Although incidences for generalized vacuolation were increased in the low- and mid-dose 

3534 male groups, the incidences in the high-dose groups were not significantly increased compared 

3535 with controls (Table 4-12). The study authors identified a LOAEL of 586 mg/kg-day for 

3536 centrilobular vacuolation in male B6C3F1 mice. The NOAEL for males was considered by the 

3537 investigators to be between 226 and 587 mg/kg-day. 
96

3538 

Table 4-12. Incidences of histopathologic changes in livers of male and 

female B6C3F1 mice exposed to dichloromethane in drinking water for 

90 days 

Lesion, by sex Controls Low dose Mid dose High dose

Males—n per groupa 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 

 12b (86) 

 3 

 7 

 2 

 0 

 13b (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 groupa 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 

 5c (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 authors; Fisher’s exact test for comparison with control 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 authors; Fisher’s exact test for comparison with control 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 

3541 4.2.1.2. Toxicity Studies of Chronic Oral Exposures: Hepatic Effects and Carcinogenicity 

3542 Longer-term (up to 2-year) oral exposure studies in mice and rats are summarized in 

3543 Table 4-13 and described in more detail below. These studies provide additional information 

3544 pertaining to hepatotoxicity and carcinogenicity. 

3545 

3546 

3547 
97

Table 4-13. Studies of chronic oral dichloromethane 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) 

 B6C3F1 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 

3550 4.2.1.2.1. Chronic oral exposure in F344 rats (Serota et al., 1986a). Treatment with 

3551 dichloromethane did not induce adverse clinical signs or affect survival in the F344 rats (Serota 

3552 et al., 1986a). BWs of rats in the 125 and 250 mg/kg-day groups were generally lower than in 

3553 controls throughout the study. The authors stated that the differences, although small, were 

3554 statistically significant, but the data were not shown in the published report. Water consumption 

3555 was lower throughout the study in both sexes of rats from the 125 and 250 mg/kg-day groups 

3556 relative to controls; food consumption was also lower in these groups during the first 13 weeks 

3557 of treatment. Mean hematocrit, hemoglobin, and red blood cell count were increased in both 

3558 sexes at dichloromethane levels of 50, 125, and 250 mg/kg-day for 52 and 78 weeks. Half of 

3559 these increases were reported to be statistically significant, but the report did not provide the 

3560 numerical values or specify which parameters were significant. Clinical chemistry results 

3561 showed decreases in alkaline phosphatase (AP), creatinine, blood urea nitrogen, total protein, and 

3562 cholesterol in both sexes at 250 mg/kg-day, and most of these changes were statistically 
98

3563 significant at one or both of the intervals evaluated. (Significant parameters not specified and the 

3564 mean group values were not presented in the published report.) No significant deviations in 

3565 urinary parameters were observed. Organ weights were not significantly affected by treatment 

3566 with dichloromethane. 

3567 No treatment-related histopathological effects were noted in the tissues examined except 

3568 for the liver (Serota et al., 1986a). Examination of liver sections showed a dose-related positive 

3569 trend (positive Cochran-Armitage trend test) in the incidences of foci/areas of cellular alteration 

3570 in treated F344 rats (Table 4-14). Comparisons of incidences with control incidences indicated 

3571 statistically significant elevations at all dose levels except 5 mg/kg-day. These liver changes 

3572 

Table 4-14. Incidences of nonneoplastic liver changes and liver tumors in 

male and female F344 rats exposed to dichloromethane in drinking water for 

2 years 

Target dose (mg/kg-day)

Controls

 0a

5 50 125 250 

Trend 

p-valueb

250 with 

recoveryc

Males—n per groupd 76 34 38 35 41 15

Estimated mean intake (mg/kg-day) 0 6 52 125 235 232 

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 groupd 67 29 41 38 34 20 

Estimated mean intake (mg/kg-day) 0 6 58 136 263 239 

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)fp < 0.01 2 (10)f

3573 

a

3574 Two control groups combined. 

b

3575 Cochran-Armitage trend test was used for trend test of liver foci/areas of alteration. For tumor mortality-unadjusted 

3576 analysis, a Cochran-Armitage trend test was used, and, for tumor mortality-adjusted analyses, tumor prevalence 

3577 analytic method by Dinse and Lagakos (1982) was used. Similar results were seen in these two analyses. 

c

3578 Recovery group was exposed for 78 weeks and then had a 26-week period without dichloromethane exposure; 

3579 n = 15 for nonneoplastic lesions and n = 17 for neoplastic lesions. 

d

3580 Number available at terminal sacrifice; starting with 135 controls (combining both control groups) and 85 per sex 

3581 per dose group except recovery group (n = 25); subtracted 5, 10, and 20 per group (except for recovery group) 

3582 sacrificed at 25, 52, and 78 weeks, respectively, and subtracted unscheduled deaths, which ranged from 5 to 19 per 

3583 group. 

e

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

f

3585 Significantly (p < 0.05) different from controls with Fisher’s exact test, mortality-unadjusted and mortality-adjusted

3586 analyses. 

3587 

3588 Source: Serota et al. (1986a). 
99

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

3590 treated with dichloromethane also showed an increased incidence of fatty change, but incidence 

3591 data for this lesion were not presented in the published report. The recovery group also showed 

3592 an increased incidence of areas of cellular alterations, but the fatty changes were less pronounced 

3593 than in the 250 mg/kg-day group dosed for 104 weeks. The authors indicate that 5 mg/kg-day 

3594 was a NOAEL and 50 mg/kg-day was a LOAEL for nonneoplastic liver changes in male and 

3595 female F344 rats exposed to dichloromethane in drinking water for 2 years. 

3596 Dichloromethane-exposed male rats showed no statistically significant increased 

3597 incidence of liver tumors. In females, there was a positive trend for increasing incidence of 

3598 hepatocellular carcinoma or neoplastic nodules with increasing dose (Table 4-14) (Serota et al., 

3599 1986a). Statistically significant increases in tumor incidences were observed in the 50 and 

3600 250 mg/kg-day groups (incidence rates of 10% and 14%, respectively) but not in the 125 mg/kg3601 day group (incidence rate of 3%). Incidence was also increased (10%) in a group exposed for 

3602 78 weeks followed by a 26-week period of no exposure. The characterization of malignant 

3603 potential of the nodules was not described, however, and no trend was seen in the data limited to 

3604 hepatocellular carcinomas. The incidence of hepatocellular carcinoma or neoplastic nodules in 

3605 this control group (0%) was lower than that seen in historical controls from the same laboratory 

3606 (324 female F344 rats; 4 with carcinoma, 21 with neoplastic nodules; 25/324 = 7.7%). 

3607 

3608 4.2.1.2.2. Chronic oral exposure in B6C3F1 mice (Serota et al., 1986b; Hazelton Laboratories,

3609 1983). A 2-year drinking water study similar to the previously described study in F344 rats was 

3610 also conducted in B6C3F1 mice (Serota et al., 1986b; Hazelton Laboratories, 1983). The mice 

3611 received target doses of 0, 60, 125, 185, or 250 mg/kg-day of dichloromethane in deionized 

3612 drinking water for 24 months. Treatment groups consisted of 100 female mice in the low-dose 

3613 group and 50 in the remaining treatment groups. There were 200, 100, 100, and 125 male mice 

3614 (low- to high-dose groups) in the treated groups. One hundred females (in two groups of 50) and 

3615 125 males (in two groups of 60 and 65 mice) served as controls. (The authors do not state why 

3616 two groups of control mice were used, other than to say that the design was used due to the high 

3617 and erratic incidence of liver tumors in historical control B6C3F1 mice.) Based on water 

3618 consumption and BW measurements, mean intakes were reported to be 61, 124, 177, and 

3619 234 mg/kg-day for males and 59, 118, 172, and 238 mg/kg-day for females. Endpoints 

3620 examined included clinical signs, BW and water consumption, hematology at weeks 52 and 104, 

3621 and gross and microscopic examinations of tissues and organs at termination. All tissues from 

3622 the control and 250 mg/kg-day groups were examined microscopically, as well as the livers and 

3623 neoplasms from all groups and the eyes of all males from all groups. 

3624 Throughout the 2-year study, mice from both control and treatment groups exhibited a 

3625 high incidence of convulsions (Serota et al., 1986b; Hazelton Laboratories, 1983). The 

3626 convulsions were noted only during handling for BW determinations, and efforts to establish a 
100

3627 basis for this response were unsuccessful. The incidence of convulsions did not correlate with an 

3628 increased mortality rate. Survival to 104 weeks was high (82% in males and 78% in females), 

3629 and no evidence for exposure-related negative effects on survival were found. Exposure had no 

3630 significant effect on BW or water consumption. Mean leukocyte count was significantly 

3631 elevated in males and females dosed with 250 mg/kg-day dichloromethane for 52 weeks, but the 

3632 authors indicated that the mean values were within the normal historical range for the laboratory. 

3633 Treatment-related nonproliferative histopathologic effects were restricted to the liver and 

3634 consisted of a marginal increase in the amount of Oil Red O-positive material in the liver of 

3635 males and females dosed with 250 mg/kg-day (group incidences for this lesion, however, were 

3636 not presented in the published report). The results indicate that 185 mg/kg-day was a NOAEL 

3637 and 250 mg/kg-day was a LOAEL for marginally increased amounts of fat in livers of male and 

3638 female B6C3F1 mice. 

3639 Incidences for proliferative hepatocellular changes in female mice were not presented in 

3640 the published reports (Serota et al., 1986b; Hazelton Laboratories, 1983), but it was reported that 

3641 exposed female mice did not show increased incidences of proliferative hepatocellular lesions. 

3642 In the male B6C3F1 mice, incidences for hepatic focal hyperplasia showed no evidence of an 

3643 exposure-related effect (Table 4-15). The authors interpret the data regarding adenomas alone or 

3644 carcinomas alone as showing no significantly elevated incidence compared with controls. The 

3645 trend tests for each of these outcomes were 0.172 and 0.147 (Hazelton Laboratories, 1983), 

3646 respectively, and none of the comparisons between individual exposure groups and the controls 

3647 was statistically significant at the chosen Bonferroni-corrected level of <0.01. However, 

3648 exposed male mice showed increased combined incidences of hepatocellular adenomas and 

3649 carcinomas, with a linear trend p-value = 0.058 and individual p-values of <0.05. 

3650 

Table 4-15. Incidences for focal hyperplasia and tumors in the liver of male 

B6C3F1 mice exposed to dichloromethane in drinking water for 2 years 

Target dose (mg/kg-day)

Controls

 0a 60 125 185 250 

Trend 

p-valueb

n per groupc 125 200 100 99 125 

Estimated mean intake (mg/kg-day) 0 61 124 177 234 

Number (%) with 

Focal hyperplasiad

 

 10 (8) 14 (7) 

 

 4 (4) 10 (10) 13 (10) not reported

Hepatocellular adenoma

 mortality-adjusted percent and p-valuee

 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-valuee

 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-valuee

 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 

3651 
101

a

3652 Two control groups combined. 

b

3653 Cochran-Armitage trend test (source: Hazelton Laboratories [1983]) 

c

3654 Number at start of treatment. 

d

3655 Some mice with hyperplasia also had hepatocellular neoplasms, but the exact number was unspecified by Serota et 

3656 al. (1986b). 

e

3657 Percent calculated based on number at risk, using Kaplan-Meier estimation, taking into account mortality losses; p3658 value for comparison with control group, using asymptotic normal test (source: Hazelton Laboratories [1983]). 3659 

3660 Sources: Serota et al. (1986b); Hazelton Laboratories (1983). 

3661 

3662 

3663 Serota et al. (1986b) noted, in summary, that slight increases in proliferative 

3664 hepatocellular lesions were found in exposed male but not female B6C3F1 mice and that the 

3665 increases were not dose related and were within the range of historical control incidences. The 

3666 average incidence of hepatocellular adenomas and carcinomas in 354 control male B6C3F1 mice 

3667 in the laboratory in which the experiment was performed was 17.8% with a range of 5–40% 

3668 (Serota et al., 1986b). Serota et al. (1986b) concluded that dichloromethane “did not induce a 

3669 treatment-related carcinogenic response in B6C3F1 mice” under the conditions of this study. An 

3670 alternative conclusion, as determined by EPA, is that dichloromethane induced a carcinogenic 

3671 response in male B6C3F1 mice as evidenced by small, but statistically significant, increases in 

3672 hepatocellular adenomas and carcinomas at dose levels of 125, 185, and 250 mg/kg-day but not 

3673 at 60 mg/kg-day and by a marginally increased trend test for combined hepatocellular adenomas 

3674 and carcinomas. Results for the highest dose group (no effect on BW, histologic findings 

3675 restricted to mild histologic changes in the liver [vacuolation], and a slight, but statistically 

3676 significant, increase in incidence in liver tumors in males only) indicate that this mouse study 

3677 may not have included the maximum tolerated dose. 

3678 

3679 4.2.1.2.3. Chronic oral exposure in Sprague-Dawley rats and Swiss mice (Maltoni et al., 

3680 1988). Maltoni et al. (1988) conducted gavage carcinogenicity studies in Sprague-Dawley rats 

3681 and in Swiss mice. Groups of rats (50/sex/dose level) were gavaged with dichloromethane 

3682 (99.9% pure) in olive oil at dose levels of 0 (olive oil), 100, or 500 mg/kg-day, 4–5 days/week 

3683 for 64 weeks. This dosing regime was also used for groups of Swiss mice (50/sex/dose level 

3684 plus 60/sex as controls). Endpoints monitored included clinical signs, BW, and full necropsy at 

3685 sacrifice (when spontaneous death occurred). For each animal sacrificed, histopathologic 

3686 examinations were performed on the following organs: brain and cerebellum, zymbal glands, 

3687 interscapular brown fat, salivary glands, tongue, thymus and mediastinal lymph nodes, lungs, 

3688 liver, kidneys, adrenals, spleen, pancreas, esophagus, stomach, intestine, bladder, uterus, gonads, 

3689 and any other organs with gross lesions. High mortality was observed in male and female high3690 dose rats (data not shown) and achieved significance (p < 0.01) in males. The increased 

3691 mortality became evident after 36 weeks of treatment and led to the termination of treatment at 

3692 week 64. Explanation of the mortality was not provided by the study authors. As with the rats, 
102

3693 high mortality occurred in male and female mice from the high-dose group (p < 0.01), and the 

3694 exposure was terminated after 64 weeks. 

3695 Little information is provided regarding nonneoplastic effects (Maltoni et al., 1988). 

3696 Treatment with dichloromethane did not affect BW in the Sprague-Dawley rats. A reduction in 

3697 BW was apparent in treated mice after 36–40 weeks of treatment, but no data were shown to 

3698 determine the magnitude of the effect. The lack of reporting of nonneoplastic findings from the 

3699 histopathologic examinations precludes assigning NOAELs and LOAELs for possible 

3700 nonneoplastic effects in these studies 

3701 The Maltoni et al. (1988) studies of Sprague-Dawley rats and Swiss mice did not find 

3702 distinct exposure-related carcinogenic responses following gavage exposure to dichloromethane 

3703 at dose levels up to 500 mg/kg-day, although the early termination of the study (at 64 weeks) 

3704 limits the interpretation of this finding. Dichloromethane exposure was not related to the 

3705 percentage of either study animal bearing benign and malignant tumors or bearing malignant 

3706 tumors or to the number of total malignant tumors per 100 animals. High-dose female rats 

3707 showed an increased incidence in malignant mammary tumors, mainly due to adenocarcinomas 

3708 (8, 6, and 18% in the control, 100, and 500 mg/kg dose groups, respectively; the number of 

3709 animals examined was not provided), but the increase was not statistically significant. A dose3710 related increase, although not statistically significant, in pulmonary adenomas was observed in 

3711 male mice (5, 12, and 18% in control, 100, and 500 mg/kg-day groups, respectively). When 

3712 mortality was taken into account, high-dose male mice that died in the period ranging from 52 to 

3713 78 weeks were reported to show a statistically significantly (p < 0.05) elevated incidence for 

3714 pulmonary tumors (1/14, 4/21, and 7/24 in control, 100, and 500 mg/kg-day groups, 

3715 respectively). Details of this analysis were not provided. EPA applied a Fisher’s exact test to 

3716 these incidences and determined a p-value of 0.11 for the comparison of the 500 mg/kg-day 

3717 group (7/24) to the controls (1/14). 

3718 

3719 4.2.2. Inhalation Exposure: Overview of Noncancer and Cancer Effects 

3720 Inhalation dichloromethane exposure studies in rats and mice, using subchronic and 

3721 chronic durations, identify the CNS, liver, and lungs as potential toxicity targets. Data from 

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

3723 effects of dichloromethane than are rats and mice. 

3724 Increased incidences of nonneoplastic liver lesions were observed in Sprague-Dawley 

3725 rats exposed to ≥500 ppm for 2 years (Nitschke et al., 1988a; Burek et al., 1984), F344 rats 

3726 exposed to concentrations ≥1,000 ppm for 2 years (Mennear et al., 1988; NTP, 1986), and 

3727 B6C3F1 mice exposed to ≥2,000 ppm for 2 years (Mennear et al., 1988; NTP, 1986). 

3728 Two-year inhalation exposure studies at concentrations of 2,000 or 4,000 ppm 

3729 dichloromethane produced increased incidences of lung and liver tumors in B6C3F1 mice 

3730 (Mennear et al., 1988; NTP, 1986). Additional studies examining mechanistic issues regarding 
103

3731 this effect are described in sections 4.5.2 and 4.5.3 (Maronpot et al., 1995; Foley et al., 1993; 

3732 Kari et al., 1993). Significantly increased incidences of benign mammary tumors (primarily 

3733 fibroadenomas) were also observed in male and female F344/N rats exposed by inhalation to 

3734 2,000 or 4,000 ppm for 2 years (Mennear et al., 1988; NTP, 1986). In the male rats, the 

3735 incidence of fibromas or sarcomas originating from the subcutaneous tissue around the 

3736 mammary gland was also increased, but the difference was not statistically significant. In other 

3737 studies in Sprague-Dawley rats with exposures of 50–500 ppm (Nitschke et al., 1988a) and 500–

3738 3,500 ppm (Burek et al., 1984), the incidence of benign mammary tumors was not increased, but 

3739 in females the number of tumors per tumor-bearing rat increased at the higher dose levels. 

3740 No obvious clinical signs of neurological impairment were observed in the 2-year 

3741 bioassays involving exposure concentrations up to 2,000 ppm in F344 rats (Mennear et al., 1988; 

3742 NTP, 1986) or 3,500 ppm in Sprague-Dawley rats (Nitschke et al., 1988a; Burek et al., 1984). In 

3743 B6C3F1mice exposed to 4,000 ppm in B6C3F1, there was some evidence of hyperactivity during 

3744 the first year of the study and lethargy during the second year, with female mice appearing to be 

3745 more sensitive (Mennear et al., 1988; NTP, 1986). Studies that evaluated batteries of 

3746 neurobehavioral endpoints following subchronic or chronic inhalation exposure are restricted to 

3747 one in which no effects were observed more than 64 hours postexposure in an observational 

3748 battery, a test of hind-limb grip strength, a battery of evoked potentials, or histologic 

3749 examinations of brain, spinal cord, or peripheral nerves in F344 rats exposed to concentrations 

3750 up to 2,000 ppm for 13 weeks (Mattsson et al., 1990) (see section 4.2.3) 

3751 No effects on reproductive performance were found in a two-generation reproductive 

3752 toxicity study with F344 rats exposed to concentrations up to 1,500 ppm for 14 and 17 weeks 

3753 before mating of the F0 and F1 generations, respectively (Nitschke et al., 1988b) (described 

3754 more completely in section 4.3). Developmental effects following exposure of Long-Evans rats 

3755 to 4,500 ppm for 14 days prior to mating and during gestation (or during gestation alone) 

3756 included decreased offspring weight at birth and changed behavioral habituation of the offspring 

3757 to novel environments (Bornschein et al., 1980; Hardin and Manson, 1980) (see section 4.3 for 

3758 more details). In standard developmental toxicity studies involving exposure to 1,250 ppm on 

3759 GDs 6–15, no adverse effects on fetal development were found in Swiss-Webster mice or 

3760 Sprague-Dawley rats (Schwetz et al., 1975) (see section 4.3). 

3761 

3762 4.2.2.1. Toxicity Studies of Subchronic Inhalation Exposures: General, Renal, and Hepatic 

3763 Effects 

3764 Data pertaining to general (e.g., BW, mortality), hepatic, and renal effects from several 

3765 inhalation exposure studies in various species, with exposure periods of 3–6 months, are 

3766 described below. (Studies providing detailed neurological data are described separately in 

3767 section 4.4.3.) The earliest study involved several different species, with exposures of 

3768 5,000 ppm for up to 6 months (Heppel et al., 1944). Two 14-week studies in dogs, monkeys, 
104

3769 rats, and mice were conducted with exposures at 0, 1,000, and 5,000 ppm (Haun et al., 1972, 

3770 1971; Weinstein et al., 1972) and at 0, 25, and 100 ppm (Haun et al., 1972). Neurological effects 

3771 and hepatic degeneration were seen at the 1,000 ppm dose. In the lower-dose portion of the 

3772 Haun et al. (1972) study in mice, decreased cytochrome P-450 levels in liver microsomes and 

3773 some histopathologic liver changes (fat stains and cytoplasmic vacuolation) were seen at 

3774 100 ppm but more obvious adverse effects were not observed. Leuschner et al. (1984) reported 

3775 data from a high exposure (10,000 ppm) 90-day study of rats; beagle dogs were also included in 

3776 this study, at an exposure level of 5,000 ppm. No evidence of toxicity was reported by the 

3777 authors of this study. In a 13-week exposure study conducted by NTP (1986), decreased BWs 

3778 and increased incidence of foreign body pneumonia were seen at 8,400 ppm in F344 rats, and 

3779 histologic changes in the liver in B6C3F1 mice were seen at 4,200 ppm. 

3780 The first experimental study of dichloromethane exposure included dogs, rabbits, guinea 

3781 pigs, and rats, with an exposure of approximately 5,000 ppm for 7 hours/day, 5 days per week 

3782 for up to 6 months (Heppel et al., 1944). The strains of the animals, the comparability between 

3783 exposed and unexposed group (in terms of sex distribution and other attributes), and process by 

3784 which animals were chosen for histologic examination are not clearly described in the report. 

3785 Exposed animals included adult dogs (1 male and 5 females), juvenile dogs (1 male and 1 female 

3786 born in the exposure chamber and exposed daily from birth), adult rabbits (2 males and 

3787 2 females), guinea pigs (14 males), and rats (15 males and 6 females). The nonexposed control 

3788 group included 14 guinea pigs, 28 rats, 4 rabbits, and an unspecified number of dogs. Exposure 

3789 produced no significant effects on BWs except in the guinea pigs; after 131 exposures, average 

3790 BWs were 0.820 and 1.025 kg for exposed and control guinea pigs, respectively. Three exposed 

3791 guinea pigs died after 35, 90, and 96 exposures. No other deaths occurred, except for one 

3792 exposed female rat that died after 22 exposures and giving birth to a litter. Autopsy showed 

3793 thrombi in the renal vessels associated with marked cortical infarction. No adverse clinical signs 

3794 of toxicity (such as decreased activity or incoordination) were observed in exposed animals 

3795 during the study. Urinalysis, hematology tests, and tests of liver function performed on dogs 

3796 during the study showed no treatment-related effects. At termination, gross and microscopic 

3797 examination of the major organs showed no pathological changes after exposure to 5,000 ppm 

3798 dichloromethane, with the exception that two of the exposed guinea pigs that died showed 

3799 extensive pneumonia associated with moderate centrilobular fatty degeneration of the liver. The 

3800 results indicate that 5,000 ppm was a NOAEL for nonneoplastic systemic effects in dogs, rabbits, 

3801 and rats exposed 7 hours/day, 5 days/week for up to 6 months. The findings of three deaths (two 

3802 with pulmonary congestion and centrilobular fatty degeneration) and 20% decreased average 

3803 BW among the 14 exposed guinea pigs indicates that 5,000 ppm was a LOAEL in this species.

3804 Haun et al. (1972, 1971) and Weinstein et al. (1972) reported results from studies in 

3805 which groups of 8 female beagle dogs, 4 female rhesus monkeys, 20 male Sprague-Dawley rats, 

3806 and 380 female ICR mice were continuously exposed to 0, 1,000, or 5,000 ppm dichloromethane 
105

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

3808 were scheduled to be made on animals that died or were sacrificed during or at termination of the 

3809 study. At 5,000 ppm, obvious nervous system effects (e.g., incoordination, lethargy) were 

3810 observed in dogs, monkeys, and mice. At 1,000 ppm, these effects were most apparent in dogs 

3811 and monkeys (Haun et al., 1971). Food consumption was reduced in all species at 5,000 ppm 

3812 and in dogs and monkeys at 1,000 ppm. All exposed animals either lost weight or showed 

3813 markedly decreased BW gains compared with controls. For example, rats exposed to 1,000 or 

3814 5,000 ppm for 14 weeks showed average BWs that were roughly 10 and 20% lower than control 

3815 values. Significant numbers of dogs (4) and mice (123), as well as 1 monkey, died within the 

3816 first 3 weeks of exposure to 5,000 ppm. Because of this high mortality, all surviving 5,000 ppm 

3817 animals were sacrificed at 4 weeks of exposure, except for one half (10) of the rats that went on 

3818 to survive the 14-week exposure period. At 1,000 ppm, six of eight dogs died by 7 weeks, at 

3819 which time the remaining two were sacrificed. Monkeys, rats, and all but a few mice survived 

3820 exposure to 1,000 ppm for 14 weeks. 

3821 Gross examination of tissues showed yellow, fatty livers in dogs that died during 

3822 exposure to 1,000 or 5,000 ppm, “borderline” liver changes in 3 monkeys exposed to 5,000 ppm, 

3823 and mottled liver changes in 4/10 rats exposed to 5,000 ppm for 14 weeks (Haun et al., 1971). 

3824 Comprehensive reporting of the histologic findings from this study were not available, but Haun 

3825 et al. (1972) reported that the primary target organ was the liver and that in some exposed 

3826 animals the kidney was also affected. Light and electron microscopy of liver sections from 

3827 groups of 4–10 mice sacrificed after 1, 4, 8, and 12 hours and 1, 2, 3, 4, 6, and 7 days of 

3828 exposure to 5,000 ppm showed hepatocytes with balloon degeneration (dissociation of 

3829 polyribosomes and swelling of rough endoplasmic reticulum) as early as 12 hours of exposure 

3830 (Weinstein et al., 1972). The degeneration peaked in severity after 2 days of exposure and, 

3831 subsequently, partially reversed in severity. Information on possible histopathologic changes in 

3832 mice exposed to 1,000 ppm was not provided. 

3833 The results from this study demonstrate that dogs and mice were more sensitive than 

3834 were rats and monkeys to lethal effects, nervous system depression, and possibly liver effects 

3835 from continuous exposure to 1,000 or 5,000 ppm. The results indicate that continuous exposure 

3836 to 1,000 ppm was an adverse effect level for mortality and effects on the nervous system and 

3837 liver in dogs (exposed for up to 4 weeks) and for BW changes in rats (exposed for 14 weeks). 

3838 The 5,000 ppm level induced mortality in beagle dogs, ICR mice, and rhesus monkeys (but not 

3839 in Sprague-Dawley rats); obvious nervous system effects in dogs, mice, monkeys, and rats; and 

3840 gross liver changes in dogs, mice, monkeys, and rats. 

3841 Haun et al. (1972) also conducted studies with groups of 20 mice, 20 rats, 16 dogs, and 

3842 4 monkeys exposed continuously to 0, 25, or 100 ppm dichloromethane for 100 days (14 weeks). 

3843 The animals presumably were of the same strains and sexes as those used in the studies involving 

3844 exposure to 1,000 or 5,000 ppm dichloromethane (Haun et al., 1972, 1971; Weinstein et al., 
106

3845 1972). All animals underwent necropsy and histopathologic evaluation at termination of the 

3846 exposure, but a list of the tissues examined and incidence or severity data were not presented in 

3847 the report. Hematology and clinical chemistry variables (including COHb levels) were measured 

3848 in blood samples collected from dogs and monkeys at biweekly or monthly intervals during 

3849 exposure. COHb levels were elevated in a dose-related manner in monkeys and peaked at about 

3850 5% (approximately 0.8% pre-exposure) after 6 weeks of exposure. COHb levels in dogs were 

3851 unaffected by the 25 ppm exposure level and rose to about 2% (from about 0.6%) from 

3852 week 4 on in high-dose dogs. Additional groups of mice were included for assessment of 

3853 hexobarbital sleep times at monthly intervals; levels of cytochromes P-450, P-420, and b5 in liver 

3854 microsomes at monthly intervals; and spontaneous physical activity at several intervals during 

3855 the study. 

3856 No clinical signs of toxicity or alterations in weight gain were seen in any of the species 

3857 examined. In dogs and monkeys, hematology and clinical chemistry results throughout the study 

3858 and at termination were unremarkable, as were the results of the gross and histopathologic 

3859 examinations. In mice exposed to 100 ppm, CYP levels in liver microsomes were significantly 

3860 decreased (compared with control values) after 30, 60, and 90 days of exposure to 100 ppm, 

3861 whereas levels of cytochrome b5 and P-420 decreased after 30 days and increased after 90 days 

3862 of exposure. At 25 ppm, no significant differences from control were seen in mouse liver levels 

3863 of cytochromes. Mice exposed to 25 ppm showed no histopathologic changes, while histologic 

3864 changes in mice at 100 ppm were restricted to positive fat stains and some cytoplasmic 

3865 vacuolation in the liver. In rats at both exposure levels, the livers showed positive staining for 

3866 increased fat, and the kidneys showed evidence of nonspecific tubular degenerative and 

3867 regenerative changes. Haun et al. (1972) indicate that no distinctively adverse effects were 

3868 found in monkeys, dogs, rats, or mice continuously exposed to 25 or 100 ppm for up to 

3869 14 weeks. Decreased CYP levels in liver microsomes and some histopathologic liver changes 

3870 (fat stains and cytoplasmic vacuolation) were seen at the 100 ppm dose. 

3871 Leuschner et al. (1984) exposed Sprague-Dawley rats (20/sex/dose level) to 0 or 

3872 10,000 ppm and beagle dogs (3/sex/dose level) to 0 or 5,000 ppm dichloromethane in whole3873 body exposure chambers. Exposure periods were 6 hours/day for 90 consecutive days. 

3874 Endpoints evaluated in both species included clinical signs, food and water consumption, BW, 

3875 hematology, clinical chemistry, urinalysis, and gross and microscopic evaluation of 27 organs at 

3876 termination. Electrocardiography and blood pressure measurements were also done in dogs. 

3877 The only significant effect observed in rats was a slight redness of the conjunctiva 

3878 1–10 hours after each exposure. In dogs, compound-related effects were restricted to slight 

3879 sedation throughout the exposure period and slight erythema lasting up to 10 hours after 

3880 exposure. In this 90-day study involving daily 6-hour exposures, 10,000 and 5,000 ppm were 

3881 NOAELs for behavioral, clinical chemistry, hematologic, and histologic signs of toxicity in 

3882 Sprague-Dawley rats and beagle dogs, respectively. 
107

3883 NTP (1986) exposed groups of F344 rats and B6C3F1 mice (10/sex/dose level) to target 

3884 concentrations of 0, 525, 1,050, 2,100, 4,200, or 8,400 ppm dichloromethane, 6 hours/day, 

3885 5 days/week for 13 weeks in whole-body exposure chambers. Endpoints monitored included 

3886 clinical signs, BW, and necropsy at termination. Comprehensive sets of tissues and organs in 

3887 control and high-dose animals were histologically examined; tissues from the lower dose groups 

3888 were examined to determine the no-observed-effect level. One male and one female rat from the 

3889 8,400 ppm exposure group died before the end of the study, but the cause of death was not 

3890 discussed. The final mean BWs of 8,400 ppm male and female rats were reduced by 23 and 

3891 11%, respectively, relative to controls. Foreign-body pneumonia was present in 4/10 male and 

3892 6/10 female rats exposed to 8,400 ppm and in 1/10 female rat from the 4,200 ppm exposure 

3893 group. The liver lipid/liver weight ratios for 8,400 ppm rats of both sexes and 4,200 ppm female 

3894 rats were significantly lower than in controls. In mice, 4/10 males and 2/10 females exposed to 

3895 8,400 ppm died before the end of the study, and these deaths were considered treatment related. 

3896 Histologic changes in exposed mice consisted of hepatic centrilobular hydropic degeneration (of 

3897 minimal to mild severity) in 3/10 males and 8/10 females at 8,400 ppm and in 9/10 females from 

3898 the 4,200 ppm exposure group. Histologic changes in the 2,100 ppm mouse group were not 

3899 mentioned. The liver lipid/liver weight ratio for the high-dose female mice was significantly 

3900 lower than in controls. In this 13-week study involving 6-hour exposure periods for 

3901 5 days/week, 4,200 ppm was a NOAEL and 8,400 ppm was a LOAEL for decreased BWs and 

3902 increased incidence of foreign-body pneumonia in F344 rats. In B6C3F1 mice, 2,100 ppm was a 

3903 NOAEL and 4,200 ppm was a LOAEL for histologic changes in the liver. 

3904 

3905 4.2.2.2. Toxicity Studies from Chronic Inhalation Exposures 

3906 Chronic inhalation exposure studies are summarized in Table 4-16. Details of each study 

3907 are described below, with the results pertaining to nonneoplastic and neoplastic effects 

3908 summarized in the following sections. 

3909 
108

3910 

Table 4-16. Studies of chronic inhalation dichloromethane 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) 

 B6C3F1 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 

3913 4.2.2.2.1. Chronic inhalation exposure in F344/N rats (Mennear et al., 1988; NTP, 1986). 

3914 NTP conducted a 2-year inhalation exposure study in F344/N rats (Mennear et al., 1988; NTP, 

3915 1986). The rats (50/sex/exposure level) were exposed to dichloromethane (>99% pure) by 

3916 inhalation in exposure chambers, 6 hours/day, 5 days/week for 2 years. Exposure concentrations 

3917 were 0, 1,000, 2,000, or 4,000 ppm. Mean daily concentrations never exceeded 110% of target 

3918 and were <90% of target in only 23 of 1,476 analyses. Endpoints monitored included clinical 

3919 signs, mortality, and gross and microscopic examinations of 32 tissues at study termination. 

3920 Clinical examinations were conducted weekly for 3.5 months and biweekly until month 8. After 

3921 8 months, the animals were clinically examined and palpated for tumors and masses monthly 

3922 until the end of the study. 

3923 Dichloromethane exposure did not significantly alter BW gain or terminal BWs 

3924 (Mennear et al., 1988; NTP, 1986). Survival of male rats was low in all exposed groups and the 
109

3925 control group, and no significant exposure-related differences were apparent. Most deaths 

3926 occurred during the last 16 weeks of the study. Survival at week 86 was 36/50, 39/50, 37/50, and 

3927 33/50 for the control, 1,000, 2,000, and 4,000 ppm groups, respectively. In female rats, there 

3928 was a trend towards decreased survival, and the survival of high-dose female rats was 

3929 significantly reduced, possibly due to leukemia. Survival in the females at 86 weeks was 30/50, 

3930 22/50, 22/50, and 15/50 for the control, 1,000, 2,000, and 4,000 ppm groups, respectively. 

3931 Nonneoplastic lesions with statistically significantly elevated incidences, compared with 

3932 controls, included hepatocyte cytoplasmic vacuolation and necrosis and liver hemosiderosis in 

3933 males and females; renal tubular cell degeneration in males and females; splenic fibrosis in 

3934 males; and nasal cavity squamous metaplasia in females (Table 4-17). The results indicate that 

3935 1,000 ppm (6 hours/day, 5 days/week) was a LOAEL for nonneoplastic liver changes 

3936 (hepatocyte cytoplasmic vacuolation and necrosis, hepatic hemosiderosis) in male and female 

3937 F344/N rats. A NOAEL was not established because effects were observed at the lowest dose. 
110

3938 

Table 4-17. Incidences of nonneoplastic histologic changes in male and 

female F344/N rats exposed to dichloromethane by inhalation (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 groupb50 50 50 50 

Number (%)c with 

 Liver changes 

 Hepatocyte cytoplasmic vacuolation 

 

 8 (16) 26 (53)d22 (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 (34) 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 groupe50 50 50 50 

Number (%)c with 

Liver changes 

 Hepatocyte cytoplasmic vacuolation 10 (20) 43 (86)d44 (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 (40) 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/m3, 2,000 ppm = 6,947 mg/m3, 4,000 ppm = 13,894 mg/m3. 

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

c

Percentages were based on the number of tissues examined microscopically per group. 

d

Statistical significance not reported in publications but significantly (p ≤ 0.05) different from control as calculated 

by Fisher’s exact test. 

eNumber of females necropsied per group; only 49 4,000 ppm kidneys and spleens were examined microscopically.

Sources: Mennear et al. (1988); NTP (1986, Appendix B, Tables C1 and C2). 

3939 

3940 Incidences of mammary fibroadenomas were significantly increased in 4,000 ppm males 

3941 and 2,000 and 4,000 ppm females, compared with controls (Table 4-18). Similar patterns were 

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

3943 subcutaneous tissue fibroma or sarcoma was seen in 1/50, 1/50, 2/50, and 5/50 rats in the 0, 

3944 1,000, 2,000, and 4,000 ppm groups, respectively, but these lesions were not seen in females. 

3945 Incidences of female rats with liver neoplastic nodules or carcinomas (combined) showed a 

3946 significant trend test after survival adjustment only, but the incidences at the two highest dose 

3947 levels were not significantly increased relative to the control (Table 4-18). 

3948 Incidences for mononuclear cell leukemias in mid- and high-dose female rats were 

3949 statistically significant after a survival-adjustment analysis. However, Mennear et al. (1988) 

3950 considered the relationship between exposure to dichloromethane and mononuclear cell leukemia 
111

3951 to be equivocal, based on the fact that most male rats had leukemia (34/50, 26/50, 32/50, and 

3952 35/50 in controls, 1,000, 2,000, and 4,000 ppm rats, respectively). Other neoplasms that had 

3953 increased incidences included mesotheliomas (predominantly in the tunica vaginalis) in males 

3954 (0/50, 2/50, 5/50, and 4/50 in controls, 1,000, 2,000, and 4,000 ppm rats, respectively). This 

3955 lesion was not considered to be related to dichloromethane exposure, because the concurrent 

3956 control incidence (0/50) for this neoplasm was low relative to earlier inhalation studies 

3957 conducted at this laboratory (4/100, 4%) and in other NTP studies with male F344/N rats 

3958 (44/1,727) (mean historical percentage across NTP studies = 3 ± 2%). 

3959 NTP (1986) concluded that there was “some evidence of carcinogenicity of 

3960 dichloromethane” in male F344/N rats as shown by increased incidence of benign mammary 

3961 gland tumors and “clear evidence of carcinogenicity” of dichloromethane in female F344/N rats 

3962 as shown by increased incidence of benign mammary gland tumors. The summary of the hepatic 

3963 effects in rats in the NTP (1986) report also notes the positive trend in the incidence of 

3964 hepatocellular neoplastic nodules or carcinomas in females, which “may have been due to 

3965 dichloromethane exposure.” 

3966 

3967 

3968 
112

Table 4-18. Incidences of selected neoplastic lesions in male and female F344/N rats exposed to dichloromethane by inhalation (6 hours/day, 5 days/week) for 2 years 

Exposure (ppm) a

 0 (Controls) 1,000 2,000 

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

Males 

n per group 50 50 50 50 

Liver—Neoplastic nodule or hepatocellular carcinoma 2 (4) (10) 3 (6) (13) 4 (8) (19) 1 

Liver—hepatocellular carcinoma 2 (4) (10) 1 (2) (4) 2 (4) (10) 1 

 Lung—Bronchoalveolar adenoma or carcinoma 1 1 (2) 2 (4) 1 

Mammary gland 

Adenoma, adenocarcinoma, or carcinoma Subcutaneous tissue fibroma or sarcoma Fibroadenoma 

 Mammary gland or subcutaneous tissue adenoma, 

fibroadenoma, fibroma, or sarcoma 

 0  1 

0 

 1

(0)(2)(0)(2)

(6)(0)(6)

 0 

 1  0  1 

(0)(2)(0)(2)

(6)(0)(6)

 0 

 2  2  4

(0)(4)(4)(8)

 (9)(12)(21)

 1  5  1  9e

Brain (carcinoma, not otherwise specified, invasive) 0 (0) 1 (2) 0 (0) 0 

Females 

n per group 50 50 50 50 

Liver—Neoplastic nodule or hepatocellular carcinoma 2 (4) (7) 1 (2) (2) 4 (8) (14) 5 

Liver—hepatocellular carcinoma 0 (0) (0) 0 (0) (0) 1 (2) (4) 0 

 Lung—Bronchoalveolar adenoma or carcinoma 1 (2) 1 (2) 0 (0) 0 

Mammary gland 

Adenocarcinoma or carcinoma Adenoma, adenocarcinoma, or carcinoma Fibroadenoma 

 Mammary gland adenoma, fibroadenoma, or 

adenocarcinoma 

 1 

 1  5  6 

(2)(2)

(10)(12)

(16)(18)

 2 

 2 11e13 

(4)(4)

(22)(26)

(41)(44)

 2 

 2 13e 14e

(4)(4)

(26)(28)

(44)(45)

 0  1 22e23e

Brain (carcinoma, not otherwise specified, invasive, and 

oligodendroglioma)f

 1 (2) 0 (0) 2 (4) 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 examined microscopically per group; for males, 49 livers and lungs were examined micros

1,000 ppm groups, and only 49 brains were examined microscopically in the 4,000 ppm group. For comparison, incidence in historicaNTP (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

3969 4.2.2.2.2. Chronic inhalation exposure in B6C3F1 mice (Mennear et al., 1988; NTP, 1986). A 

3970 2-year inhalation exposure study in B6C3F1 mice, similar to that in F344/N rats, was also 

3971 conducted by NTP (Mennear et al., 1988; NTP, 1986). The mice (50/sex/exposure level) were 

3972 exposed to dichloromethane (>99% pure) by inhalation at concentrations of 0, 2,000, or 

3973 4,000 ppm in exposure chambers 6 hours/day, 5 days/week for 2 years. As with the study in rats, 

3974 mean daily concentrations in the mice never exceeded 110% of target and were <90% of target in 

3975 only 23 of 1,476 analyses. Endpoints monitored included clinical signs, mortality, and gross and 

3976 microscopic examinations of 32 tissues at study termination. Clinical examinations were 

3977 conducted weekly for 3.5 months and biweekly until month 8. After 8 months, the animals were 

3978 clinically examined and palpated monthly for tumors and masses until the end of the study. 

3979 The BW of 4,000 ppm males was comparable to controls until week 90 and 8–11% below 

3980 controls thereafter. The BW of 4,000 ppm females was 0–8% lower than that of controls from 

3981 week 51 to 95 and 17% lower at study termination. No information was provided regarding food 

3982 consumption during the study. Male and female mice from the high-dose groups (4,000 ppm) 

3983 were hyperactive during the first year of the study; during the second year, high-dose females 

3984 appeared lethargic. Exposure was associated with decreased survivability of both male and 

3985 female mice (males: 39/50, 24/50, and 11/50 and females: 25/50, 25/50, and 8/50 in controls, 

3986 2,000 ppm, and 4,000 ppm at 104 weeks, respectively). In 4,000 ppm mice, statistically 

3987 significant incidences of nonneoplastic lesions were found in the liver (cytologic degeneration), 

3988 testes (atrophy), ovary and uterus (atrophy), kidneys (tubule casts in males only), stomach 

3989 (dilatation), and spleen (splenic follicles in males only) (Table 4-19). In 2,000 ppm mice, the 

3990 only nonneoplastic lesions showing statistically significantly elevated incidences were ovarian 

3991 atrophy, renal tubule casts, and hepatocyte degeneration in female mice (Table 4-19). The 

3992 results indicate that 2,000 ppm, the lowest exposure level, was a LOAEL for nonneoplastic 

3993 changes in the ovaries, kidneys, and livers of female B6C3F1 mice. A NOAEL was not 

3994 established because effects occurred at the lowest exposure level. 

3995 
114

3996 

Table 4-19. Incidences of nonneoplastic histologic changes in B6C3F1

mice exposed to dichloromethane by inhalation (6 hours/day, 

5 days/week) for 2 years 

Exposure (ppm) a

Lesion, by sex 

Controls 

0 2,000 4,000 

Males: n per groupb 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 groupe

 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/m3, 4,000 ppm = 13,894 mg/m3. 

bNumber of male mice necropsied per group. The number biopsied in the 0, 2,000, and 4,000 ppm 

dose groups was 50, 49, and 49 for liver; 50, 49, and 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 examined microscopically per group. 

d

Statistical significance not reported in publications but significantly different (p ≤ 0.05) from 

control as calculated by EPA using Fisher’s exact test. 

eNumber 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, and 43 for ovaries; 50, 48, and 47 for uterus; 49, 

48, and 47 for renal tubule; 49, 47, and 48 for stomach; and 49, 48, and 47 for spleen. 

Sources: Mennear et al. (1988); NTP (1986, Appendix C, Tables D1 and D2). 

3997 

3998 

3999 At both exposure levels, statistically significantly elevated incidences were found for 

4000 hepatocellular adenomas (males only); hepatocellular carcinomas; hepatocellular adenomas and 

4001 carcinomas, combined; bronchoalveolar adenomas; bronchoalveolar carcinomas; and 

4002 bronchoalveolar adenomas and carcinomas (Table 4-20). Statistically significant positive trend 

4003 tests were found for each of these tumor types in female mice. The trend tests were significant 

4004 for the liver tumors in male mice after life-table adjustment for reduced survival. The only other 

4005 statistically significant carcinogenic response was for increased incidence of hemangiosarcomas 

4006 or combined hemangiomas and hemangiosarcomas in male mice exposed to 4,000 ppm. NTP 
115

4007 (1986) concluded that the elevated incidences of liver and lung tumors provided clear evidence 

4008 of carcinogenicity in male and female B6C3F1 mice. 

4009 

Table 4-20. Incidences of neoplastic lesions in male and female B6C3F1 mice 

exposed to dichloromethane by inhalation (6 hours/day, 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-value

d

Males 

Liver 

Hepatocellular adenoma 10 (20) (23) 14 (29) (47) 14 (29) (68) 0.19 

Hepatocellular 13 (26) (30) 15 (30) (44) 26e

(53) (76) 0.004 

Hepatocellular adenoma or carcinoma 22 (44) (48) 24 (49) (67) 33e(67) (93) 0.013 

Lung 

Bronchoalveolar adenoma 3 (6) (8) 19e

(38) (56) 24e(48) (79) <0.001 

Bronchoalveolar carcinoma 2 (4) (5) 10e

(20) (34) 28e(56) (93) <0.001 

Bronchoalveolar adenoma or carcinoma 5 (10) (12) 27e

(54) (74) 40e(80) (100) <0.001 

Mammary adenocarcinomaf – – – 

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) 22e(46) (83) <0.001 

Hepatocellular carcinoma 1 (1) (4) 11 (23) (34) 32e(67) (97) <0.001 

Hepatocellular adenoma or carcinoma 3 (6) (10) 16e

(33) (48) 40e(83) (100) <0.001 

Lung 

Bronchoalveolar adenoma 2 (4) (7) 23e(48) (58) 28e(58) (91) <0.001 

Bronchoalveolar carcinoma 1 (1) (4) 13e

(27) (46) 29e(60) (92) <0.001 

Bronchoalveolar adenoma or carcinoma 3 (6) (11) 30e(63) (83) 41e(85) (100) <0.001 

Mammary adenocarcinoma 2 (4) (8) 3 (6) (10) 0 (0) (0) 0.21 

Hemangioma or hemangiosarcoma, 

combinedf

 – – – 

a

2,000 ppm = 6,947 mg/m3, 4,000 ppm = 13,894 mg/m3. 

b

Percentages based on the number of tissues examined microscopically per group; for males, 49 livers were examined in 

the 2,000 and 4,000 ppm groups; for females, only 48 livers and lungs and 49 mammary glands were microscopically 

examined in the 2,000 and 4,000 ppm groups. For comparison, incidence in historical controls reported in NTP (1986) 

were 28% for male liver tumors, 31% for male lung tumors, 5% for female liver tumors, and 10% for female lung 

tumors. 

cMortality-adjusted percentage. 

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

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

fData not reported. 

Sources: Mennear et al. (1988); NTP, (1986, Appendix E, Tables E3 and E4). 

4010 

4011 

4012 4.2.2.2.3. Chronic inhalation exposure in Syrian hamsters (Burek et al., 1984). Burek et al. 

4013 (1984) conducted a chronic toxicity and carcinogenicity study in rats and hamsters. In the 

4014 hamster study, groups of 95 Syrian golden hamsters of each sex were exposed to 0 (filtered air), 

4015 500, 1,500, or 3,500 ppm dichloromethane (>99% pure) under dynamic airflow conditions in 

4016 whole-body exposure chambers 6 hours/day, 5 days/week for 2 years. Exposure started when the 
116

4017 animals were approximately 8 weeks of age. Interim sacrifices were conducted at 6, 12, and 

4018 18 months. The hamsters were observed daily during exposure days and were palpated monthly 

4019 for palpable masses starting the third month of the study. BWs were monitored weekly for the 

4020 first 8 weeks of the study and monthly thereafter. Hematologic determinations included packed 

4021 cell volume, total erythrocyte counts, total red blood cells, differential leukocyte counts, and 

4022 hemoglobin concentration. The mean corpuscular volume, mean corpuscular hemoglobin, and 

4023 MCHC were also determined. A reticulocyte count was also performed on all animals at the 18-

4024 month kill and on 10 animals/sex/dose at 24 months. Clinical chemistry determinations included 

4025 serum AP and ALT activities, blood urea nitrogen levels, and total protein and albumin. Urinary 

4026 parameters measured were specific gravity, pH, glucose, ketones, bilirubin, occult blood, protein, 

4027 and urobilinogen. Hematology, clinical chemistries, and urinalysis were performed at interim 

4028 sacrifices and at termination. COHb was measured after a single 6-hour exposure and following 

4029 22 months of exposure. Gross and microscopic examinations were conducted on all tissues. In 

4030 addition, the weights of the brain, heart, liver, kidneys, and testes were recorded. 

4031 In the study using Syrian hamsters (Burek et al., 1984), hamsters were exposed to 

4032 analytical concentrations of dichloromethane of 510 ± 27, 1,510 ± 62, and 3,472 ± 144 ppm for 

4033 the target concentrations of 500, 1,500, and 3,500 ppm, respectively. No exposure-related 

4034 clinical signs were observed in the hamsters throughout the study. Significantly decreased 

4035 mortality was observed in females exposed to 3,500 ppm from the 13th through the 24th month 

4036 and from the 20th to the 24th month in females exposed to 1,500 ppm. Exposure to 

4037 dichloromethane had no significant effect on BW or on mean organ weights. Regarding 

4038 hematology parameters (actual data were not shown), Burek et al. (1984) stated that a few 

4039 statistically significant changes occurred, but no obvious pattern could be discerned and most 

4040 values were within the expected range for the animals. There were no exposure-related 

4041 alterations in clinical chemistry or urinalysis values. Male and female hamsters in all dose 

4042 groups had significantly elevated COHb values after a single 6-hour exposure and after 

4043 22 months of exposure, but at both time points there was no dose-response relationship above the 

4044 first dose level and no apparent significant differences in the magnitude of the changes between 

4045 the two time points. For example, mean values (± SD) for percentage COHb in male hamsters 

4046 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 

4047 control through 3,500 ppm groups, respectively. Similar values were obtained for females at 

4048 22 months and for males and females after the first day of exposure. Pathological evaluation of 

4049 hamsters showed a lack of evidence of definite target organ toxicity. Specific observations 

4050 mentioned by the authors included a trend of increasing hemosiderin in the liver of male 

4051 hamsters at 6 and 12 months; decreased amyloid deposit in organs, such as the liver, kidneys, 

4052 adrenal, and thyroid glands in exposed animals; and fewer biliary cysts in the liver. Increased 

4053 hepatic hemosiderin at the 12 month sacrifice was observed in 1/5, 1/5, 3/5, and 5/5 male 

4054 hamsters in the control through 3,500 ppm groups, respectively. No exposure-related increased 
117

4055 incidences of hepatic hemosiderin, or other liver effects, were reported for the terminal sacrifice. 

4056 The exposure-related decreases in geriatric changes (i.e., amyloid deposits and biliary cysts) 

4057 were more prominent in females and were associated with the increased survivability in the 

4058 exposed female hamsters compared with controls. The results indicate that 3,500 ppm was a 

4059 NOAEL for adverse changes in clinical chemistry and hematological variables, as well as for 

4060 nonneoplastic histologic changes in tissues, in male and female Syrian golden hamsters. A 

4061 LOAEL was not established, based on the lack of adverse changes in clinical chemistry and 

4062 hematological variables as well as the absence of nonneoplastic histologic changes in tissues, in 

4063 male and female Syrian golden hamsters. 

4064 Evaluation of the total number of hamsters with a tumor, the number with a benign 

4065 tumor, or the number with a malignant tumor revealed no exposure-related differences in male 

4066 hamsters. In the high-dose female group, there was a statistically significant increase in the total 

4067 number of benign tumors at any tissue site (the report did not specify which sites), but this was 

4068 considered to be secondary to the increased survival of this group. Incidences of male or female 

4069 hamsters with tumors in specific tissues were not statistically significantly elevated in exposed 

4070 groups compared with control incidences. The results indicate that no statistically significant, 

4071 exposure-related carcinogenic responses occurred in male or female Syrian golden hamsters 

4072 exposed (6 hours/day, 5 days/week) to up to 3,500 ppm dichloromethane for 2 years. 

4073 

4074 4.2.2.2.4. Chronic inhalation exposure in Sprague-Dawley rats (Burek et al., 1984). In the rat 

4075 study, groups of 92–97 Sprague-Dawley rats of each sex were exposed (similar to the hamster 

4076 study described in the previous section) to 0, 500, 1,500, or 3,500 ppm dichloromethane 

4077 6 hours/day, 5 days/week for 2 years (Burek et al., 1984). Rats were approximately 8 weeks old 

4078 when exposure started. Interim sacrifices were conducted at 6, 12, 15, and 18 months. 

4079 Endpoints monitored in rats were the same as in hamsters except that total protein and albumin in 

4080 blood were not determined in rats. In addition to measurement at scheduled sacrifices, serum 

4081 ALT activity was also measured after 30 days of exposure. COHb was measured after 6, 11, 18, 

4082 and 21 months of exposure. Bone marrow cells were collected for cytogenetic studies from five 

4083 rats/sex/dose after 6 months of exposure. The scope of the pathological examinations of the rats 

4084 was the same as in the hamster study. 

4085 No significant exposure-related signs of toxicity were observed in the rats during the 

4086 study. A significant increase in mortality was seen in high-dose female rats from the 18th to the 

4087 24th month of exposure, and this appeared to be exposure-related. Exposure to dichloromethane 

4088 had no significant effect on BW gain in either males or females. The only exposure-related 

4089 alterations in organ weights was a significant increase in both absolute and relative liver weight 

4090 in high-dose males at the 18-month interim kill and a significant increase in relative liver weight 

4091 in high-dose females also at 18 months. Statistically significant changes in hematologic 

4092 parameters were restricted to increased mean corpuscular volume and mean corpuscular 
118

4093 hemoglobin values at 15 months in males. The clinical chemistry tests revealed no significant 

4094 exposure-related effects. Male and female rats in all exposed groups had significantly elevated 

4095 COHb values at all time points, but no dose-response relationship was apparent. For example, 

4096 mean (± SD) values for percentage COHb after 21 months of exposure were 0.4 (± 0.7), 

4097 12.8 (± 2.6), 14.8 (± 4.4), and 12.2 (± 5.7) for the control through 3,500 ppm female rat groups, 

4098 respectively. Exposure-related statistically significant increases in incidences of nonneoplastic 

4099 lesions were restricted to the liver (Table 4-21). The incidences of males or females with 

4100 hepatocellular vacuolation consistent with fatty change increased as the exposure concentration 

4101 increased. Hepatocellular necrosis occurred at elevated incidences in male rats exposed to 

4102 1,500 or 3,500 ppm, compared with controls, but this endpoint was not reported in the female 

4103 data. Liver lesions were initially observed after 12 months of treatment. There was some 

4104 evidence that exposure at the two highest levels provided some inhibition of the age-related 

4105 glomerulonephropathy observed in the control rats at termination. The results indicate that the 

4106 lowest exposure level, 500 ppm, was a LOAEL for fatty changes in the liver of male and female 

4107 Sprague-Dawley rats and that exposure to ≥1,500 ppm induced hepatocellular necrosis in males.
119

4108 

Table 4-21. Incidences of selected neoplastic histologic changes in male and 

female Sprague-Dawley rats exposed to dichloromethane 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 2b (2) 8 (8) 10 (10)c 11 (11)c

Coagulation necrosis --d

 -- -- -- 

Hepatic vacuolation (fatty change) 16b (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 

 70b (76) 

 92b,e (100) 

62 (65) 

91 (96) 

53 (56)c

93 (98) 

39 (40)c

90 (93) 

Mammary changes 

Rats with benign mammary tumors 7b (8) 3 (3) 7 (7) 14 (14) 

Total number of benign mammary tumors 8 6 11 17 

Number of tumors per tumor-bearing ratf 1.1 2.0 1.6 1.2 

Females—n per group 96 95 96 97

Number (%) with 

Liver changes 

 

Hepatocellular necrosis – – – – 

Coagulation necrosis 1b (1) 0 (0) 2 (2) 7 (7) 

Hepatic vacuolation (fatty change) 33b (34) 49 (52)c 56 (58)c 63 (65)c

Foci of altered hepatocytes 35b

 (37) 36 (38) 27 (28) 50 (52)c

Foci of altered hepatocytes, basophilic 3b (3) 0 (0) 4 (4) 10 (10) 

Area of altered hepatocytes 19b (20) 24 (25) 28 (29) 35 (36)c

Multinucleated hepatocytes 7b (7) 36 (38)c 34 (35)c 29 (30)c

Glomerulonephropathy 

 Severe 

 Any degree 

 5 (5) 

62b

 (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 ratf

 2.1 2.7 3.1 3.5 

a

500 ppm = 1,737 mg/m3, 1,500 ppm = 5,210 mg/m3, 3,500 ppm = 12,158 mg/m3. 

b

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

c

Significantly higher than control incidence by Fisher’s exact test. 

d

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

eBurek et al. (1984) reported that 93/92 male mice had glomerulonephropathy in the kidney in the control group; the 

incidence was corrected to 92/92. 

fCalculated by EPA. 

Source: Burek et al. (1984). 

4109 

4110 

4111 In females, an increasing trend was seen in the incidence of foci or areas of altered 

4112 hepatocytes. Female rats in all exposed groups showed increased incidence of multinucleated 

4113 hepatocytes in the centrilobular region, compared with controls, but there was no evidence of 
120

4114 increasing incidence or severity with increasing exposure level (Table 4-21). The foci and areas 

4115 were apparent after 12 months and their number and size increased thereafter, but incidences for 

4116 neoplastic nodules in the liver or hepatocellular carcinomas were not increased in any exposure 

4117 group. A statistically significant increased incidence of salivary gland sarcomas was reported for 

4118 male rats exposed to 3,500 ppm. Burek et al. (1984) considered this finding unusual and 

4119 inconsistent with other existing data because the primary target organ for dichloromethane seems 

4120 to be the liver. Incidences of rats with benign mammary gland tumors were not statistically 

4121 significantly higher in exposed male or female groups compared with controls, and exposed male 

4122 and female groups showed no significantly increased incidences for malignant mammary gland 

4123 tumors. The average number of benign mammary tumors per tumor-bearing rat increased with 

4124 increasing exposure level. In females, the values were 2.1, 2.7, 3.1, and 3.5 in the control 

4125 through 3,500 ppm groups, respectively; males showed a similar response with increasing 

4126 exposure level, albeit to a lesser extent (Table 4-21). Burek et al. (1984) concluded that the 

4127 significance of this benign mammary tumor response (i.e., increase in number of tumors per 

4128 tumor-bearing rat) was unknown but speculated that the predisposition of this strain of rats 

4129 (historical control incidences of female with benign mammary tumors normally exceeded 80%) 

4130 plus the high exposure to dichloromethane may have resulted in the response. 

4131 

4132 4.2.2.2.5. Chronic inhalation exposure in Sprague-Dawley rats (Nitschke et al., 1988a). 

4133 Nitschke et al. (1988a) examined the toxicity and carcinogenicity of lower concentrations of 

4134 dichloromethane in Sprague-Dawley rats. Groups of 90 male and 90 female rats were exposed to 

4135 0, 50, 200, or 500 ppm dichloromethane (>99.5% pure) 6 hours/day, 5 days/week for 2 years. 

4136 Interim sacrifices were conducted at 6, 12, 15, and 18 months (five rats/sex/interval). An 

4137 additional group of 30 female rats was exposed to 500 ppm for 12 months and then exposed to 

4138 room air for up to an additional 12 months, and another group of 30 female rats was exposed to 

4139 room air for the first 12 months, followed by exposure to 500 ppm for the last 12 months of the 

4140 study. These latter groups were included to examine temporal relationships between exposure 

4141 and potential carcinogenic response. All groups of rats were examined daily for signs of toxicity 

4142 and all rats were examined for palpable masses prior to the initial exposure and at monthly 

4143 intervals after the first 12 months. BW was checked twice a month for the first 3 months and 

4144 monthly thereafter. Blood samples were collected at interim sacrifices and analyzed for total 

4145 bilirubin, cholesterol, triglycerides, potassium, estradiol, follicle-stimulating hormone, and 

4146 luteinizing hormone levels. In addition, COHb was determined at multiple times in blood 

collected from the tail vein. DNA synthesis (incorporation of 3

4147 H-thymidine as a measure of 

4148 cellular proliferation) was measured in the liver of separate groups of female rats after exposure 

4149 to the various concentrations for 6 and 12 months (four females/exposure group per interval). 

4150 All rats were subjected to a complete necropsy, and sections from most tissues were processed 

4151 for microscopic examination. 
121

4152 Exposure to dichloromethane at any of the exposure levels did not significantly alter 

4153 mortality rates, BWs, organ weights, clinical chemistry values, or plasma hormone levels 

4154 (Nitschke et al., 1988a). Blood COHb was elevated in a dose-related manner but not in an 

4155 exposure duration-related fashion, suggesting lack of accumulation with repeated exposures. For 

4156 example, mean (± SD) values for percentage COHb were 2.2 (± 1.3), 6.5 (± 1.1), 12.5 (± 0.8), 

4157 and 13.7 (± 0.6) for male rats in the control through 500 ppm groups, respectively, at the 

4158 terminal sacrifice and were similarly affected at the 6-month and 12-month intervals (e.g., 

4159 respective values for males were 0.3(± 0.7), 2.8 (± 0.3), 9.6 (± 1.2), and 12.7 (± 1.6) at the 

4160 12-month sacrifice). 

4161 The results of the thymidine incorporation experiment revealed no detectable alteration in 

4162 the rate of liver DNA synthesis in the exposed groups compared with controls. Statistically 

4163 significantly increased incidences of nonneoplastic liver lesions (hepatic vacuolation and 

4164 multinucleated hepatocytes) occurred only in females in the 500 ppm group (Table 4-22). Male 

4165 rat incidence for hepatocyte vacuolation was elevated at 500 ppm but not to a statistically 

4166 significant degree. In the group of female rats exposed for only 12 months to 500 ppm, 

4167 significantly increased incidences of nonneoplastic lesions, compared with controls, were 

4168 restricted to liver cytoplasmic vacuolization (16/25 = 64%) and multinucleated hepatocytes 

4169 (9/25= 36%) in rats exposed during the first 12 months of the study; rats exposed only during the 

4170 last 12 months of the study showed no elevated incidences of the liver lesions. 

4171 

Table 4-22. Incidences of selected nonneoplastic histologic changes in male 

and female Sprague-Dawley rats exposed to dichloromethane by inhalation 

(6 hours/day, 5 days/week) for 2 years 

Exposure (ppm) a

Lesion, by sex 

Controls 

0 50 200 500 

Trend 

p-valueb

Late 

500c

Early 

500c

Males—n per group 70 70 70 70 NAd 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 (17) 27 (39)f <0.0001 3 (12) 9 (36)f

a

50 ppm = 174 mg/m3, 200 ppm = 695 mg/m3, 500 ppm = 1,737 mg/m3. 

bCochran-Armitage trend test. 

cLate 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.

dNA= there were no male rats in these exposure groups. 

e

– = Incidences not reported. 

f

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

Source: Nitschke et al. (1988a). 

4172 

4173 
122

4174 A few fibrosarcomas or undifferentiated sarcomas in the mammary gland were seen in 

4175 the exposed rats, but these incidences were not statistically significant (Table 4-23). 

4176 Significantly increased incidences of rats with neoplastic lesions were restricted to benign 

4177 mammary tumors in female rats exposed for 2 years to 200 ppm compared with controls 

4178 (61/69 = 88%) (Table 4-23). However, significantly elevated incidences of this tumor type were 

4179 not observed in 500 ppm females, and the 200 ppm incidence was within the range for historical 

4180 control values for benign mammary tumors in female Sprague-Dawley rats (79–82%) from two 

4181 other chronic toxicity/carcinogenicity studies from the same laboratory. A slight, but statistically 

4182 significant, increase in the number of palpable masses in subcutaneous or mammary regions (at 

4183 23 months) per tumor-bearing rat was observed only in the 500 ppm female group. The numbers 

4184 of benign mammary tumors per tumor-bearing rat were slightly elevated in the exposed groups 

4185 compared with control groups, but no statistical analysis of this variable was performed. In 

4186 female rats exposed to 500 ppm (during the first or second 12 months of the study), slight but 

4187 statistically significant elevations were found in the number of palpable masses in subcutaneous 

4188 or mammary regions per tumor-bearing rat; the numbers of benign mammary tumors per tumor4189 bearing rat were slightly elevated compared with those of controls, but statistical analysis of this 

4190 variable was not performed. 

4191 A statistically significant increased incidence of brain or CNS tumors was not observed, 

4192 but six astrocytoma or glioma (mixed glial cell) tumors were seen in the exposed groups (4 in 

4193 males, 2 in females). The authors concluded that there was no distinct exposure-related 

4194 malignant carcinogenic response in male or female Sprague-Dawley rats exposed (6 hours/day, 

4195 5 days/week) to up to 500 ppm dichloromethane for 2 years (Nitschke et al., 1988a). 

4196 
123

Table 4-23. Incidences of selected neoplastic histologic changes in male and 

female Sprague-Dawley rats exposed to dichloromethane by inhalation 

(6 hours/day, 5 days/week) for 2 years 

Exposure (ppm) a

Lesion, by sex 

Controls

0 50 200 500 

Late 

500b

Early 

500b

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 

sarcomad

 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 2.2e2.3e 2.7e

Number with benign tumors 52 (75) 58 (84) 61f

 (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/m3, 200 ppm = 695 mg/m3, 500 ppm = 1,737 mg/m3

bLate 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. No males were included in these exposure groups. 

c

Percentages were based on the number of tissues examined microscopically per group. In males, 69 lungs were 

examined microscopically in the 50 ppm groups, and only 57, 65, 59, and 64 mammary glands were examined in 

the control, 50, 200, and 500 ppm groups, respectively. In females, 69 mammary glands were examined 

microscopically in the control, 50, 200, and 500 ppm groups. 

dEPA summed across these three tumors, assuming no overlap. 

e

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

f

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

Source: Nitschke et al. (1988a). 

4197 
124

4198 4.2.2.2.6. Chronic inhalation exposure in Sprague-Dawley rats (Maltoni et al., 1988). Maltoni 

4199 et al. (1988) conducted an inhalation exposure study in Sprague-Dawley rats. Two groups of 

4200 female rats (54–60/dose) were exposed to 0 or 100 ppm dichloromethane for 104 weeks. The 

4201 exposure period was 4 hours/day, 4 days/week for 7 weeks and then 7 hours/day, 5 days/week 

4202 for 97 weeks. Endpoints monitored included clinical signs, BW, and full necropsy at sacrifice 

4203 (when spontaneous death occurred). For each animal sacrificed, histopathologic examinations 

4204 were performed on the following organs: brain and cerebellum, zymbal glands, interscapular 

4205 brown fat, salivary glands, tongue, thymus and mediastinal lymph nodes, lungs, liver, kidneys, 

4206 adrenals, spleen, pancreas, esophagus, stomach, intestine, bladder, uterus, gonads, and any other 

4207 organs with gross lesions. 

4208 There was no evidence of increased mortality in the exposed group, and there was no 

4209 effect on BW (Maltoni et al., 1988). Little information was provided regarding nonneoplastic 

4210 effects, precluding assignment of NOAELs and LOAELs for possible nonneoplastic effects in 

4211 this study. Dichloromethane exposure was not related to the percentage of rats with benign 

4212 tumors and malignant tumors, malignant tumors, or the number of total malignant tumors per 

4213 100 animals. The percentage of rats with benign mammary tumors was 40.0% in controls and 

4214 64.8% in the exposed group, and the percentage of malignant mammary tumors was 3.3 and 

4215 5.5% in controls and exposed, respectively. Neither of these differences was statistically 

4216 significant. 

4217 

4218 4.3. REPRODUCTIVE/DEVELOPMENTAL STUDIES—ORAL AND INHALATION 

4219 Reproductive and development studies of dichloromethane exposure are summarized in 

4220 Table 4-24 and described in detail below. No effects on reproductive performance were 

4221 observed in a 90-day gavage study in Charles River CD rats with doses up to 225 mg/kg-day 

4222 (General Electric Co., 1976) or in a two-generation reproductive toxicity study with F344 rats 

4223 exposed to concentrations up to 1,500 ppm for 14 or 17 weeks before mating of the F0 and 

4224 F1 generations, respectively, as well as during the F1 gestational period (GDs 0–21) (Nitschke et 

4225 al., 1988b). Reproductive parameters (e.g., number of litters, implants/litter, live fetuses/litter, 

percent dead/litter, percent resorbed/litter, or fertility index4

4226 ) were also examined in a study in 

4227 male Swiss-Webster mice administered dichloromethane (250 or 500 mg/kg) by subcutaneous 

4228 injection three times/week for 4 weeks, and in a similar study involving inhalation exposure to 0, 

4229 100, 150, or 200 ppm dichloromethane; no statistically significant effects were seen in either 

4230 protocol, although some evidence of a decrease in fertility index was seen in the 150 and 

4231 200 ppm groups (Raje et al., 1988). 

4232 

 

4

Fertility index defined as number of females impregnated divided by total number of females mated times 100. 
125

4233 

Table 4-24. Summary of studies of reproductive and developmental effects of dichloromethane exp

animals 

Species and n Exposure dose Exposure period Results 

Oral and Gavage

Charles River rats (males and females), 10 per sex per dose group 

0, 25, 75, 225 mg/kg (gavage) 

90 days before mating (10 days between last exposure and mating period) 

No effects on fertility index, number of pups pelitter, pup survival, or F1 BW, hematology, and clinical chemistry tests (up to 90 days of age) 

Swiss-Webster mice (males), 20 per group 

0, 250, 500 mg/kg (subcutaneous injection), 3× per 

week 

4 weeks prior to mating (1 week between last exposure and mating period) 

No effects on fertility index, number of litters, implants per litter, live fetuses per litter, resorprate; no testicular effects 

F344 rats (females), 17–21 per dose group 

0, 337.5, 450 mg/kg-day (gavage) 

GDs 6–19 Decreased maternal weight gain; no effect on 

resorption rate, number of live litters, implantslive pups, or pup weight 

Inhalation

F344 rats (males and females, two generation), 30 per sex per dose group (F0 and F1) 

0, 100, 500, 1,500 ppm, 6 hours/day 

14 weeks prior to mating (F0), GDs 0–21, and 17 weeks prior to mating, beginning PND 4, (F1) 

No effect on fertility index, litter size, neonatal

survival, growth rates, or histopathologic lesion

Swiss-Webster mice (males), 20 per group 

0, 100, 150, 200 ppm, 2 hours/day 

6 weeks, prior to mating (2 days between last exposure and mating period) 

Fertility index decreased in 150 and 200 ppm group (statistical significance depends on test used); no effects on number of litters, implants litter, live fetuses per litter, resorption rate; no testicular effects. 

Long-Evans rats (female), 16–21 per dose group 

0, 4,500 ppm 12–14 days before 

mating and/or GDs 1–17 

Gestational exposure resulted in increased absoand relative maternal liver weight, decreased feBW 

Long-Evans rats (female), 16–21 per dose group 

0, 4,500 ppm 12–14 days before 

mating and/or GDs 1–17 

Altered rate of behavioral habituation to novel environment (at 4 days of age). No effect on crawling (at 10 days), movement in photocell c(15 days), use of running wheel (45–108 days)shock avoidance (4 months). 
126

Table 4-24. Summary of studies of reproductive and developmental effects of dichloromethane exp

animals 

Species and n Exposure dose Exposure period Results 

Swiss-Webster mice (females), 30–40 per group 

0, 1,250 ppm, 7 hours/day 

GDs 6–15 Increased incidence of extra center of ossificati

in sternum, increased (~10%) maternal blood COHb, increased maternal weight, increased maternal absolute liver weight 

Sprague-Dawley rats (females), 20–35 per group 

0, 1,250 ppm, 7 hours/day 

GDs 6–15 Decreased incidence of lumbar ribs or spurs, 

increased incidence of delayed ossification of sternebrae, increased (~10%) maternal blood COHb, increased maternal absolute liver weigh

4234 4235 4236 4237 
127

4238 Following exposure of pregnant F344 rats to gavage doses of up to 450 mg/kg-day on 

4239 GDs 6–19, maternal weight gain was decreased, but no effects were found on the number of 

4240 resorption sites, pup survivability, or pup weights at postnatal days (PNDs) 1 or 6 (Narotsky and 

4241 Kavlock, 1995). The developmental effects following exposure of Long-Evans rats to 

4242 4,500 ppm for 14 days prior to mating and during gestation (or during gestation alone) were 

4243 decreased offspring weight at birth and changed behavioral habituation of the offspring to novel 

4244 environments (Bornschein et al., 1980; Hardin and Manson, 1980) (see section 4.3.2 for more 

4245 details). In standard developmental toxicity studies involving exposure to 1,250 ppm on GDs 6–

4246 15, no adverse effects on fetal development were found in Swiss-Webster mice or Sprague4247 Dawley rats, but the incidence of minor skeletal variants (e.g., delayed ossification of sternebrae) 

4248 was increased. (Schwetz et al., 1975) (see section 4.3.2). 

4249 

4250 4.3.1. Reproductive Toxicity Studies 

4251 4.3.1.1. Oral (Gavage) Studies

4252 In a study sponsored by the General Electric Co. (1976), Charles River CD rats 

4253 (10/sex/dose level) were administered 0, 25, 75, or 225 mg/kg-day dichloromethane by gavage in 

4254 water for 90 days. The test material was dichloromethane (of unspecified purity) purchased from 

4255 Dow Chemical Company. At approximately 100 days of age, the rats were mated 1 to 1 to 

4256 produce the F1 generation. F1 rats (15/sex/dose level) received the same treatment as F0 for 

4257 90 days, at which time they were sacrificed and necropsied. Comprehensive sets of 24 tissues 

4258 from 10 male and 10 female F1 rats from the control and 225 mg/kg-day groups were examined 

4259 microscopically after embedding, sectioning, and staining. F1 rats were monitored for clinical 

4260 signs, BW effects, and food consumption. Reproductive parameters examined were fertility 

4261 index, number of pups per litter, and pup survival. F1 rats also underwent hematology and 

4262 clinical chemistry tests and urinalysis at 1, 2, and 3 months of the study and ophthalmoscopic 

4263 examination at 3 months. There were no significant compound-related alterations in any of the 

4264 endpoints monitored. 

4265 Raje et al. (1988) administered dichloromethane (250 or 500 mg/kg) by subcutaneous 

4266 injection three times per week for 4 weeks to male Swiss-Webster mice (20/group). Mating with 

4267 unexposed females started 1 week after the last exposure and continued for 2 weeks. After the 

4268 mating period, the males were sacrificed, and the testes were examined microscopically. On 

4269 GD 17, the females were sacrificed and the uterine horns examined for live, dead, or resorbed 

4270 fetuses. The authors reported that exposure to dichloromethane had no statistically significant 

4271 effects on number of litters, implants/litter, live fetuses/litter, percent dead/litter, percent 

4272 resorbed/litter, or fertility index. Examination of the testes showed no significant alterations 

4273 compared with controls. 

4274 

4275 
128

4276 4.3.1.2. Inhalation Studies

4277 Nitschke et al. (1988b) conducted a two-generation reproductive toxicity study in rats. 

4278 Groups of F344 rats (30/sex/dose level) were exposed by inhalation in whole-body chambers to 

4279 0, 100, 500, or 1,500 ppm dichloromethane (99.86% pure) 6 hours/day, 5 days/week for 

4280 14 weeks and then mated to produce the F1 generation. Exposure of dams continued after 

4281 mating on GDs 0–21 but was interrupted until PND 4. After weaning 30 randomly selected F1 

4282 pups/sex/dose level were exposed as the parental generation for 17 weeks and subsequently 

4283 mated to produce the F2 generation. The results showed no statistically significant exposure4284 related changes in reproductive performance indices (fertility, litter size), neonatal survival, 

4285 growth rates, or histopathologic lesions in F1 (Table 4-25) or F2 weanlings sacrificed at time of 

4286 weaning. According to the authors none of the values in Table 4-25 was significantly different 

4287 from control values (α = 0.05). 

4288 

Table 4-25. Reproductive outcomes in F344 rats exposed to 

dichloromethane by inhalation for 14 weeks prior to mating and from 

GDs 0–21 

Exposure (ppm)a

0 100 500 1,500 

Fertility indexb 77% 77% 63% 87% 

Gestation indexc 100% 100% 100% 100% 

Gestation survival indexd 99.6% 100% 100% 96.6% 

 4-day survival indexe

 91.0% 95.2% 98.5% 98.6% 

 28-day survival indexf 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/m3, 500 ppm = 1,737 mg/m3 , 1,500 ppm = 5,210 mg/m3, 

bNumber of females delivering a litter expressed as a percentage of females placed with a male. 

cNumber of females delivering a live litter expressed as a percentage of the number of females delivering 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 

4290 
129

4291 Raje et al. (1988) exposed groups of male Swiss-Webster mice (20/group) to 0, 100, 150, 

4292 or 200 ppm dichloromethane (HPLC grade, JT Baker Chemical Co.) in inhalation chambers for 

4293 2 hours/day, 5 days/week for 6 weeks. Mating with unexposed females started 2 days after the 

4294 last exposure. As in the subcutaneous injection protocol described in the previous section, after 

4295 the 2-week mating period, the males were sacrificed and the females were sacrificed on GD 17. 

4296 Exposure of the male mice to dichloromethane had no statistically significant effects on number 

4297 of litters, implants/litter, live fetuses/litter, percent dead/litter, or percent resorbed/litter, and no 

4298 significant alterations in the testes were noted. The fertility index was 95%, 95%, 80%, and 80% 

4299 in the control, 100, 150 and 200 ppm groups, respectively. This decrease was not statistically 

4300 significant as reported by the authors. Details of the statistical analyses were not provided. The 

4301 overall Chi-square p-value was 0.27. Using a Cochran-Armitage exact trend test on these data, 

4302 EPA calculated a one-sided p-value of 0.059. Individual p-values for the comparison of each 

4303 group with the control group were 0.97, 0.17, and 0.17 for the 100, 150 and 200 ppm groups, 

4304 respectively. The results for the combined 150 and 200 ppm groups were statistically different 

4305 from the combined controls and 100 ppm group (Fisher’s exact test, one-sided p-value = 0.048), 

4306 suggesting a NOAEL of 100 ppm and LOAEL of 150 ppm. 

4307 

4308 4.3.2. Developmental Toxicity Studies 

4309 The metabolism of dichloromethane into CO by CYP2E1 raises concerns pertaining to 

4310 developmental neurotoxicity. Gestational exposure to CO results in developmental toxicity and 

4311 there are reports indicating that exposures as low as 75 ppm CO can result in significant 

4312 neurological effects in offspring (Giustino et al., 1999). Neurobehavioral deficits in offspring 

4313 include impaired avoidance behavior (De Salvia et al., 1995) and memory (Giustino et al., 1999). 

4314 Neurochemical changes, such as abnormal dopaminergic function (Cagiano et al., 1998) and 

4315 disruption of neuronal proliferation (Fechter, 1987), have also been observed. Oral and 

4316 inhalation dichloromethane exposure studies have demonstrated increased blood CO levels (see 

4317 section 3.3). In addition, increased blood CO levels were seen in rat fetuses exposed through 

4318 maternal inhalation to 500 ppm dichloromethane on GD 21 (Anders and Sunram, 1982), and 

4319 placental transfer of dichloromethane also occurs (Withey and Karpinski, 1985; Anders and 

4320 Sunram, 1982) 

4321 

4322 4.3.2.1. Oral (Gavage) Studies and Culture Studies

4323 Narotsky and Kavlock (1995) evaluated developmental effects of dichloromethane 

4324 (99.9% pure) in F344 rats (17–21/dose group) treated with 0, 337.5, or 450 mg/kg-day 

4325 dichloromethane by gavage in corn oil on GDs 6–19. Dams were weighed on GDs 6, 8, 10, 13, 

4326 16, and 20 and allowed to deliver naturally. They were sacrificed on PND 6 to count uterine 

4327 implantation sites. Pups were grossly examined for developmental abnormalities and weighed 

4328 on PNDs 1, 3, and 6. Dead pups or pups with no gross abnormalities were sacrificed and 
130

4329 examined for soft tissue abnormalities. Maternal weight gain during pregnancy was significantly 

4330 reduced in high-dose dams (by 33%, as estimated from Figure 5 of the paper); this group also 

4331 exhibited rales and nasal congestion. Treatment with dichloromethane did not induce resorptions 

4332 or alter the number of live litters on PND 1 or 6, the number of implants, the number of live pups 

4333 on PND 1 or 6, or pup weight per litter. No gross or soft tissue abnormalities were observed. 

4334 Rat embryos in culture medium were exposed to 0, 3.46, 6.54, 9.79, or 11.88 μmol/mL 

4335 dichloromethane for 40 hours. At the end of the exposure, embryos were observed for 

4336 development of yolk sac vasculature, crown-rump length, total embryonic protein content, and 

4337 number of somite pairs. A concentration of dichloromethane of 6.54 μmol/mL of culture 

4338 medium resulted in decreased crown-rump length, decreased somite number, and decreased 

4339 amount of protein per embryo, whereas no effects were seen at 3.46 μmol/mL (Brown-Woodman 

4340 et al., 1998). A time-course experiment conducted with a concentration of dichloromethane of 

4341 9.22 μmol/mL showed that marked differences in growth and development from controls were 

4342 not significant until about 8 hours of culture. Brown-Woodman et al. (1998) noted that the 

4343 concentrations that caused embryotoxicity in this study were much higher than those found in 

4344 individuals studied under controlled exposure conditions and comparable to those found in 

4345 postmortem blood after fatal inhalation. 

4346 

4347 4.3.2.2. Inhalation Studies

4348 Schwetz et al. (1975) exposed pregnant Swiss-Webster mice (30–40/group) and Sprague4349 Dawley rats (20–35/group) by inhalation in whole-body chambers to 0 or 1,250 ppm 

4350 dichloromethane (97.86% pure) 7 hours/day on GDs 6–15. Maternal BWs were recorded on 

4351 GDs 6, 10, and 16 and on the day of sacrifice (GD 18 for mice, GD 21 for rats). At sacrifice, 

4352 uterine horns were excised and examined for fetal position and number of live, dead, or absorbed 

4353 fetuses. Fetuses were observed for gross, soft tissue, and skeletal abnormalities. The only 

4354 effects seen on developing fetuses were changes in the incidence of minor skeletal variants. In 

4355 rats, the incidence of lumbar ribs or spurs was significantly decreased compared with controls, 

4356 whereas the incidence of delayed ossification of sternebrae was significantly greater than in 

4357 controls. In mice, a significant number of litters contained pups with a single extra center of 

4358 ossification in the sternum. Exposure to dichloromethane produced significantly elevated blood 

4359 COHb content in dams of both species (approximately 9–10% after 10 exposures versus 1–2% in 

4360 controls). BWs in exposed mouse dams were significantly increased (11–15%) compared with 

4361 those in controls but were not affected in exposed rat dams. Mean absolute liver weights of 

4362 exposed dams of both species were significantly elevated compared with controls, but mean 

4363 relative liver weights were not affected. The results indicate that 1,250 ppm was a LOAEL for 

4364 minimal maternal effects (increased COHb and increased absolute liver weight) and a LOAEL 

4365 for adverse effects on the fetuses. 
131

4366 Hardin and Manson (1980) conducted a study in female Long-Evans rats to determine 

4367 whether exposure before and during gestation is more detrimental to reproductive outcome than 

4368 exposure either before or during gestation alone. Four groups of 16–21 rats were formed in 

4369 which the rats were exposed by inhalation in whole-body chambers to 4,500 ppm 

4370 dichloromethane (technical grade, >97% pure) 6 hours/day for 12–14 days before breeding 

4371 and/or on GDs 1–17 or were exposed to filtered air. Maternal BWs were measured every 4 days. 

4372 Dams were euthanized on GD 21 and livers and uteri removed. Livers were weighed, and 

4373 uterine horns were examined for fetal position and number of live, dead, or absorbed fetuses. 

4374 Fetuses were observed for gross, soft-tissue, and skeletal abnormalities. Exposure during 

4375 gestation (with or without pre-gestation exposure) significantly increased maternal liver weight 

4376 (absolute and relative) by about 10–12 and 9–12%, respectively, and decreased fetal BW by 

4377 about 9–10% relative to those exposed to filtered air during gestation. None of the groups 

4378 showed significant alterations in the incidence of gross, external, skeletal, or soft-tissue 

4379 anomalies. Using the same study design and exposure level, Bornschein et al. (1980) observed 

4380 behavioral activities at various ages. Assessed activities included head movement/pivoting when 

4381 placed in a novel environment (4 days of age), limited crawling (10 days), movement in a 

4382 photocell cage (15 days), use of running wheel (45–108 days), and shock avoidance (4 months). 

4383 Exposure during gestation (with or without pre-gestation exposure) caused altered rates of 

4384 behavioral habituation to novel environments in the pups tested as early as 10 days of age that 

4385 were still present at 150 days of age. Growth, food and water consumption, wheel running 

4386 activity, and avoidance learning were not significantly affected by exposure to dichloromethane. 

4387 The results indicate that 4,500 ppm was a LOAEL for maternal effects (10% increased absolute 

4388 and relative liver weight) and for effects on the fetuses (10% decreased fetal BW and altered 

4389 behavioral habituation to novel environments). 

4390 In a study of early-life (including gestational) exposures, Maltoni et al. (1988) exposed 

4391 54 pregnant Sprague-Dawley rats to 100 ppm dichloromethane via inhalation 4 hours/day, 

4392 5 days/week for 7 weeks, followed by 7 hours/day, 5 days/week for 97 weeks. Exposure 

4393 apparently started on GD 12. Groups of 60 male and 69 female newborns continued to be 

4394 exposed after birth to 60 ppm dichloromethane 4 hours/day, 5 days/week for 7 weeks, followed 

4395 by exposure 7 hours/day, 5 days/week for 97 weeks. Additional groups of 60 male and 

4396 70 female newborn were exposed after birth to 60 ppm dichloromethane 4 hours/day, 

4397 5 days/week for 7 weeks and then for 7 hours/day, 5 days/week for 8 weeks. BWs were 

4398 measured every 2 weeks during exposure and every 8 weeks thereafter. At the end of exposure, 

4399 animals were sacrificed and histologic examinations were performed on 20 tissue types. 

4400 Early life exposures of Sprague-Dawley rats to dichloromethane (Maltoni et al., 1988) 

4401 did not affect mortality or BW in any group. Also, there was no significant effect of exposure to 

4402 dichloromethane on the percentage of animals with benign and malignant tumors and malignant 

4403 tumors, the number of malignant tumors per 100 animals, or the percentage of animals with 
132

4404 benign mammary tumors, malignant mammary tumors, leukemias, pheochromocytomas, and 

4405 pheochromoblastomas. The results provide no evidence that gestational exposure to 100 ppm 

4406 dichloromethane during early life stages of development increases the susceptibility of Sprague4407 Dawley rats to the potential carcinogenicity of dichloromethane, but further conclusions from 

4408 these results are precluded because the study included only one exposure level that was below 

4409 the maximum tolerated dose for adult Sprague-Dawley rats. Experiments comparing cancer 

4410 responses from early-life exposures with adult exposures are not available for F344 rats or 

4411 B6C3F1 mice, the strains of animals in which carcinogenic responses to dichloromethane have 

4412 been observed. 

4413 In summary, the potential for gestational exposure to CO, resulting from maternal 

4414 dichloromethane exposure via oral and inhalation routes, raises concerns regarding 

4415 neurodevelopmental effects. In addition, dichloromethane transfer across the placenta has also 

4416 been seen in inhalation exposure studies in rats (Withey and Karpinski, 1985; Anders and 

4417 Sunram, 1982). Although few developmental effects were observed at high exposures of 

4418 dichloromethane (Bornschein et al., 1980; Schwetz et al., 1975), there are no studies that have 

4419 thoroughly evaluated neurobehavioral and neurochemical changes resulting from gestational 

4420 dichloromethane exposure. The available data identify changes of behavior habituation at 

4421 4,500 ppm (Bornschein et al., 1980) and increases in COHb at 1,250 ppm (Schwetz et al., 1975). 

4422 The behavioral changes observed at 4,500 ppm indicate developmental neurotoxic effects. No 

4423 other neurological endpoints have been evaluated in the available developmental studies of 

4424 dichloromethane, but increases in blood COHb strongly suggest that dichloromethane is being 

4425 metabolized to CO. Gestational exposure to CO can result in significant neurological effects in 

4426 offspring, including neurobehavioral deficits (De Salvia et al., 1995), memory effects (Giustino 

4427 et al., 1999), and neurochemical changes (Cagiano et al., 1998; Fechter, 1987). As a result, it is 

4428 unknown if developmental neurotoxicity could occur at lower exposures to dichloromethane. 

4429 

4430 4.4. OTHER DURATION- OR ENDPOINT-SPECIFIC STUDIES 

4431 4.4.1. Short-term (2-Week) Studies of General and Hepatic Effects in Animals 

4432 Two short-term (2-week) studies examined hepatic and renal effects of dichloromethane 

4433 exposure in F344 rats (Berman et al., 1995) and CD-1 mice (Condie et al., 1983). Berman et al. 

4434 (1995) administered dichloromethane by gavage in corn oil for up to 14 days to groups of eight 

4435 female F344 rats at dose levels of 0, 34, 101, 337, or 1,012 mg/kg-day. Starting at day 4, deaths 

4436 occurred in the 1,012 mg/kg-day exposure group, with seven of eight rats dying before the end of 

4437 the 14-day exposure period. In the dose groups that did not experience this high mortality, 

4438 incidences of increased necrotic hepatocytes were 0/8, 0/8, 0/8, and 3/8 for the 0, 34, 101, and 

4439 337 mg/kg-day groups, respectively. The increase in liver lesions was not accompanied by 

4440 increases in serum activities of ALT or AST. Kidneys, spleen, and thymus were also 

4441 histopathologically examined in this study, but none showed exposure-related lesions. The 
133

4442 results indicate that 101 mg/kg-day was a NOAEL and 337 mg/kg-day was a LOAEL for 

4443 increased incidence of degenerative lesions in female rats exposed for 14 days. In a companion 

4444 study with groups of eight female F344 rats that were given single doses of 0, 101, 337, 1,012, or 

4445 1,889 mg/kg-day, incidences of rats with increased necrotic hepatocytes were 1/8, 0/8, 8/8, 7/8, 

4446 and 8/8, respectively (Berman et al., 1995). 

4447 Condie et al. (1983) detected exposure-related liver lesions in a 14-day gavage study in 

4448 which dichloromethane in corn oil was administered to male CD-1 mice at dose levels of 0, 133, 

4449 333, or 665 mg/kg-day. Incidences of mice with minimal or slight cytoplasmic vacuolation were 

4450 1/16, 0/5, 3/5, and 4/5 for the control through high-dose groups, respectively. The kidneys were 

4451 also examined histopathologically in this study but showed no exposure-related lesions. No 

4452 other tissues were prepared for histologic examination. Blood urea nitrogen, serum creatinine, 

4453 and serum ALT activities were not significantly altered by exposure. All dose levels 

4454 significantly reduced to the same extent the active transport of p-aminohippurate into renal 

4455 cortical slices in vitro, a measure of proximal tubule function. The results most clearly identify 

4456 133 mg/kg-day as a NOAEL and 333 mg/kg-day as a LOAEL for increased incidence of 

4457 hepatocyte vacuolation in male mice. 

4458 

4459 4.4.2. Immunotoxicity Studies in Animals 

4460 Aranyi et al. (1986) studied the effects of acute inhalation exposures to 50 or 100 ppm 

4461 dichloromethane on two measures of immune response (susceptibility to respiratory infection 

4462 and mortality due to Streptoccocus zooepidemicus exposure and ability of pulmonary 

4463 macrophages to clear infection with Klebsiella pneumoniae). Female CD1 mice that were 5–

4464 7 weeks of age at the start of the exposure portion of the experiment were used for both assays. 

4465 Up to five replicate groups of about 30 mice were challenged with viable S. zooepidemicus

4466 during simultaneous exposure to dichloromethane or to filtered air. Deaths were recorded over a 

14-day observation period. Clearance of 35 4467 S-labeled K. pneumoniae by pulmonary macrophages 

4468 was determined by measuring the ratio of the viable bacterial counts to the radioactive counts in 

4469 each animal’s lungs 3 hours after infection; 18 animals were used per dose group. A single 3-

4470 hour exposure to 100 ppm dichloromethane significantly increased the susceptibility to 

4471 respiratory infection and greater mortality following exposure to S. zooepidemicus (p ≤ 0.01). 

4472 Twenty-six deaths occurred in 140 (18.6%) mice challenged during a 3-hour exposure to 100 

4473 ppm dichloromethane; in contrast, nine deaths occurred in 140 mice (6.4%) exposed to filtered 

4474 air. The 3-hour exposure to 100 ppm dichloromethane was associated with a statistically 

4475 significant (p ≤ 0.001) 12% decrease in pulmonary bactericidal activity (91.6 and 79.6% of 

4476 bacteria killed in controls and 100 ppm group, respectively). No difference was seen in either 

4477 mortality rate or bactericidal activity in experiments using a single 3-hour exposure to 50 ppm or 

4478 3-hour exposures to 40 ppm dichloromethane repeated daily for 5 days compared with control 

4479 animals exposed to filtered air. These results suggest that 3-hour exposure to 50 ppm 
134

4480 dichloromethane was a NOAEL and 100 ppm was a LOAEL for decreased immunological 

4481 competence (immunosuppression) in CD-1 mice. 

4482 Aranyi et al. (1986) also conducted a similar set of experiments with 13 other chemicals 

4483 (acetaldehyde, acrolein, propylene oxide, chloroform, methyl chloroform, carbon tetrachloride, 

4484 allyl chloride, benzene, phenol, monochlorobenzene, benzyl chloride, perchloroethylene, and 

4485 ethylene trichloride). Perchloroethylene and ethylene trichloride were the only chemicals in this 

4486 group for which an increased mortality risk from streptococcal pneumonia was seen (mortality 

4487 risk 15.0 and 31.4% in controls and 50 ppm exposure groups for perchloroethylene and 13.4 and 

4488 58.1% in controls and 50 ppm exposure groups for ethylene trichloride). Decreased bactericidal 

4489 activity was also seen with acetaldehyde, acrolein, methyl chloroform, allyl chloride, benzene, 

4490 benzyl chloride, perchloroethylene, and ethylene trichloride at one or more exposures. Results 

4491 from several chemicals suggest that 5 days of exposure results in greater decrease in bactericidal 

4492 activity (i.e., acetaldehyde, acrolein, and benzene), and others (e.g., perchloroethylene) suggest 

4493 that 5 days of exposure does not result in greater suppression than a single exposure period. 

4494 There was considerable variation in both measures of immune response among the 

4495 controls in the experiments (Aranyi et al., 1986). Among the controls in the experiments with 

4496 the 13 chemicals other than dichloromethane, mortality in the streptococcal infectivity model 

ranged from 5.7–22.1%, with a mean of 12.7%.54497 Bactericidal activity in the klebsiella model 

4498 among controls ranged from 67.9–94.7%, with a mean of 81.8%. The number of bacteria 

4499 deposited in the lung in an inhalation bacterial infectivity model can show considerable variation, 

4500 (i.e., between 750 to 1,500 viable streptococcus or klebsiella organisms, [Ehrlich, 1980]). 

4501 Therefore, concurrent controls are particularly import due to the variation in preparation and 

4502 aerosol administration of the bacteria in these assays. 

4503 Warbrick et al. (2003) evaluated immunocompetence in male and female Sprague4504 Dawley rats by measuring the immunoglobulin M (IgM) antibody responses following 

4505 immunization with sheep red blood cells in addition to hematological parameters and 

4506 histopathology of the spleen, thymus, lungs, and liver. Groups of rats (8/sex/dose level) were 

4507 exposed to 0 or 5,000 ppm dichloromethane 6 hours/day, 5 days/week for 28 days. Rats injected 

4508 with cyclophosphamide served as positive controls. Five days before sacrifice (day 23 of 

4509 exposure) all rats were injected with sheep red blood cells. IgM levels in response to the sheep 

4510 red blood cells were comparable between dichloromethane-exposed and air-exposed rats, 

4511 indicating that dichloromethane did not produce immunosuppression in the animals under these 

4512 exposure conditions. Cyclophosphamide-treated animals had significantly lower levels of IgM 

4513 in the blood serum, indicating immunosuppression. Rats exposed to dichloromethane showed 

4514 reduced response to sound, piloerection, and hunched posture during exposures. Neither BW 

4515 gain nor the hematological parameters monitored were significantly affected by exposure to 

 

5EPA did not include the duplicate assay of perchloroethylene in calculating this summary statistic. If this 

additional assay is included, the mortality risk ranges from 5.7–45.7%, with a mean of 15.0%. 
135

4516 dichloromethane. Relative and absolute liver weights were significantly increased in females, 

4517 but not in males. Relative spleen weight was reduced in females, and no significant changes 

4518 were seen in the weight of the thymus and lungs. Histopathology of the tissues examined was 

4519 unremarkable. Exposure to 5,000 ppm dichloromethane did not affect antibody production to the 

4520 challenge with sheep red blood cells. 

4521 In the 2-year drinking water study (Serota et al., 1986a, b) and 2-year inhalation study 

4522 (Nitschke et al., 1988a), histopathologic analyses were conducted on the lymph nodes, thymus, 

4523 and spleen among several other organs, and no significant changes were noted. 

4524 In summary, one study (Aranyi et al., 1986) demonstrated evidence of 

4525 immunosuppression, including increased risk of streptococcal-pneumonia-related mortality and 

4526 decreased clearance of klebsiella bacteria following a single dichloromethane exposure at 

4527 100 ppm for 3 hours in CD-1 mice. The streptococcal and klebsiella bacterial inhalation assays 

4528 are models of respiratory infection that test for local immune effects associated with inhalation 

4529 exposure rather than systemic immunosuppression. The NOAEL identified in this study was 

4530 50 ppm. In contrast, in a functional immune assay of systemic immunosuppression conducted in 

4531 rats, Warbrick et al. (2003) did not observe changes in the antibody response to sheep red blood 

4532 cells in a 28-day inhalation exposure to 5,000 ppm dichloromethane. Histopathologic analyses 

4533 of immune system organs in chronic exposure studies for B6C3F1 mice and F344 rats (Nitschke 

4534 et al., 1988a; Serota et al., 1986a, b) revealed no changes from controls. However, no assays of 

4535 functional immunity were included in these chronic studies. The limited database for 

4536 dichloromethane does not suggest systemic immunosuppression, but the Aranyi et al. (1986) 

4537 study provides evidence of route-specific local immunosuppression from acute exposure studies 

4538 in CD1 mice. Due to the acute exposure duration used in Aryani et al. (1986), the immune 

4539 effects of short-term or chronic exposure to dichloromethane are unclear. 

4540 

4541 4.4.3. Neurotoxicology Studies in Animals 

4542 Neurological evaluations in animals during and after exposure to dichloromethane have 

4543 resulted in CNS depressant effects similar to other chlorinated solvents (e.g., trichloroethylene, 

4544 perchloroethylene) and ethanol. Overall, there are decreased motor activity, impaired memory, 

4545 and changes in responses to sensory stimuli. Neurobehavioral, neurophysiological, and 

4546 neurochemical/ neuropathological studies have been used to characterize the effects of 

4547 dichloromethane on the CNS. A brief overview of these types of studies is provided below, 

4548 followed by a detailed description of individual studies. 

4549 Neurobehavioral studies with dichloromethane used protocols to measure changes in 

4550 spontaneous motor activity, a functional observational battery (FOB) test (to evaluate gross 

4551 neurobehavioral deficits), and a task developed to assess learning and memory. The FOB 

4552 protocol includes various autonomic parameters, neuromuscular parameters, sensorimotor 

4553 parameters, excitability measures, and activity. Learning and memory changes with 
136

4554 dichloromethane were studied by using a passive avoidance task. The oral and inhalation studies 

4555 that examined neurobehavioral endpoints are summarized in Table 4-26. 

4556 Neurophysiological studies with dichloromethane exposure consisted of measuring 

4557 evoked responses in response to sensory stimuli. In these studies, animals were implanted with 

4558 electrodes over the brain region that responds to the particular stimuli. For example, an electrode 

4559 would be implanted over the visual cortex in an animal presented with a visual stimulus. Once 

4560 the stimulus is presented to the animal, an evoked response is elicited from the brain region and 

4561 transmitted to the implanted electrode. During administration of a chemical, if there is a 

4562 significant change in the magnitude, shape, and latency (among other measures) in the evoked 

4563 response, then the chemical is considered to produce neurological effects. A summary of studies 

4564 examining dichloromethane exposure and neurophysiological changes is shown in Table 4-27. 

4565 In neurochemical/neuropathological studies with dichloromethane, animals were first 

4566 exposed to dichloromethane (via oral, inhalation, or injection), and then the brains were 

4567 removed. Changes in excitatory neurotransmitters, such as glutamate and acetylcholine, and the 

4568 inhibitory neurotransmitter, GABA, were measured. Additionally, dopamine and serotonin 

4569 levels, which are associated with addiction and mood, were also measured. Other parameters 

4570 that were measured included DNA/protein content and regional brain changes in the cerebellum 

4571 and hippocampus. Measurement of neurochemical changes provides mechanistic information, 

4572 and neurobehavioral and neurophysiological effects can be correlated to these results. Table 

4573 4-28 summarizes studies of neurochemical changes and dichloromethane. 
137

Inhalation exposure

Spontaneous activity 

 NMRI mouse, male 400–2,500 ppm 1 hour Initial increase in activity followed by a 

pronounced decrease at exposures 600 ppm and higher 

Kje

 Rat, male 5,000 ppm 1 hour, every other day for 10

days 

Decreased spontaneous locomotor activity 

Hep

 Wistar rat, male 500 ppm 6 hours/day, 6 days Increased preening frequency Sav ICR mouse, female 5,000 ppm Continuous, 7 days Increased spontaneous activity in first 

few hours and then decreased activity 

We

 Sprague-Dawley rat, male 1,000, 5,000 ppm Continuous, 14 weeks No neurobehavioral changes Hau

 ICR mouse, female 1,000, 5,000 ppm Continuous, 14 weeks Incoordination, lethargy Hau Beagle dog, female 1,000, 5,000 ppm Continuous, 14 weeks Incoordination, lethargy Hau Rhesus monkey, female 1,000, 5,000 ppm Continuous, 14 weeks Incoordination, lethargy Hau

 ICR mouse, female 25, 100 ppm Continuous, 14 weeks Increased spontaneous activity at 25 ppm Tho

Functional observational battery 

 F344 rat, male and female 50, 200, 2,000 ppm 6 hours/day, 5 days/week, 13 

weeks + 65 hours exposure free 

No effects observed on FOB, grip strength 

Ma

Learning and memory

Swiss-Webster mouse, male 

47,000 ppm Approximately 20 seconds + 

1 hour exposure free before training; retested at days 1, 2, and 4 

Significant decrease in learning and recall ability 

Ale

4574 

Table 4-26. Studies of neurobehavioral changes from dichloromethane, by route of exposure and ty

Species Exposure(s) Duration Neurobehavioral effect 

Oral and gavage exposure 

Functional observational battery 

 F344 rat, female 101, 337, 1,012, 

1,889 mg/kg, gavage 

Acute—evaluated 4 and 24 hours after dosing 

FOB neuromuscular and sensorimotor parameters significantly different from controls at 1,012 and 1,889 mg/kg (337 mg/kg = NOAEL) 

Mo

 F344 rat, female 34, 101, 337, 1,012 

mg/kg-day, gavage 

14 day—evaluated on days 4, 9, and 15. 

All FOB parameters (except activity) significantly affected from day 4 at doses 

of 337 and 1,012 mg/kg-day 

Mo
138

4575 4576 

a 4577 SEP = somatosensory-evoked potential; FEP = flash-evoked potential; BAER = brainstem-auditory-evoked response; CAEP = cortical

4578 potential; i.p. = intraperitoneal. 4579 

Table 4-27. Studies of neurophysiological changes as measured by evoked potentials resulting from 

dichloromethane, by route of exposure 

Species Exposure(s) Duration SEPsa measured Effect 

Intraperitoneal

Long-Evans rat, male 57.5, 115, 230, 

460 mg/kg, i.p.a

Acute; tested at 15 minutes, 1 hour, and 5 hours after dosing 

FEPa Significant changes in FEPs were no

in animals dosed 115 mg/kg and higFEP changes time and dose depende

Inhalation Exposure 

F344 rat, male 5,000, 10,000, 

15,000 ppm 

Acute, 1 hour; tested during exposure 

Electroencephalogram, BAERa, CAEPa, FEP, 

SEP 

Significant changes in SEP, FEP, BAER, and CAEP responses at all exposures; slight recovery noted at 1 hour after exposure 

F344 rat, male and female 50, 200, 

2,000 ppm 

Subchronic, 6 hour/day, 5 days/week, 13 weeks; tested 65 hours after last exposure 

FEP, CAEP, BAER, SEP 

No significant changes noted in any evoked potential measurements 
139

4580 

a 4581 ↑ = increase; ↓ = decrease.

Table 4-28. Studies of neurochemical changes from dichloromethane, by route of exposure 

Species and sex Exposure Duration Regions Effecta 

Oral exposure

Sprague-Dawley rat, male 534 mg/kg Acute, single dose; evaluated 

2 hours after dosed 

Hippocampus, medulla, midbrain, hypothalamus 

↑ acetylcholine in hippocampus ↑ dopamine and serotonin in med↓ norepinephrine in midbrain ↓ norepinephrine and serotonin ihypothalamus 

Inhalation exposure 

Wistar rat, male 1,000 ppm TWA (basal 

exposure of 100 ppm + 2,800 ppm, 1 hour peak exposures at hours 1 and 4) 

6 hours/day, 5 days/week, 2 weeks 

Cerebrum, cerebellum 

↑ NADPH diaphorase, succinate dehydrogenase in cerebrum ↑ cerebral RNA ↓ succinate dehydrogenase in cerebellum 

Wistar rat, male 1,000 ppm TWA 6 hours/day, 5 days/week, 

2 weeks + 7 days exposure free 

Cerebrum, cerebellum 

↓ succinate dehydrogenase in boregions 

Wistar rat, male 1,000 ppm 6 hours/day, 5 days/week, 

2 weeks 

Cerebrum, cerebellum 

↑ acid proteinase ↓ succinate dehydrogenase in cerebellum 

Wistar rat, male 1,000 ppm 6 hours/day, 5 days/week, 

2 weeks + 7 days exposure free 

Cerebrum ↓ cerebral RNA 

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 ppm) No effect on luteinizing hormone release 

Mongolian gerbil, male and female 

210, 350 ppm Continuous (24 hours/day), 

3 months + 4 months exposure free 

Hippocampus, cerebellum cerebral cortex 

↓ DNA concentration per wet wehippocampus (210, 350 ppm) ancerebellar hemispheres (350 ppm↑ astroglial proteins in frontal ansensory motor cerebral cortex 

Mongolian gerbil, male and female 

210 ppm Continuous (24 hours/day), 

3 months 

Frontal cortex, cerebellum 

↓ glutamate, GABA, phosphoethanolamine in frontal ↑ glutamate, GABA in posterior cerebellar vermis 

Mongolian gerbil, male and female 

210 ppm Continuous (24 hours/day), 

3 months + 4 months exposure free 

Hippocampus, olfactory bulbs, cerebral cortex 

↓ DNA concentration per wet wehippocampus only 
140

4582 4.4.3.1. Neurotoxicology Studies—Oral Exposures 

4583 Three studies evaluated the neurotoxic potential of dichloromethane by either 

4584 administering the solvent orally or by injection; two of these studies (Herr and Boyes, 1997; 

4585 Kanada et al., 1994) only evaluated acute effects (2–5 hours) from single-dose exposures. 

4586 Observed neurological effects included decreased spontaneous activity (Moser et al., 1995), 

4587 changes in flash-evoked potential (FEP) measurements (Herr and Boyes, 1997), and changes in 

4588 catecholamine levels in the brain (Kanada et al., 1994). 

4589 Moser et al. (1995) conducted neurobehavioral evaluations in female F344 rats following 

4590 an acute or 14-day oral administration of dichloromethane. A FOB protocol was utilized to 

4591 determine changes in autonomic parameters (lacrimation, salivation, pupil response, urination, 

4592 defecation), neuromuscular parameters (gait, righting reflex, forelimb and hind-limb grip 

4593 strength, landing foot splay), sensorimotor parameters (tail pinch, click response, touch 

4594 response), excitability measures (handling reactivity, arousal, clonic, and/or tonic movements), 

4595 and activity (rearing, motor activity). A baseline FOB was performed on all rats prior to initial 

4596 dichloromethane administration. After dichloromethane administration, a FOB was conducted at 

4597 selected time points followed by a motor activity test in a maze. In the acute study, rats were 

4598 dosed with 0, 101, 337, 1,012, or 1,889 mg/kg dichloromethane. At 4 and 24 hours after the 

4599 administered dose, rats were tested for the neurological parameters. Significant changes in the 

4600 neuromuscular and sensorimotor parameters were observed and occurred mostly in rats 

4601 administered with the highest dose. These significant changes were only observed at the 4-hour 

4602 time point and not when measured at 24 hours. The NOAEL identified by the authors for this 

4603 study was 337 mg/kg, based on no observable changes in the FOB. In the 14-day study, rats 

4604 were administered 0, 34, 101, 337, or 1,012 mg/kg-day. FOB testing was conducted on days 4 

4605 and 9 (before the daily dose) and approximately 24 hours after the last (14th) dose. With the 

4606 exception of the activity measurements, all other neurobehavioral parameters (neuromuscular, 

4607 sensorimotor, autonomic, excitability) were significantly affected from the 4th day through the 

4608 entire 14-day exposure cycle. The NOAEL identified for the 14-day study was 101 mg/kg-day, 

4609 based on FOB changes associated with the dichloromethane exposure. 

4610 A single dose acute neurophysiology study by Herr and Boyes (1997) evaluated the effect 

4611 of dichloromethane on FEPs in adult male Long-Evans rats. Rats were implanted with epidural 

4612 electrodes over the visual cortex area. After placement in an enclosed rectangular mirror 

4613 chamber, PEPs were stimulated with a 10 μsec flash. Baseline FEPs were collected and rats 

4614 were injected intraperitoneally with 0 (corn oil, n = 16), 57.5 (n = 15), 115 (n = 15), 230 (n = 14), 

4615 or 460 (n = 15) mg/kg dichloromethane. Animals were retested at 15 minutes, 1 hour, and 

4616 5 hours after injection. Amplitude decreases in the early FEP components were observed. The 

4617 FEP amplitude changes were time and dose dependent with maximal effects at 15 minutes after 

4618 dichloromethane dosage. All of the waveform amplitudes returned to control levels when 

4619 measured at the 1-hour time point for all doses tested. Response latencies were still different 
141

4620 from controls when measured 5 hours after dosing, but the effect was less pronounced than at the 

4621 15-minute and 1-hour time points. In this study, 57.5 mg/kg did not produce any significant 

4622 changes in the FEP measures as compared to control and was considered this study’s NOAEL. 

4623 The LOAEL was 115 mg/kg based on changes in the FEP amplitudes. 

4624 Kanada et al. (1994) examined the effect of dichloromethane on acetylcholine and 

4625 catecholamines (dopamine, norepinephrine, serotonin) and their metabolites in the midbrain, 

4626 hypothalamus, hippocampus, and medulla from male Sprague-Dawley rats (4–5/group) in a 

4627 neurochemical/neuropathology study. The rats were sacrificed 2 hours after a single gavage dose 

4628 of 0 or 534 mg/kg of undiluted dichloromethane. Administration of dichloromethane 

4629 significantly increased the concentration of acetylcholine in the hippocampus by approximately 

4630 10% and increased dopamine and serotonin levels in the medulla by approximately 75%. 

4631 Dichloromethane decreased norepinephrine levels in the midbrain and hypothalamus by 12–

4632 15%, and serotonin levels were decreased in the hypothalamus by approximately 30%. There 

4633 was a trend toward decreased dopamine in the hypothalamus, but the variability between the 

4634 animals was so high that the effect was not significant. (These values for the percent changes 

4635 were estimated by EPA from the figures presented in the paper.) The authors speculated that 

4636 increased acetylcholine release associated with exposure to dichloromethane and other solvents 

4637 may originate from the nerve terminals. 

4638 

4639 4.4.3.2. Neurotoxicology Studies—Inhalational Exposure 

4640 The database pertaining to neurotoxic effects from inhalation exposure to 

4641 dichloromethane is considerably larger than the oral exposure database. Acute (less than 1 day) 

4642 and short-term (1–14 days) exposures resulted in an initial increase in spontaneous activity 

4643 followed by a decrease for exposures between 500 and 2,500 ppm (Kjellstrand et al., 1985; 

4644 Savolainen et al., 1977). Higher (5,000 ppm) acute and short-term exposures resulted in 

4645 decreased spontaneous activity and lethargy (Weinstein et al., 1972; Heppel and Neal, 1944). 

4646 Longer-term exposures (up to 14 weeks) produced decreased motor activity and lethargy in 

4647 several animals at 1,000 and 5,000 ppm (Haun et al., 1971), and exposures at 25 ppm for 

4648 14 weeks produced significant increases in activity in mice, starting at week 9. CNS depression 

4649 was evidenced by decreased responses in the auditory, visual, and somatosensory regions of the 

4650 brain in a study of sensory-evoked potential effects in 12 adult male F344 rats exposed to 0, 

4651 5,000, 10,000, and 15,000 ppm for 1 hour periods (Rebert et al., 1989). Altered learning and 

4652 memory abilities were demonstrated in young (3-, 5-, and 8-week-old) male Swiss-Webster mice 

4653 exposed to 168 mg/L (~47,000 ppm) dichloromethane for approximately 20 seconds (until there 

4654 was a loss of the righting reflex) (Alexeef and Kilgore, 1983). 

4655 
142

4656 4.4.3.2.1. Inhalational exposure—neurobehavioral studies.

4657 Spontaneous motor activity—acute and short-term studies 

4658 Heppel and Neal (1944) evaluated the neurological effects of 5,000 ppm dichloromethane 

4659 in five male rats by measuring changes in spontaneous activity during and after exposure. The 

4660 five rats were not randomly selected, since the investigators chose to pick out the most active 

4661 animals in the litter. During the 1-hour testing runs, rats were placed in a rotating drum. 

4662 Spontaneous activity was reported as the number of drum revolutions/hour. Twenty control test 

4663 runs (1 run/day) were conducted prior to dichloromethane exposure runs. After the pre-exposure 

4664 period, rats were exposed to 5,000 ppm dichloromethane every other day for 1 hour, and activity 

4665 was measured in the same manner as in the control runs. Once dichloromethane exposure was 

4666 stopped, the animals were allowed to recover for 30 minutes and a second 1-hour test run was 

4667 performed to evaluate spontaneous activity during recovery. On nonexposure days, spontaneous 

4668 activity was also measured in1-hour intervals to compare to the pre-exposure period. A total of 

4669 five dichloromethane exposures, five postexposure, and five nonexposure trials were conducted 

4670 over 10 days. Spontaneous activity significantly declined (p < 0.01, Fisher’s t-test) during 

4671 exposure to 5,000 ppm dichloromethane in comparison to nonexposure days. The average 

4672 number of revolutions for all five rats over the test runs was 576 on nonexposure days and 59 

4673 revolutions during dichloromethane exposure. 

4674 Weinstein et al. (1972) continuously exposed female ICR mice to 5,000 ppm 

4675 dichloromethane for up to 7 days. Clinical behavioral observations of the mice were made 

4676 during dichloromethane exposure. Within the first few hours of exposure, spontaneous activity 

4677 increased in comparison to control animals. After 24 hours of continuous exposure, there was a 

4678 considerable decrease in spontaneous activity as noted by observation only. The mice also 

4679 appeared to be very lethargic and had a hunched posture and a rough hair coat, which are all 

4680 signs of CNS depressive effects in rodents. These effects became progressively worse until after 

4681 96 hours of exposure, where many mice resumed normal activity. After the 7-day exposure, 

4682 mice were nearly as active as the control animals but had a rougher coat and were judged to be 

4683 emaciated and dehydrated. 

4684 Male Wistar rats exposed to 500 ppm dichloromethane 6 hours/day for 6 days exhibited 

4685 an increase in preening frequency and time 1 hour after the last exposure relative to controls 

4686 (Savolainen et al., 1977). However, there were no significant changes in other types of 

4687 spontaneous activity. 

4688 In the study by Kjellstrand et al. (1985), male NMRI mice were exposed to 

4689 dichloromethane concentrations ranging from 400 to 2,500 ppm. At concentrations of 600 ppm 

4690 and higher, exposures for 1 hour produced a biphasic pattern of activity characterized by an 

4691 initial increase in activity (as high as 200% of preexposure motor activity at 2,200 ppm, as 

4692 estimated from Figure 6 in Kjellstrand et al. [1985]) during exposure followed by a decreased 

4693 that reached the lowest point 1–2 hours after the end of exposure (as low as 40% motor activity 
143

4694 at 2,200 ppm, in comparison to preexposure, as estimated from Figure 6 in Kjellstrand et al. 

4695 [1985]). Motor activity returned to normal levels after the decreased activity observed 1–2 hours 

4696 after exposure was stopped and indicated that the effect was reversible in this study design. 

4697 

4698 Spontaneous motor activity—subchronic (14 week) studies 

4699 Haun et al. (1971) reported results from studies in which female beagle dogs, female 

4700 rhesus monkeys, male Sprague-Dawley rats, and female ICR mice were continuously exposed to 

4701 0, 1,000, or 5,000 ppm dichloromethane for up to 14 weeks in whole-body exposure chambers. 

4702 Gross and histopathologic examinations were made on animals that died or were sacrificed 

4703 during or at termination of the study. At 5,000 ppm, obvious nervous system effects (e.g., 

4704 incoordination, lethargy) were most apparent in dogs and also observed in monkeys and mice. 

4705 Rats did not demonstrate any of these sedative effects. At 1,000 ppm, these effects were 

4706 observed to a lesser extent in monkeys and mice, but dogs still displayed prominent CNS 

4707 depressive behavior. Histopathologic analysis revealed edema of the brain in three dogs that 

4708 died during exposure to 5,000 ppm dichloromethane. No other gross brain-related changes were 

4709 reported. The results indicate that continuous exposure to 1,000 ppm was an adverse effect level 

4710 for mortality and effects on the nervous system and liver in dogs (exposed for up to 4 weeks) and 

4711 for BW changes in rats (exposed for 14 weeks). The 5,000 ppm level induced mortality in 

4712 beagle dogs, ICR mice, and rhesus monkeys (but not Sprague-Dawley rats); obvious nervous 

4713 system effects in dogs, mice, monkeys, and rats; and gross liver changes in dogs, mice, monkeys, 

4714 and rats. 

4715 In the study by Thomas et al. (1972), female ICR mice were exposed continuously to 0, 

4716 25, or 100 ppm dichloromethane for 14 weeks. Spontaneous activity of mice was evaluated by 

4717 using closed circuit television for monitoring. Mice were evaluated in daily 2-hour testing 

4718 sessions. The 25 and 100 ppm exposure groups were tested for 2 weeks prior to the onset of 

4719 dichloromethane exposure. Starting at week 9, mice exposed to 25 ppm dichloromethane 

4720 exhibited increases in spontaneous activity, but no quantitative measurements or statistical 

4721 analysis were reported. The authors stated that no significant effect was observed in the group 

4722 exposed to 100 ppm. 

4723 

4724 FOB—subchronic (13 week) study

4725 Only one study, a 13-week inhalation study in F344 rats (Mattsson et al., 1990) has 

4726 conducted an FOB testing paradigm following a subchronic exposure to dichloromethane. 

4727 Groups of rats (12/sex/exposure level) were exposed to 0, 50, 200, or 2,000 ppm 

4728 dichloromethane 6 hours/day, 5 days/week for 13 weeks. An additional group of rats was 

4729 exposed to 135 ppm CO to induce approximately 10% COHb, approximately the level produced 

4730 by saturation of oxidative metabolism of dichloromethane. After the 13 weeks of exposure 

4731 (beginning 65 hours after the last exposure), rats were subject to an FOB to evaluate any 
144

4732 neurobehavioral changes from the dichloromethane exposure. Autonomic parameters were first 

4733 characterized and then the rat was placed in a clear plastic box to evaluate locomotor activity and 

4734 then responsiveness to touch, sharp noise, and tail pinch. Hind-limb grip strength was also 

4735 measured by using a strain gauge. All animals were examined clinically at weekly intervals and 

4736 were tested at the end of the exposure period by FOB, grip strength, BW, temperature, and 

4737 sensory-evoked potentials. No exposure-related effects were observed on the FOB, grip 

4738 strength, or sensory-evoked potentials. No histopathologic changes were noted in brains, spinal 

4739 cords, or peripheral nerves from the high-dose dichloromethane group compared with control 

4740 animals. In the absence of changes, lower concentrations were not examined. 

4741 

4742 Learning and memory—acute study 

4743 In a study by Alexeef and Kilgore (1983), a learning and memory evaluation was 

4744 conducted following acute exposure to dichloromethane. Mice were exposed to 168 mg/L 

4745 (~47,000 ppm) dichloromethane and were tested for learning ability by using a passive4746 avoidance conditioning task. Male Swiss-Webster mice (3, 5, and 8 weeks old) were used in this 

4747 study. In the passive avoidance task, mice were placed on a metal platform that extended into a 

4748 hole. If the mouse went into the hole (a darkened area, which would be the preferred area for the 

4749 mouse), it received a foot shock. Prior to the training session, mice were exposed to either air or 

4750 ~47,000 ppm dichloromethane. Animals were exposed to dichloromethane until there was a loss 

4751 of the righting reflex, which would take about 20 seconds on average, and then placed back in 

4752 their home cage. One hour after exposure, animals were trained to learn the passive avoidance 

4753 task. A mouse was considered to have learned the task once it remained on the platform for at 

4754 least 30 seconds without entering the hole. Mice were then tested for recollection of the task at 

4755 either 1, 2, or 4 days after the initial training session. In the learning phase of the task, 74% of 

4756 the control mice retained the task in comparison to 59% of the dichloromethane-exposed group, 

4757 indicating the significant effect of dichloromethane on learning. There was also an age-related 

4758 effect since exposed 3-week-old mice were less likely to recall the task than five- or eight-week4759 old mice. There was no difference in task recall between the 5- and 8-week-old mice. 

4760 Dichloromethane, at the exposure used in the study, was demonstrated to be non-analgesic, since 

4761 pain-response times were comparable to those in air-exposed animals in the hot-plate pain test, 

4762 and therefore the results of the passive avoidance test were not confounded by potential analgesic 

4763 effects. As a result, it is demonstrated that exposure to an acute and high concentration of 

4764 dichloromethane alters learning ability in mice. 

4765 

4766 4.4.3.2.2. Inhalational exposure—neurophysiological studies. The effect of dichloromethane 

4767 on sensory stimuli was evaluated by measuring sensory-evoked responses during an acute 

4768 exposure (Rebert et al., 1989) and following a subchronic (13 week) exposure (Mattsson et al., 

4769 1990). Rebert et al. (1989) evaluated the effects of dichloromethane on sensory-evoked 
145

4770 potentials (auditory, visual, and somatosensory) in F344 rats exposed to 0, 5,000, 10,000, and 

4771 15,000 ppm dichloromethane for 1 hour in a head-only exposure chamber. Twelve adult male 

4772 rats were implanted with chronic epidural electrodes placed over the visual and somatosensory 

4773 cortices. Each rat served as its own control, with a 1-week recovery period between testing 

4774 sessions. During each testing session, spontaneous electroencephalograms were recorded. 

4775 Additionally, brainstem-auditory-evoked responses (BAERs) (tone stimulus), cortical-auditory4776 evoked potentials (CAEPs) (click stimulus), FEPs (flash stimulus), and somatosensory-evoked 

4777 potentials (SEPs) (tail current stimulus) were measured in response to the stimuli. 

4778 Dichloromethane decreased the SEP response to the tail current stimulus, and earlier components 

4779 of the FEP response were attenuated and eventually eliminated with increasing exposures. The 

4780 BAER response profile was also significantly altered. Dichloromethane completely abolished 

4781 the CAEP at all concentrations tested. Slight recovery of this response was noted approximately 

4782 1 hour after exposure. The collective results strongly suggest a CNS depressive profile for 

4783 dichloromethane and indicate that this chemical affects the auditory, visual, and somatosensory 

4784 regions of the brain. 

4785 In a subchronic exposure study, male and female F344 rats were exposed to 

4786 dichloromethane for 6 hours/day, 5 days/week for 13 weeks (Mattsson et al., 1990). Twelve 

4787 animals of each sex were selected for exposure to 0, 50, 200, or 2,000 ppm dichloromethane or 

4788 135 ppm CO. For electrophysiological measures, rats were surgically implanted with epidural 

4789 electrodes 10 weeks after the onset of exposure. Electrodes were placed over the somatosensory, 

4790 visual, and cerebellar region. Electrophysiological measures that were recorded included FEP 

4791 measurements, cortical flick fusion responses, CAEPs, BAERs, and SEPS recorded from the 

4792 sensory (SEP-S) and cerebellar (SEP-C) regions. None of these measures were significantly 

4793 altered by any dichloromethane or CO treatment in this study. However, it should be noted that 

4794 all of the electrophysiological measures were conducted at least 65 hours after the last 

4795 dichloromethane exposure. As a result, it can be concluded that a subchronic exposure to 

4796 dichloromethane did not result in persistent changes in any of the neurophysiological measures 

4797 that were evaluated in this study. It is not known if any neurological compensation occurred 

4798 since SEP measurements were not taken during actual dichloromethane exposure in this 

4799 subchronic study. 

4800 Based on these two studies, the significant changes noted in several SEP measures during 

4801 dichloromethane exposure were not observed after a subchronic exposure where animals were 

4802 tested at least 65 hours after the last exposure. As a result, it is difficult to ascertain if tolerance 

4803 is developed to the dichloromethane-mediated changes in sensory potentials during an acute 

4804 exposure or if these effects are still maintained during repeated exposure, since measurements 

4805 were not taken during the subchronic exposure. 
146

4806 4.4.3.2.3. Inhalational exposure—neurochemistry and neuropathology studies. The studies 

4807 evaluating specific neurochemical changes in relation to dichloromethane exposure include 

4808 studies of effects of short-term (3-day to 2-week) exposures (Fuxe et al., 1984; Savolainen et al., 

4809 1981) and subchronic (3-month) exposures (Karlsson et al., 1987; Briving et al., 1986; 

4810 Rosengren et al., 1986). 

4811 Savolainen et al. (1981) examined three different exposure schemes in male Wistar rats. 

4812 The rats were exposed to 500, 1,000, or 1,000 ppm TWA dichloromethane for 6 hours/day, 

4813 5 days/week for 2 weeks. (Note: The abstract of this paper describes the exposures as 500, 1,000, 

4814 and 100 ppm TWA, but, based on information in the body of the paper, the abstract appears to be 

4815 incorrect.) The 1,000 ppm TWA exposure consisted of a basal 100 ppm exposure with two 

4816 2,800 ppm 1-hour peak concentrations (at 1 and 4 hours) resulting in a time-weighted exposure 

4817 of 1,000 ppm. Brains were removed from rats at the end of study and analyzed. The 1,000 ppm 

4818 TWA group displayed increases in cerebral RNA. Other changes noted for this group in the 

4819 cerebrum included significant increases in NADPH diaphorase and succinate dehydrogenase 

4820 activity. In the 1,000 ppm constant exposure group, acid proteinase activity was below the levels 

4821 observed in control animals in the first week but increased to levels above control animals in the 

4822 second week. In the cerebellum, there were no changes in RNA concentration, and there was a 

4823 decrease in succinate dehydrogenase activity in both the 1,000 and 1,000 ppm TWA groups. 

4824 After a 7-day withdrawal, RNA levels in the cerebrum were significantly lower in the 1,000 ppm 

4825 group. Succinate dehydrogenase levels remained lowered in the 1,000 ppm TWA group after the 

4826 7-day exposure-free period. No significant effects were seen at 500 ppm. 

4827 Fuxe et al. (1984) evaluated changes in brain catecholamine levels after a 3-day exposure 

4828 to dichloromethane, using male Sprague-Dawley rats. Rats were exposed to 70, 300, and 

4829 1,000 ppm dichloromethane 6 hours/day for 3 consecutive days. Additional groups of rats were 

4830 exposed to the same levels of dichloromethane and given intraperitoneal injections of the 

4831 tyrosine hydroxylase inhibitor, α-methyl-dl-p-tyrosine methyl ester (H44/68), 2 hours prior to 

4832 sacrifice. Brains were removed, stained, and evaluated for catecholamine changes 16–18 hours 

4833 after the last exposure. Catecholamine levels were measured in the hypothalamus, frontal cortex, 

4834 and caudate nucleus among other brain regions. At all exposures, there was a significant 

4835 decrease by approximately 10–15% of catecholamine concentrations in the posterior 

4836 periventricular region of the hypothalamus. In the medial part of the caudate nucleus, which is 

4837 involved in memory processes, catecholamine levels were significantly higher (12%) in the 

4838 70 ppm group but significantly lower in the 300 ppm (1%) and 1,000 ppm (8%) groups 

4839 compared with controls. The impact of dichloromethane was also evaluated on the 

4840 hypothalamic-pituitary gonadal axis. The hypothalamus regulates secretion of reproductive 

4841 hormones, such as follicle-stimulating hormone and luteinizing hormone. The levels of the 

4842 hormone release were not significantly changed with dichloromethane exposure. However, 

4843 when rats were dosed concurrently with H44/68 and dichloromethane, statistically significant 
147

4844 inversely dose-related increases in luteinizing hormone levels were observed (330, 233, and 

4845 172% higher than controls in the 70, 300, and 1,000 ppm groups, respectively). The study 

4846 overall demonstrates significant changes in catecholamine levels in the hypothalamus and 

4847 caudate nucleus. No significant changes in catecholamine levels in the frontal cortex were 

4848 reported. Catecholamine level changes in the hypothalamus did not appear to significantly affect 

4849 hormone release; however, decreased catecholamine levels in the caudate nucleus at higher 

4850 exposures may lead to memory and learning impairment. 

4851 A series of studies were conducted in male and female Mongolian gerbils exposed 

4852 continuously to 210 ppm (Karlsson et al., 1987; Briving et al., 1986), 350 ppm, or 700 ppm 

4853 (Rosengren et al., 1986) dichloromethane for 3 months, followed by a 4-month exposure-free 

4854 period. High mortality rates occurred at 350 ppm (6/10 males and 3/10 females by 71 days) and 

4855 700 ppm (10/10 males and 9/10 females by 52 days). Rosengren et al. (1986) monitored two 

4856 astroglial proteins, S-100 and GFA, as well as DNA concentrations in the brain. Decreased 

4857 DNA concentrations were noted in the hippocampus at both the 210 and 350 ppm exposures. At 

4858 350 ppm, there was also decreased DNA concentration in the cerebellar hemispheres, indicating 

4859 a decreased cell density in these regions, probably due to cell loss. Increased astroglial proteins 

4860 were found in the frontal and sensory motor cerebral cortex, which directly correlated to the 

4861 astrogliosis that was observed in those areas. Up-regulation of these astroglial proteins is a good 

4862 indicator of neuronal injury (Rosengren et al., 1986). 

4863 Karlsson et al. (1987) measured DNA concentrations in different regions of the gerbil 

4864 brain. After the solvent-free exposure period, brains were removed and the olfactory bulbs and 

4865 cerebral cortices were dissected. Brain weights and weights of the dissected brain regions were 

4866 the same between control and dichloromethane-exposed animals. The total protein concentration 

4867 per wet weight was not significantly different between dichloromethane-exposed and control 

4868 animals. However, DNA concentrations per wet weight were significantly decreased in the 

4869 hippocampus after dichloromethane exposure. No other examined regions demonstrated 

4870 significant changes in DNA concentrations after dichloromethane exposure. This selective DNA 

4871 concentration decrease observed in the hippocampus is a sign of neurotoxicity and may possibly 

4872 explain why some studies have noted memory and learning deficits with dichloromethane 

4873 exposure. In a companion paper, in which only the 210 ppm level was tested, it was found that 

4874 exposure to dichloromethane decreased the levels of glutamate, γ-aminobutyric acid, and 

4875 phosphoethanolamine in the frontal cortex, while glutamine and γ-aminobutyric acid were 

4876 increased in the posterior cerebellar vermis (Briving et al., 1986). Increased levels of glutamate 

4877 in the posterior cerebellar vermis could reflect an activation of astrocytic glia, since glutamine 

4878 synthetase is localized exclusively in astrocytes. The gerbils did not have a solvent-free 

4879 exposure period as in the other two studies (Karlsson et al., 1987; Rosengren et al., 1986). The 

4880 exposure regime in these studies did not affect BW or brain weight. Furthermore, the 

4881 neurochemical changes observed in these studies were not attributed to formation of CO. 
148

4882 Neurological changes have been investigated by measuring changes in neurotransmitter 

4883 levels and changes in neurotransmitter localization. Changes in catecholamine levels in the 

4884 caudate nucleus after an acute exposure (Fuxe et al., 1984) as well as decreased DNA content in 

4885 the hippocampus after a subchronic dichloromethane exposure (Rosengren et al., 1986) suggest 

4886 that memory functions are altered since both brain regions are associated with learning and 

4887 memory. The results from Fuxe et al. (1984) directly correlated with the finding that learning 

4888 and memory were impaired in mice after an acute (single) and very high exposure (47,000 ppm) 

4889 to dichloromethane (Alexeef and Kilgore, 1983). Additionally, changes in the hippocampus also 

4890 suggest memory effects after a long-term, continual exposure to dichloromethane, although no 

4891 conclusive evidence has been presented to date. In another subchronic, continuous exposure to 

4892 350 ppm dichloromethane for 3 months, decreased DNA concentration was observed in the 

4893 cerebellar hemispheres of Mongolian gerbils and is suggestive of cell loss (Rosengren et al., 

4894 1986). However, in a 2-week exposure study in male Wistar rats, RNA changes were not noted 

4895 in the cerebellum, although enzyme activity was significantly decreased in this region (but was 

4896 increased in the cerebrum) (Savolainen et al., 1981). These results suggest that the cerebellum is 

4897 a target for dichloromethane. Noted neurobehavioral effects that may be linked to impaired 

4898 cerebellar function include changes in motor activity and impaired neuromuscular function 

4899 (Moser et al., 1995). 

4900 

4901 4.5. MECHANISTIC DATA AND OTHER STUDIES IN SUPPORT OF THE MODE OF 

4902 ACTION 

4903 4.5.1. Genotoxicity Studies 

4904 4.5.1.1. In Vitro Genotoxicity Assays 

4905 Bacterial, Yeast, and Fungi mutagenicity assays 

4906 Numerous in vitro studies have demonstrated dichloromethane as being mutagenic in 

4907 bacterial assays, yeast, and fungi, and several studies provide evidence that the genotoxic action 

4908 of dichloromethane in bacterial systems is enhanced in the presence of GSH (e.g., Dillon et al., 

4909 1992; Their et al., 1993; Oda et al., 1996; DeMarini et al., 1997; Pegram et al., 1997) (Table 4-

4910 29). Considering the results are primarily dependent on the presence of GSH, activation likely 

4911 involves the GST-T1 metabolic pathway, which produces two proposed DNA-reactive 

4912 metabolites, S-(chloromethyl)glutathione and formaldehyde. 

4913 Dichloromethane induced mutations in Salmonella typhimurium strains containing GSH 

4914 (e.g. TA100, TA98). These effects were not markedly influenced by the addition of exogenous 

4915 mammalian liver fractions, suggesting that endogenous metabolism in these strains was 

4916 sufficient to activate dichloromethane (Green, 1983; Jongen et al., 1982; 1978; Gocke et al., 

4917 1981). In support of this hypothesis, dichloromethane exposure of NG-11, a glutathione4918 deficient variant of S. typhimurium strain TA100, produced twofold fewer base-pair mutations 
149

Table 4-29. Results from in vitro genotoxicity assays of dichloromethane with bacteria, yeast, or fu

Results 

Assay Test system Concentration(s) 

Without metabolic 

activation 

–S9 

With metabolic 

activation 

+S9 

Bacteria 

Reverse mutation S. typhimurium 

TA98a, TA100a

6-hour exposure to 0, 7,000, and 14,000 ppm 

+ + Jo

Reverse mutation S. typhimurium 

TA98, TA100 

Up to 3,600 μg/plate + ++ Go

Reverse mutation S. typhimurium 

TA1535b, TA1537b, 

TA1538b

Up to 3,600 μg/plate – – Go

Reverse mutation S. typhimurium 

TA100 

6-hour exposure to 0, 7,000, and 14,000 ppm 

+ ++ Jo

Reverse mutation S. typhimurium 

TA100 

Up to 84,000 ppm, 3-day exposure 

+ + Gr

Reverse mutation S. typhimurium 

TA100, TA1535, TA1950a, 

E. coli WU361089a

10 μL/plate + for TA100, TA1950, 

WU361089 – for TA1535

Not determined O

(1

Reverse mutation S. typhimurium 

TA100 

+ Not 

determined 

Ze

Reverse mutation S. typhimurium 

TA100, NG54c

2- and 6-hour exposures to 0, 2,500, 5,000, 7,500, and 10,000 ppm 

+ + Di

Reverse mutation S. typhimurium 

TA100, TA1535 and TA1538 (+GSTA 1-1 and GSTP 1-1) 

0, 50, 100, and 200 μL/plate + for TA100 

– for TA1535, TA1538

Not determined Si

Reverse mutation S. typhimurium 

TA1535 (+GST 5–5), TA1535 (wild type) 

0–2.0 mM/plate + for TA1535 (+GST 5–5)

– for TA1535 (wild type) 

Not determined Pe

et 

Reverse mutation S. typhimurium 

TA100, TA100/NG-11d

0, 30, 60, 130 mM/plate ++ for TA100 

+ for TA100/NG-11 

Not determined Gr

Reverse mutation S. typhimurium TA100, 

RSJ100c

Up to 24,000 ppm + for TA100 

+ for RSJ100 

+ for TA100 

+ for RSJ100 

De
150

Table 4-29. Results from in vitro genotoxicity assays of dichloromethane with bacteria, yeast, or fu

Results 

Assay Test system Concentration(s) 

Without metabolic 

activation 

–S9 

With metabolic 

activation 

+S9 

Forward mutation S. typhimurium BA13 0–130 μmol/plate +++ + Ro

(1

Gene mutation S. typhimurium

TA1535/pSK1002c, 

NM5004c

0, 2.5, 5.0, 10, 20 mM + NM5004 

– TA1535/ pSK100 2 

Not determined Od

Prophage induction 

E. coli K-39 (λ) 10 μL/plate + Not determined O

(1

Reverse mutation E. coli WP2 uvra 

pKM101 

2- and 6-hour exposures to 6,300, 12,500, 25,000, and 50,000 ppm 

+ + Di

Forward mutation E. coli K12 0, 30, 60, 130 mM/plate – + Gr

Forward mutation E. coli Uvr+, UvrB– 20,000 ppm + Not determined Zi

Fungi and yeasts

Mitotic segregation 

Aspergillus nidulans Up to 8,000 ppm + only at 4,000 ppm; no 

dose-response relationship established 

Not determined Cr

Gene conversion and recombination

Saccharomyces cerevisiae 

Up to 209 mM + Not determined Ca

a bacterial strains that have GSH (e.g. TA100, TA 98) b bacterial strains that do not have GSH (e.g. TA1535) c bacterial strains engineered to have more GSH activity than wild type d bacterial strains engineered to have less GSH activity than wild type 
151

4919 compared with exposure of strain TA100, which produces normal levels of GSH. Furthermore, 

4920 this difference was not apparent when the culture medium contained 1 mM 

4921 GSH (Graves et al., 1994a). 

4922 In contrast to strain TA100, S. typhimurium strains TA1535, TA1537, and TA1538 

4923 (strains deficient in GSH) did not develop base-pair mutations in response to dichloromethane 

4924 exposure (Gocke et al., 1981; Osterman-Golkar et al., 1983; Simula et al., 1993; Thier et al., 

4925 1993; Pegram et al., 1997). However, when strain TA1535 was transfected with rat GST-T1, 

4926 dichloromethane induced base-pair reverse mutations (DeMarini et al., 1997; Pegram et al., 

4927 1997; Thier et al., 1993). A 60-fold higher concentration of dichloromethane was needed to 

4928 induce a response (i.e., a sixfold increase over background levels in reverse mutations) in S. 

4929 typhimurium strain TA100 than in TA1535 transfected with rat GST-T1 (DeMarini et al., 1997). 

4930 This study also included several trihalomethanes; dichloromethane was several fold less 

4931 genotoxic than dibromochloromethane or bromoform, but was similar in potency to 

4932 bromodichloromethane (DeMarini et al., 1997; Pegram et al., 1997). The authors suggest that 

4933 these results support a role of GST-T1 in the mutagenicity of the trihalomethanes. 

4934 The mutagenic effects of dichloromethane have also been examined in fungi and yeasts 

4935 assays with both systems reporting positive results. Fungi assays were positive for mitotic 

4936 segregation in Asperigillus ridulans (Crebelli et al., 1988) but there was not a dose response 

4937 relationship as only the 4,000 ppm dichloromethane exposure was positive (exposure up to 8,000 

4938 ppm). A yeast assay was positive for gene conversion and recombination in Saccharomyces 

4939 cerevisiae for concentrations up to 209 mM (Callen et al., 1980). 

4940 

4941 Mammalian assays 

4942 In the in vitro mammalian system studies conducted with murine cell lines (Table 4-30), 

4943 dichloromethane was negative for producing point mutations in the mouse lymphoma L5178Y 

4944 cell line (Thilagar et al., 1984), but was positive in producing single stranded DNA breaks in 

4945 mouse Clara cells (Graves et al., 1995) and mouse hepatocytes (1994b). Given that exposure to 

4946 dichloromethane results specifically in lung and liver tumors, this pattern is not surprising. 

4947 Additionally, GST is localized in the nucleus of hepatocytes and lung cells in the mouse 

4948 (Mainwaring et al., 1996), which would also increase sensitivity of these particular cell fractions 

4949 to genotoxic effects of dichloromethane. DNA single strand breaks (SSBs) were induced at 

4950 lower concentrations in mouse hepatocytes (0.5 mM) than in rat hepatocytes (30 mM). The 

4951 extent of DNA damage was shown to be reduced to the background level seen in control (no 

4952 exposure) conditions by pretreating the cells with buthionine sulfoxime to deplete cellular levels 

4953 of GSH and thus inhibit dichloromethane metabolism via the GST pathway (Graves et al. 1995; 

4954 1994b). Similar results were seen in mouse lung Clara cells. Freshly isolated Clara cells from 

4955 the lungs of B6C3F1 mice also showed significantly increased, concentration-dependent amounts 

4956 of DNA SSBs when incubated in vitro for 2 hours in the presence of 5–60 mM dichloromethane. 
152

4957 Pretreatment with buthionine sulphoximine before Clara-cell isolation or the presence of 

4958 buthionine sulphoximine in the culture medium decreased the amount of in vitro DNA damage 

4959 induced. 

4960 In a series of experiments with freshly isolated hepatocytes from multiple species (Table 

4961 4-30), DNA-protein cross-links were detected in hepatocytes of B6C3F1 mice but not in 

4962 hepatocytes of F344 rats, Syrian golden hamsters, or three human subjects, following 2-hour in 

4963 vitro exposure to concentrations ranging from 0.5–5 mM dichloromethane (Casanova et al., 

4964 1997). Within the range of concentrations tested, DNA-protein cross-links in mouse hepatocytes 

4965 appeared to increase with increasing concentration of dichloromethane. 

4966 Negative results for dichloromethane were predominantly seen in in vitro test systems 

4967 that used rat or hamster cell lines with low or no GST activity (Table 4-30). Several genotoxic 

4968 endpoints, including DNA and protein synthesis (Garrett and Lewtas, 1983), chromosomal 

4969 aberrations or sister chromatid exchanges (Thilagar et al., 1984; Thilagar and Kumaroo, 1983; 

4970 Jongen et al., 1981), unscheduled DNA synthesis (Thilagar et al., 1984; Andrae and Wolff, 1983; 

4971 Jongen et al., 1981), and mutations (Thilagar et al., 1984; Jongen et al., 1981) were evaluated in 

4972 these cell lines. In contrast, positive results (DNA-protein cross links and DNA SSBs) were 

4973 observed when mouse liver cytosol was included in Chinese hamster ovary (CHO) cells (Graves 

4974 et al., 1994b; Graves et al., 1995). Dichloromethane also induced hypoxanthine-guanine 

4975 phosphoribosyl transferase (HPRT) gene mutations in CHO cells when they were incubated with 

4976 GST-competent mouse liver cytosol preparations (Graves et al., 1996). 

4977 The instability of the S-(chloromethyl)glutathione-adducts presents considerable 

4978 challenges to studies of these products (Hashmi et al., 1994). Kayser and Vuilleumier (2001), 

4979 however, demonstrated the formation of DNA adducts with radiolabeled dichloromethane in calf 

4980 thymus DNA in the presence of dichloromethane dehalogenase/GST purified from a bacterial 

4981 source (Methylophilus sp. strain DM11) and GSH (Table 4-30). The type of adduct could not be 

4982 identified because of low yield, but it was determined that guanine was more actively 

4983 incorporated than cytosine, adenine or thymine by at least 2 fold in the presence of GST4984 activated dichloromethane, indicating a base specificity for these adducts. Incubation of calf 

4985 thymus DNA with formaldehyde and GSH, however, did not result in detectable DNA adduct 

4986 formation. In another study, Marsch et al. (2004) further evaluated the presence of adducts in 

4987 calf thymus DNA in the presence of dichloromethane and human (GSTT1-1), rat (GST 5-5) or 

4988 bacterial (DM11) GST (Marsch et al., 2004). This study found that all three enzymes yielded a 

4989 similar pattern of adduct formation, forming primarily with guanine and to a lesser extent with 

4990 cytosine, adenine, and thymine (2-3 fold less than guanine), consistent with the results reported 

4991 by Kayser and Vuilleumier (2001). High levels of guanosine-specific adducts were also seen 

4992 with S-(1-acetoxymethyl)glutathione, a compound that is structurally similar, but more stable, 

4993 than S-(chloromethyl)glutathione (Marsch et al., 2001). These findings indicate that the S4994 (chloromethyl)glutathione intermediate formed by GSH conjugation has mutagenic potential and 
153

4995 is likely responsible, at least in part, for the mutagenic response observed following 

4996 dichloromethane exposure. 

4997 In studies with human cell lines or isolated cells, positive results were reported for sister 

4998 chromatid exchanges and chromosomal aberrations (Thilagar et al., 1994) and in the 

4999 micronucleus test (Doherty et al., 1996). Negative results with human cells were seen in the 

5000 unscheduled DNA synthesis assays (Perocco and Prodi, 1981; Jongen et al., 1981), DNA SSBs, 

5001 and DNA-protein cross-links (Graves et al., 1995; Casanova et al., 1997). 

5002 Dichloromethane-induced DNA damage (comet assay) was examined in primary cultures 

5003 of human lung epithelial cells collected by brush biopsy from four healthy volunteers (Landi et 

5004 al., 2003). This study was designed to assess the genotoxicity of four thrihalomethanes 

5005 (chloroform, bromodichloromethane, dibromochloromethane and bromoform), with 

5006 dichloromethane included because of its known activation by GST-T1. Two of the subjects were 

of the GST-T1+

 genotype, and two were of the GST-T1– genotype.65007 The cells had been frozen, 

5008 and GST activity was not detected in the cultured cells. DNA damage was reported to occur in 

the combined GST-T1–5009 samples (tail extent moment 7.1, 13.7 and 15.3 in the 10, 100 and 1000 

μM dichloromethane groups, respectively), but not in the combined GST-T1+

5010 samples (tail 

5011 extent moment 8.1, 11.5 and 10.4 in the 10, 100 and 1000 μM dichloromethane groups, 

5012 respectively). This pattern was not seen across the individual samples, however, as only one 

5013 sample exhibited a clear dose-response gradient. Given the absence of GST activity, an analysis 

5014 combining the four samples could provide a more informative picture of the dose-response 

5015 relation between dichloromethane (and the other compounds) studied and DNA damage. For 

5016 dichloromethane, values of 9.4, 7.6, 12.6, and 12.9 were seen in the 0, the 10, 100 and 1000 μM 

5017 groups, respectively. This pattern was similar to that seen with chloroform (9.4, 6.9, 11.4, and 

5018 12.7 in the 0, the 10, 100 and 1000 μM groups, respectively), but weaker than the pattern for 

5019 bromoform (9.4, 12.5, 15.8, and 18.2 in the 0, the 10, 100 and 1000 μM groups, respectively), 

5020 and much weaker than for bromodichloromethane (9.4, 25.2, 28.5, and 39.1 in the 0, the 10, 100 

and 1000 μM groups, respectively).75021 No dose-response gradient was seen with 

5022 dibromochloromethane (9.4, 6.5, 8.1 and 8.0 in the 0, the 10, 100 and 1000 μM groups, 

5023 respectively). This relative pattern is also seen in the estimated slopes (beta coefficient for the 

5024 change in tail extent moment per unit increase in μM concentration): 0.0, 0.003, 0.004., 0.006 

5025 and 0.02 for dibromochloromethane, dichloromethane, chloroform, bromoform, and 

5026 bromodichloromethane, respectively (statistical significance not reported). 

 

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 from 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

5027 A stronger and more consistent response was seen under the same experimental 

5028 conditions with bromodichloromethane, but dibromochloromethane resulted in no increase in 

5029 DNA damage in any of the donor cells at any concentration tested. 

5030 Several studies have examined patterns of mutations or DNA damage with 

5031 dichloromethane and formaldehyde to assess the relative role of S-(chloromethyl)glutathione and 

5032 formaldehyde in the observed genotoxicity. In a study in CHO cells incubated with 

5033 dichoromethane (0.3% plus mouse liver cytosol), 2.5-fold increases in DNA-protein cross-links 

5034 that are indicative of formaldehyde exposure were observed, compared with a 25-fold increase 

5035 when 1 mM formaldehyde was added directly to cultures. Both treatments induced a comparable 

5036 degree of DNA SSBs (Graves and Green, 1996). In a subsequent study, Graves et al. (1996) 

5037 compared the mutational spectra induced by dichloromethane to that induced by direct addition 

5038 of formaldehyde or 1,2-dibromoethane (a chemical known to act through a glutathionyl 

5039 conjugate metabolite) at the HPRT locus in CHO cells. The mutations induced by 

5040 dichloromethane and 1,2-dibromoethane were predominantly GC to AT transitions, while all six 

5041 formaldehyde-induced mutants sequenced were single base transversions. This provided further 

5042 evidence that the S-(chloromethyl)glutathione intermediate may be primarily responsible for 

5043 dichloromethane genotoxicity. In contrast, Hu et al. (2006) found evidence of significant 

5044 amounts of formaldehyde formation following dichloromethane exposure in the cytosol of V79 

5045 (hamster) cells transfected with the murine GSTT1 gene compared to the parent cell line. In 

5046 accordance with this, they observed concentration-dependent increases in DNA-protein 

5047 crosslinks in the GSTT1 transfected cells using the comet assay with and without proteinase K 

5048 treatment that frees DNA from crosslinks and allows DNA migration. These findings are 

5049 consistent with those by Casanova et al. (1997), who performed a comparison of the amounts of 

5050 DNA-protein and RNA-formaldehyde crosslinks formed following dichloromethane exposure in 

5051 hepatocytes isolated from mice, rats, hamsters, and human GSTT1 genetic variants. Only DNA5052 protein crosslinks were observed in mouse hepatocytes, but RNA-formaldehyde crosslinks were 

5053 found in all species, which were highest in the mouse hepatocytes, and were 4-, 7-, and 14-fold 

5054 higher than rats, humans, and hamsters. These results showed that human hepatocytes can 

5055 metabolize dichloromethane to formaldehyde, resulting in RNA-formaldehyde crosslinks. In 

5056 addition, the results indicate, that there is considerable variation among species, and that the 

5057 human variation in the GSTT1 gene can also affect the amount of formaldehyde produced. The 

5058 authors also noted that comparing results following ectopic addition of formaldehyde directly to 

5059 cells with results following dichloromethane metabolism in situ can be misleading, as the 

5060 formaldehyde produced internally may reside in different locations intracellularly, potentially 

5061 affecting the capability of interacting with DNA. These results show that, while most studies 

5062 indicate the importance of the S-(chloromethyl)glutathione intermediate in mediating genotoxic 

5063 damage following dichloromethane exposure, DNA damage resulting from formaldehyde 

5064 formation should also be considered. 
155

Table 4-30. Results from in vitro genotoxicity assays of dichloromethane with mammalian systemof test 

Assay Test system Concentrations Results 

Mouse

Point mutation Mouse lymphoma 

L5178Y cells 

Not provided Negative Thil

 DNA SSBs by 

 alkaline elution 

Mouse hepatocytes 0, 0.4, 3.0, 5.5 mM Positive at 0.4 mM Grav

 DNA SSBs by 

 alkaline elution 

Mouse Clara cells 0, 5, 10, 30, 60 mM Positive, but DNA damage was reduced by incubating 

in the presence of GSH depletory 

Grav

DNA-protein cross-links 

Mouse hepatocytes 0.5–5 mM Positive Casa

Rat 

Unscheduled DNA synthesis 

Rat hepatocytes Up to 16 mM 

(measured); 30 mM (nominal) 

Negative And

Unscheduled DNA synthesis 

Rat hepatocytes Not provided Marginally positive Thil

 DNA SSBs by  alkaline elution 

Rat hepatocytes 0, 30, 90, 90 mM Positive at 30 mM Grav

DNA-protein cross-links 

Rat hepatocytes 0.5–5 mM Negative Casa

Hamster with GST activity from mouse 

DNA-protein cross-links 

Chinese hamster ovary cells 

60 mM Positive with mouse liver cytosol (negative without) 

at much higher concentrations of dichloromethane (60 mM) than formaldehyde (0.5–4 mM). 

Grav

HPRTa mutation 

analysis 

Chinese hamster ovary cells 

2,500 ppm Positive with mouse liver cytosol Grav

DNA SSBs and DNA-protein cross-links 

Chinese hamster ovary cells 

3,000 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 (in 

absence of mouse liver cytosol) was positive at 0.5mM for both DNA SSBs and DNA-protein crosslinks; Chinese hamster ovary cell cultures were suspended. 

Grav
156

Table 4-30. Results from in vitro genotoxicity assays of dichloromethane with mammalian systemof test 

Assay Test system Concentrations Results 

DNA-protein cross-links 

Syrian golden hamster hepatocytes 

0.5–5 mM Negative Casa

Comet Assay V79 hamster cells 

transfected with mouse GSTT1 

2.5, 5, 10 mM A significant, dose-dependent increase in DNA 

damage resulting from DNA-protein crosslinks in V79 cells transfected with mouse GSTT1 compared to parental cells 

Hu e

Hamster without GST activity from mouse

Forward mutation Chinese hamster 

epithelial cells 

5,000, 10,000, 30,000, 50,000 ppm 

Negative Jong

Unscheduled DNA synthesis 

Chinese hamster epithelial cells 

5,000, 10,000, 30,000, 50,000 ppm 

Negative Jong

Sister chromatid exchange 

Chinese hamster epithelial cells 

5,000, 10,000, 20,000, 30,000, and 40,000 ppm

Weak positive with or without rat-liver microsomal system 

Jong

Chromosomal aberrations 

Chinese hamster ovary cells 

Not provided Positive, independent of rat liver S9 Thil

(198

Sister chromatid exchange 

Chinese hamster ovary cells 

Not provided Negative with or without rat liver S9 Thil

(198

DNA and protein synthesis 

Chinese hamster ovary cells 

Up to 1,000 μg/mL Negative Garr

DNA SSBs by alkaline elution 

Hamster hepatocytes 0.4–90 mM Negative Grav

Calf 

DNA Adducts Calf thymus DNA 50 mM Positive in the presence of bacterial GST DM11 and 

dichloromethane dehalogenase. Adducts primarily formed with the guanine residues. 

Kay(200

DNA Adducts Calf thymus DNA 0 -8.0 umol (0 – 60 mM) Positive in the presence of bacterial GST DM11, rat

GST5-5, and human GSTT1-1. Adducts primarily formed with the guanine residues. 

Mar

Human 

Unscheduled DNA synthesis 

Human peripheral lymphocytes 

250, 500, 1,000 ppm Negative with or without rat liver S9 Pero

Unscheduled DNA synthesis 

Primary human fibroblast 

5,000, 10,000, 30,000, 50,000 ppm 

Negative Jong
157

Table 4-30. Results from in vitro genotoxicity assays of dichloromethane with mammalian systemof test 

Assay Test system Concentrations Results 

Sister chromatid exchange 

Human peripheral lymphocytes 

Not provided Weak positive Thil

Chromosomal aberrations 

Human peripheral lymphocytes 

Not provided Positive Thil

DNA SSBs by alkaline elution 

Human hepatocytes Up to 120 mM Negative at concentrations between 5 and 120 mM Grav

Micronucleus test Human AHH-1, 

MCL-5, h2E1 cell lines

Up to 10 mM Positive in all three cell lines Doh

DNA-protein cross-links 

Mouse, rat, hamster, human hepatocytes 

0.5–5 mM Negative Casa

DNA damage by comet assay 

Primary human lung epithelial cells 

10, 100, 1,000 μM Weak trend, independent of GST activity (GST 

enzymatic activity not present in the cultured cells) 

Land

aHPRT = hypoxanthine-guanine phosphoribosyl transferase 

5065 
158

5066 4.5.1.2. In Vivo Genotoxicity Assays 

5067 Genotoxicity findings in Drosophila melanogaster assays are mixed (Table 4-31). A 

5068 study of gene mutation in D. melanogaster showed a marginal increase in sex-linked recessive

5069 deaths following oral exposure (Gocke et al., 1981). An additional feeding study (RodriguezArnaiz, 1998) reported a positive response in the somatic w/w+

5070 assay. A third study of 

5071 D. melanogaster (Kramers et al., 1991) found no evidence of increased sex-linked recessive 

5072 deaths, somatic mutations, or recombinations following exposure to airborne dichloromethane. 

5073 

Table 4-31. Results from in vivo genotoxicity assays of dichloromethane in 

insects 

Assay Test system Doses Result Reference 

Gene 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 

5075 Some in vivo studies investigating certain genotoxic endpoints in mice exposed to 

5076 dichloromethane produced negative results (Table 4-32). Unscheduled DNA synthesis was not 

5077 induced in hepatocytes from mice (and rats) after 2- or 6-hour inhalation exposure to 

5078 concentrations that were carcinogenic in the NTP (1986) mouse bioassay (Trueman and Ashby, 

5079 1987) or other exposure routes (Lefevre and Ashby, 1989). Although positive results were not 

5080 observed in the unscheduled DNA synthesis studies, it is generally recognized that this assay is 

5081 not sensitive for detecting genotoxic chemicals (Eastmond et al., 2009; Madle et al., 1994). 

5082 Distinct, unequivocal cytogenetic effects (e.g., induction of micronuclei, sister chromatid 

5083 exchanges, or chromosome aberrations) were not consistently found in bone marrow or 

5084 erythrocytes in several studies of mice after acute oral exposures (Sheldon et al., 1987) or 

5085 parenteral exposures (Westbrook-Collins et al., 1990; Gocke et al., 1981). However, 

5086 tumorigenic effects in mice are generally localized to the liver and lung (due to high GST 

5087 activity) and therefore it is not surprising that genotoxic effects were, for the most part, not 

5088 observed in the bone marrow or erythrocytes (cell types with minimal GST activity). Crebelli et 

5089 al. (1999) stated that genotoxic effects induced by halogenated hydrocarbons (such as 

5090 dichloromethane) are not very effective in inducing micronucleus formation in mouse bone 

5091 marrow and a negative bone marrow micronucleus assay should not offset the consistently 

5092 positive in vitro results (Dearfield and Moore, 2005) 

5093 When genotoxic endpoints were examined in the cancer target tissues (liver and lung) in 

5094 mice exposed to dichloromethane, positive results were consistently reported (Table 4-32). 
159

5095 These findings provide supporting evidence that GST-pathway metabolites may be key actors in 

5096 the genotoxic effects and carcinogenic mode of action for dichloromethane. Increased sister 

5097 chromatid exchanges were found in lung cells and peripheral lymphocytes from mice exposed by 

5098 inhalation for 2 weeks to 8,000 ppm or for 12 weeks to 2,000 ppm (Allen et al., 1990). Under 

5099 the same exposure conditions, increased chromosomal aberrations in lung and bone cells and 

5100 micronuclei in peripheral red blood cells also were found (Allen et al., 1990). DNA-protein 

5101 cross-links were detected in mouse hepatocytes, but not in lung cells, after a 3-day inhalation 

5102 exposure to 4000 ppm (Casanova et al., 1992) and between 500 and 4000 ppm (Casanova et 

5103 al.,1996). DNA damage, detected as increased DNA SSBs, was observed in liver and lung tissue 

5104 of B6C3F1 mice immediately following 3-hour exposures (Graves et al., 1995). The DNA 

5105 damage was not detectable 2 hours after in vivo exposure, indicating that DNA repair occurs 

5106 rapidly. Pretreatment of mice with buthionine sulphoximine, a GSH depletor, caused a decrease, 

5107 to levels seen in controls, in the amount of DNA damage detected immediately after in vivo 

5108 exposure in liver and lung tissue, indicating GSH involvement in the genotoxic process. DNA 

5109 damage (detected by the comet assay) was also reported in liver and lung tissues from male CD5110 1 mice sacrificed 24 hours after administration of a single oral dose of 1,720 mg/kg of 

5111 dichloromethane (Sasaki et al., 1998). In this study, DNA damage in lung and liver was not 

5112 detected 3 hours after dose administration, and no DNA damage occurred at either time point in 

5113 several other tissues in which a carcinogenic response was not seen in chronic animal cancer 

5114 bioassays (e.g., stomach, kidney, bone marrow). 

5115 Formation of DNA adducts was evaluated in male and female B6C3F1 mice as well as in 

5116 male F344 rats (Watanabe et al., 2007). Animals were administered 5 mg/kg, i.p., of 

5117 radiolabeled dichloromethane and sacrificed at 1 or 8 hours after administration. The kidneys 

5118 and livers were removed and the DNA was isolated from these tissues to evaluate formation of 

5119 DNA adducts. At the administered dose, DNA adducts were not detected. 

5120 Other studies in mice have looked for mutations in specific oncogenes (K-ras or H-ras) 

5121 (Devereux et al., 1993) or in a tumor suppressor gene (p53) (Hegi et al., 1993) in liver or lung 

5122 tumors from dichloromethane-exposed mice. These studies have not demonstrated exposure5123 related patterns of mutations in these genes, although it should be noted that the statistical power 

5124 of this analysis for the lung tumors is limited (discussed further in sections 4.5.2 and 4.5.3). 

5125 
160

Table 4-32. Results from in vivo genotoxicity assays of dichloromethane in mice 

Assay Test system Route and dose Duration Results 

Micronucleus test Mouse bone marrow 425, 850, or 1,700 mg/kg 2 doses Negative at all doses Micronucleus test Mouse bone marrow Gavage, 

1,250, 2,500, and 4,000 mg/kg 

single dose Negative at all doses 

Micronucleus test Mouse peripheral red 

blood cells 

Inhalation 6 hr/day, 5 d/wk, 0, 4,000, 8,000 ppm 

2 wk Positive at 4,000 and 

8,000 ppm 

Micronucleus test Mouse peripheral red 

blood cells 

Inhalation, 6 hr/day, 5 d/wk, 0, 2,000 ppm 

12 weeks Positive at 2,000 ppm 

DNA synthesis Mouse liver Gavage, 1,000 mg/kg; 

inhalation, 4,000 ppm 

single dose; 2 hours 

Negative in both oral inhalation studies 

Unscheduled DNA synthesis Mouse hepatocytes Inhalation, 2,000 and 

4,000 ppm. 

 2 or 6 hours Negative 

Sister chromatid exchange Mouse bone marrow Intraperitoneal, 100, 1,000, 

1,500, 2,000 mg/kg 

single dose Negative 

Sister chromatid exchange Mouse bone marrow Subcutaneous, 0, 2,500, 

5,000 mg/kg 

single dose Negative at all doses 

Sister chromatid exchange Mouse lung cells and 

peripheral lymphocytes 

Inhalation 6 hr/day, 5 d/wk, 0, 4,000, 8,000 ppm 

2 weeks Positive at 8,000 ppm 

Sister chromatid exchange Mouse lung cells Inhalation 6 hr/day, 5 d/wk, 

0, 2,000 ppm 

12 weeks Positive at 2,000 ppm 

Chromosome aberrations Mouse bone marrow Intraperitoneal, 100, 1,000, 

1,500, 2,000 mg/kg 

single dose Negative 

Chromosome aberrations Mouse bone marrow Subcutaneous, 0, 2,500, 

5,000 mg/kg 

single dose Negative 
161

Table 4-32. Results from in vivo genotoxicity assays of dichloromethane in mice 

Assay Test system Route and dose Duration Results 

Chromosome aberrations Mouse lung and bone 

marrow cells 

Inhalation, 6 hr/day, 5 d/wk, 

0, 4,000, 8,000 ppm 

2 weeks Positive at 8,000 ppm 

DNA-protein cross-links Mouse liver and lung 

cells 

Inhalation, 6 hr/day, 3 days, 4,000 ppm 

3 days Positive in mouse liv

at 4,000 ppm; negatimouse lung cells 

DNA-protein cross-links Mouse liver and lung 

cells 

Inhalation, 6 hr/day, 150, 500, 1,500, 3,000, 4,000 ppm 

3 days Positive in mouse liv

at 500–4,000 ppm; nin mouse lung cells 

DNA single strand breaks by 

alkaline elution 

Mouse hepatocytes Inhalation, 2,000 and 

4,000 ppm 

3 or 6 hours Positive at 4,000 ppm 

3 and 6 hours 

DNA single strand breaks by 

alkaline elution 

Mouse liver and lung 

homogenate 

Liver: inhalation, 2,000, 4,000, 6,000, 8,000 ppm Lung: inhalation, 1,000, 2,000, 4,000, 6,000 ppm 

3 hours 3 hours 

Liver: positive at 4,08,000 ppm Lung: positive at 2,04,000 ppm 

DNA damage by comet assay Mouse liver and lung 

cells 

Gavage, 1,720 mg/kg; organs harvested at 0 (control), 3, and 24 hours 

single dose Positive only at 24 h

after dosing 

DNA damage by comet assay Mouse stomach, urinary 

bladder, kidney, brain, bone marrow 

Gavage, 1,720 mg/kg; organs harvested at 0 (control), 3, and 24 hours 

single dose Negative 3 or 24 hr a

dosing 

DNA adducts Mouse liver and kidney 

cells 

Intraperitoneal, 5 mg/kg Single dose Negative 

Kras and Hras oncogenes Mouse liver and lung 

tumors 

0, 2000 ppm Up to 104 weeks No difference in mut

profile between contdichloromethane-indliver tumors; number spontaneous lung tum(n=4) limits comparithis site 

p53 tumor suppressor gene Mouse liver and lung 

tumors 

0, 2000 ppm Up to 104 weeks Loss of heterozygoci

infrequently seen 
162

5126 Results from in vivo studies in other mammals (i.e., rats and hamsters) of hepatocyte 

5127 sensitivity to dichloromethane induction of DNA SSBs (Table 4-33) are consistent with 

5128 interspecies differences in the induction of liver tumors in the inhalation cancer bioassays. A 

5129 gavage study in rats reported the presence of DNA SSBs with a dose of 1,275 mg/kg (Kitchin 

5130 and Brown, 1989). The other available studies, however, did not find any genotoxicity following 

5131 dichloromethane exposure. No increase in unscheduled DNA synthesis in rat hepatocytes was 

5132 seen following inhalation dichloromethane exposure of 2 to 6 hours at 2000 or 4000 ppm 

5133 (Trueman and Ashby, 1987), exposure by gavage up to 1000 mg/kg (Trueman and Ashby, 1987), 

5134 or intraperitoneal exposure of 400 mg/kg (Mirsalis et al., 1989). DNA adducts were not detected 

5135 in the livers and kidneys of male F344 rats dosed with 5 mg/kg dichloromethane, i.p. (Watanabe 

5136 et al., 2007). DNA SSBs were significantly increased in hepatocytes isolated from B6C3F1 mice 

5137 exposed to 4,831 ppm (4,000 ppm nominal) for 6 hours but were not increased in hepatocytes 

5138 from Sprague-Dawley rats exposed to 4,527 ppm (4,000 ppm nominal) for 6 hours (Graves et al., 

5139 1994b). Results from in vivo interspecies comparisons of dichloromethane induction of DNA5140 protein cross-links in hepatocytes (expected products of the GSH pathway) are also consistent 

5141 with the hypothesis that the mouse is more sensitive than other mammalian species due to greater 

5142 activity of the GST pathway. DNA-protein cross-links were formed in the liver of mice, but not 

5143 hamsters, following in vivo exposure to air concentrations ranging from 500 to 4,000 ppm, 

5144 6 hours/day for 3 days (Casanova et al., 1996). The absence of a genotoxic response in the rat 

5145 and hamster is consistent with considerably lower GST activity and therefore, these mammalian 

5146 systems would be expected to be less sensitive at detecting genotoxic effects than the studies 

5147 conducted in mice. 

5148 Table 4-34 compares results from studies of mice and rats in which comparable tissue5149 specific endpoints were examined in in vivo genotoxicity assays. Several of the endpoints that 

5150 were positive in mice (e.g., sister chromatid exchange, DNA-protein cross-links, comet assay) 

5151 have not been examined in the rat. Unscheduled DNA synthesis has been demonstrated in 

5152 mouse, but not in rat hepatocytes. In contrast to the positive results seen in mouse inhalation 

5153 exposure studies, DNA SSB induction was not seen in rat inhalation studies but was seen in an 

5154 oral gavage study. 

5155 In summary, the available data provide evidence for mutagenicity of dichloromethane. 

5156 Most of the in vitro bacterial assays with GST activity showed positive results when there was 

5157 GST activity. Non-positive results were reported only in bacterial assays with low GST activity; 

5158 in experiments where GST was added, positive results were then observed. Evaluation of the in 

5159 vitro mammalian studies also demonstrates consistency of the requirement for GST for 

5160 observation of genotoxic effects. In rat and hamster cell lines where GST activity is significantly 

5161 less than mouse, primarily negative results were reported following dichloromethane exposure. 

5162 However, when mouse liver cytosol or transfected mouse GST were included in these same cell 

5163 lines, mutagenic effects were reported after dichloromethane exposure. In mouse cell lines, 
163

5164 positive results were obtained in Clara cells but no effects were observed in a mouse lymphoma 

5165 cell line, which is consistent with the absence of tumors in this site for mice. The results of in 

5166 vivo mutagenicity in mice also provide support for the site-specificity of the observed tumors. 

5167 Assays using mouse bone marrow were all negative. However, micronuclei and sister chromatid 

5168 exchange tests in peripheral blood produced a positive response at high doses. With the 

5169 exception of one study of unscheduled DNA synthesis in hepatocytes, numerous site-specific 

5170 studies in either the liver or lung were also positive at various doses. These liver and lung 

5171 studies included chromosomal aberrations, SSBs and sister chromatid exchanges, and DNA5172 protein cross-links and correspond to genotoxic and mutagenic effects associated with 

5173 metabolites from the GST pathway. 

5174 
164

5175 

Table 4-33. Results from in vivo genotoxicity assays of dichloromethane in rats and hamsters 

Assay Test system Route and dose Duration Results 

Unscheduled DNA synthesis Rat hepatocytes Gavage, 100, 500, 

1,000 mg/kg 

Liver harvested 4 and 12 hours after dosing 

Negative 4 or 12 houdosing 

Unscheduled DNA synthesis Rat hepatocytes Inhalation, 2 or 6 hours, 

2,000 and 4,000 ppm 

2 or 6 hours Negative at both 

concentrations and exdurations 

Unscheduled DNA synthesis Rat hepatocytes Intraperitoneal, single dose, 

400 mg/kg 

Single dose Negative 48 hours aft

dosing 

DNA SSBs by alkaline elution Rat hepatocytes Inhalation, 3 or 6 hours, 

2,000 and 4,000 ppm 

3 or 6 hours Negative at all 

concentrations and timpoints 

DNA SSBs by alkaline elution Rat liver homogenate Gavage, 2 doses, 425 mg/kg 

and 1,275 mg/kg, administered 4 and 21 hours before liver harvesting 

4 or 21 hours (time between dosing and liver harvesting) 

Positive at 1,275 mg/

DNA SSBs by alkaline elution Rat liver and lung 

homogenate 

Liver: inhalation, 4,000, 5,000 ppm Lung: inhalation, 4,000 ppm 

3 hours 3 hours 

Negative for both livelung at all concentrat

DNA-protein cross-links Hamster liver and lung 

cells 

Inhalation, 6 hr/day, 500, 1,500, 4,000 ppm 

3 days Negative at all 

concentrations 

DNA adducts Rat liver and kidney 

cells 

Intraperitoneal, 5 mg/kg Single dose Negative 
165

Table 4-34. Comparison of in vivo dichloromethane genotoxicity assays targeted to lung or liver cell

 Studies in mice Studies in rats 

Assay Test system Route, dose (duration) Results Reference Test system Route, dose (duration) 

DNA synthesis 

Liver Gavage, 1,000 mg/kg; 

inhalation, 4,000 ppm (2 hours) 

Negative in oral and inhalation studies 

Lefevre and Ashby (1989) 

 

Unscheduled DNA synthesis 

Hepatocytes Inhalation, 2,000 and 

4,000 ppm. (2 or 6 hours) 

Negative Trueman 

and Ashby (1987) 

Hepatocytes Inhalation, 2,000 and 

4,000 ppm (2 or 6 hours) 

Neg

Unscheduled DNA synthesis 

Hepatocytes Intraperitoneal, 

400 mg/kg 

Neg

Sister chromatid exchange 

Lung cells Inhalation 6 hr/day, 

5 days/wk, 0, 4,000, 8,000 ppm (2 weeks) Inhalation 6 hr/day, 5 days/wk, 0, 2,000 ppm (12 weeks) 

Positive at 8,000 ppm Positive at 2,000 ppm 

Allen et al. (1990) 

 

Chromosome aberrations 

Lung cells Inhalation, 6 hr/day, 

5 days/wk, 0, 4,000, 8,000 ppm (2 weeks) 

Positive at 8,000 ppm 

Allen et al. (1990) 

 

DNA-protein cross-links 

Liver and 

lung cells 

Inhalation, 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 liver 4,000 ppm Positive in liver at 500–4,000 ppm; both studies negative 

in lung 

Casanova et al. (1992) 

 

DNA SSBs by alkaline elution 

Hepatocytes Inhalation, 2,000 and 

4,000 ppm (3 or 6 hours) 

Positive at 4,000 ppm 

Graves et al. (1994b) 

Hepatocytes Inhalation, 3 or 6 hours, 

2,000 and 4,000 ppm 

Negconand 

DNA SSBs by alkaline elution 

Liver and 

lung homogenate 

Liver: inhalation, 2,000, 

4,000, 6,000, 8,000 ppm (3 hours) Lung: inhalation, 1,000, 2,000, 4,000, 6,000 ppm (3 hours) 

Liver: Positive at 4,000–8,000 ppm Lung: Positive at 2,000–4,000 ppm 

Graves et al. (1995) 

Liver and 

lung homogenate 

Liver: inhalation, 4,000, 5,000 ppm Lung: inhalation, 4,000 ppm 

Negliveall conand 

DNA SSBs by alkaline elution 

Liver homogenate

Gavage, 425 mg/kg and 1,275 mg/kg 

Pos1,2
166

Table 4-34. Comparison of in vivo dichloromethane genotoxicity assays targeted to lung or liver cell

 Studies in mice Studies in rats 

Assay Test system Route, dose (duration) Results Reference Test system Route, dose (duration) 

DNA damage by comet assay 

Liver and 

lung cells 

Gavage, 1,720 mg/kg; organs harvested at 0 (control), 3, and 24 hours 

Positive only at 24 hours after dosing 

Sasaki et al. (1998) 

 

DNA adducts Liver and 

kidney cells 

Intraperitoneal, 5 mg/kg Negative Watanabe et 

al. (2007) 

Liver and kidney cells 

Intraperitoneal, 5 mg/kg Neg

5176 
167

5177 4.5.2. Mechanistic Studies of Liver Effects 

5178 One of the major target organs from dichloromethane exposure is the liver, and several 

5179 studies have focused on examining the potential mechanisms producing liver tumors. This 

5180 section summarizes the primary mechanistic studies that were conducted in order to examine the 

5181 hepatic tumors produced by dichloromethane in mice. A parallel set of studies, discussed in the 

5182 next section, focus on potential mechanisms that produce lung tumors. Briefly, 

5183 dichloromethane-induced liver tumors first appeared in mice after 52 weeks of exposure 

5184 (Maronpot et al., 1995; Kari et al., 1993), which was when tumors began to appear in control 

5185 mice, indicating a similar time course in tumor formation between treated and untreated groups. 

5186 Onset of liver tumor formation is not preceded by liver cell proliferation (Casanova et al., 1996; 

5187 Foley et al., 1993). Further mechanistic studies were conducted to assay the tumor for 

5188 significant changes in proto-oncogene activation and tumor suppressor gene inactivation 

5189 (Maronpot et al., 1995; Devereux et al., 1993; Hegi et al., 1993). A second subset of mechanistic 

5190 studies was conducted to elucidate the reason that mice are the most sensitive species to liver 

5191 tumors and if other species exhibited changes in liver function (Thier et al., 1998; Reitz et al., 

5192 1989). It was found that mice have the highest level of GST-T1 catalytic activity but that 

5193 humans, rats, and hamsters among other species also metabolize dichloromethane in the liver to a 

5194 GST conjugate. In contrast, there has been little research focusing on the mechanisms through 

5195 which nonneoplastic hepatic effects (seen most strongly in the rat) are produced, and the role of 

5196 the parent material, metabolites of the CYP2E1 pathway, metabolites of the GST pathway, or 

5197 some combination of parent material and metabolites is not known. 

5198 

5199 Liver tumor characterization studies 

5200 Several studies have examined the time course of appearance of liver tumors in B6C3F1

5201 mice exposed to 2,000 or 4,000 ppm and possible links between hepatic nonneoplastic 

5202 cytotoxicity, enhanced hepatic cell proliferation, and the development of liver tumors (Casanova 

5203 et al., 1996; Maronpot et al., 1995; Foley et al., 1993; Kari et al., 1993). The studies provide no 

5204 clear evidence for a sustained liver cell proliferation response to dichloromethane that can be 

5205 linked to the development of dichloromethane-induced liver tumors. Additionally, a few studies 

5206 have examined if dichloromethane-induced liver tumors are the result of proto-oncogene 

5207 activation and tumor suppressor gene inactivation (Maronpot et al., 1995; Devereux et al., 1993; 

5208 Hegi et al., 1993). 

5209 Kari et al. (1993) (also summarized by Maronpot et al. [1995]) reported data from 

5210 6 groups of 68 female B6C3F1 mice exposed to six ”stop-exposure“ protocols of differing 

5211 durations and sequences, with each exposure concentration standardized at 2,000 ppm for 

5212 6 hours per day, 5 days per week. The six stop-exposure protocols were 26 weeks of exposure 

5213 followed by 78 weeks without exposure, 78 weeks without exposure followed by 26 weeks of 

5214 exposure, 52 weeks without exposure followed by 52 weeks with exposure, 52 weeks exposed 
168

5215 followed by 52 weeks without exposure, 78 weeks exposed followed by 26 weeks without 

5216 exposure, and 26 weeks without exposure followed by 78 weeks with exposure. A control group 

5217 (no exposure, 104 weeks duration) and a maximum exposure (104 weeks duration) group were 

5218 also included. Exposure for 26 weeks did not result in an increased incidence of liver tumors 

5219 (adenomas or carcinomas). Respective percentages of animals with liver tumors were 27% 

5220 (18/67), 40% (27/67), and 34% (23/67) for the controls, early 26-week exposure and late 26-

5221 week exposure groups, respectively. Exposure to 2,000 ppm for 52 weeks (followed by no 

5222 exposure until 104 weeks), 78 weeks (either early or late exposure periods), or 104 weeks 

5223 produced increased incidence of mice with liver tumors (p < 0.05), but this increase was not seen 

5224 in the 52-week late exposure group. Respective percentages of animals with liver tumors 

5225 (adenomas or carcinomas combined) were 44% (28/64), 31% (21/67), 62% (42/68), 48% (32/67) 

5226 and 69% (47/68) for the 52 (early exposure), 52 (late exposure), 78 (early exposure), 78 (late 

5227 exposure), and 104 week exposure periods, respectively. With the 78 week exposures, the 

5228 difference in the liver tumor incidence between the early and late exposure periods was 

5229 statistically significant (p < 0.01). A greater increase in multiplicity of liver tumors was also 

5230 seen with the early 78-week exposure period. These data suggest that 52 weeks of exposure was 

5231 required to increase the incidence of liver tumors in mice, that early exposure was more effective 

5232 than late exposure, and that the increased risk continued after cessation of exposure. 

5233 Histopathologic examination of liver tissue at interim killings at eight time periods (13, 

5234 26, 52, 68, 75, 78, 83, or 91 weeks) of exposure to 2,000 ppm (n = 20 mice per killing) found no 

5235 evidence of nonneoplastic cytotoxicity that preceded the appearance of proliferative neoplastic 

5236 liver lesions. Incidences of mice with liver adenomas or carcinomas were elevated (between 40–

5237 60%) at five of the six interim killings after 52 weeks. The incidence rates at each time period 

5238 were 0/20 (0%) at 13 weeks, 1/20 (5%) at 26 weeks, 8/20 (40%) at 52 weeks, 4/26 (15%) at 

5239 68 weeks, 13/20 (65%) at 75 weeks, 12/19 (63%) at 78 weeks, 8/20 (40%) at 83 weeks, and 

5240 20/30 (66%) at 91 weeks. The collected liver lesion data identify no exposure-related increased 

5241 incidence of nonneoplastic liver lesions that could be temporally linked to liver tumor 

5242 development. Liver tumors first appeared at about the same time in control and exposed animals 

5243 (52 weeks). 

5244 Foley et al. (1993) examined indices of cell proliferation in livers of female B6C3F1 mice 

5245 exposed to 1,000, 2,000, 4,000, or 8,000 ppm dichloromethane (6 hours/day, 5 days/week) for 1, 

5246 2, 3, or 4 weeks or to 2,000 ppm for 13, 26, 52, or 78 weeks but found no evidence for sustained 

5247 cell proliferation with prolonged exposure to dichloromethane. To label liver cells in S phase, 

5248 tritiated thymidine (1- to 4-week exposure protocols) or bromodeoxyuridine (13- to 78-week 

5249 protocols) was administered subcutaneously via an osmotic mini-pump for 6 days prior to 

5250 killing. Labeled hepatocytes in liver sections (from 10 mice in each exposure/duration group) 

5251 were counted to assess the number of cells in S-phase per 1,000 cells. S-phase labeling indices 

5252 in livers of exposed mice at most killings were equivalent to or less than those in control mice. 
169

5253 A transient increase in S-phase labeling index of about two- to fivefold over controls was 

5254 observed at the 2-week killing of mice exposed to 1,000, 4,000, or 8,000 ppm. Because of the 

5255 transient nature and small magnitude of the response, it is not expected to be of significance to 

5256 the promotion of liver tumors in chronically exposed mice. Foley et al. (1993) also compared 

5257 cell proliferation labeling indices in foci of cellular alteration and non-affected liver regions in 

5258 control and exposed mice but found no significant difference between control and exposed mice. 

5259 S-phase labeling was accomplished by immunohistologic staining for proliferating cell nuclear 

5260 antigen in liver sections from 24 control mice and 15 exposed mice, with livers showing foci of 

5261 cellular alteration. In both control and exposed livers, the labeling index was about four- to 

5262 fivefold higher in foci of cellular alteration than in surrounding unaffected liver tissue. 

5263 In mice exposed to 2,000 ppm for 13–78 weeks, relative liver weights were statistically 

5264 significantly elevated compared with controls; about 10% increased at 13 and 26 weeks and 

5265 about 30–40% increased at 52 and 78 weeks. Histologic changes in liver sections of 2,000 ppm 

5266 mice exposed for 13–78 weeks were restricted to hepatocellular hypertrophy (observed at all 

5267 killing intervals) and preneoplastic (foci of cellular alteration) and neoplastic (adenoma and 

5268 carcinoma) lesions. No signs of liver tissue degeneration were found. Adenoma and focus of 

5269 alteration were first detected at 26 weeks (2/10 versus 0/10 in controls). At 52 weeks, 4/10 

5270 exposed mice had proliferative lesions (1 focus, 1 adenoma, and 2 carcinomas), compared with 

5271 1/10 in controls (1 adenoma). At 78 weeks, 7/10 exposed mice had proliferative lesions (2 foci, 

5272 3 adenomas, 6 carcinomas) compared with 1/10 in controls (1 adenoma). In summary, the 

5273 results indicate that inhalation exposure to 2,000 ppm dichloromethane produced an increase 

5274 incidence of liver tumors in female B6C3F1 mice. No evidence was found for sustained cell 

5275 proliferation or liver tissue degeneration in response to dichloromethane exposure, but exposure 

5276 was associated with relative liver weight increases and hepatocellular hypertrophy. 

5277 Casanova et al. (1996) found no clear evidence of increased cell proliferation in the livers 

5278 of male B6C3F1 mice exposed to dichloromethane concentrations >1,500 ppm 6 hours/day for 

5279 3 days. Three or four groups of three mice were exposed to 146, 498, 1,553, or 3,923 ppm 

unlabeled dichloromethane for 2 days and then exposed to [14 5280 C]-labeled dichloromethane for 

5281 6 hours on the third day. Radiolabel incorporated into liver DNA deoxyribonucleosides was 

5282 measured as an index of cell proliferation. Radiolabel incorporated into liver DNA 

5283 deoxyribonucleosides increased approximately fivefold from 146 to 1,553 ppm, but further 

5284 increases were not apparent at 3,923 ppm. (In contrast, as described in section 4.5.3, radiolabel 

5285 incorporation into lung DNA deoxyribonucleosides displayed a 27-fold increase over this 

5286 concentration range.) The small magnitude of the increase in radiolabel incorporation into liver 

5287 DNA deoxyribonucleosides with increasing exposure concentration suggests that little, if any, 

5288 enhanced cell proliferation occurred in the liver in response to dichloromethane exposure. 

5289 Devereux et al. (1993) (also reported in Maronpot et al. [1995]) analyzed liver tumors in 

5290 female B6C3F1 mice for the presence of activated H-ras oncogenes. Fifty dichloromethane-
170

5291 induced and 49 spontaneous liver tumors were screened for H-ras mutations. There was a 

5292 relatively high frequency of activated H-ras among the nonexposed B6C3F1 mice: 67% of the 

5293 spontaneous tumors and 76% of the dichloromethane-induced tumors contained mutations in the 

5294 H-ras gene. Overall, the mutation profile of the dichloromethane-induced tumors did not 

5295 significantly differ from the spontaneous tumors. 

5296 Similarly, Hegi et al. (1993) analyzed the liver tumors from female B6C3F1 mice for 

5297 inactivation of the tumor suppressor genes, p53 and Rb-1. Half of the liver tumors used for this 

5298 study had an activated H-ras oncogene. Twenty liver tumors (15 carcinomas and 5 adenomas) 

5299 were screened for loss of heterozygosity (LOH) on chromosome 11 and 14, which is associated 

5300 with malignant conversion of the p53 gene (chromosome 11) and the Rb-1 gene (chromosome 

5301 14). Only one tumor out of 20 contained a LOH at chromosome 14 and no dichloromethane5302 induced liver tumors contained a LOH at chromosome 11. 

5303 

5304 Liver metabolic studies 

5305 As described in detail in section 3.3, GST-T1 enzymatic activity and distribution is 

5306 variable among species, and there is also considerable intraspecies variability among humans. In 

5307 summary, Reitz et al. (1989) demonstrated a greater metabolic activity with respect to 

5308 dichloromethane in livers of B6C3F1 mice compare with F344 rats, Syrian golden hamsters, and 

5309 humans. The rates of in vitro metabolism by the GST pathway were about 4-, 12-, and 20-fold 

5310 greater in B6C3F1 mouse liver samples than in F344 rat, human, and Syrian golden hamster 

5311 samples, respectively (Reitz et al., 1989). A more recent study characterized the 

5312 dichloromethane metabolic capacity specifically of hepatic GST-T1 (Thier et al., 1998). 

5313 Enzymatic activities of GST-T1 in liver from F344 rats, B6C3F1 mice, Syrian golden hamsters, 

5314 and humans with three different GST-T1 phenotypes (nonconjugators, low conjugators, high 

5315 conjugators) showed the following order with dichloromethane as a substrate: mouse >> rat > 

5316 human high conjugators>human low conjugators > hamster > human nonconjugators. 

5317 

5318 4.5.3. Mechanistic Studies of Lung Effects 

5319 The finding of increased lung tumors in B6C3F1 mice exposed to dichloromethane 

5320 (Mennear et al., 1988; NTP, 1986) has stimulated a number of studies designed to examine the 

5321 mechanism for dichloromethane-induced lung tumors in this animal. The lung tumor mechanism 

5322 studies were conducted with B6C3F1 mice, and the frequency of lung tumors in control animals 

5323 was very low. Time-course studies for lung tumor development were conducted, and it was 

5324 found that the onset of lung tumor development was much shorter than liver tumors (Kari et al., 

5325 1993) (reported in Maronpot et al. [1995]). As a result, it is hypothesized that a potential 

5326 common mechanism independent of liver metabolism is producing tumors in the lung. As with 

5327 the liver tumors, there were no significant increases in mutations for the K-ras oncogene 

5328 (Devereux et al., 1993) or the p53 and Rb-1 tumor suppressor genes (Hegi et al., 1993). 
171

5329 Additionally, the Clara cells, which are non-ciliary secretory cells found in the primary 

5330 bronchioles of the lung, are selectively targeted after dichloromethane exposure. Acute 

5331 dichloromethane exposure produces Clara cell vacuolization, which is not sustained with long5332 term exposure (Foster et al., 1992). There is a correlation between the acute effects on the Clara 

5333 cell and the lung tumors from chronic exposure to dichloromethane (Kari et al., 1993). 

5334 However, the exact mechanism for producing these lung effects is not completely understood at 

5335 this point. Provided below is a summary of the studies examining the potential mechanisms for 

5336 producing lung tumors resulting from dichloromethane exposure. 

5337 

5338 Lung tumor characterization studies 

5339 Kari et al. (1993) (also summarized in Maronpot et al. [1995]) demonstrated that only 26 

5340 weeks of exposure to 2,000 ppm was necessary to produce significantly increased incidence of 

5341 female B6C3F1 mice with lung tumors. In the six “stop-exposure” protocol experiment 

5342 (26 weeks exposure followed by 78 weeks without exposure, 78 weeks without exposure 

5343 followed by 26 weeks exposure, 52 weeks without exposure followed by 52 weeks with 

5344 exposure, 52 weeks exposed followed by 52 weeks without exposure, 78 weeks exposed 

5345 followed by 26 weeks without exposure, and 26 weeks without exposure followed by 78 weeks 

5346 with exposure), early but not late exposure for 26 or 52 weeks resulted in an increased incidence 

5347 of animals with lung tumors (adenoma or carcinomas). Respective percentages of animals with 

5348 lung tumors were 7.5% (5/67), 31% (21/68), 4% (3/67), 63% (40/63), and 15% (10/67) for the 

5349 controls, early 26-, late 26-, early 52-, and late 52-week exposure groups, respectively. With the 

5350 78-week exposures, both the early and late exposure regimens produced an increased incidence 

5351 of lung tumors compared with controls (56% [38/68] and 19% [13/68], respectively), compared 

5352 with the incidence of 63% (42/67) seen in the group exposed for the full 104 weeks. Thus a 

5353 plateauing of risk was seen, with similar incidence rates seen with the early 52-week, early 

5354 78-week, and 104-week exposure periods. The difference in the lung tumor incidence between 

5355 the early and late exposure periods of similar duration was statistically significant (p < 0.01) for 

5356 the 26-, 52-, and 78-week duration protocols. A greater increase in multiplicity of lung tumors 

5357 was also seen with the early 78-week exposure period. As with the liver tumor data from the 

5358 same series of experiments, these data suggest that early exposure was more effective than late 

5359 exposure and that the increased risk continued after cessation of exposure. 

5360 Histopathologic examination of lung tissue from mice killed at 13, 26, 52, 68, 75, 78, 83, 

5361 or 91 weeks of exposure to 2,000 ppm (n = 20 mice per killing) found no evidence of 

5362 nonneoplastic cytotoxicity that preceded the appearance of proliferative neoplastic lung lesions. 

5363 In contrast, incidences of mice with lung adenomas or carcinomas (combined) were elevated at 

5364 interim killings ≥52 weeks; incidences for the interim killings of mice exposed to 2,000 ppm 

5365 (6 hours per day, 5 days per week) between 13 and 91 weeks were 0/20 (0%) at 13 weeks, 

5366 0/20 (9%) at 26 weeks, 6/20 (30%) at 52 weeks, 6/26 (23%) at 68 weeks, 8/20 (40%) at 
172

5367 75 weeks, 9/19 (47%) at 78 weeks, 10/20 (50%) at 83 weeks, and 14/30 (47%) at 91 weeks, 

5368 respectively. Lung hyperplasia was found at an increased incidence only at 91 weeks, well after 

5369 the 26- and 52-week periods that induced increased incidences of mice with lung tumors. 

5370 Kanno et al. (1993) found no evidence for histologic changes or increased cell 

5371 proliferation in lung tissue of female B6C3F1 mice exposed to 2,000 or 8,000 ppm 

5372 dichloromethane for 1, 2, 3, or 4 weeks, compared with control mice, or in mice exposed to 

5373 2,000 ppm for 13 or 26 weeks. Osmotic mini-pumps were used to deliver tritiated thymidine and 

5374 label cells undergoing replicative DNA synthesis over 6-day periods before killing. Labeled 

5375 cells undergoing rapid DNA synthesis and cell proliferation were assessed in sections of 

5376 proximal and terminal bronchioles and alveoli of lungs from groups of 5 mice exposed for 1–

5377 4 weeks or 10 mice exposed for 13 or 26 weeks. There were no exposure-related histopathologic 

5378 or labeling index changes in the alveoli, but lower labeling indices were found in the bronchiolar 

5379 epithelium of exposed mice compared with controls. 

5380 The combined results from the Kari et al. (1993) and Kanno et al. (1993) studies are 

5381 consistent with the hypothesis that dichloromethane-induced lung tumors in B6C3F1 mice are not 

5382 preceded by overt cytotoxicity, enhanced and sustained cell proliferation, or hyperplasia in the 

5383 lung. Two other studies (Casanova et al., 1996; Foster et al., 1992), however, have reported 

5384 evidence for enhanced cell proliferation in lungs of B6C3F1 mice exposed for acute durations to 

5385 airborne dichloromethane. Only one of these studies (Foster et al., 1992), however, looked for 

5386 sustained cell proliferation in the lung with prolonged exposure. In agreement with the results 

5387 from Kanno et al. (1993), no evidence was found for sustained cell proliferation in lungs with 

5388 prolonged exposure to dichloromethane at concentrations demonstrated to induce lung tumors in 

5389 mice. 

5390 Casanova et al. (1996) detected evidence of increased cell proliferation in the lungs of 

5391 male B6C3F1 mice exposed to dichloromethane concentrations >1,500 ppm 6 hours/day for 

5392 3 days. Three or four groups of three mice were exposed to 146, 498, 1,553, or 3,923 ppm 

unlabeled dichloromethane for 2 days and then exposed to [14 5393 C]-labeled dichloromethane for 

5394 6 hours on the third day. Radiolabel incorporated into lung DNA deoxyribonucleosides (after 

5395 removal of DNA-protein cross links containing radiolabeled formaldehyde) was measured as an 

5396 index of cell proliferation. Radiolabel incorporation into lung DNA deoxyribonucleosides 

5397 increased with increasing exposure concentration, with the amount increasing by about 27-fold 

5398 between 146 and 3,923 ppm. In hamsters that did not develop tumors in response to chronic 

5399 inhalation exposure to 3,500 ppm dichloromethane (Burek et al., 1984), no evidence for 

5400 enhanced radiolabel incorporation into lung DNA deoxyribonucleosides was found following 

5401 acute exposure (Casanova et al., 1996). 

5402 Devereux et al. (1993) (also summarized in Maronpot et al. [1995]) analyzed lung tumors 

5403 in female B6C3F1 mice for the presence of activated K-ras oncogenes. Fifty-four 

5404 dichloromethane-induced and 17 spontaneous lung tumors (7 from the NTP [1986] study and 10 
173

5405 from a study in C57BL/6 × C34F1 mice reported by Candrian et al. [1991]) were screened for K5406 ras mutations. Twenty percent of the dichloromethane-induced tumors and 24% of the 

5407 spontaneous tumors contained mutations in the K-ras gene. Devereux et al. (1993) stated that 

5408 there may be a significant difference in the incidence of K-ras activation between spontaneous 

5409 and dichloromethane-induced tumors. However, the small number of the spontaneous tumors 

5410 that were available for the study limits the conclusions that can be made from the results. 

5411 Hegi et al. (1993) analyzed the lung tumors from female B6C3F1 mice for inactivation of 

5412 the tumor suppressor genes, p53 and Rb-1. Forty-nine dichloromethane-induced lung 

5413 carcinomas, five lung adenomas, and seven spontaneous lung carcinomas were screened for 

5414 LOH on mouse chromosome 11 and 14, which is associated with malignant conversion of the 

5415 p53 gene (chromosome 11) and the Rb-1 gene (chromosome 14). Fourteen percent (n = 7) of the 

5416 dichloromethane-induced lung carcinomas exhibited LOH at chromosome 11. No p53 mutations 

5417 were detected in the seven spontaneous lung tumors or the five dichloromethane-induced lung 

5418 adenomas. Only three dichloromethane-induced tumors exhibited LOH at chromosome 14. The 

5419 authors noted that inactivation of the p53 and Rb-1 tumor suppressor genes infrequently occur in 

5420 lung and liver tumors. 

5421 

5422 Clara cell studies 

5423 Foster et al. (1992) found enhanced cell proliferation in bronchiolar cells and, to a lesser 

5424 degree, alveolar cells in the lungs of male B6C3F1 mice exposed for acute durations (2, 5, 8, or 

5425 9 days) to 4,000 ppm dichloromethane (6 hours/day, 5 days/week), but the response was less 

5426 distinct after subchronic durations of exposure (89, 92, or 93 days). To measure cell 

5427 proliferation, mice (n = 5 per exposure-duration group) were given subcutaneous doses of 

5428 tritiated thymidine for five consecutive days prior to killing. Labeled cells in bronchiolar or 

5429 alveolar epithelium in lung sections were counted to assess the number of cells in S phase per 

5430 1,000 cells. Counts of bronchiolar epithelium cells in S phase in exposed mice sacrificed on 

5431 days 2, 5, 8, and 9 were approximately 2-, 15-, 3-, and 5-fold higher, respectively, than those of 

5432 unexposed mice at day 0 of the experiment. In exposed mice sacrificed on days 89, 92, and 93, 

5433 less than twofold increases in bronchiolar epithelium cell labeling were observed. Increased cell 

5434 labeling was found in alveolar epithelium only on day 8 (about seven- to eightfold increase) and 

5435 day 9 (about fourfold increase). Vacuolation of the Clara cells of the bronchiolar epithelium was 

5436 observed on day 2 (scored as ++, majority of cells affected), day 9 (+++, virtually all the cells 

5437 affected), and day 44 (+, moderate effect to cells) but was not apparent on days 5, 8, 40, 43, 89, 

5438 92, or 93. No hyperplasia of the bronchiolar epithelium or changes to Type II alveolar epithelial 

5439 cells were observed in the lungs of any of the exposed mice at any time point. The appearance 

5440 and disappearance of the Clara cell vacuolation were generally correlated with the appearance 

5441 and disappearance of enhanced cell proliferation in the bronchiolar epithelium; enhanced cell 

5442 proliferation was observed on days 2, 5, 8, and 9 (along with appearance of Clara cell 
174

5443 vacuolation on days 2 and 9) but was not observed on days 89, 92, and 93 when Clara cell 

5444 lesions also were not observed. This suggests that cell proliferation was enhanced in response to 

5445 Clara cell damage but was not sustained with repeated exposure to dichloromethane. 

5446 Currently, a mechanistic connection has not been established between the acute effects of 

5447 dichloromethane on Clara cells in the bronchiolar epithelium and the development of lung 

5448 tumors in B6C3F1 mice exposed by inhalation to concentrations ≥2,000 ppm dichloromethane 

5449 for 2 years (NTP, 1986) or for 26 weeks followed by no exposure through 2 years (Maronpot et 

5450 al., 1995; Kari et al., 1993). There appears to be a concordance between exposure concentrations 

5451 inducing acute Clara cell vacuolation (≥2,000 ppm) and those inducing lung tumors 

5452 (≥2,000 ppm). However, transient acute Clara cell vacuolation does not appear to progress to 

5453 necrosis or lead to sustained cell proliferation (which could promote the growth of tumor5454 initiated cells) and appears to be dependent on CYP metabolism of dichloromethane (see the 

5455 following paragraphs discussing pertinent findings reported by Foster et al. [1994, 1992]). In 

5456 contrast, there is consistent and specific evidence for an association between the formation of 

5457 DNA-reactive GST-pathway metabolites and the formation of lung and liver tumors from 

5458 inhalation exposure (see sections 4.5.2 and 4.7.3).

5459 Foster et al. (1992) noted that the appearance and disappearance of Clara cell vacuolation 

5460 in mouse lungs showed concordance with temporal patterns for immunologic staining for 

5461 CYP2B1 and 2B2 levels in lung sections. A similar temporal pattern was reported for CYP2B1 

5462 and 2B2 monooxygenase activities (ethoxycoumarin O-dealkylation or aldrin epoxidation) 

5463 assayed in lung microsomes. When there was a marked decrease in CYP2B1 and 2B2 staining 

5464 (e.g., on day 5) or monooxygenase activities, the lesion was not present. Similarly, the 

5465 appearance of the lesion was preceded (the day before) by the recovery of monooxygenase 

5466 activities or immunologic staining close to control levels. These patterns suggested to Foster et 

5467 al. (1992) that Clara cells may have developed tolerance to dichloromethane due to inactivation 

5468 of a CYP isozyme. 

5469 In subsequent studies, increased percentages of vacuolated bronchiolar epithelium cells 

5470 were noted in mice exposed to 2,000 ppm (26.3 ± 6.7%) or 4,000 ppm (64.8 ± 12.8%), but 

5471 vacuolated cells were not observed in bronchiolar epithelium of lung sections from control mice 

5472 or mice exposed to 125, 250, 500, or 1,000 ppm (Foster et al., 1994). Pretreatment with the CYP 

5473 inhibitor, piperonyl butoxide, counteracted the 2,000 ppm effect (2.4 ± 3.6% vacuolated cells), 

5474 whereas GSH-depleted mice showed no statistically significant change in percentage of 

5475 vacuolated cells (32.7 ± 16.9%) compared with the mean percentage in mice exposed to 

5476 2,000 ppm without pretreatment. No consistent, statistically significant, exposure-related 

5477 changes were found in cytosolic GST metabolic activities (with dichloromethane as substrate) or 

5478 microsomal CYP monooxygenase activities (ethoxycoumarin O-dealkylation), but mean 

5479 cytosolic levels of nonprotein sulfhydryl compounds were elevated in lungs of mice exposed to 

5480 1,000 and 2,000 ppm (134.6 ± 17.1 and 146.4 ± 6.7 nmol/mg protein, respectively) compared 
175

5481 with control levels (109.5 ± 7.6 nmol/mg protein). Increased cell proliferation was found in 

5482 cultured Clara cells isolated from 4,000 ppm exposed mice compared with nonexposed mice; 

5483 respective values for percentage of cells in S phase were 18.97 ± 1.18 and 2.02 ± 0.86% (Foster 

5484 et al., 1994). 

5485 Results from the studies by Foster et al. (1994, 1992) indicate that 6-hour exposures of 

5486 B6C3F1 mice to dichloromethane concentrations ≥2,000 ppm caused transient Clara cell 

5487 vacuolation in the bronchiolar epithelium, which was not consistently observed following 

5488 repeated exposures. With repeated exposure to 4,000 ppm, the Clara cell vacuolation did not 

5489 progress to necrosis, and no hyperplasia of the bronchiolar epithelium was found after up to 

5490 13 weeks of exposure. The transient Clara cell vacuolation was decreased by CYP inhibition 

5491 with piperonyl butoxide and was unaffected by GSH depletion, indicating that a CYP metabolite 

5492 was involved. Clara cell vacuolation was not found after five consecutive daily 6-hour 

5493 exposures to 4,000 ppm but reappeared after 2 days without exposure followed by two additional 

5494 consecutive daily exposures (day 9). With repeated exposure, the lesion was detected at a 

5495 diminished severity on day 44 (but was not found on day 40 or 43) and on day 93 (but was not 

5496 found on day 89 or 92). The temporal pattern of Clara cell vacuolation with repeated exposure 

5497 was reflected in the occurrence of transiently decreased CYP metabolic activity after the 

5498 appearance of vacuolation. Foster et al. (1994, 1992) proposed that the diminishment of severity, 

5499 or the disappearance, of the Clara cell vacuolation with repeated exposure was due to the 

5500 development of a tolerance to dichloromethane, linked with a decrease of CYP metabolism of 

5501 dichloromethane. 

5502 

5503 4.5.4. Mechanistic Studies of Neurological Effects

5504 Several neurobehavioral studies (see section 4.4.3 for a complete summary) have 

5505 demonstrated that dichloromethane exposure results in decreased spontaneous motor activity 

5506 with pronounced lethargy at high concentrations (1,000 ppm or greater). These effects, 

5507 combined with the observation that dichloromethane impairs learning and memory (Alexeef and 

5508 Kilgore, 1983) and affects production of evoked responses to sensory stimuli (Herr and Boyes, 

5509 1997; Rebert et al., 1989), indicate that dichloromethane produces significant neurological 

5510 effects. The mechanisms behind producing these changes have been examined by measuring 

5511 changes in neurotransmitter levels and changes in neurotransmitter localization. Specific brain 

5512 regions (e.g., hippocampus, caudate nucleus, cerebellum) were analyzed to determine if 

5513 dichloromethane-induced behavioral effects, such as learning and memory (hippocampus, 

5514 caudate nucleus) and movement (cerebellum), are resulting from pathological changes in these 

5515 regions. Changes in neurotransmitter levels were also monitored to see if there was any 

5516 correlation in behavior and neurochemical changes. Summaries of these studies are provided 

5517 below. It is not yet known if dichloromethane directly interacts with neuronal receptors, as has 

5518 been demonstrated for toluene and ethanol, two other solvents with neurobehavioral and 
176

5519 neurophysiological profiles that are similar to those of dichloromethane (for a review see Bowen 

5520 et al. [2006]). 

5521 Kanada et al. (1994) examined the effect of dichloromethane on acetylcholine and 

5522 catecholamines (dopamine, norepinephrine, serotonin) and their metabolites in the midbrain, 

5523 hypothalamus, hippocampus, and medulla from male Sprague-Dawley rats (four to five per 

5524 group). The rats were sacrificed 2 hours after a single gavage dose of 0 or 534 mg/kg of 

5525 undiluted dichloromethane. Administration of dichloromethane significantly increased the 

5526 concentration of acetylcholine in the hippocampus and increased dopamine and serotonin levels 

5527 in the medulla. Dichloromethane decreased norepinephrine levels in the midbrain, and 

5528 hypothalamus and serotonin levels were decreased in the hypothalamus. There was a trend 

5529 toward decreased dopamine in the hypothalamus, but the variability between the animals was so 

5530 high that the effect was not significant. The authors speculated that increased acetylcholine 

5531 release from dichloromethane administration may be due to decreased acetylcholine release from 

5532 the nerve terminals. It is unclear as to how these neurochemical changes could be correlated to 

5533 the neurobehavioral changes observed after dichloromethane exposure. 

5534 In a 2-week exposure study, male Wistar rats were exposed to dichloromethane at 500 or 

5535 1,000 ppm (6 hours/day, 5 days/week for 1 or 2 weeks) or 1,000 ppm TWA (1 hour at 100 ppm, 

5536 1 hour peak at 2,800 ppm, 1 hour at 100 ppm, repeated immediately, 5 days/week for 1 or 

5537 2 weeks) (Savolainen et al., 1981). Brains were removed from rats at the end of the study and 

5538 analyzed. The 1,000 ppm TWA group displayed increases in cerebral RNA. Other changes 

5539 noted for this group in the cerebrum included significant increases in NADPH-diaphorase and 

5540 succinate dehydrogenase activity. These changes suggest increased neural activity to possibly 

5541 offset the overall inhibitory effect of dichloromethane in the CNS. It could also possibly explain 

5542 why lethargic effects decrease with continued dichloromethane exposure, and this result 

5543 demonstrates a neuroprotective mechanism resulting from dichloromethane exposure. After a 

5544 7-day withdrawal, RNA levels in the cerebrum were significantly lower in the 1,000 ppm group. 

5545 Succinate dehydrogenase levels remained lowered in the 1,000 ppm TWA group after the 7-day 

5546 exposure-free period. 

5547 Changes in brain catecholamine levels after a subacute exposure to dichloromethane were 

5548 evaluated using male Sprague-Dawley rats (Fuxe et al., 1984). Rats were exposed to 70, 300, 

5549 and 1,000 ppm dichloromethane, 6 hours/day for 3 consecutive days. At all exposures, there was 

5550 a significant decrease of catecholamine concentrations in the posterior periventricular region of 

5551 the hypothalamus. The impact of dichloromethane was also evaluated on the hypothalamic5552 pituitary gonadal axis. The hypothalamus regulates secretion of reproductive hormones, such as 

5553 follicle-stimulating hormone and luteinizing hormone. The levels of the hormone release were 

5554 not significantly changed with dichloromethane exposure. In the caudate nucleus, which is 

5555 involved in memory processes, the catecholamine level initially increased (at 70 ppm) and then 

5556 was lower (1,000 ppm) in comparison to the control. The study overall demonstrates significant 
177

5557 changes in catecholamine levels in the hypothalamus and caudate nucleus. Catecholamine level 

5558 changes in the hypothalamus did not have any significant effect on hormonal release and 

5559 decreased catecholamine levels in the caudate nucleus at higher exposures may lead to memory 

5560 and learning impairment. 

5561 A series of studies were conducted in male and female Mongolian gerbils exposed 

5562 continuously to ≥210 ppm dichloromethane for 3 months, followed by a 4-month exposure-free 

5563 period (Karlsson et al., 1987; Briving et al., 1986; Rosengren et al., 1986). Decreased DNA 

5564 concentrations were noted in the hippocampus at both the 210 and 350 ppm exposures. At 

5565 350 ppm, there was also decreased DNA concentration in the cerebellar hemispheres, indicating 

5566 a decreased cell density in these regions probably due to cell loss (Rosengren et al., 1986). 

5567 These findings indicate that the cerebellum, which is the section of the brain that regulates motor 

5568 control, is a target for dichloromethane. In the same study, increased astroglial proteins were 

5569 found in the frontal and sensory motor cerebral cortex, which directly correlated to the 

5570 astrogliosis that was observed in those areas. Up-regulation of these astroglial proteins is a good 

5571 indicator of neuronal injury (Rosengren et al., 1986). 

5572 Karlsson et al. (1987) measured DNA concentrations in different regions of the gerbil 

5573 brain. The total brain protein concentration per wet weight was not significantly different 

5574 between dichloromethane-exposed and control animals. However, DNA concentrations per wet 

5575 weight were significantly decreased in the hippocampus after dichloromethane exposure. No 

5576 other examined regions demonstrated significant changes in DNA concentrations after 

5577 dichloromethane exposure. Therefore, this result indicates that the hippocampus, which plays a 

5578 role in the formation of new memories, is another target for dichloromethane in the CNS. This 

5579 selective DNA concentration decrease observed in the hippocampus is a sign of neurotoxicity as 

5580 noted by the authors and may possibly explain why some studies have noted memory and 

5581 learning deficits with dichloromethane exposure. 

5582 At a 210 ppm exposure, Briving et al. (1986) observed that dichloromethane decreased 

5583 glutamate, γ-aminobutyric acid, and phosphoethanolamine levels in the frontal cortex, while 

5584 glutamate and γ-aminobutyric acid were increased in the posterior cerebellar vermis. Increased 

5585 levels of glutamate in the posterior cerebellar vermis could reflect an activation of astrocytic glia,

5586 since glutamine synthetase is localized exclusively in astrocytes. 

5587 

5588 4.6. SYNTHESIS OF MAJOR NONCANCER EFFECTS 

5589 4.6.1. Oral Exposures 

5590 4.6.1.1. Summary of Human Data 

5591 Information on noncancer effects in humans exposed orally to dichloromethane are 

5592 restricted to case reports of neurological impairment, liver and kidney effects (as severe as organ 

5593 failure), and gastrointestinal irritation in individuals who ingested amounts ranging from about 

5594 25 to 300 mL (Chang et al., 1999; Hughes and Tracey, 1993). Neurological effects with these 
178

5595 individuals consisted of general CNS depressive symptoms, such as drowsiness, confusion, 

5596 headache, and dizziness. Hemoglobinuria has been noted as a kidney effect associated with 

5597 ingestions. 

5598 

5599 4.6.1.2. Summary of Animal Data 

5600 Acute oral or intraperitoneal administration of dichloromethane in animals has resulted in 

5601 several significant effects. General activity and function were affected as evidenced by 

5602 decreased neuromuscular activity (Moser et al., 1995). Additionally, decreased sensorimotor 

5603 function was detected through measurement of evoked potentials (Herr and Boyes, 1997) and by 

5604 using the FOB (Moser et al., 1995). Neurochemical changes (e.g. acetylcholine, dopamine, 

5605 norepinephrine, serotonin) were detected 2 hours after oral dosage of dichloromethane within 

5606 specific parts of the brain. It should be noted that all the acute effects that were observed after 

5607 oral or intraperitoneal administration occurred within 5 hours after dosage. No other significant 

5608 organ effects were noted after a single acute oral exposure, but in oral pharmacokinetic studies it 

5609 is known that dichloromethane is primarily distributed to the liver, lungs, and kidneys (Angelo et 

5610 al., 1986a). 

5611 Results from short-term, subchronic, and chronic oral toxicity studies in laboratory 

5612 animals are summarized in Table 4-35. The data indicate that rats may be more sensitive than 

5613 mice to nonneoplastic liver effects from orally administered dichloromethane, as evidenced by 

5614 lower NOAELs and LOAELs, with more severe liver effects in rats. The most frequently 

5615 observed liver effect was hepatocyte vacuolation, seen with drinking water exposure (90 days) in 

5616 F344 rats at ≥166 mg/kg-day and B6C3F1 mice at 586 mg/kg-day (Kirschman et al., 1986) and

5617 with gavage exposure (14 days) in CD-1 mice at 333 mg/kg-day (Condie et al., 1983). 

5618 Hepatocyte degeneration or necrosis was observed in female F344 rats exposed in drinking water 

5619 for 90 days to 1,469 mg/kg-day (Kirschman et al., 1986) and in female F344 rats exposed by 

5620 gavage for 14 days to 337 mg/kg-day (Berman et al., 1995), but was not seen in a 90-day 

5621 drinking water study in B6C3F1 mice exposed to doses as high as 2,030 mg/kg-day (Kirschman 

5622 et al., 1986). In the chronic-duration (2-year) study, liver effects were described as 

5623 nonneoplastic foci and areas of alteration in F344 rats exposed to drinking water doses between 

5624 50 and 250 mg/kg-day; an increased incidence of fatty changes in the liver was also noted but the 

5625 incidence of the latter was not provided (Serota et al., 1986a). These effects were considered to 

5626 be nonneoplastic for several reasons. Serota et al., (1986b) observed a dose-related increased 

5627 incidence of 0, 65, 92, 97, 98 and 100% in male rats and 51, 41, 73, 89, 91 and 85% in female 

5628 rats for the 0, 5, 50, 125, 250 and 250 with recovery groups, respectively. Evidence for liver 

5629 tumors has been reported in female rats only. Specifically, evidence for liver tumors in rats 

5630 includes a small number of hepatocellular carcinomas observed in female rats at 50 and 250 

5631 mg/kg-day, which reached statistical significance (for trend and for individual pairwise 

5632 comparisons) only with the combined grouping of neoplastic nodules and hepatocellular 
179

5633 carcinomas. In male rats, only one hepatocellular carcinoma was observed in all of the exposure 

5634 groups (compared to 4 in the controls), and the incidence of neoplastic nodules and 

5635 hepatocellular carcinomas was higher in controls (16%) than in any exposure group (16, 3, 0, 6, 

5636 5 and 13% for the 0, 5, 50,125, 250 mg/kg-day and 250 with recovery groups, respectively). The 

5637 authors (Serota et al., 1986a) did not elaborate on the characterization of the altered foci. 

5638 However, the characterization of altered foci could range from a focal change in fat distribution 

5639 (nonneoplastic effect) to enzyme altered foci which are generally considered a precursor to 

5640 tumor formation (Goodman et al., 1994). Serota et al (1986a) reported an increased incidence of 

5641 fatty change in the liver at doses of 50 mg/kg-day and higher, but the incidence was not reported. 

5642 In addition, a 90-day study (Kirschman et al., 1986) demonstrated that increased fatty deposits 

5643 were present in the hepatocyte vacuoles. Therefore, the altered foci (i.e. change in fat 

5644 distribution) observed by Serota et al., (1986b) may represent a precursor to fatty liver changes 

5645 which is considered a nonneoplastic effect. Taken together, the data support the conclusion that 

5646 the altered foci were nonneoplastic. 

5647 The NOAEL and LOAEL, 101 and 337 mg/kg-day, for altered neurological functions in 

5648 female F344 rats (as reported by Moser et al. [1995]) were identical to those reported by Berman 

5649 et al. (1995) for hepatocyte necrosis in female F344 rats. In the 90-day (Kirschman et al., 1986) 

5650 and 104-week (Serota et al., 1986a, b) drinking water studies, no obvious clinical signs of 

5651 neurological impairment were observed in rats or mice at exposure levels that induced liver 

5652 effects (see Table 4-35), but this study did not include a standardized neurological testing 

5653 battery. 
180

Table 4-35. NOAELs and LOAELs in selected animal studies involving oral exposure to dichlorofor short-term, subchronic, or chronic durations 

Type of effect and 

exposure, reference Species and exposure details Results 

Hepatic, 14-day gavage 

Berman et al. (1995) F344 rat, female, 8/dose group

0, 34, 101, 337, 1,012 mg/kg-day 

Hepatocyte necrosis 

Condie et al. (1983) CD-1 mouse, male, 5/group for 

histological examinations; 8/group for blood urea nitrogen, serum creatinine, and serum glutamate-pyruvate transaminase; 0, 133, 333, 665 mg/kgday 

Hepatocyte vacuolation (minimal to mild in 3/5) 

Hepatic, 90-day drinking water 

Kirschman et al. (1986) F344 rat, male and female; 15/sex/group; 

males 0, 166, 420, 1,200 mg/kg-day females 0, 209, 607, 1,469 mg/kg-day 

Hepatic vacuolation (generalized, centrilobular, orperiportal, at lowest dose, in 10/15 males and 13/15 females compared with 1/15 males and 6/15 females in controls) 

Kirschman et al. (1986) B6C3F1 mouse, male and female, 

males 0, 226, 587, 1,911 mg/kg-day females 0, 231, 586, 2,030 mg/kg-day 

Hepatic vacuolation (increased severity of centrilobular fatty 

change in mid- and high-dose groups compared with controls) 

Hepatic, 104-week drinking water 

Serota et al. (1986a) F344 rat, male and female, 

0, 5, 50, 125, 250 mg/kg-day 

Liver foci/areas of alteration (considered nonneoplastic histologic changes); fatty liver changes also seen at same 

doses but incidence data not reported; no evidence that 

increased altered foci progresses to liver tumor formation 

Serota et al. (1986b) Hazelton Laboratories (1983) 

B6C3F1 mouse, male and female, 

0, 60, 125, 185, 250 mg/kg-day 

Some evidence of fatty liver; marginal increase in the Oil 

Red-O-positive material in the liver 

Neurologic, 14 day 

Moser et al. (1995) F344 rat, female, 

0, 34, 101, 337, 1,012 mg/kg-day 

FOB 24 hours postexposure: altered autonomic, neuromuscular, and sensorimotor and excitability measures 

Reproductive 

General Electric Co. (1976) 

Charles River CD rat, male and female, gavage for 90 d before mating (10 days between last exposure and mating period); 0, 25, 75, 225 mg/kg-day; F1 offspring received same treatment as parents for 90 days 

Reproductive performance of F0 and histologic examination of tissues from F1 offspring 
181

Table 4-35. NOAELs and LOAELs in selected animal studies involving oral exposure to dichlorofor short-term, subchronic, or chronic durations 

Type of effect and 

exposure, reference Species and exposure details Results 

Raje et al. (1988) Swiss-Webster mouse, male, 0, 250, 

500 mg/kg (subcutaneous injection), 3× per week, 4 weeks prior to mating 

with nonexposed females (1 week between last exposure and mating period)

No statistically significant effects on testes, number of litters, live fetuses/litter, percent dead fetuses/litter, percent resorbed/litter, or fertility index 

Developmental 

Narotsky and Kavlock (1995) 

F344 rat, pregnant female, gavage on GDs 6–19; 0, 338, 450 mg/kg-day 

Maternal: weight gain depression Fetal: no effects on pup survival, resorptions, pup weight 

Other developmental 

No studies 
182

5654 Results from a limited number of studies do not provide evidence for effects on 

5655 reproductive or developmental endpoints (Table 4-35). No effects on pup survival, resorptions, 

5656 or pup weight were found following exposure of pregnant F344 rats to doses as high as 

5657 450 mg/kg-day on GDs 6–19, a dose that depressed maternal weight gain (Narotsky and 

5658 Kavlock, 1995), and no effects on reproductive performance endpoints (fertility index, number 

5659 of pups per litter, pup survival) were found in Charles River CD rats exposed for 90 days before 

5660 mating to doses as high as 225 mg/kg-day. There are no oral exposure studies focusing on 

5661 neurobehavioral effects or other developmental outcomes. 

5662 

5663 4.6.2. Inhalation Exposures 

5664 4.6.2.1. Summary of Human Data 

5665 As discussed in sections 4.1.3.1 and 4.1.3.2, acute inhalation exposure of humans to 

5666 dichloromethane has been associated with cardiovascular impairments due to decreased oxygen 

5667 availability from COHb formation and neurological impairment from interaction of 

5668 dichloromethane with nervous system membranes. Results from studies of acutely exposed 

5669 human subjects indicate that acute neurobehavioral deficits, measured, for example, by 

5670 psychomotor tasks, tests of hand-eye coordination, visual evoked response changes, and auditory 

5671 vigilance, may occur at concentrations >200 ppm with 4–8 hours of exposure (Bos et al., 2006; 

5672 American Conference of Governmental Industrial Hygienists [ACGIH], 2001; ATSDR, 2000; 

5673 Cherry et al., 1983; Putz et al., 1979; Gamberale et al., 1975; Winneke, 1974). 

5674 The clinical and workplace studies of noncancer health effects of chronic 

5675 dichloromethane exposure have examined markers of disease and specific clinical endpoints 

5676 relating to cardiac, neurological disease, hepatic function, and reproductive health. As 

5677 summarized in section 4.1.2.9, the limited available data do not provide evidence of cardiac 

5678 damage related to dichloromethane exposure in occupationally exposed workers (Hearne and 

5679 Pifer, 1999; Tomenson et al., 1997; Gibbs et al., 1996; Lanes et al., 1993; Ott et al., 1983d; 

5680 Cherry et al., 1981). Relatively little is known about the long-term neurological effects of 

5681 chronic exposures, although there are studies that provide some evidence of an increased 

5682 prevalence of neurological symptoms among workers with average exposures of 75–100 ppm 

5683 (Cherry et al., 1981), long-term effects on some neurological measures (i.e., possible detriments 

5684 in attention and reaction time in complex tasks) in workers whose past exposures were in the 

5685 100–200 ppm range (Lash et al., 1991), and an increased risk of suicide in worker cohort studies 

5686 (Hearne and Pifer, 1999; Gibbs, 1992). Given the suggestions from these studies and their 

5687 limitations (particularly with respect to sample size and power considerations), the statement that 

5688 there are no long-term neurological effects of chronic exposures to dichloromethane cannot be 

5689 made with confidence. With respect to markers of hepatic damage, three studies measured 

5690 serum enzyme and bilirubin levels in workers exposed to dichloromethane (Soden, 1993; 

5691 Kolodner et al., 1990; Ott et al., 1983c). There is some evidence of increasing levels of serum 
183

5692 bilirubin with increasing dichloromethane exposure (Kolodner et al., 1990; Ott et al., 1983c), but 

5693 there are no consistent patterns with respect to the other hepatic enzymes examined (serum γ5694 glutamyl transferase, serum AST, serum ALT). Thus these studies do not provide clear evidence 

5695 of hepatic damage in dichloromethane exposed workers to the extent that this damage could be 

5696 detected by these serologic measures. 

5697 Only limited, and somewhat indirect, evidence pertaining to immune-related effects of 

5698 dichloromethane in humans is available. No risk of the broad category of infection- and parasite5699 related mortality was reported by Hearne and Pifer (1999), but there was some evidence of an 

5700 increased risk of influenza and pneumonia-related mortality at two cellulose triacetate fiber 

5701 production work sites in Maryland and South Carolina (Gibbs, 1992). 

5702 Few studies have been conducted pertaining to reproductive effects (i.e., spontaneous 

5703 abortion, low birth weight, or oligospermia) of dichloromethane exposure from workplace 

5704 settings (Wells et al., 1989; Kelly, 1988; Taskinen et al., 1986) or environmental settings (Bell et 

5705 al., 1991). Of these, the data pertaining to spontaneous abortion provide the strongest evidence 

5706 of an adverse effect of dichloromethane exposure. The limitations of the only study (Taskinen et 

5707 al., 1986) pertaining to this outcome, however, do not allow firm conclusions to be made 

5708 regarding dichloromethane and risk of spontaneous abortion in humans. 

5709 

5710 4.6.2.2. Summary of Animal Studies 

5711 Acute and short-term (up to 7 days) inhalational exposure to dichloromethane has 

5712 resulted in neurological and hepatocellular changes. Several neurological-mediated parameters 

5713 were reported, including decreased activity (Kjellstrand et al., 1985; Weinstein et al., 1972; 

5714 Heppel and Neal, 1944), impairment of learning and memory (Alexeef and Kilgore, 1983), and 

5715 changes in responses to sensory stimuli (Rebert et al., 1989). Although learning and memory 

5716 properties were impaired in one acute exposure (47,000 ppm until loss of righting reflex), it 

5717 should be noted that this effect has not been characterized by using other learning and memory 

5718 tasks nor any other exposure paradigms. In a 3-day exposure to dichloromethane (70, 300, or 

5719 1,000 ppm 6 hours/day), it was found that in the rat brain there were changes in catecholamine 

5720 (dopamine, serotonin, norepinephrine) in the hypothalamus and caudate nucleus (Fuxe et al., 

5721 1984). The catecholamine level changes did not affect hormonal release, which is a primary 

5722 function of the hypothalamus. 

5723 Another acute exposure study examined immunological response as measured by 

5724 increased streptoccocal pneumonia-related mortality and decreased bactericidal activity of 

5725 pulmonary macrophages in CD-1 mice following a single 3-hour exposure to dichloromethane at 

5726 100 ppm (Aranyi et al., 1986). No effects were seen at 50 ppm. A 4-week inhalation exposure 

5727 to 5,000 ppm dichloromethane in rats did not result in changes in immune response as measured 

5728 by the sheep red blood cell assay (Warbrick et al., 2003). These studies suggest a localized, 

5729 portal-of-entry effect within the lung, without evidence of systemic immunosuppression. 
184

5730 Mouse hepatocytes showed balloon degeneration (dissociation of polyribosomes and 

5731 swelling of rough endoplasmic reticulum) within 12 hours of exposure to 5,000 ppm (Weinstein 

5732 et al., 1972). A subacute exposure in Wistar rats to 500 ppm dichloromethane 6 hours/day for 6 

5733 days resulted in increased hemochrome content in liver microsomal CYP (Savolainen et al., 

5734 1977). 

5735 Results pertaining to liver, lung, and neurological effects from longer (>7 days) 

5736 subchronic and chronic inhalation toxicity studies in laboratory animals are summarized in 

5737 Table 4-36; reproductive and developmental studies are summarized in Table 4-37.
185

Table 4-36. NOAELs and LOAELs in animal studies involving inhalation exposure to dichloromsubchronic or chronic durations, hepatic, pulmonary, and neurologic effects 

Type of effect and NOA

exposure period, 

reference Species and exposure details Results 

Hepatic, 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 (continuous exposure; 14 weeks) 

Fatty liver at 1,000 ppm in dogs, “borderline” liver changes in monkey at 5,000 ppm, mottled liver changes in rats at 5,000 ppm Decreased movement and lethargy at 1,000 ppm in dogs, mice, and monkey. 

Not idenNot iden1,000 Not iden

Haun 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 mice at 100 ppm Increased fatty liver content at 25, 100 ppm in rats 

25; Not idenNot iden25 

Leuscher et al. (1984) Sprague-Dawley rat, male and female, 

(20/sex/group) - 0, 1,000 ppm (6 hours/day; 90 days) Beagle, male and female (3/sex/group) - 0, 5,000 ppm 

No liver effects noted 1,000 

5,000 

NTP (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 at 4,200 (females); 8,400 (males); decreased BW by 23% and 11% in males and females at 8,400 ppm compared with controls; one male and one female died at 8,400 ppm before the end of the study. 

4,200 

NTP (1986) B6C3F1 mouse, 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 4,200 females) and 8,400 (males); decreased lipid:liver weight ratios at 8,400 (females); at 8,400 ppm, 4/10 males and 2/10 females died before end of study. 

2,100 

Hepatic, 2 years (6 hours/day, 5 day/week) Mennear et al. (1988); NTP (1986) 

F344/N rat, male and female 0, 1,000, 2,000, 4,000 ppm 

Hepatocyte vacuolation and necrosis Hemosiderosis in liver Renal tubular degeneration 

Not idenNot iden1,000 
186

Table 4-36. NOAELs and LOAELs in animal studies involving inhalation exposure to dichloromsubchronic or chronic durations, hepatic, pulmonary, and neurologic effects 

Type of effect and NOA

exposure period, 

reference Species and exposure details Results 

Mennear et al. (1988); NTP (1986) 

B6C3F1 mouse, male and female 

0, 2,000, 4,000 ppm 

Hepatocyte degeneration Renal tubule casts 

Not idenNot iden

Burek et al. (1984) Syrian golden hamster, male and female 

0, 500, 1,500, 3,500 ppm 

No effects on histologic, clinical chemistry, urinalytic, and hematologic variables no obvious clinical signs of toxicity 

3,500 

Burek et al. (1984) Sprague-Dawley rat, male and female 

0, 500, 1,500, 3,500 ppm 

Hepatocyte vacuolation (M and F) Hepatocyte necrosis (M only), no obvious clinical signs of toxicity) 

Not iden500 

Nitschke et al. (1988a) Sprague-Dawley rat, male and female 

0, 50, 200, 500 ppm 

Hepatocyte vacuolation significantly increased in females; non-significant increase in males at 500 

ppm (31% in controls and 40% in 500 ppm group). 

200 

Pulmonary, 13 weeks (6 hours/day, 5 days/week)

NTP (1986) F344 rat, male and female 

0, 525, 1,050, 2,100, 4,200, 8,400 ppm 

Foreign body pneumonia 4,200 

Foster et al. (1992) B6C3F1 mouse, male and female 

0, 4,000 ppm 

Clara cell vacuolation Not iden

Neurological, 14 days

Savolainen et al. (1981) Wistar rats, male 

500, 1,000, 1,000 TWA (100 + 2,800 1-hour peaksa) ppm (6 hours/day, 

5 days/week, 2 weeks) 

Increased RNA in cerebrum at 1,000 ppm; increased enzymatic activitiesb in cerebrum and 

cerebellum at 1,000 ppm TWA 

500 

Neurological, 13–14 weeks 

Mattsson et al. (1990) F344 rat, male and female 

0, 50, 200, 2,000 ppm (6 hours/day, 5 days/week) 

No exposure-related effects on an observational battery, hind-limb grip strength, a battery of evoked potentials, or histology of brain, spinal cord, peripheral nerves; measured 64 hours postexposure

2,000 

Haun 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 iden

Not iden1,000 Not iden
187

Table 4-36. NOAELs and LOAELs in animal studies involving inhalation exposure to dichloromsubchronic or chronic durations, hepatic, pulmonary, and neurologic effects 

Type of effect and NOA

exposure period, 

reference Species and exposure details Results 

Karlsson et al. (1987) Briving et al. (1986) Rosengren et al. (1986) 

Mongolian gerbils, male and female 210, 350, 700 ppm (continuous exposure, followed by 4 month exposure-free period) 

Astrogliosis in frontal and sensory motor cerebral cortex suggested by increases in astroglial proteins; cell loss in cerebellar regions; decreased DNA in hippocampus; neurochemical changes observed at all exposures 

Not iden

Thomas et al. (1972) ICR mice, female 

0, 25, 100 ppm, continuous 

Increased spontaneous activity observed at 25 ppm but not 100 ppm 

Not iden

CoHb, 13-14 weeks

Haun 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 ppm for monkeys and 100 ppm for beagles 

Not iden

COHb, 2 years (6 hours/day, 5 day/week)

Burek et al. (1984) Syrian golden hamster, male and female 

0, 500, 1,500, 3,500 ppm 

About 30% COHb in each exposed group 

Burek et al. (1984) Sprague-Dawley rat, male and female 

0, 500, 1,500, 3,500 ppm 

About 12–14% COHb in each exposed group 

Nitschke et al. (1988a) Sprague-Dawley rat, male and female 

0, 50, 200, 500 ppm 

COHb 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 and γ-aminobutyric acid increased in posterior cerebellar v

5738 5739 5740 
188

5741 

Table 4-37. NOAELs and LOAELs in selected animal studies involving inhalation exposure to 

dichloromethane, reproductive and developmental effects 

Type of effect and exposure NOA

period, reference Species and exposure details Results 

Reproductive 

Nitschke et al. (1988b) F344 rat, male and female, F0: 6 hr/d, 

5 d/wk for 14 wk before mating and GDs 0 to 21; F1: 6 hr/d, 5 d/wk, beginning PND 4 for 17 wk before mating; 0, 100, 500, 1,500 ppm 

No statistically significant effects on fertility or litter size, neonatal survival, growth rates, or histopathologic lesions in F1 or F2 weanlings 

1,500 

Mennear et al. (1988); NTP (1986) 

B6C3F1 mouse; 0, 2,000 or 4,000 ppm, 

6 hours/day, 5 days/week for 2 years 

Testicular atrophy Ovarian atrophy (considered secondary to hepatic effects) 

2,000 Not iden

Raje et al. (1988) Swiss-Webster mouse, male, 2 hr/d, 

5 d/wk for 6 wk before mating with nonexposed females; 0, 100, 150, 200 ppm 

No statistically significant effects on testes, number of litters, live fetuses/litter, percent dead fetuses/litter, percent resorbed/litter Fertility index was lower in 150 and 200 ppm groups (80%) compared with controls and 100 ppm groups (95%) (statistical significance depends on test used). 

200 100 

Developmental 

Schwetz et al. (1975) Swiss-Webster mouse, pregnant female, 

7 hr/d, GDs 6–15; 0, 1,250 ppm 

Maternal effects: 9–10% COHb; increased absolute, not relative, liver weight, increased maternal weight (11–15%). Fetal effects: increased litters with extra center of ossification in sternum 

Not iden1,250 

Schwetz et al. (1975) Sprague-Dawley rat, pregnant female, 

7 hr/d, GDs 6–15; 0, 1,250 ppm 

Maternal: 9–10% COHb; increased absolute, not relative, liver weight Fetal: increased incidence of delayed ossification of sternebrae 

Not iden1,250 

Other developmental 

Bornschein et al. (1980); Hardin and 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 ppm 

Maternal (both protocols): increased absolute and relative liver weight (~10%) Fetal/offspring: decreased fetal BW (~10%); changed behavioral habituation to novel environments; no changes in gross, skeletal, or soft-tissue anomalies 

Not idenNot iden
189

5742 Hepatic centrilobular degeneration was observed in several studies containing different 

5743 species and inhalational exposures. This effect was observed in guinea pigs exposed to 

5744 5,000 ppm (7 hours/day) for 6 months (Heppel et al., 1944). Monkeys, rats, and mice 

5745 continuously exposed (24 hours/day) to 5,000 ppm dichloromethane for 14 weeks also had 

5746 increased centrilobular degeneration (Haun et al., 1972, 1971). This effect was also observed at 

5747 lower exposures when mice were exposed to 4,200 ppm for 6 hours/day for 13 weeks (NTP, 

5748 1986) and in dogs exposed to 1,000 ppm for 24 hours/day for up to 14 weeks (Haun et al., 1972, 

5749 1971). 

5750 Increased incidences of histologic hepatic lesions were not found in F344 rats exposed to 

5751 4,200 or 8,400 ppm 6 hours/day for 13 weeks (NTP, 1986) or in Sprague-Dawley rats exposed to 

5752 10,000 ppm 6 hours/day for 90 days (Leuschner et al., 1984). Hepatic lesions were also not 

5753 observed in beagle dogs exposed to 5,000 ppm 6 hours/day for 90 days (Leuschner et al., 1984) 

5754 or in dogs, monkeys, rats, and mice exposed to 25 or 100 ppm for 24 hours/day for up to 

5755 14 weeks (Haun et al., 1972). Heppel et al. (1944) also demonstrated absence of hepatic lesions 

5756 in unspecified strains of monkeys, rabbits, and rats exposed to 10,000 ppm 4 hours/day for up to 

5757 8 weeks and in unspecified strains of dogs, rabbits, and rats exposed to 5,000 ppm 7 hours/day 

5758 for up to 6 months. 

5759 Gross neurological impairments were observed in several laboratory species with 

5760 repeated exposure to 10,000 ppm for 4 hours/day for 8 weeks (Heppel et al., 1944) or to 1,000 or 

5761 5,000 ppm for 24 hours/day for 14 weeks (Haun et al., 1972, 1971). Dogs exposed to 5,000 ppm 

5762 6 hours/day for 90 days showed slight sedation during exposures, but Sprague-Dawley rats 

5763 exposed to 10,000 ppm for 90 days did not (Leuschner et al., 1984). In F344 rats exposed to 

5764 concentrations up to 2,000 ppm 6 hours/day for 13 weeks, no effects were observed on an 

5765 observational battery, hind-limb grip strength, a battery of evoked potentials, or histology of the 

5766 brain, spinal cord, or peripheral nerves; these tests were conducted beginning 65 hours or more 

5767 after the last exposure (Mattsson et al., 1990). 

5768 Exposure-related nonneoplastic effects on the lungs reported in the subchronic studies 

5769 were restricted to foreign body pneumonia in rats exposed to 8,400 ppm 6 hours/day for 

5770 13 weeks (NTP, 1986), Clara cell vacuolation in mice exposed to 4,000 ppm 6 hours/day for 

5771 13 weeks (Foster et al., 1992), and pulmonary congestion in guinea pigs exposed to 5,000 ppm 

5772 7 hours/day for 6 months (Heppel et al., 1944). 

5773 The chronic duration inhalation studies were conducted at lower exposure levels than the 

5774 short-term and subchronic studies and provide results indicating that the liver is the most 

5775 sensitive target for noncancer toxicity in rats and mice (Table 4-36). Life-time exposure was 

5776 associated with hepatocyte vacuolation and necrosis in F344 rats exposed to 1,000 ppm 

5777 6 hours/day (Mennear et al., 1988; NTP, 1986), hepatocyte vacuolation in Sprague-Dawley rats 

5778 exposed to 500 ppm 6 hours/day (Nitschke et al., 1988a; Burek et al., 1984), and hepatocyte 

5779 degeneration in B6C3F1 mice exposed to 2,000 ppm 6 hours/day (lower concentrations were not 
190

5780 tested in mice) (Mennear et al., 1988; NTP, 1986). As shown in Tables 4-36 and 4-37, other 

5781 effects observed include renal tubular degenerations in F344 rats and B6C3F1 mice at 2,000 ppm, 

5782 testicular atrophy in B6C3F1 mice at 4,000 ppm, and ovarian atrophy in B6C3F1 mice at 

5783 2,000 ppm (considered secondary to hepatic effects). No exposure-related increased incidences 

5784 of nonneoplastic lung lesions were found in any of the chronic studies (Table 4-36). 

5785 In comparison to rats and mice, Syrian golden hamsters are less sensitive to the chronic 

5786 inhalation toxicity of dichloromethane. No exposure-related changes were found in 

5787 comprehensive sets of histologic, clinical chemistry, urinalytic, and hematologic variables 

5788 measured in hamsters exposed for 2 years to 500, 1,500, or 3,500 ppm for 6 hours/day, with the 

5789 exception that mean COHb percentages were about 30% in each of these groups compared with 

5790 a mean value of about 3% for the controls (Burek et al., 1984). 

5791 The reproductive and developmental studies are limiting in terms of the exposure 

5792 regimen used, with two of the developmental studies using only a single, relatively high daily 

5793 exposure over the gestational period (1,250 ppm, GD 6-15 in Schwetz et al. [1975] and 4,500 

5794 ppm, GD 1-17 in Hardin and Manson [1980] and Bornschein [1980]). No significant effects on 

5795 reproductive performance variables were found in a two-generation reproduction assay with 

5796 F344 rats exposed to concentrations as high as 1,500 ppm (Nitschke et al., 1988b). No effects on 

5797 most of the measures of reproductive performance were observed in male mice exposed to 

5798 200 ppm for 2 hours/day for 6 weeks before mating to nonexposed females. Fertility index was 

5799 reduced in the 150 and 200 ppm groups, but the statistical significance of this effect varied 

5800 considerably depending on the statistical test used in this analysis (Raje et al., 1988). No adverse 

5801 effects on fetal development were found following exposure of pregnant Swiss-Webster mice or 

5802 Sprague-Dawley rats to 1,250 ppm 6 hours/day on GDs 6–15 (Schwetz et al., 1975). Following 

5803 exposure of female Long-Evans rats to 4,500 ppm (6 hours/day) for 14 days before breeding plus 

5804 during gestation or during gestation alone, a 10% decrease in fetal BW and changed behavioral 

5805 habituation of the offspring to novel environments were seen (Bornschein et al., 1980; Hardin 

5806 and Manson, 1980). No exposure-related changes in gross, skeletal, or soft-tissue anomalies 

5807 were found. 

5808 

5809 4.6.3. Mode of Action Information 

5810 4.6.3.1. Mode of Action for Nonneoplastic Liver Effects 

5811 Studies of chronically exposed rats, both by the oral route and the inhalation route, 

5812 identified liver changes as the most sensitive exposure-related noncancer effect associated with 

5813 exposure to dichloromethane (Tables 4-35 to 4-37). The liver changes included increased 

5814 incidence of liver foci/areas of alteration and hepatocyte vacuolation in rats and degenerative 

5815 liver effects in rats, guinea pigs, monkeys, and mice. 

5816 The mode of action by which dichloromethane induces these nonneoplastic hepatic 

5817 effects is unknown. The determination of whether or not these effects are due to the parent 
191

5818 material, metabolites of the CYP2E1 pathway, metabolites of the GST pathway, or some 

5819 combination of parent material and metabolites has not been elucidated. The available data 

5820 indicate that rats may be more sensitive than mice to the noncancer hepatotoxicity, but a 

5821 mechanistic explanation of this possible interspecies difference is not currently available. 

5822 

5823 4.6.3.2. Mode of Action for Nonneoplastic Lung Effects 

5824 Single 6-hour inhalation exposures to concentrations ≥2,000 ppm dichloromethane 

5825 produced a transient vacuolation of Clara cells in the bronchiolar epithelium of B6C3F1 mice. 

5826 Vacuolization of the Clara cells disappeared or was diminished with repeated exposure and was 

5827 correlated with subsequent transient diminishment of CYP metabolic activity. CYP inhibition 

5828 with piperonyl butoxide counteracted the vacuolation observed in the Clara cells (Foster et al., 

5829 1994, 1992). With repeated exposure to 4,000 ppm (up to 13 weeks), the Clara cell vacuolation 

5830 did not appear to progress to necrosis, and no hyperplasia of the bronchiolar epithelium was 

5831 found. Foster et al. (1994, 1992) proposed that the diminished severity or disappearance of Clara 

5832 cell vacuolation with repeated exposure was due to the development of tolerance to 

5833 dichloromethane, linked with a transient decrease of CYP metabolism of dichloromethane. The 

5834 available data suggests that CYP metabolism of dichloromethane may be involved in the mode 

5835 of action for the acute effects of dichloromethane on the bronchiolar epithelium of mice. 

5836 Mode of action research attention on lung effects from chronic exposure to 

5837 dichloromethane has focused on neoplastic effect; nonneoplastic lung effects have received 

5838 relatively little attention. No exposure-related increased incidences of nonneoplastic lung lesions 

5839 (including epithelial hyperplasia) were found in any of the chronic studies listed in Table 4-36, 

5840 but chronic inhalation exposure of B6C3F1 mice to concentrations ≥2,000 ppm has consistently 

5841 been shown to induce lung tumors in several studies (Kari et al., 1993; NTP, 1986). In a study 

5842 that included interim sacrifices at 13, 26, 52, 68, 75, 78, 83, and 91 weeks of B6C3F1 mice 

5843 exposed to 2,000 ppm, hyperplasia of lung epithelium (the only nonneoplastic lung lesion found) 

5844 was found in only three of the eight interim sacrifices (68, 78, and 91 weeks) and was only 

5845 statistically significantly elevated at 91 weeks (5/30 versus 0/15 in controls) (Kari et al., 1993). 

5846 

5847 4.6.3.3. Mode of Action for Neurological Effects 

5848 Results from studies of acutely exposed human subjects indicate that mild 

5849 neurobehavioral deficits may occur at air concentrations >200 ppm with 4–8 hours of exposure 

5850 (Bos et al., 2006; ACGIH, 2001; ATSDR, 2000; Cherry et al., 1983; Putz et al., 1979; Gamberale 

5851 et al., 1975; Winneke, 1974). Acute high-dose exposures also resulted in gross neurological 

5852 impairments in several laboratory species (Haun et al., 1972, 1971; Heppel et al., 1944). 

5853 Exposure of F344 rats to concentrations up to 2,000 ppm 6 hours/day for 13 weeks produced no 

5854 effects on an observational battery, hind-limb grip strength, a battery of evoked potentials, or 

5855 histology of the brain, spinal cord, or peripheral nerves (Mattsson et al., 1990). However, oral 
192

5856 exposures have been shown to alter autonomic, neuromuscular, and sensorimotor functions have 

5857 been observed in F344 rats exposed to gavage doses ≥337 mg/kg-day for 14 days (Moser et al., 

5858 1995). 

5859 Dichloromethane may be metabolized by the CYP2E1 enzyme to CO (Guengerich, 1997; 

5860 Hashmi et al., 1994; Gargas et al., 1986). Many of the acute human exposure studies evaluated if 

5861 CO was the primary metabolite responsible for producing the CNS depressant effects observed 

5862 during dichloromethane exposure. Overall, at lower exposures and acute durations, it appears 

5863 that CO is the primary mediator of the neurobehavioral effects. Putz et al. (1979) demonstrated 

5864 that similar neurobehavioral deficits were present when an equivalent COHb blood level (and 

5865 CO exposure) was achieved between CO and dichloromethane exposures. Incidentally, after a 

5866 longer duration, neurobehavioral deficits are more pronounced with dichloromethane exposure in 

5867 comparison to CO exposure alone. This additional increase in the CNS depressive effects is 

5868 most likely due to the saturation of the CYP2E1 metabolic pathway. In humans, saturation of the 

5869 CYP2E1 metabolic pathway was seen at approximately 400–500 ppm after a 1-hour exposure 

5870 (Ott et al., 1983e). CYP2E1 pathway saturation with dichloromethane has also been noted in 

5871 hamsters (Burek et al., 1984) and in rats (McKenna et al, 1982; Nitschke et al., 1988a). It is 

5872 highly probable that initially CYP2E1 is metabolizing dichloromethane to CO, which results in 

5873 the neurological effects. However, at higher concentrations (greater than 500 ppm) and for 

5874 longer durations (greater than 3 hours), the CYP2E1 pathway is most likely saturated. As a 

5875 result, either the remaining dichloromethane could be metabolized by the GST pathway or the 

5876 parent compound is producing the effects itself. 

5877 Once the CYP2E1 enzyme is saturated, it is unknown whether dichloromethane or a 

5878 GST-T1 pathway metabolite (e.g., formaldehyde) mediates the resulting neurological effects. 

5879 Based on the available literature on other solvents, such as toluene and perchloroethylene (for a 

5880 review see Bowen et al. [2006]), it can be hypothesized that once the CYP2E1 enzyme is 

5881 saturated dichloromethane or a GST metabolite may interact directly with excitatory and 

5882 inhibitory receptors, such as the NMDA, GABA, dopamine, and serotonin receptors among other 

5883 targets, to produce the resulting neurobehavioral effects. This hypothesis is supported by the 

5884 evidence that changes in relation to dichloromethane exposures in glutamate, GABA, dopamine, 

5885 serotonin, acetylcholine, and other neurotransmitters are found in the brain (Kanada et al., 1994; 

5886 Briving et al., 1986; Fuxe et al., 1984). Additionally, several neurobehavioral effects, such as 

5887 decreased spontaneous motor activity, deficits in learning and memory, and deficits in FOB 

5888 parameters are similar to other more characterized solvents such as toluene. However, more 

5889 comprehensive studies specifically designed to determine the mode of action for 

5890 dichloromethane-induced impairment of neurological functions have not been conducted. 

5891 

5892 4.6.3.4. Mode of Action for Reproductive and Developmental Effects
193

5893 No significant effects on reproductive performance variables were found in a two5894 generation reproduction assay with F344 rats exposed to concentrations as high as 1,500 ppm 

5895 (Nitschke et al., 1988b), and no effects were seen on most of the measures of reproductive 

5896 performance examined in a study of male mice exposed to 200 ppm for 2 hours/day for 6 weeks 

5897 before mating to nonexposed females (Raje et al., 1988). In the mouse study, fertility index 

5898 (number of females impregnated divided by total number of females mated times 100) was 

5899 reduced in the 150 and 200 ppm groups (Raje et al., 1988), but the statistical significance of this 

5900 effect varied considerably depending on the statistical test used in the analysis. Mechanistic 

5901 studies of dichloromethane or its metabolites that would provide mode of action information on 

5902 reproductive effects in the male are not available. 

5903 The mode of action for developmental effects can be hypothesized to involve the 

5904 CYP2E1 pathway and, specifically, the production of CO. CO is a known developmental 

5905 neurotoxicant. Demonstrated effects include neurobehavioral deficits and neurochemical 

5906 changes (Giustino et al., 1999; Cagiano et al., 1998; De Salvia et al., 1995; Fechter, 1987). In 

5907 addition, placental transfer of dichloromethane has been demonstrated with inhalation exposure 

5908 (Withey and Karpinski, 1985; Anders and Sunram, 1982). Pups exposed in utero to high 

5909 concentrations of dichloromethane (4,500 ppm) demonstrated neurobehavioral-related changes 

5910 in comparison to air-exposed animals (Bornschein et al., 1980). This observed effect coupled 

5911 with the known developmental neurotoxicological effects produced by CO suggests that the 

5912 CYP2E1 metabolic pathway is involved in producing observed and suspected 

5913 neurodevelopmental effects. In humans, CYP2E1 activity in the brain occurs earlier in gestation 

5914 than it does in the liver, with activity in the brain seen in the first trimester (Johnsrud et al., 2003; 

5915 Brzezinski et al., 1999). Thus, the direct effects of dichloromethane in fetal circulation, as well 

5916 as the effects of CO and the effects of the CYP2E1-related metabolism in the fetal liver and the 

5917 fetal brain, may be relevant to the risk of developmental effects in humans. Mechanistic studies 

5918 of dichloromethane or its metabolites that would provide mode of action information on other 

5919 noted developmental effects such as delayed ossification (Schwetz et al., 1975) are not available. 

5920 

5921 4.6.3.5. Mode of Action for Immunotoxicity

5922 Evidence of a localized immunosuppressive effect in the lung resulting from inhalation 

5923 dichloromethane exposure was seen in an acute exposure (3 hours, 100 ppm) study in CD-1 mice 

5924 (Aranyi et al., 1986). The lung infectivity assay used in this study examined response to 

5925 bacterial challenges (i.e., risk of streptococcal-pneumonia-related mortality and clearance of 

5926 Klebsiella bacteria). The innate immune response plays an important role in limiting the initial 

5927 lung burden of bacteria through the activity of macrophages, neutrophils, and dendritic cells, and 

5928 alveolar macrophages are particularly important in the response to respiratory infections 

5929 (Marriott and Dockrell, 2007). The adaptive response develops from several days up to several 

5930 weeks following infection, so that an effective immune response in a lung infectivity assay 
194

5931 requires multiple immune mechanisms and in particular cooperation of macrophages, 

5932 neutrophils, and T cells along with the appropriate cytokines (Selgrade and Gilmour, 2006). 

5933 Although immunosuppression in the Streptococcal and Klebsiella infectivity models has been 

5934 reported in the acute exposure scenarios tested in Aranyi et al. (1986), mechanistic studies of 

5935 dichloromethane or its metabolites that would provide mode of action information on the 

5936 immune system cells or function have not been performed. 

5937 

5938 4.7. EVALUATION OF CARCINOGENICITY 

5939 4.7.1. Summary of Overall Weight of Evidence 

5940 Following U.S. EPA (2005a) Guidelines for Carcinogen Risk Assessment, 

5941 dichloromethane is “likely to be carcinogenic in humans” by the inhalation and oral routes of 

5942 exposure, based predominantly on evidence of carcinogenicity at two sites in 2-year bioassays in 

5943 B6C3F1 mice (liver and lung tumors with inhalation exposure in both sexes, liver tumors with 

5944 drinking water exposure in males only). In addition, evidence of a trend for increased risk of 

5945 liver tumors (described as neoplastic nodule or hepatocellular carcinoma) was seen in female 

5946 F344 rats exposed via drinking water (p < 0.01) (Serota et al., 1986a) or inhalation (p = 0.08) 

5947 (NTP, 1986). However, the potential malignant characterization of the nodules was not 

5948 described, and no trend was seen in the data limited to hepatocellular carcinomas. Additional 

5949 evidence of the tumorigenic potential of dichloromethane comes from the observation of an 

5950 increase in benign mammary tumors following inhalation exposure (NTP, 1986; Burek et al., 

5951 1986b; Nitschke et al. 1988a). An inhalation study (exposures of 0, 50, 200, and 500 ppm) also 

5952 reported the presence of another relatively rare tumor in rats, astrocytoma or glioma (mixed glial 

5953 cell) tumors (Nitschke et al., 1988a). This collection of studies in the rat does not provide 

5954 evidence for a carcinogenic response that is as strong as that seen in the mouse. Taken together, 

5955 however, the rat data provide supporting evidence of carcinogenicity. Studies in humans found 

5956 some evidence linking occupational exposure to dichloromethane and increased risk for some 

5957 specific cancers, including brain cancer (Hearne and Pifer, 1999; Tomenson et al., 1997; 

5958 Heineman et al., 1994) and liver cancer (Lanes et al., 1993, 1990). 

5959 The proposed mode of action for dichloromethane-induced liver tumors is through a 

5960 mutagenic mode of carcinogenic action. Mode of action data indicate that dichloromethane5961 induced DNA damage in cancer target tissues of mice involves DNA-reactive metabolites 

5962 produced via a metabolic pathway initially catalyzed by GST-T1. Evidence of mutagenicity 

5963 includes in vitro bacterial and mammalian assays as well as in vivo mammalian system assays, 

5964 although mutational events in critical genes (tumor suppressor genes, oncogenes) leading to 

5965 tumor initiation and tumor promotion have not been established. This metabolic pathway has 

5966 been found in human tissues, albeit at lower activities than in mouse tissues; therefore, the cancer 

5967 results in animals are considered relevant to humans. 

5968 
195

5969 4.7.2. Synthesis of Human, Animal, and Other Supporting Evidence 

5970 Section 4.1.2 reviewed the results, strengths, and limitations of epidemiological research 

5971 of dichloromethane and cancer, including cohort and case-control studies. The available 

5972 epidemiologic studies provide some evidence of an association between dichloromethane and 

5973 brain cancer and liver cancer, but the available data are limited. 

5974 Two small cohort studies with relatively good exposure metrics and relatively long 

5975 follow-up periods (mean over 25 years) reported an increased risk of brain cancer, with SMRs of 

5976 1.45 (95% CI 0.40–3.72) in Tomenson et al. (1997) and 2.2 (95% CI 0.79–4.69) in Cohort 1 of 

5977 Hearne and Pifer (1999). Cohort 1 is an inception cohort, following workers from the beginning 

5978 of their employment, which is methodologically more robust than Cohort 2, which only included 

5979 workers who were working between 1964 and 1970. These observations are supported by the 

5980 data from a case-control study of brain cancer that reported relatively strong trends with 

5981 increasing probability, duration, and intensity measures of exposure but not with a cumulative 

5982 exposure measure. This difference could reflect a relatively more valid measure of relevant 

5983 exposures in the brain from the intensity measure, as suggested by the study in rats reported by 

5984 Savolainen et al. (1981) in which dichloromethane levels in the brain were much higher with a 

5985 higher intensity exposure scenario compared with a constant exposure period with an equivalent 

5986 TWA (see section 3.2). A statistically significant increased incidence of brain or CNS tumors 

5987 has not been observed in any of the animal cancer bioassays, but a 2-year study using relatively 

5988 low exposure levels (0, 50, 200, and 500 ppm) in Sprague-Dawley rats observed a total of six 

5989 astrocytoma or glioma (mixed glial cell) tumors in the exposed groups (in females, the incidence 

5990 was 0, 0, 0, and 2 in the 0, 50, 200, and 500 ppm exposure groups; in males, the incidence was 0, 

5991 1, 2, and 1 in the 0, 50, 200, and 500 ppm exposure groups; sample size of each group was 

5992 70 rats). These tumors are exceedingly rare in rats, and there are few examples of statistically 

5993 significant trends in animal bioassays (Sills et al., 1999). These cancers were not seen in two 

5994 other studies in rats, both involving higher doses (1,000–4,000 ppm) (NTP, 1986; Burek et al., 

5995 1984), or in a high dose (2,000–4,000 ppm) study in mice (NTP, 1986). 

5996 With respect to epidemiologic studies of liver and biliary duct cancer, the highest 

5997 exposure cohort, based in the Rock Hill, South Carolina, triacetate fiber production plant, 

5998 suggested an increased risk of liver cancer, with an SMR of 2.98 (95% CI 0.81, 7.63) in the latest 

5999 study update (Lanes et al., 1993). This observation was based on four cases; an earlier analysis 

6000 in this cohort reported an SMR of 5.75 (95% CI 1.82, 13.8), based on these same four cases but 

6001 with a shorter follow-up period (and thus a lower number of expected cases) (Lanes et al., 1990). 

6002 No other cohort study has reported an increased risk of liver cancer mortality, although it should 

6003 be noted that there is no other inception cohort study of a population with exposure levels similar 

6004 to those of the Rock Hill plant, and no data from a case-control study of liver cancer are 

6005 available pertaining to dichloromethane exposure. 
196

6006 The primary limitation of all of the available dichloromethane cohort studies is the 

6007 limited statistical power for the estimation of effects relating to relatively rare cancers (such as 

6008 brain cancer, liver cancer, and leukemia). Limitations with respect to studies of other cancers 

6009 can also be noted. With respect to breast cancer, the only cohort that included a significant 

6010 percentage of women had limited exposure information (analysis was based on a dichotomous 

6011 exposure variable) and co-exposure to other solvents (Blair et al., 1998). The only breast cancer 

6012 case-control study available used death certificate data to classify disease and occupational 

6013 exposure (Cantor et al., 1995), which is likely to result in significant misclassification; exposure 

6014 misclassification in particular would be expected to result in an attenuated measure of 

6015 association (Rothman and Greenland, 1998). No studies of adult leukemia and dichloromethane 

6016 exposure and only one study of childhood leukemia (acute lymphoblastic leukemia) in relation to 

6017 maternal occupational dichloromethane exposure were found. Thus, EPA views the 

6018 epidemiologic data pertaining to breast cancer and leukemia as inadequate to assess carcinogenic 

6019 potential. 

6020 In addition to the epidemiologic studies, several dichloromethane cancer bioassays in 

6021 animals are available. In the only oral exposure cancer bioassay involving lifetime exposure, 

6022 increases in incidence of liver adenomas and carcinomas were observed in male (trend p-value = 

6023 0.058) but not female B6C3F1 mice exposed for 2 years (Table 4-38) (Serota et al., 1986b; 

6024 Hazelton Laboratories, 1983). The authors concluded that these increases were “within the 

6025 normal fluctuation of this type of tumor incidence,” noting that there was no dose-related trend 

6026 and that most of the individual group paired tests were not statistically significant after use of a 

6027 Bonferroni correction factor. [The trend p-value and pairwise test p-values were not given in the 

6028 Serota et al. (1986b) paper but can be found in the full report (Hazelton Laboratories, 1983)]. 

6029 However, the trend p-value for these results is of borderline statistical significance and it may not 

6030 be reasonable to apply a correction for multiple comparisons given the lack of independence of 

6031 the groups and given a specific focus on the liver as a target organ. In Syrian golden hamsters 

6032 exposed to 500, 1,500, or 3,500 ppm for 2 years, no statistically significantly increased 

6033 incidences of tumors were found in any tissues (Burek et al., 1984). 

6034 
197

6035 

Table 4-38. Incidence of liver tumors in male B6C3F1 mice exposed 

to dichloromethane in a 2-year oral exposure (drinking water) studya

Estimated mean intake 

(mg/kg-day)a

Controls 

0 61 124 177 234 

Number of male miceb125c200 100 99 125 

Trend 

p-valued

Number of cancers (%) 

Hepatocellular adenoma or 

carcinoma

 Mortality-adjusted percente

 Mortality-adjusted p-valuee

 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 

aTarget doses were 60, 125, 185, and 250 mg/kg-day from the lowest dose group (excluding controls) to 

the highest dose group, respectively. 

bNo significant increases in females were found, but incidence data were not reported. 

cTwo control groups combined. The incidence in control groups 1 and 2 were 20 and 23%, respectively. 

dCochran-Armitage trend test (source: Hazelton Laboratories [1983]). 

eMortality-adjusted percent calculated, based on number at risk, using Kaplan-Meier estimation, taking into 

account mortality losses; p-value for comparison with control group, using asymptotic normal test 

(source: Hazelton Laboratories [1983]). 

Sources: Serota et al. (1986b); Hazelton Laboratories (1983). 

6036 

6037 

6038 In a similar study in F344 rats (Serota et al., 1986a), no increased incidence of liver 

6039 tumors was seen in male rats, and the pattern in female rats was characterized by a jagged 

6040 stepped pattern of increasing incidence of hepatocellular carcinoma or neoplastic nodule 

6041 (Table 4-39). Information was not provided which would allow characterization of the nodules 

6042 as benign or malignant. Statistically significant increases in incidences were observed in the 50 

6043 and 250 mg/kg-day groups (incidence rates of 0, 3, 10, 3, and 14%, respectively, for the 0, 5, 50, 

6044 125, and 250 mg/kg-day groups) and in the group exposed to 250 mg/kg-day for 78 weeks 

6045 followed by a 26-week period of no exposure (incidence rate 10%). A similar pattern, but with 

6046 more sparse data, was seen for hepatocellular carcinomas, with 2 incidences in the 50 mg/kg-day 

6047 and 2 in the 250 mg/kg-day groups. The authors concluded that dichloromethane exposure did 

6048 not result in an increased incidence of liver tumors, because the increase was based on a low rate 

6049 (0%) in the controls and because of a lack of monotonicity. 
198

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-valueb

250 with 

recoveryc

Males 

Estimated mean intake (mg/kg-day) 0 6 52 125 235 232 

n per groupd 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 58 136 263 239 

n per groupd 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)ep < 0.01 2 (10)e

aTwo control groups combined. 

bCochran-Armitage trend test was used for trend test of liver foci/areas of alteration. For tumor mortalityunadjusted analyses, a Cochran-Armitage trend test was used, and, for tumor mortality-adjusted analyses, a tumor 

prevalence analytic method by Dinse and Lagakos (1982) was used. Similar results were seen in these two 

analyses. 

cRecovery group was exposed for 78 weeks and then had a 26-week period without dichloromethane exposure; 

n = 17 for neoplastic lesions. 

d

n available at terminal sacrifice; starting with 135 controls (combining both control groups) and 85 per sex per 

dose group except recovery group (n = 25); subtracted 5, 10, and 20 per group (except for recovery group) 

sacrificed at 25, 52, and 78 weeks, respectively, and subtracted unscheduled deaths, which ranged from 5 to 19 per 

group. 

e

Significantly (p < 0.05) different from controls with Fisher’s exact test, mortality-unadjusted and mortalityadjusted analyses. 

Source: Serota et al. (1986a). 

6051 

6052 

6053 Another oral (gavage) exposure study in Sprague-Dawley rats and in Swiss mice provides 

6054 limited data concerning cancer incidence because the study was terminated early (at 64 weeks) 

6055 due to high treatment-related mortality (Maltoni et al., 1988). Exposure groups included controls 

6056 (olive oil), 100, or 500 mg/kg-day 4–5 days/week. High-dose female rats showed an increased 

6057 incidence of malignant mammary tumors, mainly adenocarcinomas (8, 6, and 18% in the control, 

6058 100, and 500 mg/kg dose groups, respectively), but the increase was not statistically significant. 

6059 Data were not provided to allow an analysis accounting for differing mortality rates. A dose6060 related increase, although not statistically significant, in pulmonary adenomas was observed in 
199

6061 male mice (5, 12, and 18% in control, 100, and 500 mg/kg-day groups, respectively). When 

6062 mortality was taken into account, high-dose male mice that died in the period ranging from 52 to 

6063 78 weeks were reported to show a statistically significantly (p < 0.05) elevated incidence for 

6064 pulmonary tumors (1/14, 4/21, and 7/24 in control, 100, and 500 mg/kg-day groups, 

6065 respectively). Details of this analysis were not provided. EPA applied a Fisher’s exact test to 

6066 these incidences and determined a p-value of 0.11 for the comparison of the 500 mg/kg-day 

6067 group (7/24) to the controls (1/14). 

6068 As discussed in section 4.2, repeated inhalation exposure to concentrations of 2,000 or 

6069 4,000 ppm dichloromethane produced increased incidences of lung and liver tumors in B6C3F1

6070 mice (Mennear et al., 1988; NTP, 1986). The incidence of mortality-adjusted liver tumors across 

6071 dose groups (0, 2,000, and 4,000 ppm) increased from 48 to 67 and 93%, respectively, in male 

6072 mice (trend p-value = 0.013) and from 10 to 48 and 100% female mice (trend p-values <0.001) 

6073 (Table 4-40). For lung tumors, the mortality-adjusted incidence was 12, 74, and 100% in males 

6074 and 11, 83, and 100% in females in the 0, 2,000, and 4,000 ppm groups, respectively (trend p6075 values <0.001). Elevated incidences of lung and liver tumors in B6C3F1 mice were observed 

6076 with 52 weeks of exposure to 2,000 ppm, and lung tumors were also elevated by week 104 in 

6077 mice exposed for only 26 weeks to 2,000 ppm, followed by 78 weeks without exposure 

6078 (Maronpot et al., 1995; Kari et al., 1993). 
200

6079 6080 

Table 4-40. Incidences of selected neoplastic lesions in B6C3F1 mice exposed 

to dichloromethane by inhalation (6 hours/day, 5 days/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-valued

Males 

Liver—hepatocellular adenoma or 

carcinoma 

22 (44) (48) 24 (49) (67) 33e(67) (93) 0.013 

Lung—bronchoalveolar adenoma or 

carcinoma 

 5 (10) (12) 27e(54) (74) 40e(80) (100) <0.001 

Females 

Liver— hepatocellular adenoma or 

carcinoma 

 3 (6) (10) 16e(33) (48) 40e(83) (100) <0.001 

Lung—bronchoalveolar adenoma or 

carcinoma 

 3 (6) (11) 30e(63) (83) 41e(85) (100) <0.001 

a

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 examined 

microscopically per group; for male mice, 49 livers were examined in the 2,000 and 4,000 ppm groups; for female 

mice, 48 liver and lungs were examined. For comparison, incidence in historical controls reported in NTP (1986) 

were 28% for male liver tumors, 31% for male lung tumors, 5% for female liver tumors, and 10% for female lung 

tumors. 

cMortality-adjusted percentage. 

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

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

Sources: Mennear et al. (1988); NTP (1986). 

6081 

6082 

6083 Liver tumors are relatively rare in F344 rats, and a moderate trend of increasing incidence 

6084 of what was described as neoplastic nodules or hepatocellular carcinoma was seen in the females 

6085 (trend p-value = 0.08) but not the males in the NTP (1986) study (Table 4-41). As with the rat 

6086 oral exposure study by Serota et al., 91986a), these nodules were not characterized as benign or 

6087 malignant. There was no evidence of an increasing trend in incidence when hepatocellular 

6088 carcinomas only were considered. 

6089 Female F344 rats exposed by inhalation to 2,000 or 4,000 ppm showed significantly 

6090 increased incidences of benign mammary tumors (adenomas or fibroadenomas) (Table 4-41); the 

6091 number of benign mammary tumors per animal also increased with dichloromethane exposure in 

6092 studies in Sprague-Dawley rats at levels of 50–500 ppm (Nitschke et al., 1988a) and 500–3,500 

6093 ppm (Burek et al., 1984) (Table 4-42). Male rats in two of these studies (Nitscke et al., 1988a ; 

6094 NTP, 1986) also exhibited a low rate of sarcoma or fibrosarcoma in mammary gland or 

6095 subcutaneous tissue around the mammary gland. 

6096 
201

Table 4-41. Incidences of selected neoplastic lesions in F344/N rats exposed to dichloromethane by (6 hours/day, 5 days/week) for 2 years 

Exposure (ppm)a

 0 (Controls) 1,000 2,000 

Sex and neoplastic lesion n (%)b (%)c n (%)b (%)c n (%)b (%)c n 

Males 

Liver—Neoplastic nodule or hepatocellular carcinoma 2 (4) (10) 3 (6) (13) 4 (8) (19) 1 

Liver—hepatocellular carcinoma 2 (4) (10) 1 (2) (4) 2 (4) (10) 1 

 Lung—bronchoalveolar adenoma or carcinoma 1 1 (2) 2 (4) 1 

Mammary gland 

Adenoma, adenocarcinoma, or carcinoma Subcutaneous tissue fibroma or sarcoma Fibroadenoma 

 Mammary gland or subcutaneous tissue adenoma, 

fibroadenoma, fibroma, or sarcoma 

 0  1  0  1

(0)(2)(0)(2)

(6)(0)(6)

 0 

 1  0  1 

(0)(2)(0)(2)

(6)(0)(6)

 0 

 2  2  4

(0)(4)(4)(8)

 (9)(12)(21)

 1  5  1  9d

Females 

Liver—neoplastic nodule or hepatocellular carcinoma 2 (4) (7) 1 (2) (2) 4 (8) (14) 5 

Liver—hepatocellular carcinoma 0 (0) (0) 0 (0) (0) 1 (2) (4) 0 

 Lung—bronchoalveolar adenoma or carcinoma 1 (2) 1 (2) 0 (0) 0 

Mammary gland 

Adenocarcinoma or carcinoma Adenoma, adenocarcinoma, or carcinoma Fibroadenoma 

 Mammary gland adenoma, fibroadenoma, or 

adenocarcinoma 

 1 

 1  5  6 

(2)(2)

(10)(12)

(16)(18)

 2 

 2 11d13 

(4)(4)

(22)(26)

(41)(44)

 2 

 2 13d14d

(4)(4)

(26)(28)

(44)(45)

 0  1 22d23e

a1,000 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 examined microscopically per group; fo

were examined in the 2,000 and 4,000 ppm groups; for females, only 48 liver and lungs and 49 mammary glands were microscopicall

and 4,000 ppm groups. For comparison, incidence in historical controls reported in NTP (1986) were 1% for female liver tumors and 

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). 

Sources: Mennear et al. (1988); NTP (1986). 

6097 6098 
202

Table 4-42. Incidences of mammary gland tumors in two studies of male and 

female Sprague-Dawley rats exposed to dichloromethane by inhalation 

(6 hours/day, 5 days/week) for 2 years 

Exposure (ppm)a

 

Study, lesion 

Controls

0 50 200 500 

Late 

500b

Early 

500b 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 sarcomad

 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 tumorse 52 (74) 58 (83) 61(87)f55 (79) 23 (92) 23 (92) 

Number of benign tumors per tumorbearing rate2.0 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 

ratg

 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 165 218 245 287 

Number of tumors per tumor-bearing 

ratf

 2.1 2.7 3.1 3.5 

a

50 ppm = 174 mg/m3, 200 ppm = 695 mg/m3, 500 ppm = 1,737 mg/m3, 1,500 ppm = 5,210 mg/m3, 3,500 ppm = 

12,158 mg/m3

. 

bLate 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. 

cNo data for this exposure level in this study.

dEPA summed across these tumor types, assuming no overlap.

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 exact test (Nitschke et al., 1988a). 

gCalculated by EPA. 

Sources: Nitschke et al. (1988a); Burek et al. (1984b). 

6099 
203

6100 Supporting evidence for the carcinogenicity of dichloromethane comes from the results 

6101 of genotoxicity and mode of action studies discussed in section 4.5. A mutagenic mode of 

6102 carcinogenic action for dichloromethane involves metabolic activation by GST, as evidenced by 

6103 several observations, including the enhancement of dichloromethane mutagenic activity in 

6104 normally unresponsive S. typhimurium strain TA1535 after it is transfected with the gene for rat 

6105 GST-T1 (DeMarini et al., 1997; Thier et al., 1993); increased HPRT gene mutations and DNA 

6106 damage (DNA SSBs) in CHO cells when they are incubated with dichloromethane in the 

6107 presence of mouse liver cytosol preparations rich in GST enzymatic activities (Graves and 

6108 Green, 1996; Graves et al., 1996, 1994b); the detection of DNA damage (DNA SSBs) in liver 

6109 and lung tissue of B6C3F1 mice immediately following 6-hour inhalation exposure to 

6110 dichloromethane (2,000–8,000 ppm); and a suppression of the DNA damage when mice were 

6111 pretreated with buthionine sulphoximine, a GSH depletor (Graves et al., 1995). 

6112 Additional data from several studies indicate that dichloromethane genotoxicity is 

6113 expressed in cancer target tissues in mice following in vivo exposure. Increased sister chromatid 

6114 exchanges were observed in lung cells of B6C3F1 mice after 90 days of inhalation exposure to 

6115 2,000 ppm or 10 days of exposure to 4,000 or 8,000 ppm (Allen et al., 1990). DNA damage 

6116 (comet assay) was detected in liver and lung tissue (but not stomach, kidney, brain, or bone 

6117 marrow) 24 hours after oral administration of 1,720 mg/kg dichloromethane to CD-1 mice 

6118 (Sasaki et al., 1998). DNA-protein cross-links were observed in the liver of B6C3F1 mice but 

6119 not hamsters, following inhalation exposure to concentrations ranging from 500 to 4,000 ppm 

6120 6 hours/day for 3 days (Casanova et al., 1996, 1992). Much less is known about genotoxicity in 

6121 the liver in rats. Studies of single-strand DNA breaks in rat hepatocytes or liver homogenate 

6122 were negative, with inhalation exposures up to 5,000 ppm for 3 hours (Graves et al., 1995, 

6123 1994b), but positive results were seen in a high-dose gavage study (1,275 mg/kg) (Kitchin and 

6124 Brown, 1989). Few other specific types of genotoxicity endpoints (e.g., sister chromatid 

6125 exchange, DNA-protein cross-links) have been studied in the rat liver. 

6126 Since there are limited data on mutagenic events following oral exposure, EPA conducted 

6127 a pharmacokinetic analysis to evaluate how comparable the internal doses to the liver in the oral 

6128 bioassay (Serota et al., 1986b; Hazelton Laboratories, 1983) were to the internal doses to the 

6129 liver in the inhalation bioassay (Mennear et al., 1988; NTP, 1986). The PBTK model of Marino 

6130 et al. (2006) predicted that the average daily amount of dichloromethane metabolized via GST 

6131 per liter of liver was about 14-fold lower in mice exposed to the highest dose of 244 mg/kg-day 

6132 in the drinking water bioassay than in mice exposed to the lowest inhalation exposure of 2,000 

6133 ppm, inducing liver tumors (Table 4-43). Thus, the lower incidence of liver tumors induced by 

6134 oral doses of 244 mg/kg-day, compared with the higher incidence induced by inhalation 

6135 exposure to 2,000 ppm, is consistent with the predicted lower liver dose of GST metabolites (and 

6136 hence lower probability of DNA modification) with oral exposure. 

6137 
204

6138 

Table 4-43. Comparison of internal dose metrics in

inhalation and oral exposure scenarios, in male mice 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 of 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 one gender. 

6139 

6140 

6141 

6142 4.7.3. Mode of Action Information 

6143 4.7.3.1. Hypothesized Mode of Action 

6144 The hypothesized mode of action for dichloromethane-induced tumors is through a 

6145 mutagenic mode of carcinogenic action. Specifically, the data indicate that dichloromethane is 

6146 metabolized by GST to reactive metabolites that induce mutations in DNA leading to 

6147 carcinogenicity. Much of the experimental mode of action research has focused on the liver and 

6148 lung, the sites of tumor formation in chronic bioassays (Mennear et al., 1988; NTP, 1986; Serota 

6149 et al., 1986b, Hazleton Laboratories, 1983). The mode of action is potentially relevant to other 

6150 sites, particularly those in which GST-T1 is expressed, such as mammary tissue (Lehmann and 

6151 Wagner, 2008) and the brain (Juronen et al., 1996). 

6152 Support for the importance of GST in the hypothesized mutagenic mode of action has 

6153 been demonstrated in in vitro bacterial and mammalian assays as well as in vivo mammalian 

6154 system assays. Dichloromethane is consistently mutagenic in S. typhimurium strains with GST 

6155 capability, but did not produce mutagenic effects in non-GST S. typhimurium strains 

6156 (summarized in Section 4.5.1.1 and Table 4-29). In vitro mammalian cell studies (see Table 4-

6157 30) have consistently demonstrated genotoxic effects in CHO cell lines when a mouse liver 

6158 cytosol fraction was exogenously added and in mouse Clara cells; positive responses were seen 

6159 in studies measuring DNA-protein crosslinks, HPRT mutation analysis, and DNA SSBs. Other 
205

6160 studies have demonstrated DNA adducts with dichloromethane exposure in calf thymus DNA in 

6161 the presence of bacterial GST DM11. Negative results were seen in most of the other in vitro cell 

6162 studies using rat hepatocytes or CHO cells without mouse liver cytosol incubation. These 

6163 studies were conducted in cell lines where GST activity is considerably lower than in mouse cell 

6164 lines and therefore these results are not unexpected. 

6165 In studies with human cell lines or isolated cells, positive results were reported for sister 

6166 chromatid exchanges, chromosomal aberrations, and in the micronucleus test. In vivo studies in 

6167 mice (Section 4.5.1.2 and Table 4-32) consistently showed genotoxic effects following 

6168 dichloromethane exposure in the liver and lung, where tumors are observed. Other organs in the 

6169 mouse were evaluated and mutagenic changes were not consistenly observed. The specificity of 

6170 the observed effects support the hypothesized mode of action since these positive mutagenic 

6171 responses are seen in organs where tumor formation occurs (i.e., liver and lung) rather than in 

6172 areas that were not the site of tumors in the mouse bioassays (e.g., stomach, bladder, kidney). In 

6173 vivo genotoxicity studies in rats and hamsters (the other test systems used, see Table 4-33) were 

6174 predominantly non-positive. However, rats and hamsters have considerably lower GST activity 

6175 than the mouse and may be less sensitive to dichloromethane-induced genotoxic effects. 

6176 In vivo binding of S-(chloromethyl)glutathione, dichloromethane’s reactive GST 

6177 metabolite,to DNA was not demonstrated in one study in rats and mice using a relatively low 

6178 dose (5 mg/kg). The reactivity of the postulated DNA-reactive species and the instability of the 

6179 derived adducts presents considerable challenges to the ability to provide direct evidence of 

6180 adduct formation. Thus this lack of in vivo evidence of S-(chloromethyl)glutathione binding to 

6181 DNA does not in itself represent a basis for invalidating the proposed mode of action. 

6182 

6183 4.7.3.1.1. Experimental support for the hypothesized mode of action 

6184 Strength, consistency, and specificity of association

6185 It is hypothesized that mutagenic events lead to the development of liver and lung tumors 

6186 following dichloromethane exposure. Several observations from experimental studies support the 

6187 mutagenicity of dichloromethane and the key role of GST metabolism and the formation of 

6188 DNA-reactive GST-pathway metabolites. The GST pathway produces two metabolites of 

6189 dichloromethane, S-(chloromethyl)glutathione and formaldehyde, which are potentially reactive 

6190 with DNA and other cell macromolecules. Enhanced dichloromethane genotoxicity in bacterial 

6191 and mammalian in vitro assays with the introduction of GST metabolic capacity provides support 

6192 that GST metabolism and metabolites are involved (DeMarini et al., 1997; Graves and Green, 

6193 1996; Graves et al., 1996, 1995, 1994b; Thier et al., 1993). 

6194 In bacterial strains where GST activity was not present (e.g., TA1535, TA1538), 

6195 mutagenic effects were not reported following dichloromethane exposure (Gocke et al., 1981; 

6196 Osterman-Golkar et al., 1983; Simula et al., 1993; Oda et al., 1996). Further tests of GST6197 dependent mutagenicity were evaluated by transfecting GST into non-GST bacterial strains or 
206

6198 decreasing GST activity in GST bacterial strains (e.g. TA100). When GSTT1-1 was cloned into 

6199 bacterial strain TA1535, dichloromethane treatment resulted in reverse mutations in this new 

GST+6200 TA1535 strain and these mutations were independent of rat S9 metabolic activation (Their 

6201 et al., 1993; Pegram et al., 1997; DeMarini et al., 1997). Similarly, TA100/NG-11, a bacterial 

6202 strain with decreased GST activity in comparison to the wild-type TA100 strain, showed 

6203 significantly decreased mutagenicity (reverse mutations) following dichloromethane treatment 

6204 (Graves et al., 1994a). 

6205 In vitro mammalian genotoxicity studies also support the importance of the GST pathway 

6206 in relation to the positive effects observed following dichloromethane exposure. Positive results 

6207 in the in vitro assays were limited to experiments with the presence of GST in the cell system. 

6208 When mouse liver cytosol was added to hamster cell lines, dichloromethane induced DNA6209 protein crosslinks, DNA SSBs, and HPRT gene mutations (Graves et al., 1994b; Graves et al., 

6210 1996; Graves and Green, 1996). Additionally, in mouse Clara cells (GST is localized in the lung 

6211 cells of mice), DNA SSBs were reported following dichloromethane treatment and the extent of 

6212 DNA damage was significantly decreased when the cells were pretreated with a glutathione 

6213 depletor (Graves et al., 1995). Other studies evaluating similar genotoxic endpoints in rat or 

6214 CHO cells without modification of the low GST activity in the test system generally reported no 

6215 evidence of genotoxic events (Graves et al., 1995; Andrae and Wolff, 1983; Jongen et al., 1981; 

6216 Garrett and Lewis, 1983; Thilagar and Kumaroo, 1983). A study evaluating the genotoxic 

6217 effects of dichloromethane (up to 6 mM) in freshly isolated mouse, rat, hamster, and human 

6218 hepatocytes provides additional supporting evidence of the influence of GST activity on 

6219 mutagenicity (Casanova et al., 1997). Positive results were only observed in hepatocytes from 

6220 B6C3F1 mice; the interspecies variability in effects correlated proportionally with the enhanced 

6221 GST metabolic capacity in mice (Reitz et al., 1989). In studies with human cell lines or isolated 

6222 cells, positive results were reported for sister chromatid exchanges, chromosomal aberrations, 

6223 DNA damage, and in the micronucleus test. Negative results were obtained with human cells in 

6224 unscheduled DNA synthesis assays (Perocco and Prodi, 1981; Jongen et al., 1981) and 

6225 dichloromethane was not demonstrated to be genotoxic in studies of human hepatocytes (Graves 

6226 et al., 1995; Casanova et al., 1997). 

6227 Two of three in vivo genotoxicity studies in insects reported positive results. 

6228 Genotoxicity was observed in Drosophila for the gene mutation assay (Gocke et al., 1981) and 

6229 the somatic assay (Rodriguez-Arnaiz, 1998) when dichloromethane was administered through 

6230 the food. When Drosophila were exposed to dichloromethane via inhalation, genotoxic effects 

6231 were negative as measured through gene mutation assays (sex-linked recessive lethal, somatic 

6232 mutation and recombination) (Kramers et al., 1991). 

6233 In vivo genotoxicity studies reported DNA-protein cross links, DNA SSBs, chromosomal 

6234 aberrations, and sister chromatid exchanges in liver cells of B6C3F1 mice following acute 

6235 inhalation exposure to concentrations producing liver tumors with chronic exposure (Casanova et 
207

6236 al., 1996, 1992; Graves et al., 1995, 1994b). The formation of DNA SSBs was suppressed when 

6237 the mice were pretreated with a GSH depletor (Graves et al., 1995), providing additional support 

6238 for the involvement of GST metabolism. Increased sister chromatid exchanges and 

6239 chromosomal aberrations were found in the lungs in mice exposed to dichloromethane for 

6240 2 weeks to 8,000 ppm or for 12 weeks to 2,000 ppm. In this study, however, there was evidence 

6241 of damage at other sites, too: sister chromatid exchanges were also seen in peripheral 

6242 lymphocytes, chromosomal aberrations were seen in bone marrow and micronuclei were seen in 

6243 in peripheral red blood cells under the same exposure protocol (Allen et al., 1990). As was seen 

6244 in the liver, DNA SSBs were seen in lungs of B6C3F1 mice following acute inhalation exposure 

6245 to concentrations producing lung tumors with chronic exposure, and this effect was suppressed 

6246 with pretreatment with a GSH depletor, buthionine sulfoximine (Graves et al., 1995). Other 

6247 studies of sister chromatid exchange (Allen et al., 1990) or DNA damage detected by the comet 

6248 assay (Sasaki et al., 1998) also provide evidence of genotoxic effects specifically in lung cells in 

6249 mice. These in vivo mammaliam genotoxicity studies demonstrate site-specific effects 

6250 correlating to the dichloromethane-induced tumors in animals. Additional evidence for site 

6251 specificity comes from a study in which DNA damage (detected by the comet assay) was 

6252 enhanced in liver tissue, but not stomach, kidney, brain, or bone marrow, 24 hours after oral 

6253 administration of 1,720 mg/kg dichloromethane to CD-1 mice (Sasaki et al., 1998). 

6254 DNA reaction products (e.g., DNA adducts) produced by GST metabolites, such as S6255 (chloromethyl)glutathione, have not been identified in in vivo studies (Watanabet et al., 2007). 

6256 The authors speculated that these results are due to the instability of the reaction products 

6257 (Hashmi et al., 1994). However, adducts with nucleosides have been observed in in vitro 

6258 studies, when DNA was treated with S-(1-acetoxymethyl)glutathione, a compound structurally 

6259 similar to S-(chloromethyl)glutathione and in calf thymus DNA in the presence of 

6260 dichloromethane, but not formaldehyde, and GST (Marsch et al., 2004; Kayser and Vuilleumier, 

6261 2001). These findings indicate that the S-(chloromethyl)glutathione intermediate formed by 

6262 GSH conjugation has mutagenic potential and is likely responsible, at least in part, for the 

6263 mutagenic response observed following dichloromethane exposure. However, other studies (Hu 

6264 et al., 2006; Casanova et al., 1996) provide evidence of formaldehyde-related DNA-protein 

6265 cross-links in relation to dichloromethane exposure. These results show that, while most studies 

6266 indicate the importance of the S-(chloromethyl)glutathione intermediate in mediating genotoxic 

6267 damage following dichloromethane exposure, DNA damage resulting from formaldehyde 

6268 formation should also be considered. 

6269 Mutagenic data in critical genes leading to the initiation of dichloromethane-induced liver 

6270 or lung tumors are not available. In vivo assays evaluating mutations in tumor suppressor genes 

6271 and oncogenes reported similar frequencies of activated H-ras genes and inactivation of the 

6272 tumor suppressor genes, p53 and Rb-1 in the liver tumors seen in the nonexposed and 

6273 dichloromethane-exposed B6C3F1 mice (Devereaux et al., 1993; Hegi et al., 1993). There were 
208

6274 too few lung tumors (n=4) in controls to provide a conclusive comparison of mutation patterns 

6275 between exposed and nonexposed tumors. 

6276 

6277 Dose-response concordance 

6278 Statistically significant increases in liver tumor incidences in male and female (2,000 and 

6279 4,000 ppm) mice were observed in the inhalation bioassay in B6C3F1 mice (NTP, 1986). 

6280 Several studies provide evidence of an association between mutagenic events mediated by GST6281 pathway metabolites with the exposure levels inducing liver tumors in B6C3F1 in this study, and 

6282 concentration dependent increases in genotoxicity have been observed in in vitro and in vivo 

6283 assays. 

6284 In vitro mammalian genotoxicity studies were positive and demonstrated a dose-response 

6285 relationship for DNA protein crosslinks, DNA single stranded breaks, and DNA damage as 

6286 measured by the comet assay at concentrations ranging from 2.5 to 60 mM when mouse liver 

6287 cytosol was added or if mouse GSTT1-1 was transfected into hamster cell lines (Graves et al., 

6288 1994b; Graves et al., 1996; Hu et al., 2006). In mouse hepatocytes, DNA protein crosslinks were 

6289 observed following dichloromethane exposures ranging between 0.5 – 6.0 mM (Casanova et al., 

6290 1997). DNA-protein cross-links were detected in mouse hepatocytes incubated with 1.9 mM 

6291 dichloromethane (Casanova et al., 1997), a concentration chosen based on its correspondence to 

6292 the TWA liver concentration of dichloromethane that was predicted by the Andersen et al. 

6293 (1987) PBTK model for mice exposed by inhalation to 4,000 ppm for 6 hours (a dose that 

6294 resulted in increased liver tumor incidence in the two-year bioassay reported by NTP, 1986). 

6295 Consistent with the relative lack of liver tumor responses in Syrian golden hamsters (Burek et al., 

6296 1984) and F344 rats (NTP, 1986) with chronic exposure to 3,500 or 4,000 ppm, hepatocytes 

6297 from these strains of animals did not form detectable DNA-protein cross-links when incubated 

6298 with 1.9 mM dichloromethane (Casanova et al., 1997). .................................................................. 

6299 DNA-protein cross-links were not detected in livers of mice exposed to 146 ppm 

6300 6 hours/day for 3 days, but a concentration-dependent increase in DNA-protein cross-links was 

6301 observed in DNA from livers of mice exposed to several concentrations between 500 and 

6302 4,000 ppm (Casanova et al., 1996). Following exposure under similar conditions (concentrations 

6303 of 498, 1,553, or 3,923 ppm), DNA-protein cross-links were not detected in the livers of Syrian 

6304 golden hamsters, a species that did not develop tumors after chronic inhalation exposure to 

6305 dichloromethane (Casanova et al., 1996, 1992). Increased DNA SSBs were detected in liver 

6306 tissue of B6C3F1 mice immediately following a 6-hour inhalation exposure to dichloromethane 

6307 at concentrations ranging from 2,000 to 8,000 ppm (Graves et al., 1995), and in mouse 

6308 hepatocytes after a 3-hour exposure to 4000 (but not 2000) ppm (Graves et al., 1994b). 

6309 Statistically significant increases in the incidence of lung tumors were observed in the 

6310 inhalation chronic bioassay in male and female B6C3F1 mice exposed to 2,000 or 4,000 ppm 

6311 dichloromethane (Mennear et al., 1988; NTP, 1986). Evidence of mutagenicity at these exposure 
209

6312 levels comes from two inhalation studies (Graves et al., 1995; Allen et al., 1990). Increased 

6313 DNA SSBs were detected in lung tissue of B6C3F1 mice immediately following a 6-hour 

6314 inhalation exposure to dichloromethane at concentrations ranging from 2,000 to 8,000 ppm 

6315 (Graves et al., 1995). In the study by Allen et al. (1990), increased presence of sister chromatid 

6316 exchanges was observed in mouse lung cells following a 12-week exposures at 2000 ppm; 

6317 shorter durations of exposure (2 weeks) were positive for measures of sister chromatid exchange 

6318 and chromosome aberrations at 8000 ppm, but not at 2000 or 4000 ppm. 

6319 DNA adducts were observed and increased with dose in an in vitro preparation of calf 

6320 thymus DNA when treated with dichloromethane (5 – 60 mM) and bacterial, rat, or human GST 

6321 (Marsch et al., 2004). 

6322 

6323 Temporal relationship 

6324 Dichloromethane-induced liver and lung tumors first appeared in mice after 52 weeks of 

6325 exposure (Maronpot et al., 1995; Kari et al., 1993). The detection of DNA-protein cross-links in 

6326 the livers of B6C3F1 mice following short-term inhalation exposures to dichloromethane 

6327 concentrations that induced tumors with chronic exposure (Casanova et al., 1996, 1992) provides 

6328 temporal support for the proposed mutagenic mode of action. Additional supporting evidence 

6329 comes from observations that increased levels of DNA SSBs were detected in the liver and lungs 

6330 of B6C3F1 mice immediately following 3-hour inhalation exposure to 2,000–8,000 ppm 

6331 dichloromethane (Graves et al., 1995; 1994b). Single dose and inhalation exposure studies of 6 

6332 hours or less did not detect an effect on DNA synthesis (Lefevre and Ashby, 1989) or 

6333 unscheduled DNA synthesis (Trueman and Ashby, 1987) in mouse liver cells. 

6334 

6335 Biological plausibility and coherence 

6336 Bioactivation of a parent compound into a mutagenic metabolite resulting in cancer is a 

6337 plausible mode of action carcinogenicity in humans and is a generally accepted mode of action. 

6338 Dichloromethane-induced carcinogenicity is hypothesized to be due to metabolism of the parent 

6339 compound by the GST pathway (GST-T1) to a metabolite that is tumorigenic. The GST 

6340 metabolite, S-(chloromethyl)glutathione, formed from dichloromethane, has been characterized 

6341 as labile and highly reactive through in vitro evaluation of dichloromethane metabolism in 

hepatocytes using 13 6342 C-NMR techniques (Hashmi et al., 1994) and through an enzyme digestion 

6343 assay using calf thymus DNA and GST-T1 enzyme (Marsch et al., 2004). The hypothesis that 

6344 the formation of a mutagenic metabolite is a preliminary step resulting in carcinogenicity is 

6345 based on evidence that malignant tumors are primarily located in areas where dichloromethane is 

6346 highly metabolized by GST-T1, such as the liver and the lung, and on mutagenicity studies 

6347 indicating the importance of the GST pathway and that the lung and liver are more prone to 

6348 mutagenic effects of dichloromethane (Sasaki et al., 1998; Casanova et al., 1996, 1992; Graves 

6349 et al., 1995, 1994b). The site selectivity of the mutagenicity in liver and lung tissue as evidenced 
210

6350 by several studies suggests that the GST reactive metabolite remains in the tissue where it is 

6351 formed. Collectively, the studies support the hypothesis that dichloromethane-mediated 

6352 carcinogenicity results from a GST metabolite that produces selective DNA damage in the 

6353 tissues where the metabolite is formed, but this hypothesis is based, in part, on assumptions 

6354 regarding metabolite clearance and reactivity. DNA damage in the liver and lung as well as the 

6355 increased incidence of tumor formation resulting from dichloromethane exposure indicates 

6356 coherence of the mutagenic and carcinogenic effects and is evidence supporting a mutagenic 

6357 mode of action. 

6358 Differences in GST activity in mice compared with other species, and the interspecies 

6359 variability in genotoxic effects corresponding to interspecies variability in tumor response 

6360 support the mode of action hypothesis. DNA SSBs were not detected in liver or lung cells in rats 

6361 exposed to similar inhalation exposures that induce strand breaks in mice (Graves et al., 1995; 

6362 Graves et al., 1994b), and were detected at much lower in vitro concentrations in isolated 

6363 hepatocytes from B6C3F1 mice (0.4 mM) than in hepatocytes from Alpk:APfSD rats (30 mM) 

6364 (Graves et al., 1995, figure 3). The difference in susceptibility to carcinogenic response between 

6365 mice and rats likely reflects differences in GST metabolism. Toxicokinetic studies indicate that, 

6366 with increasing exposure levels, increasing amounts of dichloromethane are metabolized via 

6367 GST metabolism. 

6368 

6369 4.7.3.1.2. Other possible modes of action for liver or lung tumors in rodents. 

6370 Data are not available to support other possible modes of action for the liver and lung 

6371 tumors in rodents. Efforts to observe sustained cell proliferation in liver following 

6372 dichloromethane exposure of B6C3F1 mice have been unsuccessful. Groups of female B6C3F1

6373 mice were exposed to 0 or 2,000 ppm dichloromethane 6 hours/day, 5 days/week for up to 78 

6374 weeks did not exhibit enhanced cell proliferation in the liver when assessed at various intervals 

6375 during exposure (Foley et al., 1993). . 

6376 Indices of enhanced cell proliferation have been measured in the lungs of male B6C3F1

6377 mice following acute duration exposure at concentrations of about 1,500, 2,500, or 4,000 ppm 

6378 dichloromethane (6 hours/day for 2 days) but not at exposure concentrations of 150 or 500 ppm 

6379 and not in lungs of Syrian golden hamsters exposed to concentrations up to 4,000 ppm 

6380 (Casanova et al., 1996). Earlier studies showed somewhat consistent findings in that the 

6381 numbers of bronchiolar cells undergoing DNA synthesis (thymidine incorporation labeling) were 

6382 markedly increased (about 6- to 15-fold) in bronchiolar cells of B6C3F1 mice exposed to 

6383 4,000 ppm dichloromethane 6 hours/day on days 5, 8, and 9 of exposure, but no evidence for 

6384 increased cell proliferation was found after 89, 92, or 93 days of exposure (Foster et al., 1992). 

6385 The results suggest that enhanced cell proliferation is not sustained in the lung with longer-term 

6386 exposure to dichloromethane concentrations associated with lung tumor development in mice, 
211

6387 and that this mode of tumor promotion is not important in the development of dichloromethane6388 induced lung tumors. 

6389 

6390 4.7.3.2. General Conclusions About the Mode of Action for Tumors in Rodents and 

6391 Relevance to Humans 

6392 The mode of action for dichloromethane is hypothesized to involve mutagenicity via 

6393 reactive metabolites. Mechanistic evidence indicates that dichloromethane-induced DNA 

6394 damage in cancer target tissues of mice involves DNA-reactive metabolites produced via a 

6395 metabolic pathway initially catalyzed by GST. Although mutational events in critical genes 

6396 leading to tumor initiation have not been established, evidence supporting a mutagenic mode of 

6397 action includes the identification of mutagenic response (reverse mutations) in short-term 

6398 bacterial assays (with microsomal activation) and induced DNA-protein cross-links and DNA 

6399 SSBs in mammalian cell assays. There are numerous positive in vivo genotoxicity studies 

6400 specifically examining responses in the liver and/or lung; these studies included evidence of 

6401 chromosomal aberrations, SSBs and sister chromatid exchanges, and DNA-protein cross-links. 

6402 The negative in vivo genotoxicity assays are generally those that were based on a micronucleus 

6403 test using mouse bone marrow, which is expected as halogenated hydrocarbons (such as 

6404 dichloromethane) are not very effective in this type of assay (Crebelli et al., 1999; Dearfield and 

6405 Moore, 2005). 

6406 

6407 Is the hypothesized mode of action sufficiently supported in test animals? 

6408 Consistent and specific evidence for the association between the formation of DNA6409 reactive GST-pathway metabolites and the formation of liver and lung tumors from inhalation 

6410 includes (1) enhanced GST metabolic capacity in the liver and lung and enhanced localization of 

6411 GST-T1 in hepatic cell nuclei in B6C3F1 mice, compared with rats and hamsters, which do not 

6412 show strong tumor responses to chronic inhalation exposure; (2) the detection of DNA-protein 

6413 cross-links, or DNA SSBs in livers and lungs of B6C3F1 mice following acute inhalation 

6414 exposure to concentrations that produce tumors with chronic exposure; (3) suppression of the 

6415 formation of DNA SSBs in livers and lungs of B6C3F1 mice pretreated with a GSH depletor; (4) 

6416 the inability to detect DNA-protein cross-links or DNA SSBs in livers or lungs of similarly 

6417 exposed rats or hamsters (5) detection of DNA SSBs at much lower in vitro concentrations in 

6418 isolated hepatocytes from B6C3F1 mice than in hepatocytes from Alpk:APfSD rats; (6) dose6419 response concordance and a temporal relationship for the formation of DNA-protein cross-links 

6420 and DNA SSBs with the formation of liver and lung tumors in B6C3F1 mice exposed to 

6421 dichloromethane; (7) the detection of increased sister chromatid exchanges in lung cells from 

6422 CD-1 mice exposed by inhalation to dichloromethane; and (8) enhancement of dichloromethane 

6423 genotoxicity in bacterial and mammalian in vitro assays with the introduction of GST metabolic 
212

6424 capacity. However, mutations in critical genes linked to initiation of tumor cells have not been 

6425 identified. 

6426 The much weaker carcinogenic response in the liver of rats and mice to chronic drinking 

6427 water exposure (Serota et al., 1986a, b) than that noted in mice exposed by inhalation (Kari et al., 

6428 1993; NTP, 1986) is correlated with much smaller amounts of GST metabolites produced in the 

6429 liver under the exposure conditions of the oral bioassay than in the inhalation bioassay (Andersen 

6430 et al., 1987). 

6431 

6432 In conclusion, there is sufficient evidence supporting a mutagenic mode of action and to 

6433 establish the involvement of GST metabolism in the lung and liver carcinogenicity of 

6434 dichloromethane in mice. 

6435 

6436 Is the hypothesized mode of action relevant to humans? 

6437 The postulated mode of action, that dichloromethane is metabolized by GST to reactive 

6438 metabolites that induce mutations in DNA leading to carcinogenicity is possible in humans. 

6439 Mutagenicity as a mode of action for carcinogenicity in humans is generally accepted and is a 

6440 biologically plausible mechanism for tumor induction. The toxicokinetic and toxicodynamic 

6441 processes that would enable reactive metabolites to produce mutations in animal models are 

6442 biologically plausible in humans. Furthermore, the detection of the GST pathway in human 

6443 tissues indicates that the hypothesized mode of action involving reactive metabolites from this 

6444 pathway, S-(chloromethyl)glutathione and formaldehyde, is relevant to humans. 

6445 Another factor that may play a role in the apparent species differences in carcinogenicity 

6446 resulting from dichloromethane exposure is species differences in intracellular localization of 

6447 GST-T1 (Sherratt et al., 2002; Mainwaring et al., 1996). In mouse liver tissue, GST-T1 appears 

6448 to be localized in the nuclei of hepatocytes and bile-duct epithelium, but rat liver does not show 

6449 preferential nuclear localization of GST-T1. In human liver tissue, some hepatocytes show 

6450 nuclear localization of GST-T1 and others show localization in cytoplasm. Nuclear production 

6451 of S-(chloromethyl)glutathione catalyzed by GST-T1 in the nucleus is more likely than 

6452 cytoplasmic production to lead to DNA alkylation. The finding of some nuclear localization of 

6453 GST-T1 in human liver tissue supports the relevance of the hypothesized mode of action to 

6454 humans. 

6455 Comparisons in mice, rats, humans, and hamsters of GST enzyme activity in liver and 

6456 lung tissues have indicated the following rank order: mice > rats > or ≈ humans > hamsters 

6457 (Reitz et al., 1989; Thier et al., 1998). This relative ranking in GST activity corresponds to the 

6458 rank order of the strength of the association between inhalation exposure to dichloromethane and 

6459 liver tumors in long-term cancer bioassays with mice, rats, and hamsters. This relative ranking 

6460 does not preclude the relevance of the hypothesized mode of action to humans, however. 

6461 
213

6462 Which populations or lifestages can be particularly susceptible to the hypothesized mode of 

6463 action? 

6464 As discussed in section 3.3, a polymorphism of the GST-T1 gene is present in humans. 

6465 People with two functional copies of the gene (+/+) readily conjugate GSH to dichloromethane. 

6466 Individuals having only one working copy of the gene (+/–) display relatively decreased 

6467 conjugation ability. Individuals with no functional copy of the gene (–/–) do not express active 

6468 GST-T1 protein and do not metabolize dichloromethane via a GST-related pathway (Thier et al., 

1998). Thus the GST-T1+/+ 6469 (wild-type) genotype would be considered to be the more “at risk” 

6470 population; this subgroup represents approximately 30% of the U.S. population (Haber et al., 

6471 2002) but would be expected to be more common among Caucasians and African-Americans 

6472 than among Asians (Raimondi et al., 2006; Garte et al., 2001; Nelson et al., 1995) (see Table 

6473 3-3). 

6474 According to the Supplemental Guidance for Assessing Susceptibility from Early-Life 

6475 Exposure to Carcinogens (U.S. EPA, 2005b), children exposed to carcinogens with a mutagenic 

6476 mode of action are assumed to have increased early-life susceptibility. The Supplemental 

6477 Guidance (U.S. EPA, 2005b) recommends the application of age-dependent adjustment factors 

6478 (ADAFs) for carcinogens that act through a mutagenic mode of action. Although the database is 

6479 lacking in vivo evidence of specific mutagenic events following chronic exposure to 

6480 dichloromethane, the weight of the available evidence indicates that dichloromethane is acting 

6481 through a mutagenic mode of carcinogenic action. Application of ADAFs is recommended for 

6482 both the oral and inhalation routes of exposure when risks are assessed that are associated with 

6483 early-life exposure. 

6484 

6485 4.8. SUSCEPTIBLE POPULATIONS AND LIFE STAGES 

6486 4.8.1. Possible Childhood Susceptibility 

6487 In humans, hepatic CYP2E1 begins to be expressed in the second trimester (Johnsrud et 

6488 al., 2003), increases significantly in the third trimester, and continues to increase during the first

6489 year of life (Hines, 2007; Johnsrud et al., 2003; Treluyer et al., 1996; Vieira et al., 1996). In the

6490 fetal brain, however, CYP2E1 activity is seen as early as GD 50, with increasing levels seen until 

6491 at least the end of the first trimester (Brzezinski et al., 1999). Neurobehavioral effects of 

6492 dichloromethane are seen with acute exposures in adults, and the available data regarding 

6493 neurological symptom history and standardized testing suggest the possibility of longer-term 

6494 effects. The relatively high activity of CYP2E1 in the brain compared to the liver of the 

6495 developing human fetus raises the potential for neurodevelopmental effects from 

6496 dichloromethane exposure. Results from a developmental toxicity study in rats also raise 

6497 concern for possible neurodevelopmental effects. Decreased offspring weight at birth and 

6498 changed behavioral habituation of the offspring to novel environments were seen following 

6499 exposure of adult Long-Evans rats to 4,500 ppm for 14 days prior to mating and during gestation 
214

6500 (or during gestation alone) (Bornschein et al., 1980; Hardin and Manson, 1980). In the only 

6501 other animal study examining possible early-life susceptibility to dichloromethane toxicity, 

6502 Alexeef and Kilgore (1983) found that exposure of young male mice to approximately 

6503 47,000 ppm for about 20 seconds significantly impaired the ability to learn, using a passive6504 avoidance conditioning task. Three-week-old mice were more affected than 5- or 8-week-old 

6505 mice. The broad issue of childhood susceptibility to chronic neurobehavioral effects of early life 

6506 exposure represents a data gap in the understanding of the health effects of dichloromethane. 

6507 The relatively low CYP2E1 activity in the liver of infants would tend to shift metabolism 

6508 of dichloromethane to the GST pathway. This shift could affect cancer risk, given the evidence 

6509 of genotoxicity through this metabolic pathway. However, the available data in humans are not 

6510 sufficient to address the question of whether in utero or early life exposures represent a period of 

6511 increased susceptibility to potential carcinogenic effects of dichloromethane. A threefold 

6512 increased risk of childhood leukemia (acute lymphoblastic leukemia) was seen in relation to 

6513 maternal occupational exposure in the year before and during pregnancy in one population-based 

6514 case-control study (OR 3.22 [95% CI 0.88, 11.7]) for ratings of “probable or definite” exposure 

6515 compared with possible or no exposure (Infante-Rivard et al., 2005). The estimates for 

6516 categories based on concentration and frequency were similar, but there was no evidence for an 

6517 increasing risk with increasing exposure level. 

6518 Experiments comparing cancer responses from early-life exposures with those from adult 

6519 exposures are not available for F344 rats or B6C3F1 mice, the strains of animals in which 

6520 carcinogenic responses to dichloromethane have been observed (mammary gland tumors in F344 

6521 rats and liver and lung tumors in B6C3F1 mice exposed by inhalation; liver tumors in female

6522 F344 rats and male B6C3F1 mice exposed via drinking water). Animal data evaluating the effect 

6523 of age on the susceptibility to dichloromethane carcinogenicity are restricted to a bioassay in 

6524 which 54 pregnant Sprague-Dawley rats were exposed, starting on GD 12, to 100 ppm 

6525 dichloromethane 4 hours/day, 5 days/week for 7 weeks, followed by 7 hours/day, 5 days/week 

6526 for 97 weeks (Maltoni et al., 1988). Groups of 60 male and 69 female newborns continued to be 

6527 exposed after birth to 60 ppm dichloromethane 4 hours/day, 5 days/week for 7 weeks, followed 

6528 by exposure 7 hours/day, 5 days/week for 97 weeks. Additional groups of 60 male and 

6529 70 female newborns were exposed after birth to 60 ppm dichloromethane 4 hours/day, 

6530 5 days/week for 7 weeks and then for 7 hours/day, 5 days/week for 8 weeks. Endpoints 

6531 monitored included clinical signs, BW, and full necropsy at sacrifice (when spontaneous death 

6532 occurred). For each animal sacrificed, histopathologic examinations were performed on the 

6533 following organs: brain and cerebellum, zymbal glands, interscapular brown fat, salivary glands, 

6534 tongue, thymus and mediastinal lymph nodes, lungs, liver, kidneys, adrenals, spleen, pancreas, 

6535 esophagus, stomach, intestine, bladder, uterus, gonads, and any other organs with gross lesions. 

6536 There was no significant effect of exposure to dichloromethane on the incidence of benign or 

6537 malignant tumors among adults or the progeny. The results provide no evidence that Sprague-
215

6538 Dawley rats would be more sensitive to potential carcinogenic activity of dichloromethane 

6539 during early life stages. Further conclusions from these results are precluded because the study 

6540 included only one exposure level, which was below the maximum tolerated dose for adult 

6541 Sprague-Dawley rats. 

6542 

6543 4.8.2. Possible Gender Differences 

6544 The limited data available from studies in humans do not indicate that there are large 

6545 differences by gender in sensitivity to cardiovascular, neurologic, cancer, or other effects; studies 

6546 have not been conducted specifically to examine this question and so do not provide information 

6547 pertaining to smaller or more subtle differences. The available animal studies similarly do not 

6548 establish whether either gender may be more susceptible to the toxic effects of dichloromethane. 

6549 Studies of the carcinogenic effects of dichloromethane, either by inhalation or by the oral route, 

6550 have not suggested an increased susceptibility of either male or female animals. 

6551 

6552 4.8.3. Other Susceptible Populations 

6553 As discussed in section 3.3, a polymorphism exists within the GST-T1 gene in humans, 

6554 resulting in individuals with diminished, or a lack of, ability to conjugate GSH to 

6555 dichloromethane. While the possible effects of this polymorphism on the toxicity of 

6556 dichloromethane have not been directly demonstrated, it can be inferred from the proposed mode 

6557 of action that a decrease in the GST-T1 metabolic pathway would result in a decreased 

6558 generation of reactive metabolites and a decrease in any chronic effects mediated through those 

6559 metabolites (Jonsson and Johanson, 2001; El-Masri et al., 1999). 

6560 Interindividual variation in the ability to metabolize dichloromethane via GST-T1 is 

6561 associated with genetic polymorphisms in humans. Estimated U.S. population prevalence of 

6562 nonconjugators (–/– at the GST-T1 locus) is about 20%, but higher prevalences (47–64%) have 

6563 been reported for Asians (Raimondi et al., 2006; Haber et al., 2002; Garte et al., 2001; Nelson et 

6564 al., 1995). Although nonconjugators are expected to have negligible extra risk for 

6565 dichloromethane-induced cancer, the U.S. prevalences for low (+/– at the GST-T1 locus) and 

6566 high (+/+) conjugators have been estimated at 48 and 32%, respectively (Haber et al., 2002). 

6567 The liver and kidney are the most enriched tissues in GST-T1, but evidence is available for the 

6568 presence of GST-T1 in other tissues, including the brain and lung, at lower levels (Sherratt et al., 

6569 2002, 1997). 

6570 Individuals may vary in their ability to metabolize dichloromethane through the CYP2E1 

6571 pathway. Individuals with decreased CYP2E1 activity may experience decreased generation of 

6572 CO and an increased level of GST-related metabolites, following exposure to dichloromethane, 

6573 which may result in increased susceptibility to the chronic effects of dichloromethane from GST6574 related metabolites. Conversely, individuals with higher CYP2E1 activity may experience 

6575 relatively increased generation of CO at a given dichloromethane exposure level and, therefore, 
216

6576 may be more susceptible to the acute CO-related toxicity or other chronic effects of 

6577 dichloromethane. Several studies indicate a three- to sevenfold variability in CYP2E1 activity 

6578 among humans, as assessed by various types of measurements among “healthy” volunteers 

6579 (Sweeney et al., 2004; Haufroid et al., 2003; Lipscomb et al., 2003; Bernauer et al., 2000; Lucas 

6580 et al., 2001, 1999; Kim et al., 1995; Shimada et al., 1994). This variability is incorporated into 

6581 the PBTK models for dichloromethane. Factors that may induce or inhibit CYP2E1 activity 

6582 (e.g., obesity, alcohol use, diabetes) or co-exposures (i.e., to various solvents or medications) 

6583 (Lucas et al., 1999) may result in greater variation within segments of the population. This 

6584 variation in CYP2E1 activity may result in earlier saturation of this pathway and greater 

6585 exposure to the parent compound, which would be of particular relevance to neurological effects. 

6586 
217

6587 5. DOSE-RESPONSE ASSESSMENTS 

6588 

6589 

6590 5.1. ORAL REFERENCE DOSE (RfD) 

6591 5.1.1. Choice of Principal Study and Critical Effect—with Rationale and Justification 

6592 As discussed in section 4.6.1, human data for oral exposures to dichloromethane are 

6593 limited to case reports involving intentional (i.e., suicidal) or accidental, acute ingestion 

6594 exposures (Chang et al., 1999; Hughes and Tracey, 1993). Reported effects reflect frank toxicity 

6595 from very high doses, such as marked CNS depression, injury to the gastrointestinal tract, liver 

6596 and kidney failure, coma, and death. No studies of human chronic oral exposures are available. 

6597 In the absence of adequate studies evaluating possible health effects in humans repeatedly 

6598 exposed to dichloromethane via the oral route, the results from the chronic laboratory animal 

6599 studies are assumed to be relevant to humans. 

6600 The database of laboratory animal oral exposure studies includes 90-day (Kirschman et 

6601 al., 1986) and 2-year drinking water toxicity studies in F344 rats (Serota et al., 1986a) and 

6602 B6C3F1 mice (Serota et al., 1986b). A reproductive study exposed Charles River CD rats via 

6603 gavage before mating (General Electric Co., 1976), and a developmental study exposed F344 rats 

6604 via gavage during GDs 6–19 (Narotsky and Kavlock, 1995). A 14-day gavage study examined 

6605 neurotoxicity in F344 rats (Moser et al., 1995). 

6606 Hepatic effects (hepatic vacuolation, nonneoplastic liver foci) are the critical dose6607 dependent noncancer effects associated with oral exposure to dichloromethane (see Table 4-35). 

6608 The 90-day drinking water toxicity study in F344 rats (Kirschman et al., 1986) reported 

6609 significant increases in hepatocyte vacuolation and necrosis in animals dosed between 166 – 

6610 1200 mg/kg-day (males) or 200 – 1469 mg/kg-day (females). These doses were used to develop 

6611 dosing levels for the 104-week drinking water study (Serota et al., 1986a). The 104-week 

6612 drinking water study of F344 rats (Serota et al., 1986a) provides adequate data to describe dose6613 response relationships for nonneoplastic liver lesions from chronic oral exposure to 

6614 dichloromethane (e.g., includes four exposure levels and a control group). In this study, rats 

6615 dosed at 50 mg/kg-day or higher in both sexes had increased fatty livers, but quantitative data 

6616 were not provided by the authors. Liver lesions, described as foci or areas of cellular alteration, 

6617 were also seen in this study in the same dose groups in which the fatty changes had occurred. A 

6618 limitation of this study is that Serota et al (1986a) did not describe the evaluation of the altered 

6619 foci in detail. However, increases in altered foci did not correspond to tumor rate incidences in 

6620 either male or female rats. Instead, the altered foci correlated more closely to fatty liver 

6621 incidence changes for both sexes in the rats. Altered foci could range from a focal fatty change 

6622 (nonneoplastic) to an enzymatic altered foci change (neoplastic) (Goodman et al., 1994). Several 

6623 lines of evidence were considered in determining whether the lesions should be characterized as 

6624 nonneoplastic or neoplastic: 1) There is a congruence between the incidence of this lesion and 
218

6625 the incidence of the fatty liver in the study by Serota et al. (1986a); 2) At higher doses, 

6626 hepatocyte vacuolation and hepatocyte necrosis were seen (Kirschman et al., 1986; Berman et 

6627 al., 1995); and 3) there is no clear indication that these altered foci progress to liver tumors since

6628 the rate of increased foci did not correlate with liver tumor increases in either male or female 

6629 rats. Based on these observations, the altered foci were determined to be more likely to be 

6630 representative of a focal fatty change (nonneoplastic) than a neoplastic event. 

6631 The LOAELs for nonneoplastic liver lesions in rodents following repeated oral exposure 

6632 (50–586 mg/kg-day) (Table 4-35) are in the same range or below the NOAELs of 225 mg/kg-day 

6633 for reproductive performance in Charles River CD rats exposed for 90 days before mating 

6634 (General Electric Co., 1976) and 450 mg/kg-day for developmental toxicity in pregnant F344 

6635 rats exposed during gestation (Narotsky and Kavlock, 1995). The LOAEL (337 mg/kg-day) and 

6636 NOAEL (101 mg/kg-day) for mild neurological impairment in a 14-day gavage exposure study 

6637 of F344 rats (Moser et al., 1995) indicates that the threshold for neurological effects may be 

6638 similar to the threshold for liver effects. A limitation of the Moser et al. (1995) study, however, 

6639 is that the observed effects were limited to measures taken within 4 hours of exposure. 

6640 The subchronic (i.e., 90-day or less study) data were not considered in the selection of a 

6641 principal study for deriving the chronic RfD because the database contains reliable dose-response 

6642 data from a chronic study at lower doses than the 90-day study (Kirschman et al., 1986) 

6643 (conducted to provide data pertaining to relevant doses to use in the chronic study). The data 

6644 from the subchronic studies are, however, used to corroborate the findings in the chronic studies 

6645 with respect to relevant endpoints (i.e., hepatic and neurological effects). The neurotoxicity 

6646 study was not selected as the principal study due to the limited measurements to inform the 

6647 chronic exposure to dichloromethane. The rat rather than the mouse chronic bioassay (Serota et 

6648 al., 1986a) was selected as the principal study for the RfD because of the consistent evidence that 

6649 rats may be more sensitive than mice to nonneoplastic liver effects from orally administered 

6650 dichloromethane; available rat LOAELs for nonneoplastic liver lesions are lower than mouse 

6651 LOAELs (see Table 4-35). Figure 5-1 is an exposure-response array that presents NOAELs, 

6652 LOAELs, and the dose range tested, corresponding to selected health effects from the short-term 

6653 (neurotoxicological) and subchronic studies, and from the chronic, reproductive, and 

6654 developmental toxicity studies that were evaluated for use in the derivation of the RfD. 
219

Figure 5-1. Exposure Response Array for Oral Exposure to Dichloromethane.

0

200400600800

100012001400

2000

Dose (mg/kg-day)

Hepatocyte

vacuolation (F344 

rat) - Kirschman et 

al. (1986)

Hepatocytevacuolation(B6C3F1 male mice) -Kirschman et al. (1986)

Nonneoplasticliver foci (F344 rat, male and female) - Serotaet al. (1986a)

Fatty liver (B6C3F1 mice, male and female) -Serota et al. (1986b)

Neurologic, Functional Observational Battery (F344 rat, female) -Moser et al. (1995)

Reproductive Performance (CD rat, male and female) -General Electric Co. (1976)

Maternal weight gain (F344 rat, pregnant female) - Narotsky and Kavlock (1995)

CHRONIC HEPATIC SUBCHRONIC HEPATIC NEUROTOX REPRODUCTIVE AND DEVELOP

NOAELLOAEL

The vertical lrange of expostudy.Closed dots ( exposure concused in study

Reproductive organs; performance (male Swiss Webster ) -Raje et al. (1988)

6655 6656 

6657 Figure 5-1. Exposure response array for oral exposure to dichloromethane 
220

6658 5.1.2. Derivation Process for Noncancer Reference Values 

6659 The toxicity values (oral RfD and inhalation RfC) for noncancer endpoints were derived 

6660 by using rat and human PBTK models to calculate internal doses in rats from experimental 

6661 exposures and extrapolate points of departure to human equivalent exposures. Figure 5-2 

6662 illustrates the process of using the PBTK models for toxicity value derivation. The process for 

6663 the RfD and RfC is summarized below, using the example of a noncancer liver effect. 

6664 A deterministic PBTK model for dichloromethane in rats was first used to convert rat 

6665 drinking water or inhalation exposures to values of an internal liver dose metric (see Appendix C 

6666 for details of the rat PBTK model). Available models in EPA benchmark dose (BMD) software 

6667 (BMDS) version 2.0 were then fit to the liver lesion incidence data and internal liver dose data 

6668 for rats, and BMD10s and their lower 95% confidence limits associated with a 10% extra risk 

6669 (BMDL10) were calculated from each of the models. Adequacy of model fit was assessed by 

overall χ

2

6670 goodness of fit (p-value > 0.10) and examination of residuals, particularly in the region 

6671 of the benchmark response (BMR). The choice of best-fitting model was based on the lowest 

6672 Akaike’s Information Criterion [AIC] among models with adequate fits (U.S. EPA, 2000b). 

The use of a PBTK model can replace the use of the BW0.75 6673 scaling factor to account for 

6674 interspecies differences in toxicokinetics. The decision with respect to use of a scaling factor 

6675 depends on the dose metric that is used. Where PBTK models predict the concentration (in 

6676 particular, the AUC) of the proximate causative agent, a scaling factor to account for interspecies 

6677 differences is not typically used. That is, it is assumed that if the time-averaged (or steady-state)

6678 concentration of the proximate causative agent predicted by the PBTK model in the target tissue 

6679 is the same in the test species as in humans, and the test species was exposed for an equivalent 

6680 portion of its lifetime (2 years in rats and mice being equivalent to a 70-year lifetime in humans), 

6681 then the resulting risks in the two species are the same. However, when the PBTK model 

predicts the rate of production of the agent, rather than its concentration, then a BW0.75 6682 scaling 

6683 factor may be appropriate, depending on what is known or expected regarding the rate of 

6684 clearance of the agent or metabolite of interest. Two different scenarios can be considered. If 

6685 the metabolite formed is considered to be highly reactive, then it can be assumed that the rate of 

6686 clearance (i.e., disappearance due to local reactivity) for this metabolite, per volume tissue, is 

equal in rodents and humans. Thus, in that situation, as with the AUC dose metric, no BW0.75 6687 

6688 scaling factor is necessary, although differences in tissue volume fraction in humans versus rats 

6689 (as occurs for liver) should be and are accounted for by the PBTK model. However, if the 

6690 metabolite is not highly reactive, then it is expected that interspecies differences in clearance or 

removal of the toxic metabolite follow the generally assumed BW0.75 6691 scaling for rates of 

6692 metabolism and blood circulation. In this case, or in situations in which the reactivity or rate of 

6693 removal of the metabolite has not been established, it is appropriate to use a scaling factor, based 

6694 on BW ratios, to account for this difference. In the case of the noncancer liver effects of 

6695 dichloromethane, very limited information is available on the mechanism(s) involved in creating 
221

Rodent Dose Response Data

PBTK 

Model Estimates of Rodent 

Internal Dose

Rodent Internal BMDL10

95% Lower Bound Estimate of Internal 

Dose Associated with a 10% response 

BMD 

Modeling

Probabilistic

Human PBTK Model 

(What administered doses will produce a BMDL10 in a 

population?)

Monte Carlo Sampling from 

Distributions of Human Model 

Parameters

Distribution of Human Equivalent Administered 

Doses (mg/kg) or Inhalation Concentrations (mg/m3) 

(Points of Departure)

1st 5th

Divide by Uncertainty Factors for Interspecies ToxicodynamicVariability, Human Toxicodynamic Variability and Database Deficiencies )

Recommend lower percentile (e.g., 

1st) to protect sensitive individuals

Oral Reference Doses or 

Inhalation References 

Concentrations 

BMD

BMDL10

Scaling 

Factor

Human Internal

BMDL10

6696 

6697 Figure 5-2. Process for deriving noncancer oral RfDs and inhalation RfCs using rodent and human 6698 models.
222

6699 the type of hepatic damage seen. The dose metric used in the PBTK modeling is a rate of 

6700 metabolism, rather than the concentration of putative toxic metabolites, and the clearance of 

6701 these metabolites may be slower per volume tissue in the human compared with the rat. Thus the 

6702 rat internal dose metric for noncancer effects was adjusted by dividing by a pharmacokinetic 

6703 scaling factor to obtain a human-equivalent internal BMDL10 

6704 A probabilistic PBTK model for dichloromethane in humans, adapted from the model 

6705 of David et al. (2006) as described in Appendix B, was then used to calculate distributions of 

6706 chronic exposures associated with the human equivalent internal BMDL10, based on the 

6707 responses in rats. Parameters in the human PBTK model are distributions that incorporate 

6708 information about dichloromethane toxicokinetic and physiological variability and 

6709 uncertainty among humans, incorporating information about both the CYP2E1 and GST-T1 

6710 metabolic pathways (see Table 3-9 and Appendix B). Monte Carlo sampling was performed 

6711 in which each human model parameter was defined by a value randomly drawn from each 

6712 respective parameter distribution. The model was then executed by using the human internal 

6713 BMDL10 as input, and the resulting human equivalent administered dose or human equivalent 

6714 concentration (HEC) was recorded. This process was repeated for 10,000 iterations to 

6715 generate a distribution of human equivalent administered doses or concentrations. 

6716 The parameter statistics reported by David et al. (2006) include both the inter-individual 

6717 variability that would have been elucidated by the Bayesian analysis (variation between mean 

6718 values for each individual for which data were available) and uncertainty in those values. Since 

6719 EPA’s objective is to account for both population variability and parameter uncertainty, 

6720 however, these statistics were primarily used as published in David et al. (2006) (exceptions 

6721 discussed in Appendix B) to define population distributions. Assuming that these parameters are 

6722 distributed independently, ignoring the covariance that was likely represented in the actual 

6723 posterior chains, will tend to over-estimate the overall range of parameters and hence distribution 

6724 of dose metrics in the population, compared to what one would obtain if the covariance were 

6725 explicitly included. Thus if the covariance (i.e., the variance-covariance matrix) for the set of 

6726 parameters had been reported by David et al., it could have been used to narrow the predicted 

6727 distribution of internal doses, or equivalent applied doses. Lacking such information the 

6728 approach used will not under-estimate risk or over-estimate lower bounds on human equivalent 

6729 exposure levels. 

6730 From these distributions of human equivalent administered doses (or concentrations), 

6731 candidate RfDs or RfCs were derived by dividing the first percentile value (point of 

6732 departure) by uncertainty factors (UFs) to account for uncertainty about potential interspecies 

6733 toxicodynamic variability, human toxicodynamic variability, and database deficiencies. The 

6734 first percentile was chosen because it allowed generation of a stable estimate for the lower 

6735 end of the distribution while being protective of the overall human population, including 
223

6736 sensitive individuals. Choosing this lower point replaces the use of an additional UF to 

6737 account for human toxicokinetic variability. 

6738 

6739 5.1.3. Evaluation of Dose Metrics for Use in Noncancer Reference Value Derivations 

6740 There are no data to support the role of a specific metabolite in the development of 

6741 the noncancer liver lesions seen in oral and inhalation exposure studies. Four dose metrics 

6742 were examined as potential metrics for the internal dose of interest: rate of hepatic 

6743 metabolism through the CYP pathway, rate of hepatic metabolism through the GST pathway, 

6744 the combined rate of hepatic metabolism through the CYP and GST pathways, and the 

6745 concentration (area under the curve, AUC) of dichloromethane, the parent compound, in the 

6746 liver. The dose-response patterns for each of these metrics in the oral study in rats (Serota et 

6747 al., 1986a) and in two inhalation studies in rats (Nitscke et al., 1988a; Burek et al., 1984) 

6748 were examined for fit and congruence. 

6749 Using the oral exposure data, only one of the seven models, the log-logistic model, 

6750 produced an adequately fit (p > 0.10) for the GST metabolism metric and the 

6751 dichloromethane AUC metrics. Adequate model fit was seen in all of the models using the 

6752 CYP dose metric with the oral data, and using the GST, CYP, and AUC dose metrics for the 

6753 inhalation data. 

6754 A limitation in using the GST metric can be observed when comparing the oral and 

6755 inhalation responses at various exposure levels. At 200 ppm, where the GST metric is 

6756 predicted by the PBTK model to be 93 mg metabolism/L liver/day, no liver effects were 

6757 seen. In contrast, liver responses were elevated at an oral dose of 50 mg/kg-day, where the 

6758 GST metric is predicted to be 60 mg metabolism/L liver/day (see Tables 5-1 and 5-5, 

6759 respectively, for the oral and inhalation internal metrics). Thus the liver GST metric 

6760 produces an inconsistency in the dose-response relationship, with very different responses 

6761 observed depending on the route of exposure. A similar inconsistency occurs with the AUC 

6762 metric. These differences are not observed, however, when using the CYP metric. At the 

6763 200 ppm inhalation exposure, where no hepatoxicity was observed, the CYP metric is 

6764 predicted to be 660 mg/L liver/day. This internal CYP metabolism metric is less than that 

6765 predicted for the oral dose for the 50 mg/kg-day group (i.e., 872 mg metabolism/L liver/day), 

6766 in which liver effects were observed. Thus, the CYP internal metric is consistent with the 

6767 observed responses seen in the oral and inhalation exposure studies. 

6768 The GST metabolism and the AUC dose metrics did not present reasonable choices 

6769 based on model fit and consistency of response across studies at comparable dose levels. 

6770 Given these results, the combination of hepatic metabolism through the GST and the CYP 

6771 pathways would not be expected to result in an improvement to a metric based only on CYP 

6772 metabolism. The CYP-metabolism dose metric is the most consistent with the data. This 

6773 metric was selected for the subsequent RfD and RfC derivations. The lack of information on 
224

6774 mechanisms with respect to noncancer health effects represents data gaps in the 

6775 understanding of the health effects of dichloromethane. 

6776 

6777 5.1.4. Methods of Analysis—Including Models (PBTK, BMD, etc.) 

6778 PBTK models for dichloromethane in rats were described previously in section 3.5. 

6779 From the evaluation described in Appendix C, a modified model of Andersen et al. (1991) 

6780 was selected for the calculation of internal dosimetry of ingested dichloromethane in the rats 

6781 in the principal study (Serota et al., 1986a). 

6782 PBTK model simulations of the drinking water study of Serota et al. (1986a) (Table 

6783 5-1) were performed to calculate average lifetime daily internal liver doses in male and 

6784 female F344 rats. In the absence of data for group- and sex-specific BWs, reference values 

6785 were used for male and female F344 rats in chronic studies (U.S. EPA, 1988a). The mode of 

6786 action by which dichloromethane induces nonneoplastic liver effects in rodents has not 

6787 received research attention to determine the role of the parent material, metabolites of the 

6788 CYP2E1 pathway, metabolites of the GST pathway, or some combination of parent material 

6789 and metabolites. In the absence of this kind of knowledge, and considering the pattern of 

6790 response seen in the oral and inhalation studies (as described in section 5.1.3.), an internal 

6791 dose metric based on the amount of dichloromethane metabolized via the CYP pathway in 

6792 the liver (mg dichloromethane metabolized via CYP pathway per liter liver per day) was 

6793 used. Figure 5-3 shows the comparison between oral external and internal doses, using this 

6794 dose metric for the rat and for the human. 
225

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 (Serota et al., 1986a)

Rat internal liver doseb

Sex 

Nominal 

(actual) 

daily intake 

(mg/kg-day) 

Rat liver 

lesion incidencea 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%)c872.7 58.8 931.4 13.1

125 (125) 34/35 (97%)c1,433.1 236.0 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%)c977.8 66.0 1,043.8 12.6

125 (136) 34/38 (89%)c1577.0 258.7 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 affected divided by total sample size. 

b

Internal doses were estimated using a rat PBTK model 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 dichloromethane 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 dichloromethane*hrs)/L tissue.

c

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

Source: Serota et al., 1986a. 

6797 

6798 
226

6799 

6800 

6801 

6802 Figure 5-3. PBTK model-derived internal doses (mg dichloromethane metabolized via the 

6803 CYP pathway per liter liver per day) in rats and humans and their associated external 

6804 exposures (mg/kg-day), used for the derivation of RfDs. Six simulated daily drinking water 

6805 episodes are described by Reitz et al. (1997). The human metabolism rates were estimated 

6806 using a computational sample of 1000 individuals per dose, including random samples of the 

6807 three GST-T1 polymorphisms (+/+, +/-, -/-) in the current U.S. population based on data from 

6808 Haber et al. (2002). Since a different set of samples was used for each dose, some stochasticity 

6809 is evident as the human points (values) do not fall on smooth curves.

6810 

6811 The seven dichotomous dose-response models in BMDS version 2.0 were fit to the rat 

6812 liver lesion incidence data and PBTK model-derived internal dose data to derive a rat internal 

6813 BMD10, and corresponding BMDL10, associated with 10% extra risk (Table 5-2). The quantal 

6814 model is identical to the one-stage multistage model and so is not included in this set of models. 

6815 A BMR of 10% was selected because, in the absence of information regarding the magnitude of 

6816 change in a response that is thought to be minimally biologically significant, a BMR of 10% is 

6817 generally recommended, since it provides a consistent basis of comparison across assessments. 

6818 There are no additional data to suggest that the critical response has a greater sensitivity that 

6819 would warrant a lower BMR. The male rats exhibited a greater sensitivity compared to the 

6820 female rats (based on lower BMDL10 values for all of the models examined) and thus the male 

6821 data are used as the basis for the RfD derivation. The logistic model was the best fitting model 

6822 for the male incidence data, based on AIC value among models with adequate fit (U.S. 

6823 EPA,2000b). Modeling results are shown in detail in Appendix D-1). (mg DCM metabolized/L liver/day)
227

6824 

Table 5-2. BMD modeling results for incidence of noncancer liver lesions 

in male and female F344 rats exposed to dichloromethane in drinking 

water for 2 years, based liver-specific CYP metabolism dose metric (mg 

dichloromethane metabolism via CYP pathway per liter liver tissue per 

day)

Sex and modela BMD10 BMDL10 

χ

2 

goodness of fit 

p-value AIC 

Males 

Gammaa 151.73 48.93 0.62 185.33 

Logistic 85.17 61.78 0.75 183.61 

Log-logistica 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-probita 197.65 77.56 0.81 184.84 

Weibulla 117.29 48.39 0.57 185.49 

Females 

Gammaa

336.38 98.70 0.52 233.07 

Logistic 169.77 134.87 0.59 231.70 

Log-logistica

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-probita

400.95 173.57 0.60 232.80 

Weibulla

283.24 97.31 0.47 233.27 

aThese models in EPA BMDS version 2.0 were fit to the 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 restrict 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 . 

6826 The BMDL10 from the logistic model was used as the point of departure for the RfD 

6827 calculations (Table 5-3). This rat internal dose metric for noncancer effects was adjusted to 

6828 obtain a human-equivalent internal BMDL10 by dividing by a pharmacokinetic scaling factor 

based on a ratio of BWs (BWhuman/BWrat)

0.25 6829 = 4.09). This scaling factor was used because the

6830 metric is a rate of metabolism, rather than the concentration of putative toxic metabolites, and the 

6831 clearance of these metabolites may be slower per volume tissue in the human compared with the 

rat (that is, total rate of removal may scale as BW0.75, while tissue volume scales as BW1

6832 ). 

6833 The human PBTK model (adapted from David et al. [2006], as described in Appendix B), 

6834 using Monte Carlo sampling techniques, was used to calculate quantiles of human equivalent 

6835 administered oral daily doses (in mg/kg-day) associated with the internal BMDL10 values 
228

6836 (Table 5-3), as described above in section 5.1.2. The human model used parameter values 

6837 derived from Monte Carlo sampling of probability distributions for each parameter, including 

6838 MCMC-derived distributions for the metabolic parameters for the metabolism through the 

6839 CYP2E1 pathway (Vmax and Km) and a distribution of GST metabolic rate constants that is 

weighted to reflect the estimated frequency of GST-T1 genotypes (20% GST-T1–/– 6840 , 48% GSTT1+/–, and 32% GST-T1+/+ 6841 ) in the current U.S. population, based on data from Haber et al. 

6842 (2002). All simulations also included a distribution of CYP activity based on data from 

6843 Lipscomb et al. (2003). The drinking water exposures were comprised of six discrete drinking 

6844 water episodes for specified times and percentages of total daily intake (Reitz et al., 1997). The 

6845 mean and two lower points on the distributions of human equivalent administered daily doses 

derived from the Serota et al. (1986a) data for male rats, using the BMDL10 6846 from the logistic 

6847 model , are shown in Table 5-3. 

6848 

Table 5-3. RfD for dichloromethane based on PBTK model-derived 

probability distributions of human drinking water exposures 

extrapolated from nonneoplastic liver lesion incidence data for male 

rats exposed via drinking water for 2 years, based on liver-specific 

CYP metabolism dose metric (mg dichloromethane metabolized via 

CYP pathway per liter liver tissue per day) 

Human equivalent dose 

(mg/kg-day)d

Modela

Rat 

 internal

BMDL10

b

Human 

internal 

BMDL10

c

1

st

 

percentile

5

th 

percentile Mean 

Human 

 RfD 

(mg/m3)

e

Logistic 61.78 15.11 0.214 0.252 0.395 7 x 10-3

aBased on the best-fitting model from Table 5-2. 

bRat dichloromethane PBTK model-derived internal liver dose associated with 

the lower bound on 10% extra risk for developing liver foci/areas of cellular 

alteration. 

cHuman dichloromethane internal liver dose, derived by dividing the rat internal 

BMDL10 by a scaling factor of 4.09 [(BWhuman/BWrat)

0.25 ] to account for potential 

interspecies pharmacokinetic differences in the clearance of metabolites. 

d

PBTK model-derived distributions of daily average dichloromethane drinking 

water doses predicted by the PBTK model to yield an internal dose in humans 

equal to the dichloromethane internal BMDL10, 

eHuman RfD, based on male rat data, derived by dividing the 1st percentile of 

human equivalent dose value by a total UF of 30: 3 (100.5) for possible 

toxicodynamic differences between species, 3 (100.5) for variability in human 

toxicodynamic response, and 3 (100.5) for database deficiencies. The 1st

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). 

6849 
229

6850 

6851 5.1.5. RfD Derivation—Including Application of Uncertainty Factors (UFs) 

The 1st 6852 percentile point of departure is a stable estimate of the lower end of the 

6853 distribution. Use of this value associated with a sensitive human population addresses the 

6854 uncertainty associated with human toxicokinetic variability. To derive the candidate RfD based 

6855 on data from male rats, the first percentile value of the distribution of human equivalent 

6856 administered dose associated with the male rat-derived BMDL10 was divided by a composite UF 

of 30 (3 [100.5] to account for uncertainty about interspecies toxicodynamic equivalence, 3 [100.5 6857 ] 

to account for uncertainty about toxicodynamic variability in humans, and 3 [100.5 6858 ] for database 

deficiencies) (Table 5-3). The resulting RfD recommended for dichloromethane is 7 × 10–3 6859 

6860 mg/kg-day. 

6861 In deriving this RfD, factors for the following areas of uncertainty were considered: 

6862 • Uncertainty in extrapolating from laboratory animals to humans (UFA). The use of 

6863 PBTK models to extrapolate internal doses from rats to humans reduces toxicokinetic 

6864 uncertainty in extrapolating from the rat liver lesion data but does not account for the 

6865 possibility that humans may be more sensitive than rats to dichloromethane due to 

toxicodynamic differences. A UF of 3 (100.5 6866 ) to account for this toxicodynamic 

6867 uncertainty was used, as shown in Table 5-3. 

6868 

6869 • Uncertainty about variation from average humans to sensitive humans (UFH). The 

6870 probabilistic human PBTK model used in this assessment incorporates the best available 

6871 information about variability in toxicokinetic disposition of dichloromethane in humans 

6872 but does not account for humans who may be sensitive due to toxicodynamic factors. 

Thus, a UF of 3 (100.5 6873 ) was applied to account for possible toxicodynamic differences in 

6874 sensitive humans. 

6875 

6876 • Uncertainty in extrapolating from LOAELs to NOAELs (UFL). A UF for extrapolation 

6877 from a LOAEL to a NOAEL was not applied because BMD modeling was used to 

6878 determine the POD, and this factor was addressed as one of the considerations in 

6879 selecting the BMR. The BMR was selected based on the assumption that it represents a 

6880 minimum biologically significant change. 

6881 

6882 • Uncertainty in extrapolating from subchronic to chronic durations (UFS). The derived 

6883 RfD is based on results from a chronic-duration drinking water toxicity study. No cross6884 duration UF is necessary. 

6885 

6886 • Uncertainty reflecting incompleteness of the overall database (UFD). The oral database 

6887 for dichloromethane includes well-conducted lifetime drinking water studies in rats 
230

6888 (Serota et al., 1986a) and mice (Serota et al., 1986b) and a supporting subchronic study in 

6889 rats and mice (Kirschman et al., 1986). These studies provided dose-response data for 

6890 the hepatic effects of dichloromethane. The database also includes one-generation oral 

6891 reproductive toxicity (General Electric Co., 1976) and developmental toxicity (Narotsky 

6892 and Kavlock, 1995) studies that found no reproductive or developmental effects at dose 

6893 levels in the range of doses associated with liver lesions. A two-generation oral exposure 

6894 study is not available; however, a two-generation inhalation exposure study by Nitschke 

6895 et al. (1988a) reported no effect on fertility index, litter size, neonatal survival, growth 

6896 rates, or histopathologic lesions at exposures of ≥100 ppm. However, there have been no 

6897 oral exposure studies that evaluated neurobehavioral effects in offspring. This is a 

6898 relevant endpoint given the increase in blood CO (a known developmental neurotoxicant) 

6899 that occurs through the CYP2E1 metabolic pathway for dichloromethane after oral and 

6900 inhalation exposures. There are no oral exposure studies that include functional immune 

6901 assays; however, there is a 4-week inhalation study of potential systemic immunotoxicity 

6902 that found no effect of dichloromethane exposure at concentrations up to 5,000 ppm on 

6903 the antibody response to sheep red blood cells (Warbrick et al., 2003). The Warbrick et 

6904 al. (2003) data suggest that systemic immunosuppression is not a concern for 

6905 dichloromethane exposure. Route-specific local immunosuppression from acute 

6906 inhalation exposure in CD1 mice was seen in Aryani et al. (1986). The findings from 

6907 Aryani et al. (1986) were considered to be portal-of-entry effects involving local 

6908 immunosuppression within the lung (Streptococcus and Klebsiella infectivity models) 

6909 and unlikely to be observed following oral exposure. Because of concern regarding the 

6910 adequacy of available data pertaining to possible neurodevelopmental toxicity and the 

6911 lack of a two-generation reproductive study, a UFD of 3 was applied.

6912 

6913 5.1.6. Previous RfD Assessment 

6914 The previous IRIS assessment derived an RfD of 0.06 mg/kg-day based on the NOAELs 

6915 of 5.85 and 6.47 mg/kg-day for nonneoplastic liver toxicity (foci/areas of cellular alteration) in 

6916 male and female rats, respectively, in a 2-year drinking water study (Serota et al., 1986a). The 

6917 LOAELs associated with these NOAELs were 52.58 and 58.32 mg/kg-day for males and 

6918 females, respectively. The RfD of 0.06 mg/kg-day was derived by dividing the average NOAEL 

6919 of 6 mg/kg-day (for male and female rats) by a UF of 100 (10 for intraspecies variability and 10 

6920 for interspecies variability). 

6921 

6922 5.1.7. RfD Comparison Information 

 Use of the mean value (3.95 ×10–1 6923 mg/kg-day) of the human equivalent administered 

dose distribution instead of the 1st percentile, with an additional UF of 3 (100.5 6924 ) to account for 

human toxicokinetic variability, would yield an RfD of 4 × 10–3 6925 mg/kg-day. 
231

6926 Additional comparisons between the derived RfD and values developed from other 

6927 endpoints or data sets using NOAEL/LOAEL methods are shown in Table 5-4 and Figure 5-4. 

6928 NOAELs were used as comparison points of departure and were not scaled allometrically. The 

6929 point of departure for three endpoints (Serota et al., 1986a; Moser et al., 1995; Narotsky and 

6930 Kavlock) are presented in Table 5-4. For Serota et al. (1986a), this is based on BMD modeling 

6931 of a 10% increase in liver lesions using internal liver dose metric (mg dichloromethane 

6932 metabolism via CYP pathway per liter liver tissue per day) derived from a rat PBTK model. 

6933 After an allometric scaling factor of 4.09 was applied, the human internal BMDL10 was 15.11 

6934 mg/kg-day. A probabilistic human PBTK model adapted from David et al. (2006) was used to 

6935 generate a distribution of human equivalent doses from the human internal BMDL10 and the first 

6936 percentile of this distribution was used as the point of departure. For other studies, POD is based 

6937 on the lowest non-control dose in which no effect was seen. These NOAELs that were used as 

6938 points of departure and were not scaled allometrically
232

6939 

Table 5-4. Potential points of departure with applied UFs and resulting candidate RfDs 

Uncertainty Factors Appliedb

Endpoint 

PODa

(mg/kg-day) 

POD Type and 

Description 

Total 

UF UFA UFH 

UFL

UFS UFD 

Candi

RfD 

(mg/kg

Nonneoplastic liver foci, male rats 

61.78 

BMD; 10% increase in incidence of liver lesion 

30 3 3 1 1 3 7 x 1

Neurological changes (FOB), female rats 

101 

NOAEL; No effect at POD, approximate doubling of severity score of neuromuscular and sensorimotor domains 

3,000 10 10 1 10 3 3.4 × 

Maternal weight gain, female rats 338 

NOAEL; No effect at POD, approximate 33% decrease in weight gain seen at next dose 

300 10 10 1 1 3 1.1 

aPOD = point of departure. . 

bUFA = uncertainty in extrapolating from laboratory animals to humans, UFH = uncertainty about variation from average humans to se

uncertainty about extrapolating from LOAEL to NOAEL, UFS = uncertainty in extrapolating from subchronic to chronic durations, 

reflecting incompleteness of the overall database. A UF for extrapolation from a LOAEL to NOAEL (UFL) was not used for any of 

Serota et al., (1986a) study, the use of the first percentile of the human equivalent dose distribution as the point of departure replace

human toxicokinetic variability.

Bolded value is the basis for the RfD of 7 × 10–3 mg/kg-day. 
233

6940 6941 

6942 Figure 5-4. Comparison of candidate RfDs derived from selected point of departures for endpoints 

6943 presented in Table 5-4.6944 
234

6945 5.2. INHALATION REFERENCE CONCENTRATION (RfC) 

6946 5.2.1. Choice of Principal Study and Critical Effect—with Rationale and Justification 

6947 Figure 5-5 includes exposure-response arrays from some of the human studies that were 

6948 evaluated for use in the derivation of the RfC. Several acute-duration controlled exposure 

6949 studies (section 4.1.2.2) and cross-sectional occupational studies (sections 4.1.2.3 and 4.1.2.4) in 

6950 humans are available that show neurological effects from dichloromethane exposure. These 

6951 effects include an increase in prevalence of neurological symptoms among workers (Cherry et 

6952 al., 1981) and possible detriments in attention and reaction time in complex tasks among retired 

6953 workers (Lash et al., 1991). However, these studies have inadequate power for the detection of 

6954 effects with an acceptable level of precision. In addition, the Cherry et al. (1981) study is limited

6955 by the definition and documentation of neurological symptom history, and the Lash et al. (1991) 

6956 study has exposure measurements from 1974–1986, but the work histories of exposed workers 

6957 go back to the 1940s. Ott et al. (1983c) reported an increase in serum bilirubin among exposed 

6958 workers, but there was no association seen with respect to the other hepatic enzymes examined 

6959 (serum γ-glutamyl transferase, serum AST, serum ALT), and no evidence of hepatic effects was 

6960 seen in a later study of the same cohort (Soden, 1993). Because of these limitations, these 

6961 human studies of chronic exposures do not serve as an adequate basis for RfC derivation. As 

6962 discussed in section 5.2.6, however, the quantitative measures of neurological function from 

6963 Cherry et al. (1983) were used to derive a comparative RfC. 

6964 The database of experimental animal dichloromethane inhalation studies includes 

6965 numerous 90-day and 2-year studies, with data on hepatic, pulmonary, and neurological effects, 

6966 (see Table 4-36) and reproductive and developmental studies (Table 4-37) (see summary in 

6967 Section 4.6.2). NOAELs, LOAELs, and the dose range tested corresponding to selected health 

6968 effects from the chronic studies are shown in Figure 5-5, and effects seen in subchronic, 

6969 reproductive, and developmental studies are shown in Figure 5-6. The subchronic (i.e., 90-day 

6970 or less study) data were not considered in the selection of a principal study for deriving the RfC 

6971 because the database contains reliable dose-response data from the chronic study at lower doses 

6972 than the 90-day study. The data from the subchronic studies are, however, used to corroborate 

6973 the findings with respect to relevant endpoints (i.e., hepatic and neurological effects). 
235

1

10

100

1000

10000

Hepatocytevacuolation, SpragueDawley rat (Nitschke et al., 1988a)

Hepatocytenecrosis, SpragueDawley rat (Burek et al., 1984)

Hepatocytevacuolation, male and female SpragueDawley rat (Burek et al., 1984)

Hepatocytevacuolation, necrosis, hemosiderosisin liver, male and female F344 rat (Mennear et al., 1988; NTP, 1986)

Renal tubular degeneration, male and female F344 rat (Mennearet al., 1988; NTP, 1986)

Hepatocytedegeneration, male and female B6C3F1 mice (Mennear et al., 1988; NTP, 1986)

Renal tubule casts, mlae and female B6C3F1 mice (Mennear et al., 1988; NTP, 1986)

Changes in CNS measures, males (Lash et al., 1991)

Chronic neurological effects, males (Cherry et al., 1983)

Cardiac –ST segment depression, males (Ottet al., 1983c)

NOAEL LOAEL

The vertical lines = range of exposures in study.Closed dots ( ) = exposure concentrations used in study

RAT MOUSE HUMAN

Concentration (ppm)

6974 6975 6976 

6977 Figure 5-5. Exposure response array for chronic (animal) or occupational (human) inhalation expo6978 dichloromethane (log Y axis) 
236

6979 6980 

6981 Figure 5-6. Exposure response array for subacute to subchronic inhalation exposure to dichlorome6982 axis)
237

6983 Hepatic effects (hepatic vacuolation and necrosis, hemosiderosis, hepatocyte 

6984 degeneration) are the critical dose-dependent noncancer effects associated with inhalation 

6985 exposure to dichloromethane. These effects were seen in mice (Mennear et al., 1988; NTP, 

6986 1986) and rats (Mennear et al., 1988; Nitschke et al., 1988a; NTP, 1986; Burek et al., 1984), but 

6987 not in Syrian golden hamsters (Burek et al., 1984). Inhalation bioassays with Sprague-Dawley 

6988 rats identified the lowest inhalation LOAEL for nonneoplastic liver lesions in the database: 500 

6989 ppm (6 hours/day, 5 days/week for 2 years) (Nitschke et al., 1988a; Burek et al., 1984), and 

6990 Nitschke et al. (1988a) identified a NOAEL of 200 ppm in female rats. Based on the results 

6991 reviewed above, nonneoplastic liver lesions (specifically, hepatic vacuolation) in rats are 

6992 identified as the critical noncancer effect from chronic dichloromethane inhalation in animals. 

6993 Because Nitschke et al. (1988a) examined a range of exposures that included doses at the low 

6994 end of the rangecompared with the range examined in Burek et al. (1984), the former study was 

6995 selected as the principal study for derivation of a chronic inhalation RfC. 

6996 Reproductive performance (e.g., as assessed by number of litters, resorption rate, fetal 

6997 survival, and growth) was not affected in two generations of F344 rats exposed to up to 

6998 1,500 ppm for 14 or 17 weeks before mating of the F0 and F1 generations, respectively 

6999 (Nitschke et al., 1988b) or in a study of Swiss-Webster mice or Sprague-Dawley rats exposed to 

7000 1,250 ppm on GDs 6–15 (Schwetz et al., 1975). A decrease in fertility index was seen in the 150 

7001 and 200 ppm groups in a study of male Swiss-Webster mice exposed via inhalation for 6 weeks 

7002 prior to mating (Raje et al., 1988), but the statistical significance of this effect varied 

7003 considerably depending on the statistical test used in this analysis. Two types of developmental 

7004 effects, decreased offspring weight at birth and changed behavioral habituation of the offspring 

7005 to novel environments, were seen in Long-Evans rats following exposure to 4,500 ppm for 

7006 14 days prior to mating and during gestation (or during gestation alone) (Bornschein et al., 1980; 

7007 Hardin and Manson, 1980). This dose was the only exposure dose used in this study. Schwetz et 

7008 al. (1975) did not observe an adverse effect on gross development or soft tissue abnormalities in 

7009 a study involving exposure to 1,250 ppm on GD 6 in Swiss-Webster mice or Sprague-Dawley 

7010 rats, but an increase in delayed ossification of the sternebrae was seen. 

7011 Neurological impairment was not seen in lifetime rodent bioassays involving exposure to 

7012 airborne dichloromethane concentrations of ≤2,000 ppm in F344 rats (Mennear et al., 1988; 

7013 NTP, 1986), ≤3,500 ppm in Sprague-Dawley rats (Nitschke et al., 1988a; Burek et al., 1984), or 

7014 ≤4,000 ppm in B6C3F1 mice (Mennear et al., 1988; NTP, 1986). It should be noted, however, 

7015 that these studies did not include standardized neurological or neurobehavioral testing. The sole 

7016 subchronic or chronic study in which neurobehavioral batteries were utilized found no effects in 

7017 an observational battery, a test of hind-limb grip strength, a battery of evoked potentials, or 

7018 brain, spinal cord, or peripheral nerve histology in F344 rats exposed to concentrations up to 

7019 2,000 ppm for 13 weeks, with the tests performed beginning 65 hours after the last exposure 

7020 (Mattsson et al., 1990). 
238

7021 Other effects associated with lifetime inhalation exposure to dichloromethane include 

7022 renal tubular degeneration in F344 rats exposed to ≥2,000 ppm, testicular atrophy in male 

7023 B6C3F1 mice exposed to 4,000 ppm, and ovarian atrophy in female B6C3F1 mice exposed to 

7024 ≥2,000 ppm (Mennear et al., 1988; NTP, 1986). No effects on histologic, clinical chemistry, 

7025 urinalysis, or hematologic variables were found in Syrian golden hamsters exposed to 

7026 concentrations up to 3,500 ppm for 2 years, with the exception that the mean COHb percentage 

7027 of exposed hamsters was about 30%, compared with values of about 3% in controls (Burek et al., 

7028 1984). 

7029 

7030 5.2.2. Derivation Process for Reference Concentration Values 

7031 The derivation process used for the RfC parallels the process described in section 5.1.2 

7032 on the RfD derivation; consideration of dose metrics was described in section 5.1.3. As was 

7033 noted in the RfD discussion, the mechanistic issues with respect to noncancer health effects 

7034 represents data gaps in the understanding of the health effects of dichloromethane. 

7035 

7036 5.2.3. Methods of Analysis—Including Models (PBTK, BMD, etc.)

7037 The modified rat PBTK model of Andersen et al. (1991), described in Appendix C and 

7038 also used in the derivation of the RfD (Figure 5-2), was used for calculating internal dosimetry of 

7039 inhaled dichloromethane in Sprague-Dawley rats. Simulations of 6 hours/day, 5 days/week 

7040 inhalation exposures used in the Nitschke et al. (1988a) study were performed to calculate 

7041 average daily internal liver doses (Table 5-5). In the absence of data for group- and sex-specific 

7042 BWs, reference values for male and female Sprague-Dawley rats in chronic studies were used 

7043 (U.S. EPA, 1988a). 
239

7044 

Table 5-5. Incidence data for nonneoplastic liver lesions (hepatic vacuolation) and 

internal liver doses, based on various metrics, in female Sprague-Dawley rats exposed to 

dichloromethane via inhalation for 2 years (Nitschke et al., 1988a)

Rat internal liver doseb

Sex 

Exposure 

(ppm) 

Liver lesion 

incidencea 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 were 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 

aNumber affected divided by total sample size. 

b

Internal doses were estimated using a rat PBTK model using exposures reported by study authors (50 ppm = 

174 mg/m3

, 200 ppm = 695 mg/m3, and 500 ppm = 1737 mg/m3) 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 dichloromethane 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 dichloromethane*hrs)/L tissue.

c

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

Source: Nitschke et al., 1988a. 

7045 

7046 As described in section 5.1.2, the internal dose metric used was based on total hepatic 

7047 metabolism through the CYP2E1 pathway (mg dichloromethane metabolized via CYP 

7048 pathway/L liver/day). Figure 5-7 shows the comparison between inhalation external and internal 

7049 doses, using this dose metric for the rat and the human. 
240

7050 

7051 

7052 

7053 

7054 Figure 5-7. PBTK model-derived internal doses (mg dichloromethane metabolized 

7055 via the CYP pathway/L liver/day) in rats and humans versus external exposures 

7056 (ppm). Average daily doses were calculated from simulated rat exposures of 6 hours/day, 5 

7057 days/week, while simulated human exposures were continuous. The human metabolism rates 

7058 were estimated using a computational sample of 1000 individuals per dose, including random 

7059 samples of the three GST-T1 polymorphisms (+/+, +/-, -/-) in the current U.S. population based 

7060 on data from Haber et al. (2002). Since a different set of samples was used for each dose, some 

7061 stochasticity is evident as the human points (values) do not fall on smooth curves. 

7062 

7063 The seven dichotomous dose-response models available in EPA BMDS version 2.0 were 

7064 fit to the female rat liver lesion incidence of Nitschke et al. (1988a) and PBTK model-derived 

7065 internal dose data to derive rat internal BMD10 and the associated BMDL10 values (Table 5-6). 

7066 The quantal model is identical to the one-stage multistage model; therefore, it is not included in 

7067 this set of models. A BMR of 10% was selected because, in the absence of information 

7068 regarding the magnitude of change in a response that is thought to be minimally biologically 

7069 significant, a BMR of 10% is generally recommended, as it provides a consistent basis of 

7070 comparison across assessments. There are no additional data to suggest that the critical response 

7071 has a greater sensitivity that would warrant a lower BMR. The log-probit model was the best 

7072 fitting model for the female incidence data, based on AIC value among models with adequate fit. 

7073 Modeling results are shown in detail in Appendix D-2). 

7074 

(mg DCM metabolized/L liver/day)
241

7075 

Table 5-6. BMD modeling results for incidence of noncancer liver 

lesions in female Sprague-Dawley rats exposed to dichloromethane by 

inhalation for 2 years, based on liver specific CYP metabolism metric 

(mg dichloromethane metabolized via CYP pathway per liter liver 

tissue per day) 

Modela

 BMD10 BMDL10 

χ

2 

goodness of fit 

p-value AIC 

Gammaa 614.27 225.96 0.48 367.22 

Logistic 274.58 150.43 0.14 369.77 

Log-logistica 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-probita 728.96 523.94 0.98 365.82 

Weibulla 706.45 487.45 0.95 365.87 

aThese models in EPA BMDS version 2.0 were fit to the 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 slope >1, 

multistage model restrict betas ≥0; lowest degree polynomial with an adequate fit reported 

(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 internal BMDL10 7078 was obtained by 

7079 dividing this rat internal dose metric by a pharmacokinetic scaling factor based on a ratio of BWs 

7080 (scaling factor = 4.09) (Table 5-7). This scaling factor was used because the metric is a rate of 

7081 metabolism rather than the concentration of putative toxic metabolites, and the clearance of these 

7082 metabolites may be slower per volume tissue in the human compared with the rat. A probabilistic 

7083 PBTK model for dichloromethane in humans, adapted from the model of David et al. (2006) as 

7084 described in Appendix B, was then used with Monte Carlo sampling to calculate distributions of 

chronic HECs (in units of mg/m37085 ) associated with the internal BMDL10, based on the responses 

7086 in female Sprague-Dawley rats. Estimated mean, first, and fifth percentiles of this distribution 

7087 are shown in Table 5-7. 
242

7088 

Table 5-7. Inhalation RfC for dichloromethane based on PBTK modelderived probability distributions of human inhalation exposure 

extrapolated from nonneoplastic liver lesion data for female rats exposed 

via inhalation for 2 years, based on liver-specific CYP metabolism dose 

metric (mg dichloromethane metabolized via CYP pathway per liter liver 

tissue per day) 

HEC (mg/m3)

d

Modela

Rat 

internal

BMDL10

b

Human 

internal 

BMDL10

c

1

st

 

percentile

5

th 

percentile Mean 

Human RfC 

(mg/m3)

e

Log-probit 523.94 128.10 16.63 20.89 47.36 0.2

aBased on the best-fitting model from Table 5-6. 

bRat dichloromethane PBTK model-derived internal liver dose associated with lower bound 

on 10% extra risk for developing hepatocyte vacuolation. 

cHuman dichloromethane internal liver dose, derived by dividing the rat internal BMDL10 by 

a scaling factor of 4.09 [(BWhuman/BWrat)

0.25 ] to account for potential interspecies 

pharmacokinetic differences in the clearance of metabolites. 

d

PBTK model-derived distributions of long-term, daily average airborne dichloromethane 

concentrations predicted by the PBTK model to yield an internal dose in humans equal to the 

dichloromethane internal BMDL10. 

eHuman candidate RfC, based on female rat data, derived by dividing the 1st percentile of 

HEC values by a total UF of 100: 3 (100.5) for possible toxicodynamic differences between 

species, 3 (100.5) for variability in human toxicodynamic response, and 10 for database 

deficiencies. The 1st 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 

7091 5.2.4. RfC Derivation—Including Application of Uncertainty Factors (UFs) 

 The 1st 7092 percentile point of departure is a stable estimate of the lower end of the 

7093 distribution. Use of this value associated with a sensitive human population addresses the 

7094 uncertainty associated with human toxicokinetic variability. The RfC was calculated by dividing 

the first percentile of the HEC distribution in Table 5-7 by a composite UF of 100 (3 [100.5 7095 ] to 

account for uncertainty about interspecies toxicodynamic equivalence, 3 [100.5 7096 ] to account for 

7097 uncertainty about toxicodynamic variability in humans, and 10 for database deficiencies). The 

resulting RfC was 0.2 mg/m37098 based on liver lesions in female Sprague-Dawley rats in Nitschke et 

7099 al. (1988a). In deriving this RfC, factors for the following areas of uncertainty were considered: 

7100 

7101 • Uncertainty in extrapolating from laboratory animals to humans (UFA). The use of 

7102 PBTK models to extrapolate internal doses from rats to humans reduces toxicokinetic 

7103 uncertainty in extrapolating from the rat liver lesion data but does not account for the 

7104 possibility that humans may be more sensitive than rats to dichloromethane due to 

toxicodynamic differences. A UF of 3 (100.5 7105 ) to account for this toxicodynamic 

7106 uncertainty was applied, as shown previously in Table 5-7. 

7107 

7108 • Uncertainty about variation in human toxicokinetics (UFH). The probabilistic human 

7109 PBTK model used in this assessment incorporates the best available information about 

7110 variability in toxicokinetic disposition of dichloromethane in humans but does not 

7111 account for humans who may be sensitive due to toxicodynamic factors. Thus, a UF of 

3 (100.5 7112 ) was applied to account for possible toxicodynamic differences in sensitive 

7113 humans. 

7114 

7115 Uncertainty in extrapolating from LOAELs to NOAEL (UFL). A UF for extrapolation from a 

7116 LOAEL to a NOAEL was not applied because BMD modeling was used to determine the POD, 

7117 and this factor was addressed as one of the considerations in selecting the BMR. The BMR was 

7118 selected based on the assumption that it represents a minimum biologically significant change. 

7119 • Uncertainty in extrapolating from subchronic to chronic durations (UFS). The derived 

7120 RfD is based on results from a chronic-duration drinking water toxicity study. No cross7121 duration UF is necessary. 

7122 

7123 • Uncertainty reflecting incompleteness of the overall database (UFD). A UF of 10 was 

7124 selected to address the deficiencies in the dichloromethane toxicity database. The 

7125 inhalation database for dichloromethane includes several well-conducted chronic 

7126 inhalation studies. In these chronic exposure studies, the liver was identified as the most 

7127 sensitive noncancer target organ in rats (Nitschke et al., 1988a; NTP, 1986; Burek et al., 
244

7128 1984). The critical effect of hepatocyte vacuolation was corroborated in the two principal 

7129 studies (Nitschke et al., 1988a; Burek et al., 1984), which identified 500 ppm as the 

7130 lowest inhalation LOAEL for noncancer liver lesions. Gross signs of neurologic 

7131 impairment were not seen in lifetime rodent inhalation bioassays for dichloromethane at 

7132 exposure levels up to 4,000 ppm (see section 4.2.2.2 for references), and no exposure7133 related effects were observed in an observational battery, a test of hind-limb grip 

7134 strength, a battery of evoked potentials, or histologic examinations of nervous tissues in 

7135 F344 rats exposed to dichloromethane concentrations as high as 2,000 ppm (Mattson et 

7136 al., 1990). A two-generation reproductive study in F344 rats reported no effect on 

7137 fertility index, litter size, neonatal survival, growth rates, or histopathologic lesions at 

7138 exposures ≥100 ppm dichloromethane (Nitschke et al., 1988b). Fertility index (measured 

7139 by number of unexposed females impregnated by exposed males per total number of 

7140 unexposed females mated) was reduced following inhalation exposure of male mice to 

7141 150 and 200 ppm dichloromethane for 2 hours/day for 6 weeks, but the statistical 

7142 significance of this effect varied considerably depending on the statistical test used in this 

7143 analysis. (Raje et al., 1988). The available developmental studies include single-dose 

7144 studies that use relatively high exposure concentrations (1,250 ppm in Schwetz et al. 

7145 [1975]; 4,500 ppm in Hardin and Manson [1980]; and 4,500 ppm in Bornschein et 

7146 al.[1980]). In one of the single-dose studies, decreased offspring weight at birth and 

7147 changed behavioral habituation of the offspring to novel environments were seen 

7148 following exposure of adult Long-Evans rats to 4,500 ppm for 14 days prior to mating 

7149 and during gestation (or during gestation alone) (Bornschein et al., 1980; Hardin and 

7150 Manson, 1980). CO, a known developmental neurotoxicant, is produced through the 

7151 CYP2E1 metabolic pathway for dichloromethane. Schwetz et al. (1975) reported 

7152 increased concentrations (~10% higher compared with controls) in maternal blood COHb 

7153 levels in mice and rats exposed during GDs 6–15. A chronic exposure study in F344 rats 

7154 reported a dose-related increase in blood COHb in females exposed to 50, 200, and 500 

7155 ppm, beginning with the first measure taken after 6 months of exposure (Nitschke et al., 

7156 1988a). The increase was seen at the lowest exposure group (50 ppm). Anders and 

7157 Sunram (1982) reported elevated CO levels in maternal and fetal blood in rats following 

7158 exposure to 500 ppm for 1 hour on GD 21; levels were similar in the maternal and fetal 

7159 samples. Placental transfer of dichloromethane was also seen, although levels were lower 

7160 in the fetus. The results from the single dose developmental toxicity study in rats 

7161 (Bornschein et al., 1980; Hardin and Manson, 1980), in addition to the known increase in 

7162 CO, the placental transfer of dichloromethane, and the relatively high activity of CYP2E1 

7163 in the brain compared to the liver of the developing human fetus (Hines, 2007; Brzezinski 

7164 et al., 1999; Johnsrud et al., 2003), raise uncertainty regarding possible 

7165 neurodevelopmental toxicity from gestational exposure to inhaled dichloromethane. In 
245

7166 addition, Aranyi et al. (1986) demonstrated evidence of immunosuppression following a 

7167 single 100 ppm dichloromethane exposure for three hours in CD-1 mice. This study used 

7168 a functional immune assay that is directly relevant to humans (i.e., increased risk of 

7169 Streptococcal pneumonia-related mortality and decreased clearance of Klebsiella 

7170 bacteria). No effects were seen with 50 ppm exposure for either 1 or 5 days. Systemic 

7171 immunosuppression was not seen in a 4-week, 5,000 ppm inhalation exposure study, 

7172 measuring the antibody response to sheep red blood cells in Sprague-Dawley rats 

7173 (Warbrick et al., 2003). These studies suggest a localized, portal-of-entry effect within 

7174 the lung rather than a systemic immunosuppression. Therefore, in consideration of the 

7175 entire database for dichloromethane, a database UF of 10 was selected. This UF accounts 

7176 for the lack of neurodevelopmental toxicity studies and developmental toxicity studies at 

7177 low doses. 

7178 

7179 5.2.5. Previous RfC Assessment 

7180 No RfC was derived in the previous IRIS assessment. 

7181 

7182 5.2.6. RfC Comparison Information 

7183 A candidate RfC, based on a different approach to accounting for human toxicokinetic 

variability is similar to thederived RfC of 0.2 mg/m

3

7184 . Use of the mean value on the HEC 

distribution (47.36), with an additional UF of 3 (10

0.5 7185 ) to account for human toxicokinetic 

variability, would yield an RfC of 0.2 mg/m37186 . 

7187 For an additional comparison, an RfC was derived based on neurological endpoints from 

7188 human occupational exposures. Cherry et al. (1983) compared 56 exposed and 36 unexposed 

7189 workers at an acetate film manufacturing plant for dichloromethane inhalation exposure, blood 

7190 levels of dichloromethane, subjective self-reporting of general health, and two objective, 

7191 quantitative measurements of neurological function (digit symbol substitution and simple 

7192 reaction time). The exposed and unexposed individuals were matched to within 3 years of age. 

7193 The measured dichloromethane concentrations from personal breathing zone sampling of the 

7194 exposed workers ranged from 28 to 173 ppm. No information on exposure duration was given, 

7195 and Cherry et al. (1983) did not indicate if the exposure measurements were indicative of 

7196 historical exposure levels. There were no significant differences between exposed and 

7197 unexposed workers in subjective or objective measurements collected at the beginning of the 

7198 work shift on a Monday (after 2 nonworking days). Exposed workers showed a slightly slower 

7199 (but not significant) score than the control workers on a reaction time test, but the scores did not 

7200 deteriorate during the shift. These findings suggest that repeated inhalation exposures in the 

7201 range of 28–173 ppm do not result in significant effects, but the actual duration of exposure of 

7202 the workers is uncertain. In the absence of data for the mean exposure levels, the exposure range 

7203 midpoint of 101 ppm serves as a NOAEL for chronic neurological effects from dichloromethane 
246

exposure. Thus, a candidate RfC of 3.5 mg/m37204 was derived by dividing the NOAEL of 351 

mg/m37205 (101 ppm) by a composite UF of 100. A UF of 10 was applied to account for potentially 

7206 susceptible individuals in the absence of quantitative information on the variability of 

7207 neurological response to dichloromethane in the human population. A UF of 10 was applied for 

7208 database deficiencies. The duration of exposures of acetate film workers (Cherry et al., 1983) 

7209 was not reported, and a limited number of endpoints was evaluated. Further, definitive 

7210 neurological batteries were not administered in chronic-duration animal bioassays. 

7211 Another candidate RfC was developed by using the neurological data from the study by 

7212 potential long-term CNS effects in a study of retired aircraft maintenance workers (Lash et al., 

7213 1991). Retired aircraft maintenance workers, ages 55-75 years, employed in at least one of 

7214 14 targeted jobs (e.g., paint strippers) with dichloromethane exposure for 6 or more years 

7215 between 1970 and 1984 (n = 25) were compared to a like group of workers without 

7216 dichloromethane exposure (n = 21). From 1974 to 1986, when 155 measurements for 

7217 dichloromethane exposure were made, mean breathing zone TWAs ranged from 82 to 236 ppm 

7218 and averaged 225 ppm for painters and 100 ppm for mechanics; information on exposure levels 

7219 prior to this time was not provided. The evaluation included several standard neurological tests, 

7220 including physiological measurement of odor and color vision senses, auditory response 

7221 potential, hand grip strength, measures of reaction time (simple, choice, and complex), short7222 term visual memory and visual retention, attention, and spatial ability. The exposed group had a 

7223 higher score on verbal memory tasks (effect size approximately 0.45, p = 0.11) and lower score 

7224 on attention tasks (effect size approximately –0.55, p = 0.08) and complex reaction time (effect 

7225 size approximately –0.40, p = 0.18) compared with the control group. None of these differences 

7226 were statistically significant. Given the sample size, however, the power to detect a statistically 

7227 significant difference between the groups was very low (i.e., approximately 0.30 for an effect 

7228 size of 0.40 using a two-tailed alpha of 0.05) (Cohen, 1987), and these results cannot be taken as 

7229 evidence of no effect. An estimated exposure level from the study can be generated from the 

7230 midpoint value from the exposure range (82–236 ppm; mean = 159 ppm), converted to 552 

mg/m3

7231 . If these results are viewed as a LOAEL and this estimated mean exposure level of 552 

mg/m3 7232 was used, a composite UF of 1,000 would be applied for interspecies toxicodynamics 

7233 (10), extrapolation from a LOAEL to a NOAEL (10), and database uncertainties (10), resulting 

in an RfC of 0.55 mg/m37234 . 

 The value of the candidate RfC based on the data from Cherry et al. (1983), 3.5 mg/m37235 , is 

7236 approximately 15-fold higher, and the value of the candidate RfC based on the data from Lash et 

al. (1991), 0.55 mg/m3

 is approximately three times higher than thederived RfC of 0.2 mg/m37237 , 

7238 based on liver lesions in rats. The animal-derived RfC is preferable to the human-derived RfC 

7239 because of the uncertainties about the exposure durations for the workers in the Cherry et al. 

7240 (1983) study and uncertainties regarding the exposures and effect sizes in Lash et al. (1991), and 

7241 because the RfC based on the rat data is more health protective. 
247

7242 Additional comparisons among the RfC and candidate values developed from other 

7243 endpoints or data sets, using NOAEL/LOAEL methods, are shown in Table 5-8 and Figure 5-8. 
248

Table 5-8. Potential points of departure with applied UFs and resulting candidate RfCs 

UFsc

Endpoint 

Point of 

departure 

(mg/m3) a

POD Type and 

Descriptionb

Total 

UF UFA UFH UFL UFS UFD RfC (mg

Hepatocyte vacuolation, female rat 523 

BMD, 10% increase in incidence of liver lesion 

100 3 3 1 1 10 0.2 

Renal tubular degeneration; NOAEL, male rat 620 NOAEL 1000 3 10 1 1 10 2.07 Reproductive - fertility index; NOAEL, male mouse 

20.7 

No effect at POD, 16% decrease in fertility index seen at LOAEL dose 

300 3 10 1 1 10 0.07

Increased infection susceptibility (mortality risk), female mouse 

15.5 NOAEL 3,000 3 10 1 10 10 0.005 

Increased IgM production, male and female rat 17,366 NOAEL 3,000 3 10 1 10 10 1.03 

Chronic CNS effects, human male 351 NOAEL 100 1 10 1 1 3 3.51 CNS changes, human male 552 LOAEL 1,000 1 10 10 1 3 0.55 

aPOD = point of departure. For Nitschke et al. (1988a), this is based on BMD modeling of a 10% increase in liver lesions using inter

dichloromethane metabolism via CYP pathway per liter liver tissue per day) derived from a rat PBTK model. After an allometric sca

applied, the human internal BMDL10 was 128 mg/m3. A probabilistic human PBTK model adapted from David et al. (2006) was use

of human equivalent concentrations from the human internal BMDL10 and the first percentile of this distribution was used as the poin

rodent studies, the NOAEL or LOAEL concentration, in mg/m3, was adjusted to a continuous exposure taking into account hours per 

exposure. This adjusted exposure was then converted to an HEC by multiplying the value by a dosimetric adjustment factor (DAF). 

coefficients 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 humans, a DAF of 1 is recommended and was used. NOAELs or LOAELs were used as points of departure in human studies since t

already human exposures. 

bExtra risk defined for incidence data as (Incidence1 – Incidence0)/(1-Incidence0), where 1 = dose at observed increased and 0 = backgcUFA = uncertainty in extrapolating from laboratory animals to humans, UFH = uncertainty about variation from average humans to sen

uncertainty about extrapolating from LOAEL to NOAEL, and UFD = uncertainty reflecting incompleteness of the overall database. A 

subchronic 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. 
249

7244 

7245 Figure 5-8. Comparison of candidate RfCs derived from selected point of departures for endpoints 7246 Table 5-8. 
250

7247 5.3. UNCERTAINTIES IN THE ORAL REFERENCE DOSE AND INHALATION 

7248 REFERENCE CONCENTRATION 

7249 Risk assessments need to include a discussion of uncertainties associated with the derived 

7250 toxicity values. For dichloromethane, uncertainties related to inter- and intraspecies differences in

7251 toxicodynamics and database deficiencies are treated quantitatively via the UF approach (U.S. EPA, 

7252 1994b). Uncertainties in the toxicokinetic differences of dichloromethane between species and within 

7253 humans are reduced by application of the PBTK models for rats and humans. These and other areas of 

7254 uncertainty of the derived RfD and RfC are discussed below. 

7255 

7256 Adequacy of database for derivation of RfD and RfC 

7257 As summarized in sections 4.6.1.1 and 4.6.2.1, data from the available human studies on the 

7258 health effects from occupational inhalation exposures provide some, but not conclusive, evidence of 

7259 long-term health consequences of chronic dichloromethane exposure, specifically with respect to 

7260 neurologic and hepatic damage. These data are not adequate for derivation of an RfD or RfC. However,

7261 a broad range of animal toxicology data is available for the hazard assessment of dichloromethane, as 

7262 described in chapter 4. The database of oral (Table 4-35) and inhalation (Tables 4-36 and 4-37) toxicity 

7263 studies includes numerous chronic, subchronic, acute, reproductive, and developmental studies. Liver 

7264 toxicity in multiple rodent species is consistently identified as the most sensitive noncancer effect from 

7265 oral and inhalation exposure to dichloromethane. In addition to the oral and inhalation toxicity data, 

7266 there are numerous studies describing the toxicokinetics of dichloromethane. Consideration of the 

7267 available dose-response data to determine an estimate of oral exposure that is likely to be without an

7268 appreciable risk of adverse noncancer health effects over a lifetime has led to the selection of noncancer 

7269 liver lesions in the 2-year drinking water study in F344 rats (Serota et al., 1986a) as the critical effect 

7270 and principal study for deriving the RfD for dichloromethane. The critical effect selected for the 

7271 derivation of the chronic RfC is also hepatic lesions; two different studies in Sprague-Dawley rats 

7272 (Nitschke et al.,1988a; Burek et al., 1984), spanning overlapping exposures, reported data on hepatic 

7273 vacuolation, and the lower exposure study was chosen as the principal study (Nitschke et al.,1988a). 

7274 A critical data uncertainty was identified for neurodevelopmental effects. Animal bioassays 

7275 have not identified gross or microscopic effects on neural tissues from long-term exposures or single 

7276 (Schwetz et al., 1975) or multigenerational (Nitschke et al., 1988b) developmental toxicity studies. 

7277 However, behavioral changes were observed in pups born to rats exposed to high levels (4,500 ppm) of 

7278 dichloromethane (Bornschein et al., 1980; Hardin and Manson, 1980); lower exposures were not 

7279 examined in this study. Uncertainty exists as to the development of neurological effects from lower 

7280 gestational exposures in animals or humans. In addition, a critical data uncertainty has been identified 

7281 that relates to potential immunotoxicity, specifically immunosuppression seen as a localized portal–of7282 entry effect within the lung with an acute inhalation exposure. The lack of data on immune effects from 
251

7283 longer-term exposure represents a significant data gap and is of particular importance because of the 

7284 potential importance of immunosuppression with respect to response to infections and tumor 

7285 surveillance. The weight of evidence for nonneoplastic effects in humans and animals suggests that the 

7286 development of liver lesions is the most sensitive effect, with a UF applied because of the lack of 

7287 neurodevelopmental studies and, for the RfC, the uncertainty regarding immunotoxicity. 

7288 

7289 Dose-response modeling 

7290 The selection of the BMD model(s) for the quantitation of the RfD and RfC does not lead to 

7291 significant uncertainty in estimating the point of departure. It should be noted, however, that a level of 

7292 uncertainty is inherent given the lack of data in the region of the BMR. 

7293 

7294 Interspecies extrapolation of dosimetry and risk 

7295 The extrapolation of internal dichloromethane dosimetry from nonneoplastic rat responses to 

7296 human risk was accomplished using PBTK models for dichloromethane in rats and humans. 

7297 Uncertainties in rat and human dosimetry used for RfD and RfC derivation can arise from uncertainties 

7298 in the PBTK models to accurately simulate the toxicokinetics of dichloromethane for animals under 

7299 bioassay conditions and humans experiencing relatively low, chronic environmental exposures. 

7300 There is uncertainty associated with the pharmacokinetic data used for model parameter 

7301 estimation and structure validation. The data are primarily measurements of parent dichloromethane 

7302 kinetics (e.g., blood or closed-chamber air concentrations over time), rather than measurements of 

7303 metabolite levels which can be unambiguously attributed to one of the two principal metabolic pathways

7304 (GST and CYP). For the mouse model in particular, only parent dichloromethane data were used, 

7305 though exhaled amounts of CO2 and CO are available. Marino et al. (2006) did include data from mice 

7306 pre-treated with trans-1,2-dichloroethylene (tDCE), a specific CYP 2E1 inhibitor, but the authors 

7307 assumed without verification that 100% of the CYP 2E1 activity was eliminated by the inhibitor when 

7308 using those data. In contrast, Mathews et al. (1997) found that pretreatment of F344 rats by tDCE (100 

7309 mg/kg ip) only yielded 65% inhibition of CYP 2E1. If a significant fraction of the CYP 2E1 activity 

7310 was not eliminated in the dichloromethane experiments, then that activity is erroneously assigned to the 

7311 GST pathway in the parameter estimation Marino et al. (2006). 

7312 In addition to the possibility of incomplete inhibition of CYP 2E1 effecting the data 

7313 interpretation, the Michaelis-Menten rate equation used in all of the published PBTK models for 

7314 dichloromethane, including that of Marino et al. (2006), has in fact not been shown to accurately 

7315 describe the CYP 2E1-mediated metabolism of dichloromethane in the relevant concentration range. 

7316 While Michaelis-Menten kinetics usually describe CYP-mediated oxidation data quite well, the 

7317 approach of Marion et al. (2006) implicitly assumes that any metabolism not described by the Michaelis7318 Menten equation is GST-mediated. If pathway-specific metabolite data were used to define or bound 
252

7319 the ratio of GST to CYP metabolism, the resulting estimates would be less sensitive to what otherwise 

7320 might be small errors in the CYP rate equation. But in the modeling of the mouse data by Marino et al. 

7321 (2006), the fraction of total metabolism assigned to the CYP pathway depends quite strongly on the 

7322 assumed form of the CYP rate equation, along with the assumption of 100% inhibition by tDCE. In 

7323 EPA’s modeling of the rat in vivo PK data, using the same model structure and equations, a set of 

7324 parameter values could not be found which described both the parent dichloromethane kinetics and the 

7325 total amount of CO exhaled at both high and low exposure levels; in particular see panel C of Figure C-3 

7326 and note discrepancy between model and 50 mg/kg data. That the model does not describe well the 

7327 dose-dependent shift in metabolism shown by those CO data suggests that the dose-dependence of the 

7328 CYP Michaelis-Menten rate-equation is not adequate. As will be shown, an alternative equation for 

7329 CYP kinetics may fit the existing dichloromethane data better than Michaelis-Menten kinetics, with the

7330 result that a higher portion of total dichloromethane metabolism would be interpreted as being CYP7331 mediated. Thus there is uncertainty in the choice of equation for the CYP pathway, which leads to 

7332 uncertainty in the estimated GST:CYP metabolic ratio, upon which current risk predictions are based. 

7333 The potential error in assuming Michaelis-Menten kinetics for CYP-mediated oxidation of 

7334 dichloromethane is reinforced by examining the in vitro oxidative (i.e., CYP-specific) kinetics of 

7335 dichloromethane reported by Reitz et al. (1989). When extrapolated from in vitro to in vivo, the 

7336 apparent values of the oxidative saturation constant, Km, identified by Reitz et al. (1989) for mice, rats, 

7337 and humans are over 2 orders of magnitude greater than those obtained from in vivo PBTK modeling. 

7338 Part of the explanation for this apparent discrepancy lies in the disparate concentration ranges 

7339 investigated: Reitz et al. (1989) used much higher dichloromethane concentrations in vitro than those 

7340 observed in or predicted for the various in vivo pharmacokinetic studies. In particular, the oxidation of 

7341 dichloromethane could involve two oxidative processes, one with a high affinity (low Km) 

7342 corresponding to the nonlinearity observed in vivo and one with a low affinity (high Km) corresponding

7343 to the nonlinearity observed in vitro. Further, the low-affinity process would have nearly linear kinetics 

7344 in the exposure range used for the in vivo dosimetry studies and hence be difficult to distinguish from 

7345 GST-mediated metabolism unless pathway-specific metabolite data are used. One can hypothesize that 

7346 this second oxidative process is not inhibited by tDCE and hence corresponds to the 35% of oxidative 

7347 metabolism which was observed to remain in rats after tDCE treatment by Mathews et al. (1997). 

7348 The data of Reitz et al (1989) could simply indicate a second CYP with low-affinity 

7349 dichloromethane activity. However that possibility is contradicted by the results of Kim and Kim (1996) 

7350 who observed that another CYP 2E1-specific inhibitor, disulfiram, completely abolished 

7351 dichloromethane-induced increases on COHb in rats. Another possible explanation which would 

7352 support the findings observed in Kim and Kim (1996) as well as Reitz et al (1989) and the various in 

7353 vivo data is that a number of CYPs exhibit “atypical” kinetics, not described by the classic Michaelis7354 Menten equation, consistent with the enzymes having dual binding sites as proposed by Korzekwa et al 
253

7355 (1988). (Korzekwa et al. (1988) demonstrated atypical kinetics for several CYP-isozyme/substrate 

7356 pairs, but not specifically for CYP 2E1.) Figure 5-9 shows kinetic model fits to the in vitro mouse 

7357 dichloromethane oxidation kinetic data of Reitz et al. (1989), after expressing those data on a per gram 

7358 of liver basis. Both the standard Michaelis-Menten kinetic equation (solid line) and the dual-binding

7359 equation (dashed line) given by Korzekwa et al. (1988) are shown. In particular, the high-affinity (low) 

7360 Km for the dual-binding equation was set equal to that obtained by Marino et al. (2006) from their 

7361 PBTK modeling. This figure shows that the dual binding model is not only consistent with the apparent

7362 high-affinity saturation obtained from in vivo PBTK modeling (Km of Marino et al. (2006)), but also 

7363 with the apparent low-affinity (high Km) data of Reitz et al. (1989), and describes those in vitro data 

7364 better than the standard Michaelis-Menten equation. (Reitz et al. (1989) used classic Lineweaver-Burk

7365 plots to display their kinetic data; i.e., 1/reaction rate vs. 1/concentration. The systematic discrepancy 

7366 between their data and Michaelis-Menten kinetics evident in Figure 5-9 is much less obvious with that 

7367 scaling, which likely explains why they made no note of it.) 

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 

7370 Figure 5-9. Comparison of dichloromethane oxidation rate data with alternate 

7371 kinetic models. Dichloromethane (DCM) oxidation data obtained with mouse liver 

7372 microsomes by Reitz et al. (1989) (points), expressed on a per gram of liver basis, 

7373 are shown with a fitted Michaelis-Menten equation (solid line) or a fitted dual7374 binding-site equation as described by Korzekwa et al. (1988) (dashed line), where 

7375 the high affinity saturation constant of the dual-binding-site equation set equal to 

7376 the mean Km determined for mice via PBTK modeling by Marino et al. (2006). The 

7377 Km for the Michaelis-Menten equation (108 mg/L) is inconsistent with the in vivo 
254

7378 DCM dosimetry data, while the in vitro data shown here are inconsistent with the 

7379 Km estimated in vivo (0.42 mg/L) if that equation is used. 

7380 

7381 In summary regarding model equations, the current PBTK model used the standard Michaelis7382 Menten equation to describe CYP 2E1-catalyzed oxidation of small volatile organic compounds. 

7383 Analysis of the dichloromethane (pharmaco)kinetic data and evaluation of the inconsistencies describe 

7384 above suggest that an alternate equation, which would impact risk predictions, may better represent CYP 

7385 2E1-induced oxidation of dichloromethane. However, this hypothesis requires further laboratory 

7386 testing, for example, by measuring dichloromethane oxidation in a bacterial expression system where 

7387 only CYP 2E1 is expressed over a concentration range sufficient to firmly distinguish between the two 

7388 kinetic forms indicated in the figure above. Until such experiments are conducted, the existing PBTK 

7389 model remains the best available science for dose- and hence risk-extrapolation from rodents to humans. 

7390 Still, this model structure uncertainty implies uncertainty in the quantitative results obtained with the 

7391 model. Analysis of the GST-mediated metabolism of dichloromethane measured by Reitz et al. (1988) 

7392 shows that those results are within a factor of three of the GST kinetic parameters used in the current 

7393 PBTK model, indicating that the any error in the GST:CYP balance is no greater than that, a reasonable

7394 level of uncertainty. 

7395 One other component of quantitative uncertainty arises in examining the results of the Bayesian 

7396 modeling for the human PBTK model of David et al. (2006). The authors reported Bayesian posterior 

7397 statistics for the population average of each fitted parameter when calibration was performed either with 

7398 specific published data sets or the entire combined data set. While one would generally expect that the 

7399 values obtained from the combined data set should be a weighted average of the values from individual 

7400 data sets, the population mean for the liver GST activity (coefficient), KFC, was 0.852 while the values 

from the individual data sets ranged from 1.92-34.0 kg0.3 7401 /h. 

A clarification provided by D. Marino (personal communication)87402 is that the parameter bounds 

7403 stated in the text of David et al. (2006) were only applied for the analysis of the DiVincenzo and Kaplan 

7404 (1981) and the combined data set. But according to the text and distribution prior statistics specified, the 

upper bound for KFC would have been 12 kg0.3 7405 /h (mean + 2.5 standard deviations (SDs), with mean = 2 

7406 and SD = mean*CV = 2*2 = 4). The data of Andersen et al. (1991) were not used in the combined 

7407 analysis because only group average values were available from that source, rather than individual data. 

7408 Since the remaining study-specific mean KFC values were 7.95, 5.87, 34.0, and 1.92, with CVs of less 

7409 than 2, it seems unlikely that application of this upper bound would result in a value of KFC of only 

0.852 kg0.3 7410 /h. Given that there had been convergence problems with the combined data set when 

7411 parameter values were unbounded, it is possible that convergence had not actually been reached after 

 

8

 Email from Dale Marino to Glinda Cooper dated April 25, 2007. 
255

7412 parameter bounds were introduced, and a higher value for KFC would have been obtained had the chain 

7413 been continued longer. 

7414 Since the numerical average of the mean KFC values for the four data sets included in the 

7415 combined data set was 12.4 and the upper bound was 12, the impact of using an intermediate value of 

KFC, specifically the DiVincenzo and Kaplan value of 5.87 kg0.3 7416 /h was explored. Changing only the KFC

7417 is not realistic since the dichloromethane data effectively define total metabolism (sum of CYP and GST 

7418 pathways) and there is naturally a negative correlation between the predicted CYP metabolic rate and 

7419 the GST metabolic rate, required to describe this total. Therefore, it would be inconsistent with the

7420 dichloromethane data to increase KFC without adjusting the CYP metabolic rate downward, and likewise 

7421 all other parameters. The distributions for all of the fitted parameters were rescaled by the ratio of the 

7422 mean for DiVincenzo and Kaplan (1981) to the mean for the combined data set (e.g., the distribution for 

7423 KFC was multiplied by 5.87/0.852, the ratio of the two posterior means). The resulting predicted upperbound (95th and 99th percentile) GST metabolism rates for a fixed level of exposure (1 μg/m3

7424 inhaled 

7425 concentration or 1 mg/kg/day oral exposure) in the GST-T1 +/+ population increased by more than a 

7426 factor of 10. (For inhalation exposure the mean value also increased by over 10-fold, but for oral 

7427 exposure the mean increased by only 2-fold.) Since the majority of metabolism occurs via the CYP 

7428 pathway at these low levels, there is not a proportionate (i.e., over 10-fold) decrease in that rate, but 

7429 HEC and HED calculations increased by 10-30% for the mixed GST-T1 population, depending on the 

7430 route of exposure and distribution statistic compared. Thus the impact of this model uncertainty appears 

7431 to be relatively small for the noncancer assessment, but quite large for the cancer assessment. 

7432 The dose metric used in the models is the rate of metabolism to a putative toxic metabolite, 

7433 rather than the concentration average or area under the concentration curve of the metabolite, so the 

7434 model specifically fails to account for rodent-human differences in clearance or removal of the toxic 

7435 metabolite. A scaling factor based on BW ratios, was used to account for this difference. 

7436 The rat model was modified and utilized in a deterministic manner. Data were not available to 

7437 perform a hierarchical Bayesian calibration in the rat. Thus, uncertainties in the rat model predictions 

7438 had to be assessed qualitatively. To address these uncertainties, a sensitivity analysis was conducted to 

7439 determine which model parameters most influence the predictions for a given dose metric and exposure 

7440 scenario. 

7441 Sensitivity is a measure of the degree to which a given model output variable (i.e., dose metric) 

7442 is influenced by perturbation in the value of model parameters. The approach implemented was a 

7443 univariate analysis in which the value of an individual model parameter was perturbed by an amount 

7444 (∆), in the forward and reverse direction (i.e., an increase and decrease from the nominal value), and the 

7445 change in the output variable was determined. Sensitivity coefficients were calculated as follows: 

7446 
256

( )

( ) ( )

( )

'

f x

x

x

f x x f x

f x ⋅

∆

+ ∆ −

7447 ≈ (Eq. 5-1)

7448 

7449 where x is the model parameter f(x) is the output variable, ∆x is the perturbation of the parameter from 

7450 the nominal value, and f’(x) is the sensitivity coefficient. In equation 5-1, the sensitivity coefficients are 

7451 scaled to the nominal value of x and f(x) to eliminate the potential effect of units of expression. 

7452 Therefore, the sensitivity coefficient is a measure of the proportional (unitless) change in the output 

7453 variable produced by proportional change in the parameter value. Parameters that have higher 

7454 sensitivity coefficients have greater influence on the output variable. They are considered more 

7455 sensitive than parameters with lower values. The results of the sensitivity analysis are useful for 

7456 assessing uncertainty in model predictions, based on the level of confidence or uncertainty in the model 

7457 parameter(s) to which the dose metric is most sensitive. 

7458 Sensitivity coefficients for the noncancer dose metric (mg dichloromethane metabolized via 

7459 CYP-mediated pathway/L liver/day), were determined for each of the model parameters; a similar 

7460 analysis was also done for a metric based on the GST-mediated pathway. Sensitivity analyses for both 

7461 oral and inhalation exposures were performed. The exposure conditions were set to be near or just 

7462 below the lowest bioassay exposure resulting in significant increases in the critical effect. 

7463 For the CYP-mediated metabolism from oral exposure, the VLC and VSC (liver volume and 

7464 slowly perfused tissue volume, respectively) parameters exert the largest influence (Figure 5-10). The 

7465 high influence of these two parameters was due to the fact that the dose metric is a tissue-specific rate of 

7466 metabolism, the majority of CYP metabolism is attributed to the liver, and that changes in liver volume 

7467 have a greater impact on the total CYP metabolism that the individual Vmax value. For inhalation 

7468 exposures VMAXC, in addition to VLC and VSC have the highest sensitivity coefficients (Figure 5-11). 

7469 The physiological parameters (VLC and VSC) are known with a high degree of confidence (Brown et 

7470 al., 1997). VMAXC for the rat was estimated by fitting to the PK data as described in Chapter 3 and 

7471 Appendix C, subject to model structure/equation uncertainties as detailed above, and hence is known 

7472 with less certainty than the physiological parameters. That total exhaled CO, which is a specific to the 

7473 CYP pathway, is within 50% of measured levels (Fig. C-8, panel C), however, provides a similar level 

7474 of confidence in the balance between CYP and GST pathways predicted by the rat PBTK model. 
257

Oral exposure

-1.25

-1

-0.75

-0.5

-0.25

0

0.25

0.5

0.75

1

1.25

QCC

VPR

VLC

VSC

PB

VMAXC

KA

A2

KFC

Parameter

Normalized sensitivity coefficient

CYP

GST

7475 

7476 Figure 5-10. Sensitivity coefficients for long-term mass CYP- and GST-mediated 

7477 metabolites per liver volume from a daily drinking water concentration of 10 mg/L in rats. 

7478 KFC = GST-mediated metabolism rate; A2 = proportion of liver GST metabolism attributed to 

7479 the lung; KA = oral absorption rate from gut; VMAXC = CYP-mediated maximum rate of 

7480 metabolism; PB = blood:air partition coefficient; VSC = slowly perfused tissue volume; VLC = 

7481 liver volume; VPR = Ventilation perfusion ratio ; QCC = cardiac output constant. 

7482 

Inhalation exposure

-1.25

-1

-0.75

-0.5

-0.25

0

0.25

0.5

0.75

1

1.25

QCC

VPR

VLC

VSC

PB

VMAXC

A2

KFC

Parameter

Normalized sensitivity coefficient

CYP

GST

7483 

7484 Figure 5-11. Sensitivity coefficients for long-term mass CYP- and GST-mediated 

7485 metabolites per liver volume from a long-term average daily inhalation concentration of 

7486 500 ppm in rats. (KA is not included since it has no impact on inhalation dosimetry.) KFC 

7487 = GST-mediated metabolism rate; A2 = proportion of liver GST metabolism attributed to the 

7488 lung; VMAXC = CYP-mediated maximum rate of metabolism; PB = blood:air partition 

7489 coefficient; VSC = slowly perfused tissue volume; VLC = liver volume; VPR = Ventilation 

7490 perfusion ratio ; QCC = cardiac output constant. 

7491 
258

7492 In summary, the uncertainties associated with use of the rat PBTK model should not markedly 

7493 affect the values of the RfD and RfC based on the metrics considered. An additional uncertainty results 

7494 from the lack of knowledge concerning the most relevant dose metric (e.g., a specific metabolite) for the 

7495 non-cancer endpoints considered. This basic research question represents a data gap. This uncertainty 

7496 was addressed by considering different dose metrics (CYP metabolism alone, GST metabolism alone, 

7497 sum of GST and CYP, and the AUC of the parent compound). The GST metabolism and the AUC dose 

7498 metrics did not present reasonable choices based on model fit and consistency of response across studies 

7499 at comparable dose levels. Given these results, the combination of hepatic metabolism through the GST

7500 and the CYP pathways would not be expected to result in an improvement to a metric based only on 

7501 CYP metabolism. The CYP-metabolism dose metric seems to be most consistent with the data., and so 

7502 is the metric chosen for the RfD and RfC derivations. 

7503 

7504 Sensitive human populations 

7505 The potential for sensitivity to dichloromethane in a portion of the human population due to 

7506 pharmacokinetic differences was addressed quantitatively by using a human probabilistic PBTK model, 

7507 as modified by the EPA, to generate distributions of human exposures likely to result in a specified 

7508 internal BMDL10. The model and resulting distributions take into account the known non-chemical7509 specific variability in human physiology as well as total variability and uncertainty in dichloromethane7510 specific metabolic capability. The first percentile values of the distributions of human equivalent doses 

7511 (Table 5-3) and HECs (Tables 5-7) served as points of departure for candidate RfDs and RfCs, 

7512 respectively, to protect toxicokinetically sensitive individuals. Selection of the first percentile allows 

7513 generation of a stable estimate for the lower end of the distribution. The mean value of the human 

7514 equivalent oral dose in Table 5-3 was about two-fold higher than the corresponding first percentile 

7515 values, and the mean value of human equivalent inhalation concentration in Table 5-7 was 

7516 approximately three-fold higher than the first percentile value. The internal dose metric in the analyses 

7517 described in these tables was the mg dichloromethane metabolized via the CYP pathway per liter liver 

7518 per day, and thus the comparisons of the first percentile and mean values give estimates of the amount of 

7519 variability in the population to metabolize dichloromethane by the CYP metabolic pathways on a liverspecific basis. The mean: 1st 7520 percentile ratios for these distributions is attributed to the dependence of 

7521 the dose metric on hepatic blood flow rate (metabolism being flow limited). This blood flow is expected 

7522 to be highly and tightly correlated with liver volume, resulting in very similar delivery of 

dichloromethane per volume liver across the population. While the mean: 1st 7523 percentile ratios for the 

7524 oral distribution is less than the default intra-human toxicokinetic UF of 3, it is quite similar to that 

7525 obtained by Sweeney et al. (2003) for acrylonitrile, where an extensive sensitivity analysis indicated a 

99th 7526 percentile:mean ratio of less than 2.2 among several internal dose metrics. The population7527 structured distributions for physiological parameters and broadened distributions for metabolic 
259

7528 parameters used here provide a good degree of confidence that the population variability has not been 

7529 under-estimated. 

7530 The internal dose metric used in the RfD and RfC derivations was based on the rate of CYP 

7531 metabolism. GST-T1 polymorphisms could affect this rate, as the GST-T1 null genotype would be 

7532 expected to result in an increase in the metabolism through the CYP pathway, resulting in a greater 

7533 sensitivity to a CYP-related effect. The effect of GST variability on the RfD and RfC values was 

7534 examined by comparing results obtained specifically for the GST-T1 null genotype to those obtained for

7535 the population of mixed genotypes. The values for human equivalent doses and HECs were very similar 

for these two groups (e.g., mean HEC 47.36 and 47.49 for the mixed and the GST-T1-/- 7536 null genotypes, 

respectively; 1st percentile HEC 16.63 and 16.69 for the mixed and the GST-T1-/- 7537 null genotypes, 

7538 respectively), and use of this population would not result in a change in the recommended RfD or RfC. 

7539 As a further level of sensitivity analysis, we compared model predictions of the human 

7540 equivalent dose, as listed in Table 5-3, for the general population (estimates covered 0.5- to 80-year-old 

7541 male and female individuals) to three subpopulations: 1-year-old children (males and females), 70-year7542 old men, and 70-year-old women. For the general population and each subpopulation a Monte Carlo 

7543 simulation representing 10,000 individuals was conducted, and histograms of the resulting distribution

7544 of human equivalent administered doses are shown in Figure 5-12, with corresponding statistics in Table 

5-9. All groups used in these comparisons were limited to the GST-T1-/- 7545 . 

7546 The results shown above for differences in human equivalent dose values in different populations 

7547 are qualitatively what would be expected: a relatively broad distribution for the general population with 

7548 specific populations representing narrower components of that distribution. There are some differences 

7549 between men and women at 70 years of age, but neither of these would be greatly misrepresented by the 

7550 general population estimate. While 1-year-old children represent more of a distinct tail in the general 

7551 population, in this case the distribution of human equivalent concentrations in the general estimate is 

7552 lower than that seen in what would otherwise be considered a more sensitive population. This 

7553 difference most likely results from the higher specific respiration rate in children versus adults, which 

7554 allows them to eliminate more of orally ingested dichloromethane by exhalation, leading to lower 

7555 internal metabolized doses 
260

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 

7557 Figure 5-12. Frequency density of human equivalent applied doses in specific populations 

7558 in comparison to a general population (0.5- to 80-year-old males and females) estimate for 

7559 an internal dose of 15.1 mg dichloromethane metabolized by CYP per liter liver per day; all 

groups were restricted to the GST-T1 -/- 7560 population). 

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 5th percentile 1

st percentile

All agesb 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 dichloromethane 

metabolized via CYP pathway per liter liver per day (based on mean 

BMDL10 across acceptable models from Table 5-3). 

b

0.5- to 80-year-old males and females. 

7563 

7564 . 

7565 A similar comparison was made for inhalation HEC values, as shown in Figure 5-13 and Table 

7566 5-10. For HEC values, the distributions for 70-year-old men and women are both virtually 

7567 indistinguishable from the general population, and while 1-year-old children are clearly distinct, they are 
261

7568 less different than in the human equivalent administered dose comparison and in this case are more 

7569 sensitive than the population in general. As described in detail in Appendix B, the allometric alveolar 

ventilation constant, QAlvC, is about 28 L/hour-kg0.75 7570 in a 1-year-old child but averages around 14 

L/hour-kg0.75 7571 in an adult. Combining this with the difference between a BW of 10 kg in that child and 

7572 70 kg in an “average” adult, the respiration rate per kg BW is about threefold higher in the child versus 

7573 adult. As noted above, for oral exposures this leads to faster elimination by respiration in children, 

7574 while for inhalation exposures it leads to higher uptake for a given air concentration. 

7575 The lack of difference in elderly adults versus the general population in HEC values is likely due 

7576 to the fact that the rate of exposure and rates of metabolism (the latter being the key dose metric) both 

scale as BW0.75 7577 , with the scaling coefficients being either similar (respiration) or identical (metabolism) 

7578 among adults, who comprise the majority of the population, while for oral exposures the exposure rate is 

normalized to total BW and scales as BW1, while elimination routes increase as BW0.75 7579 . Moreover, oral 

7580 exposures are simulated as occurring in a series of bolus exposures (drinking episodes) during the day, 

7581 and the higher body-fat content occurring in the elderly (see Appendix B) means that such a dose that 

7582 might saturate metabolism and therefore have a higher fraction exhaled in a leaner individual will tend 

7583 to be more sequestered in fat and slowly released, resulting in a higher fraction metabolized (less 

7584 saturation of metabolism) in a more obese individual. The difference among adults of different ages for 

7585 dosimetry from oral ingestion (bolus exposure) will be greater than the difference for inhalation 

7586 exposures. More careful examination of Figure 5-13 shows that the distribution for 70-year-old women,

7587 for whom the fat fraction is estimated to be greatest, has a lower peak and higher upper tail than for the 

7588 general population. So the physiological differences do have some impact that is qualitatively consistent 

7589 with what is seen from oral exposure, given the mechanistic considerations described here. But the 

7590 impact of those differences is far less for inhalation exposure. 

7591 
262

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 

7593 Figure 5-13. Frequency density of HECs in specific populations in comparison to a 

7594 general population (0.5- to 80-year-old males and females) estimate for an internal 

7595 dose of 128.1 mg dichloromethane metabolized by CYP per liter liver per day ; all 

groups restricted to the GST-T1 -/- 7596 population). 

7597 

Table 5-10. Statistical characteristics of HECs in

specific populations of the GST-T1 -/- group

HEC (mg/m3)

a

Population Mean 5th percentile 1

st percentile

All agesb 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 dichloromethane 

metabolized via CYP pathway per liter liver per day (based on mean 

BMDL10 across acceptable models from Table 5-7). 

b

0.5- to 80-year-old males and females. 

7598 

7599 

7600 No data are available regarding toxicodynamic differences within a human population. 

7601 Therefore, a UF of 3 for possible differences in human toxicodynamic responses is intended to be 

7602 protective for sensitive individuals. 

7603 

7604 5.4. CANCER ASSESSMENT 
263

7605 5.4.1. Cancer Oral Slope Factor 

7606 5.4.1.1. Choice of Study/Data—with Rationale and Justification 

7607 No human data are available for the quantification of potential neoplastic effects from oral 

7608 exposures to dichloromethane. In the only chronic (2-year) oral exposure cancer bioassay, significant

7609 increases in the incidence of liver adenomas and carcinomas was observed in male, but not female, 

7610 B6C3F1 mice exposed by drinking water, with incidence rates of 19, 26, 30, 31, and 28% in groups with 

7611 estimated mean intakes of 0, 61, 124, 177, and 234 mg/kg-day, respectively (trend p- value = 0.058) (see 

7612 Table 4-38 for group comparisons) (Serota et al., 1986b; Hazelton Laboratories, 1983). Evidence of a 

7613 trend for increased risk of liver tumors (described as neoplastic nodule or hepatocellular carcinoma) was 

7614 seen in female F344 rats, but not males, exposed via drinking water (p < 0.01) (Serota et al., 1986a). 

7615 However, the potential malignant characterization of the nodules was not described, and no trend was 

7616 seen in the data limited to hepatocellular carcinomas. The derivation of the cancer oral slope factor 

7617 (OSF) is based on the male mouse data (Serota et al., 1986b; Hazelton Laboratories, 1983) because of 

7618 their greater sensitivity compared to female mice and to male and female rats. The study authors 

7619 concluded that these increases were “within the normal fluctuation of this type of tumor incidence”. 

7620 However, the trend p-value for these results is of borderline statistical significance and it may not be 

7621 reasonable to apply a correction for multiple comparisons given the lack of independence of the groups

7622 and given a specific focus on the liver as a target organ. The development of liver tumors in B6C3F1

7623 mice is associated with metabolite production in this tissue via the GST metabolic pathway (section 

7624 4.7.3), a pathway that also exists in humans. Modeling intake, metabolism, and elimination of 

7625 dichloromethane in mice and humans is feasible. Thus, it is reasonable to apply the best available 

7626 PBTK models to estimate equivalent internal doses in mice and humans. 

7627 

7628 5.4.1.2. Derivation of Oral Slope Factor 

7629 In a manner similar to the derivation of the noncancer toxicity values, PBTK models for 

7630 dichloromethane in mice and humans were used in the derivation of toxicity values (cancer OSF and 

7631 IUR) for cancer endpoints based on lung (for inhalation) and liver (for oral and inhalation) tumor data in 

7632 the mouse (Figure 5-14). A deterministic PBTK model for dichloromethane in mice was first used to 

7633 convert mouse drinking water or inhalation exposures to long-term daily average values of internal lung7634 specific GST metabolism (GST metabolism in lung/lung volume) or liver-specific GST metabolism 

7635 (GST metabolism in liver/liver volume). The choice of this dose metric was made based on data 

7636 pertaining to the mechanism(s) involved in the carcinogenic response, specifically data supporting the

7637 involvement of a GST metabolite(s). The evidence pertaining to the GST pathway is discussed in section

7638 4.7, and includes the enhanced genotoxicity seen in bacterial and mammalian in vitro assays with the 

7639 introduction of GST metabolic capacity (Graves et al., 1994a) and the suppression by pretreatment with
264

7640 a GSH depletory of the production of DNA SSBs seen in acute inhalation exposure to dichloromethane 

7641 in mice (Graves et al., 1995). 

7642 
265

Recommend value

Distribution of Human Cancer 

Oral Slope Factors or 

Inhalation Unit Risks 

95th 99th

Monte Carlo 

Sampling from 

Distributions of 

Human PBTK 

Model Parameters

Distribution of Human Internal 

Doses from Unit Oral Doses 

(1mg/kg) or Inhalation 

Concentrations (1ug/m3)

Probabilistic

Human PBTK 

Model

Multiply Human Tumor Risk Factor 

By Distribution of Human Internal 

Unit Doses

Rodent Dose Response Data

Rodent 

PBTK Model Estimates of Rodent 

Internal Dose

BMD 

Modeling

BMD

BMDL

Rodent Internal BMDL10

95% Lower Bound Estimate of Internal Dose Associated with a 10% response (0.1/Rodent BMDL10)

Rodent Tumor Risk Factor

(internal dose) Human Tumor Risk Factor -1

(internal dose)-1

Scaling Factor

+

Apply Age-Dependent Adjustment Factors 

(ADAFs) for early life exposure

7643 7644 

7645 Figure 5-14. Process for deriving cancer OSFs and IURs by using rodent and human PBTK model
266

7646 The multistage cancer model (using BMDS version 2.0) was then fit to the tumor 

7647 incidence data and internal dose data for rodents, and BMD10 and associated BMDL10 values (for 

7648 a BMR of 10% extra risk) were calculated. A probabilistic PBTK model for dichloromethane in 

7649 humans, adapted from David et al. (2006) (see Appendix B), was used with Monte Carlo 

7650 sampling to calculate distributions of internal lung or liver doses associated with chronic unit 

oral (1 mg/kg-day) or inhalation (1 μg/m37651 ) exposures. The resulting distribution of human 

7652 internal doses was multiplied by a human internal dose tumor risk factor (in units of reciprocal 

7653 internal dose) to generate a distribution of OSFs or IURs associated with a chronic unit oral or 

7654 inhalation exposure, respectively. 

7655 The parameter statistics reported by David et al. (2006) include both the inter-individual 

7656 variability that would have been elucidated by the Bayesian analysis (variation between mean 

7657 values for each individual for which data were available) and uncertainty in those values. Since 

7658 EPA’s objective is to account for both population variability and parameter uncertainty, 

7659 however, these statistics were primarily as-is (exceptions discussed in Appendix B) to define 

7660 population distributions. Assuming that these parameters are distributed independently, ignoring 

7661 the covariance that was likely represented in the actual posterior chains, will tend to over7662 estimate the overall range of parameters and hence distribution of dose metrics in the population, 

7663 compared to what one would obtain if the covariance were explicitly included. Thus if the 

7664 covariance (i.e., the variance-covariance matrix) for the set of parameters had been reported by 

7665 David et al., it could have been used to narrow the predicted distribution of internal doses, or 

7666 equivalent applied doses. Lacking such information the approach used will not under-estimate 

7667 risk or over-estimate lower bounds on human equivalent exposure levels. 

7668 

7669 5.4.1.3. Dose-Response Data 

7670 Data for liver tumors in male B6C3F1 mice following exposure to dichloromethane in 

7671 drinking water were used to develop oral cancer slope factors (Serota et al., 1986b; Hazelton 

7672 Laboratories, 1983). Significant increases in incidence of liver adenomas and carcinomas were 

7673 observed in male, but not female, B6C3F1 mice exposed for 2 years (Table 5-11). No significant 

7674 decreases in survival were observed in the treated groups of either sex compared with controls. 

7675 The at-risk study populations (represented by the denominators in the incidence data) were 

7676 determined by excluding all animals dying prior to 52 weeks.

7677 

7678 5.4.1.4. Dose Conversion and Extrapolation Methods: Cancer Oral Slope Factor 

7679 Dose conversion 

7680 The mouse PBTK model of Marino et al. (2006) was based on the PBTK model for 

7681 dichloromethane by Andersen et al. (1987), which was modified to include dichloromethane 

7682 metabolism in the lung compartment and kinetics of CO and COHb (Andersen et al., 1991). For 

7683 the mouse, physiological parameters and partition coefficients were adjusted to match those 
267

7684 reported in Andersen et al. (1991, 1987) and Clewell et al. (1993), respectively, while QCC, 

7685 VPR, and metabolic parameter distribution mean values were derived via MCMC model 

7686 calibration reported by Marino et al. (2006) (Appendix B). The model of Marino et al. (2006) 

7687 was used to simulate daily drinking water exposures comprising six discrete drinking water 

7688 episodes for specified times and percentage of total daily intake (Reitz et al., 1997) and to 

7689 calculate average lifetime daily internal doses for the male mouse data shown in Table 5-11. A 

first-order oral uptake rate constant (ka) of 5 hours–1 7690 was taken from Reitz et al. (1997) to 

7691 describe the uptake of dichloromethane from the gastrointestinal tract to the liver. Study-specific 

7692 BWs were not available, so reference BWs for male B6C3F1 mice in chronic studies (U.S. EPA, 

7693 1988a) were used. Based on evidence that metabolites of dichloromethane produced via the 

7694 GST pathway are primarily responsible for dichloromethane carcinogenicity in mouse liver 

7695 (summarized in section 4.7.3) and the assumption that these metabolites are sufficiently reactive 

7696 that they do not have substantial distribution outside the liver, the recommended selected internal 

7697 dose metric for liver tumors was daily mass of dichloromethane metabolized via the GST 

7698 pathway per unit volume of liver (Table 5-11). Figure 5-15 shows the comparison between 

7699 internal and external doses in the liver in mice and humans. The whole-body metabolism metric 

7700 was also examined. This metric would be more relevant under a scenario of slowly cleared 

7701 metabolites that undergo general circulation. 

7702 

Table 5-11. Incidence data for liver tumors and internal liver doses, 

based on GST metabolism dose metrics, in male B6C3F1 mice 

exposed to dichloromethane in drinking water for 2 years 

Sex 

Nominal (actual) 

daily intake 

(mg/kg-day) 

Mouse liver 

tumor incidencea

Mouse internal 

liver metabolism 

doseb 

Mouse whole body 

metabolism dosec 

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 

aHepatocellular carcinoma or adenoma, combined. Mice dying prior to 52 weeks were excluded 

from the denominators. Cochran-Armitage trend p-value = 0.058. 

bmg dichloromethane metabolized via GST pathway/L liver/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 metabolism (mg/day) + liver GST metabolism 

(mg/day)]/kg BW. Units = mg dichloromethane metabolized 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 Laboratories, 1983 

7703 
268

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 

7705 Figure 5-15. PBTK model-derived internal doses (mg dichloromethane metabolized via 

7706 the GST pathway per liter liver per day) in mice and humans and their associated 

7707 external exposures (mg/kg-day) used for the derivation of cancer OSFs based on liver 

7708 tumors in mice. Six simulated drinking water episodes are described by Reitz et al. (1997). 

7709 The human metabolism rates were estimated using a computational sample of 1000 individuals 

7710 per dose, including random samples of the three GST-T1 polymorphisms (+/+, +/-, -/-; “Human 

7711 mixed” curve) or samples restricted to the GST +/+ or +/- populations, in the current U.S. 

7712 population based on data from Haber et al. (2002). Since a different set of samples was used 

7713 for each dose, some stochasticity is evident as the human points (values) do not fall on smooth 

7714 curves. Error bars indicate the range of 5th-95th percentile for the sub-populations sampled at 

7715 select concentrations. 

7716 

7717 Dose-response modeling and extrapolation 

7718 The multistage dose-response model was fit to the mouse liver tumor incidence and 

7719 PBTK model-derived internal dose data to derive mouse internal BMD10 and BMDL10 associated 

7720 with 10% extra risk (Table 5-12). Different polynomial models, and models dropping dose 

7721 groups starting with the highest dose group were compared based on adequacy of model fit as 

assessed by overall χ

2

7722 goodness of fit (p-value > 0.10) and examination of residuals, particularly 

7723 in the region of the benchmark response (BMR). Appendix E-1 provides details of the BMD 

7724 modeling results. The mouse liver tumor risk factor (extra risk per unit internal dose) was 

7725 calculated by dividing 0.1 by the mouse BMDL10 for liver tumors. 

7726 

7727 
269

Table 5-12. BMD modeling results and tumor risk factors for internal dose metric 

associated with 10% extra risk for liver tumors in male B6C3F1 mice exposed to 

dichloromethane in drinking water for 2 years, based on liver-specific GST metabolism 

and whole body GST metabolism dose metrics

Tumor Risk Factore

Internal 

dose metric

BMDS 

modelb 

χ

2

goodness of 

fit p- value

Mouse 

BMD10

c 

Mouse 

BMDL10

c 

Allometricscaled human 

BMDL10

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 metabolized via GST pathway per liter tissue per day; 

Whole-body dose units = mg dichloromethane metabolized via GST pathway in lung and liver/kgday)

b

The multistage (MS) model in EPA BMDS version 2.0 was fit to the mouse dose-response data 

shown in Table 5-11 using internal dose metrics calculated with the mouse PBTK model. Numbers 

in parentheses indicate (1) the number of dose groups dropped in order to obtain an adequate fit and 

(2) the degree polynomial of the model. 

cBMD10 and BMDL10 refer to the BMD-model-predicted mouse internal and its 95% lower confidence 

limit, associated with a 10% extra risk for the incidence of tumors. 

dMouse BMDL10 divided 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 BMDL10 and by the allometric-scaled human BMDL10,, for the scaling = 1.0 and 

allometric-scaled risk factors, respectively. 

7728 

7729 

7730 Linear extrapolation from the internal human BMDL10 values (0.1/BMDL10) was used to 

7731 derive oral risk factors for liver tumors, based on tumor responses in male mice. Proposed key 

7732 events for dichloromethane carcinogenesis are discussed in sections 4.7 and 5.4.1.1. The linear 

7733 low-dose extrapolation approach for agents with a mutagenic mode of action was selected. 

7734 Application of allometric scaling factor 

7735 As discussed in section 4.7 and summarized in 5.4.1.2, several lines of evidence point to 

7736 the involvement of the GST metabolic pathway in the carcinogenic response seen in 

7737 dichloromethane. The role of specific metabolites has not been firmly established, however. 

7738 S-(chloromethyl)-glutathione is an intermediate to the production of formaldehyde through this 

7739 pathway (Hashmi et al., 1994). Formation of free hydrogen ion is also hypothesized, although 

7740 no direct evidence supporting this has been presented. The pattern of HPRT gene mutations seen 

7741 in CHO cells incubated with GST-complete mouse liver cytosol preparations suggest that 

7742 S-(chloromethyl)glutathione, rather than formaldehyde, is responsible for the mutagenic effects 

7743 associated with dichloromethane (Graves et al., 1996). DNA reaction products (e.g., DNA 

7744 adducts) produced by S-(chloromethyl)glutathione have not been quantified, possibly due to 

7745 potential instability of these compounds (Guengerich et al., 2003; Hashmi et al., 1994). 

7746 The question of the role of specific metabolites, and particularly how these metabolites 

7747 are transformed or removed is a key question affecting the choice of a scaling factor to be used in 

7748 conjunction with the internal dose metric based on rate of GST metabolism. If the key metabolite 

7749 is established and is known to be sufficiently reactive to not spread in systemic circulation, then 
270

7750 it can be assumed that 1) the level of reactivity and rate of clearance (i.e., disappearance due to 

7751 local reactivity) for this metabolite, per volume tissue, is equal in rodents and humans; and (2) 

7752 risk is proportional to the long-term daily average concentration of the metabolite. Under these 

7753 assumptions, rodent internal BMDL10 values based on tissue-specific dichloromethane 

7754 metabolism require no allometric scaling to account for toxicodynamic differences and predict 

7755 the corresponding level of human risk as a function of the metric (i.e., the scaling factor in Figure 

7756 5-14 was equal to 1.0). (A single metabolite is referenced, but the same argument holds in 

7757 general for more than one metabolite). Under this scenario and assumptions, humans and 

7758 rodents with the same long-term daily average metabolite formation per volume tissue (e.g., 

7759 equal internal BMDL10) should both experience the same long-term average concentration of the 

7760 metabolite when the metabolite is highly reactive and hence experience the same extra risk. 

7761 Although the evidence points to a specific metabolic pathway and to site-specific actions 

7762 resulting from a reactive metabolite that does not escape the tissue in which it is formed, some 

7763 assumptions remain concerning this hypothesis. Specifically, the active metabolite(s) have not 

7764 been established, and data pertaining to the reactivity or clearance rate of these metabolite(s) are 

7765 lacking. Quantitative measurements of adducts of interest or of the half life of relevant 

7766 compounds in humans and in mice are not available. To address the uncertainties in the 

7767 available data it may be appropriate to use a scaling factor that addresses the possibility that the 

7768 rate of clearance for the metabolite is limited by processes that are known to scale allometrically, 

7769 such as blood perfusion or enzyme activity. This case would result in use of a mouse:human 

dose-rate scaling factor of (BWhuman/BWmouse)

0.25 7770 = 7 to adjust the mouse-based BMDL10 values 

7771 downward. Using this internal dose metric (liver-specific metabolism with allometric scaling), 

7772 equivalent rodent and human internal BMDL10 values result in a human liver tumor risk factor 

7773 (0.1/BMDL10) that is assumed equal to that for the mouse, given a 70-year lifetime exposure. 

7774 Another alternative that can be used is based on an allometrically-scaled whole-body 

7775 metabolism metric. In this case, less weight is given to the evidence of site-specificity, as this 

7776 metric allows for systemic circulation of the relevant metabolites. 

7777 The cancer toxicity values derived using each of these metrics and scaling factors (i.e., 

7778 liver-specific metabolism with and without allometric-scaling and the whole-body metabolism 

7779 metric) are presented in the following tables. Considering the lack of data pertaining to 

7780 clearance rates or the actual AUC of the active carcinogenic metabolite(s) in mice and humans, 

7781 the OSF recommended by the EPA is based on the allometrically-scaled tissue-specific GST 

7782 metabolism rate dose metric. 

7783 

7784 Calculation of OSFs 

7785 The human PBTK model adapted from David et al. (2006) (see Appendix B), using 

7786 Monte Carlo sampling techniques, was used to calculate distributions of human internal dose 

7787 metrics of daily mass of dichloromethane metabolized via the liver-specific GST pathway per 
271

7788 unit volume of liver resulting from a long-term average daily drinking water dose of 1 mg/kg 

7789 dichloromethane. In another analysis of whole body metabolism, a dose metric based on the 

7790 total metabolites formed in liver and lungs via GST metabolism per BW was used. The human 

7791 model used parameter values derived from Monte Carlo sampling of probability distributions for 

7792 each parameter, including MCMC-derived distributions for the metabolic parameters (David et 

7793 al., 2006). The drinking water exposures comprised six discrete drinking-water episodes for 

7794 specified times and percentage of total daily intake (Reitz et al., 1997) (Appendix B). 

7795 The distribution of cancer OSFs shown in Table 5-13 were derived by multiplying the 

7796 human oral liver tumor risk factors by the respective distributions of human average daily 

7797 internal doses resulting from chronic, unit oral exposures of 1 mg/kg-day dichloromethane. 

7798 Because adjustments for interindividual variability are not generally used or recommended in 

7799 cancer risk analysis, the mean slope factor was selected as the recommended value to be used in 

7800 deterministic risk assessments; other values at the upper end of the distribution are also 

7801 presented. 

7802 

7803 Consideration of Sensitive Human Subpopulations 

7804 An important issue in the derivation process used by EPA, pertaining to the use of the 

7805 human PBTK model, stems from the assumption regarding the population for which the 

7806 derivation should be applied. The inclusion of the GST-T1 null subpopulation in effect dilutes 

7807 the risk that would be experienced by those who carry a GST-T1 allele, by averaging in nonresponders (i.e., the GST-T1–/– 7808 genotype). Thus, the cancer OSF was derived specifically for 

7809 carriers of the GST-T1 homozygous positive (+/+) genotype, that is the population that would be 

7810 expected to be most sensitive to the carcinogenic effects of dichloromethane given the GST7811 related dose metric under consideration. In addition, cancer values derived for a population 

7812 reflecting the estimated frequency of GST-T1 genotypes in the current U.S. population (20% 

GST-T1–/–, 48% GST-T1+/–, and 32% GST-T1+/+ 7813 , the “mixed” population) are also presented. 

7814 All simulations also included a distribution of CYP activity based on data from Lipscomb et al. 

7815 (2003). 
272

Table 5-13. Cancer OSFs for dichloromethane based on PBTK model-derived internal liver doses in B6C3F1 mice 

via drinking water for 2 years, based on liver-specific GST metabolism and whole body metabolism dose metrics, 

population genotype 

Distribution of human internal 

dichloromethane 

doses from 1 mg/kg-day exposured 

Resulting candida

OSFe (mg/kg-

Internal dose 

metric and 

scaling factora

Population 

genotypeb

Human tumor 

risk factorc Mean 

95th

percentile 

99th

percentile Mean 

95th

percentil

GST-T1+/+ 1.77 × 10–2 0.80 × 10–1 1.91 × 10–1 2.89 × 10–1 1.4 × 10–3 3.4 × 10–

Liver-specific, 

allometric-scaled Mixed 1.77 × 10–2 0.45 × 10–1 1.39 × 10–1 2.24 × 10–1 8.0 × 10–4 2.5 × 10–

GST-T1+/+ 2.53 × 10–3 0.80 × 10–1 1.91 × 10–1 2.89 × 10–1 2.0 × 10–4 4.8 × 10–

Liver-specific, scaling = 1.0 Mixed 2.53 × 10–3 0.45 × 10–1 1.39 × 10–1 2.24 × 10–1 1.2 × 10–4 3.5 × 10–

GST-T1+/+ 4.24 × 10–1 1.90 × 10–3 4.60 × 10–3 7.20 × 10–3 8.1 × 10–4 2.0 × 10–

Whole-body , allometric-scaled Mixed 4.24 × 10–1 1.08 × 10–3 3.40 × 10–3 5.49 × 10–3 4.6 × 10–4 1.4 × 10–aLiver specific dose units = mg dichloromethane metabolized via GST pathway per liter tissue per day; Whole-body dose units = m

dichloromethane metabolized via GST pathway in lung and liver/kg-day.

bGST-T1+/+ = homozygous, full enzyme activity; mixed = population reflecting estimated frequency of genotypes in current U.S. p

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 the BMR (0.1) by the allometric

human BMDL10 and the mouse BMDL10 for the allometric-scaled and scaling = 1.0 risk factors, respectively (from Table 5-12). 

dMean, 95th, and 99th percentile of the human PBTK model-derived probability distribution of daily average internal dichlorometh

resulting 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 

1 mg/kg-day. 

7816 7817 
273

7818 5.4.1.5. Oral Cancer Slope Factor 

 The recommended cancer OSF for dichloromethane is 1 × 10–3 (mg/kg-day)–1 7819 (rounded 

from 1.4 × 10–3 7820 ) , and is based on liver tumor responses in male B6C3F1 mice exposed to 

7821 dichloromethane in drinking water for 2 years (Serota et al., 1986b; Hazelton Laboratories, 

7822 1983). The OSF was derived by using a tissue-specific GST metabolism dose metric, with 

7823 allometric scaling, for the population that is presumed to have the greatest sensitivity (the GSTT1+/+ 7824 genotype). The application of ADAFs to the cancer OSF is recommended and is described 

7825 in section 5.4.4. 

7826 

7827 5.4.1.6. Alternative Derivation Based on Route-to-Route Extrapolation

7828 For comparison, alternative cancer OSFs were derived via route-to-route extrapolations 

7829 from the data for liver tumors in male and female B6C3F1 mice exposed by inhalation for 2 years 

7830 (Mennear et al., 1988; NTP, 1986). This derivation, shown in Table 5-14, uses the cancer IUR 

7831 derived in section 5.4.2.4 (see Table 5-19 for these IUR values) and the distribution of human 

7832 internal dichloromethane exposures from 1 mg/kg-day exposure using the tissue-specific GST 

7833 metabolism dose metric (mg dichloromethane metabolized via the GST pathway per liter liver 

7834 per day). The oral cancer slope factor values based on the route-to-route extrapolations from 

7835 liver tumors in mice exposed by inhalation (Table 5-14) are about one order of magnitude lower 

7836 than those based on the liver tumor responses in mice exposed via drinking water. 
274

Table 5-14. Alternative route-to-route cancer OSFs for dichloromethane extrapolated from male B6C3F1 mouse 

inhalation liver tumor incidence data using a tissue-specific GST metabolism dose metric, by population genotype 

Distribution of human internal 

dichloromethane 

doses from 1 mg/kg-day exposurec 

Resulting candidate hu

OSFd (mg/kg-day)–1

Internal dose metric 

and scaling factor Population Genotypea

Human 

tumor risk 

factorb Mean 95th percentile 99th percentile Mean 95th percentile 

p

GST-T1+/+ 1.29 × 10–3 0.80 × 10–1 1.91 × 10–1 2.89 × 10–1 1.0 × 10–4 2.5 × 10–4 3

Liver-specific, allometric-scaled Mixed 1.29 × 10–3 0.45 × 10–1 1.39 × 10–1 2.24 × 10–1 5.8 × 10–5 1.8 × 10–4 2

GST-T1+/+ 1.84× 10–4 0.80 × 10–1 1.91 × 10–1 2.89 × 10–1 1.5 × 10–5 3.5 × 10–5 5

Liver-specific, scaling = 1.0 Mixed 1.84× 10–4 0.45 × 10–1 1.39 × 10–1 2.24 × 10–1 8.3 × 10–6 2.6 × 10–5 4

GST-T1+/+ 3.03 × 10–2 1.90 × 10–3 4.60 × 10–3 7.20 × 10–3 5.8 × 10–5 1.4 × 10–4 2

Whole-body metabolism Mixed 3.03 × 10–2 1.08 × 10–3 3.40 × 10–3 5.49 × 10–3 3.3 × 10–5 1.0 × 10–4 1

aGST-T1+/+ = homozygous, full enzyme activity; mixed = population reflecting estimated frequency of genotypes in current U.S. p

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 da

by dividing the BMR (0.1) by the allometric-scaled human BMDL10 and the mouse BMDL10 for the allometric-scaled and scaling 

factors, respectively (from inhalation unit risk data, Table 5-19). 

cMean, 95th, and 99th percentile of the human PBTK model-derived probability distribution of daily average internal dichlorometh

(milligrams dichloromethane metabolized via GST pathway per liter tissue per day) resulting from chronic oral exposure of 1 mg

dDerived by multiplying the dichloromethane tumor risk factor by the PBTK model-derived probabilistic internal doses from daily 

to 1 mg/kg-day. 

7837 
275

7838 5.4.1.7. Alternative Based On Administered Dose

7839 One comparison that can be made is with an alternative OSF based on liver tumors in 

7840 mice, using the external concentrations of dichloromethane in the mouse as converted to human 

7841 equivalent doses and then applying this by using BMD modeling to obtain the BMDL10 and 

7842 resulting oral cancer risk. Mouse bioassay exposures were adjusted to human equivalent doses 

7843 as follows: 

7844 

7845 human equivalent dose = 

7846 (nominal daily intake ÷ BW scaling factor) × daily exposure adjustment factor 

7847 

where BW scaling factor = (BWhuman/BWmouse)

 0.25 7848 = 7 

7849 and 

7850 daily exposure adjustment factor = 5/7 

7851 

7852 The human equivalent administered doses for the 0, 60, 125, 185, and 250 mg/kg-day 

7853 dose groups used in the liver tumor analysis (Table 5-11) (from Serota et al. [1986b]) were 0, 

7854 6.12, 12.75, 18.87,and 25.51 mg/kg-day, respectively. The BMD modeling and OSF derived 

7855 from these values are shown in Table 5-15. The resulting OSF, based on the liver tumors in the 

7856 mouse, is approximately one order of magnitude higher than the current recommended value 

7857 obtained by using the mouse and human PBTK models. 

7858 

Table 5-15. Cancer OSF based on a human BMDL10 using administered dose for 

liver tumors in male B6C3F1 mice exposed to dichloromethane in drinking water 

for 2 years 

Sex, 

tumor type BMDS modela 

χ

2

goodness of fit 

p- value

Human 

BMD10

c 

Human 

BMDL10

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 was 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 = (BWhuman/BWmouse)

0.25 = 7 and adjusting for daily exposure by 

multiplying by 5/7 days. Numbers in parentheses indicate (1) the number of dose groups dropped in order 

to obtain an adequate fit, starting with the highest dose group, and (2) the degree polynomial of the model. 

cBMD10 and BMDL10 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. 

dCancer OSF (risk per mg/kg-day) = 0.1/human BMDL10.

7859 

7860 

7861 The administered dose methodology can be considered equivalent to using a single7862 compartment, whole-body model of dichloromethane, where the internal dose metric is the AUC 

7863 of dichloromethane itself and clearance of dichloromethane scales from mice to humans as 

BW0.75 7864 . The estimates based on the PBTK model, in contrast, use the rate of metabolism of 
276

7865 dichloromethane (GST) as the metric. Another difference is that the administered dose 

7866 methodology does not account in any way for the GST polymorphism and so might be considered 

7867 as representing the general/mixed-GST-genotype population rather than the +/+ subpopulation. 

7868 

7869 5.4.1.8. Previous IRIS Assessment: Cancer Oral Slope Factor 

 The previous IRIS assessment derived a cancer OSF of 7.5 × 10–3 (mg/kg-day)–1 7870 by the 

7871 application of the multistage model to combined incidence of hepatocellular adenomas, 

7872 carcinomas, and neoplastic nodules from two studies. These were the 2-year drinking water 

7873 study of dichloromethane in B6C3F1 mice by the Hazelton Laboratories (1983) and the 2-year 

7874 inhalation study of dichloromethane in B6C3F1 mice by NTP (1986). The slope factor was the 

arithmetic mean of two candidate slope factors, 1.2 × 10–2 (mg/kg-day)–1 7875 (Hazelton Laboratories, 

1983) and 2.6 × 10–3 (mg/kg-day)–1 7876 (NTP, 1986). Since the NTP (1986) animal data were from 

7877 inhalation exposures, the estimated inhaled doses were calculated for mice and humans 

7878 (assuming near complete uptake into lung tissues and blood) and converted to administered 

doses in units of mg/kg-day. Assumed inhalation rates of 0.0407 and 20 m37879 /day were used for 

7880 mice and humans, respectively. No adjustments were made for species differences in 

7881 metabolism or toxicokinetics. 

7882 

7883 

7884 5.4.1.9. Comparison of Cancer Oral Slope Factors Using Different Methodologies 

7885 Cancer OSFs derived using different dose metrics and assumptions are summarized in 

Table 5-16. The recommended OSF of 1 × 10–3 7886 per mg/kg-day (rounded to one significant digit) 

7887 is based on a tissue-specific GST-internal dose metric, with allometric scaling (= 7), because of 

7888 some uncertainty regarding the rate of clearance of the relevant metabolite(s) formed via the 

GST pathway. The value derived specifically for the GST-T1+/+ 7889 population is recommended to 

7890 provide protection for the population that is hypothesized to be most sensitive to the carcinogenic 

effect. The values based on the GST-T1+/+ 7891 group are approximately two-fold higher than those

7892 for the full population for the dose metrics used in this assessment (Table 5-16). Within a 

7893 genotype population, the values of the OSF among most of the various dose metrics vary by 

7894 about one to two orders of magnitude. 
277

7895 

Table 5-16. Comparison of OSFs derived using various assumptions and metrics, based on tumors in male mice 

Populationa Dose metric 

Species, 

sex Tumor 

Scaling 

factor 

Mean O

(mg/kg-d

GST-T1+/+ Tissue-specific GST-metabolism rateb Mouse, male Liver 7.0 1.4 × 1

 Tissue-specific GST-metabolism rate Mouse, male Liver 1.0 2.0 × 1

 Whole-body GST metabolism rate Mouse, male Liver 7.0 8.1 × 1 Route-to-route extrapolation, tissue-specific metabolism Mouse, male Liver 7.0 1.0 × 1 Route-to-route extrapolation, tissue-specific metabolism Mouse, male Liver 1.0 1.5 × 1 Route-to-route extrapolation, whole-body metabolism Mouse, male Liver 7.0 5.8 × 1

Mixed Tissue-specific GST-metabolism rateb Mouse, male Liver 7.0 8.0 × 1

 Tissue-specific GST-metabolism rate Mouse, male Liver 1.0 1.2 × 1

 Whole-body GST metabolism rate Mouse, male Liver 7.0 4.6 × 1 Route-to-route extrapolation, tissue-specific metabolism Mouse, male Liver 7.0 5.8 × 1 Route-to-route extrapolation, tissue-specific metabolism Mouse, male Liver 1.0 8.3 × 1 Route-to-route extrapolation, whole-body metabolism Mouse, male Liver 7.0 3.3 × 1 Applied dose (human equivalent dose) Mouse, male Liver 1.0 × 1

 1995 IRIS assessment Mouse, male Liver 7.5 × 1

aGST-T1+/+ = homozygous, full enzyme activity; Mixed = genotypes based on a population reflecting the estimated frequency of gen

U.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 × 10–3 per mg/kg-day. 

7896 
278

7897 

7898 5.4.2. Cancer Inhalation Unit Risk 

7899 5.4.2.1. Choice of Study/Data—with Rationale and Justification 

7900 As discussed in section 4.7, results from several cohort mortality studies of workers 

7901 repeatedly exposed to dichloromethane and several case-control studies provide some supporting 

7902 evidence of carcinogenicity in humans, specifically with respect to liver and brain cancer. 

7903 However, the epidemiologic studies do not provide adequate data to estimate exposure-response 

7904 relationships for dichloromethane exposure and these cancers. 

7905 Results from several bioassays provide sufficient evidence of the carcinogenicity of 

7906 dichloromethane in mice and rats exposed by inhalation, as well as adequate data to describe 

7907 dose-response relationships. As discussed in section 4.7.2, repeated inhalation exposure to 

7908 concentrations of 2,000 or 4,000 ppm dichloromethane produced increased incidences of lung 

7909 and liver tumors in male and female B6C3F1 mice (Maronpot et al., 1995; Foley et al., 1993; 

7910 Kari et al., 1993; Mennear et al., 1988; NTP, 1986). A weaker trend was seen with respect to 

7911 liver tumor incidence (described as neoplastic nodules or hepatic carcinomas) in female rats, but 

7912 this trend was not seen when limited to hepatic carcinomas (NTP, 1986). A statistically 

7913 significant increased incidence of brain tumors has not been observed in any of the animal cancer 

7914 bioassays, but a 2-year study using relatively low exposure levels (0, 50, 200, and 500 ppm) in 

7915 Sprague-Dawley rats observed a total of six astrocytoma or glioma (mixed glial cell) tumors 

7916 (combining males and females) in the exposed groups (Nitschke et al., 1988a). These tumors are 

7917 exceedingly rare in rats, and there are few examples of statistically significant trends in animal 

7918 bioassays (Sills et al., 1999). Male and female F344 rats exposed by inhalation to 2,000 or 4,000 

7919 ppm showed significantly increased incidences of benign mammary tumors (adenomas or 

7920 fibroadenomas) and the male rats also exhibited a low rate of sarcoma or fibrosarcoma in 

7921 mammary gland or subcutaneous tissue around the mammary gland (NTP, 1986). 

7922 The NTP inhalation study in B6C3F1 mice (NTP, 1986) was used to derive an IUR for 

7923 dichloromethane because of the completeness of the data, adequate sample size, and clear dose 

7924 response with respect to liver and lung tumors. The liver tumor incidence in male mice 

7925 increased from 44% in controls to 66% in the highest dose group; in females the incidence of this 

7926 tumor rose from 6 to 83%. For lung tumors, the incidence rose from 10 to 80% in males and 

7927 from 6 to 85% in females. Compelling evidence exists for the role of GST-mediated metabolism 

7928 of dichloromethane in carcinogenicity in mice (section 4.7.3), and both mice and humans possess 

7929 this metabolic pathway. Modeling intake, metabolism, and elimination of dichloromethane in 

7930 mice and humans is feasible. Thus, it is reasonable to apply the best available PBTK models to 

7931 estimate equivalent internal doses in mice and humans. 

7932 The mammary tumor data from the NTP (1986) study was also used to derive a 

7933 comparative IUR. However, the toxicokinetic or mechanistic events that might lead to 

7934 mammary gland tumor development in rats are unknown, and so a clear choice of the optimal 
279

7935 internal dose metric could not be made. Thus, this derivation is based on the average daily AUC 

7936 for dichloromethane in blood. The role of CYP- or GST-mediated metabolism in the mammary 

7937 gland is uncertain, although both GST-T1 (Lehmann and Wagner, 2008) and CYP2E1 (El-Rayes 

7938 et al., 2003; Hellmold et al., 1998) expressions have been detected in human mammary tissue. It 

7939 is also possible that some metabolites enter systemic circulation from the liver and lung, where 

7940 they are formed. 

7941 The female rat liver cancer data from the NTP (1986) inhalation study was not used to 

7942 derive an IUR because the trend was weaker than that seen in the mouse (incidence increased 

7943 from 4% in controls to 10% in the highest dose group, trend p = 0.08), and because the effect 

7944 categorization included neoplastic nodule or hepatocellular carcinoma. The brain tumor data 

7945 seen in the Nitschke et al. (1988a) study in Sprague-Dawley rats were not used to develop an 

7946 IUR because of the low incidence of this rare tumor (a total of four astrocytoma or glioma 

7947 tumors in exposed males and two in exposed females). The mechanistic issues with respect to 

7948 mammary tumors and health effects issues with respect to brain tumors represent data gaps in the 

7949 understanding of the health effects of dichloromethane and relevance of the rat data to humans. 

7950 

7951 5.4.2.2. Derivation of the Cancer Inhalation Unit Risk 

7952 The derivation of the IUR parallels the process described in section 5.4.1.2 for the cancer 

7953 OSF. Since modeling metabolism and elimination kinetics of dichloromethane in mice and 

7954 humans is feasible, it is reasonable to apply the best available PBTK models to determine 

7955 equivalent target organ doses in mice and humans. 

7956 

7957 5.4.2.3. Dose-Response Data

7958 Data for liver and lung tumors in male and female B6C3F1 mice following exposure to 

7959 airborne dichloromethane were used to develop IURs for dichloromethane (Mennear et al., 1988; 

7960 NTP, 1986). As discussed in section 5.4.1.8, the liver tumor dose-response data were also the 

7961 basis of an OSF, derived by route-to-route extrapolation using the PBTK models, to compare 

7962 with an OSF based on liver tumor data in mice exposed to dichloromethane in drinking water 

7963 (Serota et al., 1986b). In the NTP (1986) study, significant increases in incidence of liver and 

7964 lung adenomas and carcinomas were observed in both sexes of B6C3F1 mice exposed 

7965 6 hours/day, 5 days/week for 2 years (Table 5-17). Since significant decreases in survival were 

7966 observed in the treated groups of both sexes, the at-risk study populations (represented by the 

7967 denominators in the incidence data) were determined by excluding all animals dying prior 52 

7968 weeks.
280

7969 

7970 5.4.2.4. Dose Conversion and Extrapolation Methods: Cancer Inhalation Unit Risk

7971 Dose conversion 

7972 The PBTK model of Marino et al. (2006) for dichloromethane in the mouse was used to 

7973 simulate inhalation exposures of 6 hours/day, 5 days/week (Mennear et al., 1988; NTP, 1986) 

7974 and to calculate long-term daily average internal doses. Study-, group-, and sex-specific mean 

7975 BWs were used. Based on evidence that metabolites of dichloromethane produced via the GST 

7976 pathway are primarily responsible for dichloromethane carcinogenicity in mouse liver and lung 

7977 (summarized in section 4.7.3), and the assumption that these metabolites are sufficiently reactive 

7978 that they do not have substantial distribution outside these tissues, the recommended selected 

7979 internal dose metric for liver tumors and lung tumors were long-term average daily mass of 

7980 dichloromethane metabolized via the GST pathway per unit volume of liver and lung, 

7981 respectively (Table 5-17). Figure 5-16 show the comparison between inhalation external and 

7982 internal doses in the liver and lung, respectively, using this dose metric for the mouse and for the 

7983 human. A whole-body metabolism metric was also examined. This metric would be more 

7984 relevant under a scenario of slowly cleared metabolites that undergo general circulation 

7985 

Table 5-17. Incidence data for liver and lung tumors and internal doses, based on GST 

metabolism dose metrics, in male and female B6C3F1 mice exposed to dichloromethane via 

inhalation for 2 years 

Sex, 

tumor type BW (g) 

External 

dichloromethane 

concentration 

(ppm) 

Mouse 

tumor incidence

Mouse internal 

tissue dosea 

Mouse whole body 

metabolism doseb

-- 0 22/50 (44%)e Male, liver 0 0 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, lungd

32.0 4,000 40/47 (85%) 992.2 210.7 

-- 0 3/47 (6%)e

Female, liver 0 0 c

30.0 2,000 16/46 (35%) 2,453.2 104.0 

 29.0 4,000 40/46 (87%) 5,120.0 217.0 
281

Table 5-17. Incidence data for liver and lung tumors and internal doses, based on GST 

metabolism dose metrics, in male and female B6C3F1 mice exposed to dichloromethane via 

inhalation for 2 years 

Sex, 

tumor type BW (g) 

External 

dichloromethane 

concentration 

(ppm) 

Mouse 

tumor incidence

Mouse internal 

tissue dosea 

Mouse whole body 

metabolism doseb

-- 0 3/45 (6%)e 0 0 

30.0 2,000 30/46 (65%) 493.0 104.0 

Female, lungd

29.0 4,000 41/46 (89%) 1,021.8 217.0 

a

For liver tumors: mg dichloromethane metabolized via GST pathway/L liver tissue/day from 6 hours per day, 5 

days per week exposure; for lung tumors: mg dichloromethane metabolized via GST pathway/L lung tissue/day 

from 6 hours per day, 5 days per week exposure. 

bBased on the sum of dichloromethane metabolized via the GST pathway in the lung plus the liver, normalized to 

total BW (i.e., [lung GST metabolism (mg/day) + liver GST metabolism (mg/day)]/kg BW). Units = mg 

dichloromethane metabolized via GST pathway in lung and liver/kg-day. 

cHepatocellular carcinoma or adenoma. Mice dying prior to 52 weeks were excluded from the denominators. 

dBronchoalveolar carcinoma or adenoma. Mice dying prior to 52 weeks were excluded from the denominators. 

e

Statistically significant increasing trend (by incidental and life-table tests; p ≤ 0.01). 

Sources: Mennear et al., 1988; NTP, 1986 

7986 

7987 

7988 A. 

7989 

7990 

7991 B. 
282

7992 

7993 Figure 5-16. PBTK model-derived internal doses (mg dichloromethane metabolized via the 

7994 GST pathways per liter tissue per day) for liver (A) and lung (B) in mice and humans, and 

7995 their associated external exposures (ppm), used for the derivation of cancer inhalation unit 

7996 risks. Average daily doses were calculated from simulated mouse exposures of 6 hours/day, 5 

7997 days/week, while simulated human exposures were continuous. The GST metabolism rate in each 

7998 simulated human population was obtained by generating 1000 random samples from each population 

7999 (ages 0.5-80 years, males and females) for each exposure level, and calculating the average GST 

8000 metabolic rate for each sample.

8001 

8002 Dose-response modeling and extrapolation 

8003 The multistage dose-response model was fit to the mouse tumor incidence and PBTK 

8004 model-derived internal dose data to derive mouse internal BMD10 and BMDL10 values associated 

8005 with 10% extra risk (Table 5-18). Different polynomial models and models dropping dose 

8006 groups starting with the highest dose group were compared based on adequacy of model fit as 

assessed by overall χ

2

8007 goodness of fit (p-value > 0.10) and examination of residuals, particularly 

8008 in the region of the BMR (U.S. EPA, 2000c). Appendix E-2 provides details of the BMD 

8009 modeling results for the male. The mouse liver and lung tumor risk factors (extra risk per unit 

8010 internal dose) were calculated by dividing 0.1 by the mouse BMDL10 for liver and lung tumors, 

8011 respectively. 

8012 Linear extrapolation from the internal BMDL10 (0.1/BMDL10) was used to derive 

8013 inhalation risk factors for lung and liver tumors in male and female mice (Table 5-18). As 

8014 discussed in section 4.7, the linear low-dose extrapolation approach for agents with a mutagenic 

8015 mode of action was selected. 

8016 

8017 
283

8018 

Table 5-18. BMD modeling results and tumor risk factors associated with 10% extra risk for liver and lung tumors in 

male and female B6C3F1 mice exposed by inhalation to dichloromethane for 2 years, based on liver-specific GST 

metabolism and whole body GST metabolism dose metrics

Tumor Risk 

Internal dose 

metrica

BMDS 

modelb 

χ2 

goodness of fit 

p-value

Mouse 

BMD10c 

Mouse 

BMDL10c 

Allometricscaled human 

BMDL10d Scaling = 1.0

Male, liver MS (0,1) 0.40 913.9 544.4 77.8 1.84 × 10–4 Male, lung MS (0,1) 0.64 61.7 48.6 7.0 2.06 × 10–3 

Female, liver MS (0,2) 0.53 1224.1 659.7 94.2 1.52 × 10–4 

Liver-specific 

Female, lung MS (0,1) 0.87 51.2 40.7 5.8 2.46 × 10–3 

Male, liver MS (0,1) 0.40 38.7 23.1 3.3 -- 

Male, lung MS (0,1) 0.66 13.1 10.3 1.5 -- Female, liver MS (0,2) 0.53 51.9 28.0 4.0 -- 

Whole body 

Female, lung MS (0,1) 0.88 10.8 8.6 1.2 -- 

aLiver specific dose units = mg dichloromethane metabolized via GST pathway per liter tissue per day; Whole-body dose units = mg 

dichloromethane metabolized via GST pathway in lung and 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 do

metrics calculated with the mouse PBTK model. Numbers in parentheses indicate (1) the number of dose groups dropped in order to 

obtain an adequate fit and (2) the degree polynomial of the model. 

cBMD10 and BMDL10 refer 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 BMDL10 divided 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 BMDL10 and by 

allometric-scaled human BMDL10,, for the scaling = 1.0 and allometric-scaled risk factors, respectively. 

8019 8020 8021 
284

8022 Application of allometric scaling factor 

8023 As discussed in section 5.4.1.4., the choice of a scaling factor is based on the question of 

8024 the role of specific metabolites, and particularly how these metabolites are transformed or 

8025 removed. If the key metabolite is established and is known to be sufficiently reactive to not 

8026 spread in systemic circulation, then it can be assumed that 1) the level of reactivity and rate of 

8027 clearance (i.e., disappearance due to local reactivity) for this metabolite, per volume tissue, is 

8028 equal in rodents and humans; and (2) risk is proportional to the long-term daily average 

8029 concentration of the metabolite. Under these assumptions, rodent internal BMDL10 values based 

8030 on tissue-specific dichloromethane metabolism require no allometric scaling to account for 

8031 toxicodynamic differences and predict the corresponding level of human risk as a function of the 

8032 metric (i.e., the scaling factor in Figure 5-14 was equal to 1.0). (A single metabolite is 

8033 referenced, but the same argument holds in general for more than one metabolite). Under this 

8034 scenario and assumptions, humans and rodents with the same long-term daily average metabolite 

8035 formation per volume tissue (e.g., equal internal BMDL10) should both experience the same 

8036 long-term average concentration of the metabolite when the metabolite is highly reactive and 

8037 hence experience the same extra risk. However, some uncertainties remain concerning the 

8038 hypothesized role of a highly reactive metabolite in the carcinogenic effects seen with 

8039 dichloromethane. The active metabolite(s) have not been established, and data pertaining to the 

8040 reactivity or clearance rate of these metabolite(s) are lacking. For example, quantitative 

8041 measurements of adducts of interest or of the half life of relevant compounds in humans and in 

8042 mice are not available. To address these uncertainties, use of scaling factor that addresses the 

8043 possibility that the rate of clearance for the metabolite is limited by processes that are known to 

8044 scale allometrically, such as blood perfusion or enzyme activity, may be appropriate.. This case 

would result in use of a mouse:human dose-rate scaling factor of (BWhuman/BWmouse)

0.25 8045 = 7 to 

8046 adjust the mouse-based BMDL10 values downward. Using this internal dose metric (liver8047 specific metabolism with allometric scaling), equivalent rodent and human internal BMDL10

8048 values result in a human liver tumor risk factor (0.1/BMDL10) that is assumed equal to that for 

8049 the mouse, given a 70-year lifetime exposure. Another alternative that can be used is based on 

8050 an allometrically-scaled whole-body metabolism metric. In this case, less weight is given to the 

8051 evidence of site-specificity of the effects. As with the OSF derivations, the cancer toxicity 

8052 values derived using each of these metrics and scaling factors (i.e., liver-specific metabolism 

8053 with and without allometric-scaling and the whole-body metabolism metric) are presented in the 

8054 following tables. Considering the lack of data pertaining to clearance rates or the actual AUC of 

8055 the active carcinogenic metabolite(s) in mice and humans, the IUR recommended by the EPA are 

8056 based on the allometrically-scaled tissue-specific GST metabolism rate dose metric. 

8057 

8058 Calculation of IURs 
285

8059 A probabilistic PBTK model for dichloromethane in humans adapted from David et al. 

8060 (2006) (see Appendix B) was used with Monte Carlo sampling to calculate distributions of 

internal lung, liver, or blood doses associated with chronic unit inhalation (1 μg/m38061 ) exposures. 

8062 The data on which the model is based indicate that relationship between exposure and internal 

8063 dose is linear at low doses. Parameters in the human PBTK model developed by David et al. 

8064 (2006) are distributions that incorporate information about dichloromethane toxicokinetic 

8065 variability and uncertainty among humans. Monte Carlo sampling was performed in which each 

8066 human model parameter was defined by a value randomly drawn from each respective parameter 

8067 distribution. The model was then executed by using the external unit exposure as input, and the 

8068 resulting human equivalent internal dose was recorded. This process was repeated for 10,000 

8069 iterations to generate a distribution of human internal doses. 

8070 The resulting distribution of IURs shown in Table 5-19 were derived by multiplying the 

8071 human internal dose tumor risk factor (in units of reciprocal internal dose) by the respective 

8072 distributions of human average daily internal dose resulting from a chronic unit inhalation 

exposure of 1 μg/m3

8073 dichloromethane. Table 5-19 presents the analysis using the male data. 

8074 Analyses based on the female data produced very similar results, and are summarized in 

8075 Appendix F. Because adjustments for interindividual variability are not generally used or 

8076 recommended in cancer risk analysis, the mean slope factor was selected as the recommended 

8077 value to be used in deterministic risk assessments; other values at the upper end of the 

8078 distribution are also presented. As with the oral cancer slope factor derivation, the cancer IUR is 

8079 derived for a population composed entirely of carriers of the GST-T1 homozygous positive 

8080 genotype (the group that would be expected to be most sensitive to the carcinogenic effects of 

8081 dichloromethane), and a population reflecting the estimated frequency of GST-T1 genotypes in 

the current U.S. population (20% GST-T1–/–, 48% GST-T1+/–, and 32% GST-T1+/+ 8082 , the “mixed” 

8083 population). All simulations also included a distribution of CYP activity, based on data from 

8084 Lipscomb et al. (2003). 

8085 

8086 
286

8087 

Table 5-19. IURs for dichloromethane based on PBTK model-derived internal liver and lung doses in B6C3F1 male m

via inhalation for 2 years, based on liver-specific GST metabolism and whole body metabolism dose metrics, by popula

genotype 

Distribution of human internal 

dichloromethane doses from 1 μg/m3

exposured 

Resulting cand

IURe(μg

Internal dose metric 

and scaling factora

Population 

genotypeb Tumor type 

Human 

tumor risk 

factorc Mean 

95th

percentile

99th

percentile Mean 

95t

percen

GST-T1+/+ Liver 1.29 × 10–3 5.64 × 10–6 1.56 × 10–5 2.60 × 10–5 7.3 × 10–9 2.0 × GST-T1+/+ Lung 1.44 × 10–2 3.31 × 10–7 8.55 × 10–7 1.34 × 10–6 4.8 × 10–9 1.2 × Mixed Liver 1.29 × 10–3 2.62 × 10–6 8.65 × 10–6 1.45 × 10–5 3.4 × 10–9 1.1 × 

Tissue-specific, allometric-scaled 

Mixed Lung 1.44 × 10–2 1.81 × 10–7 5.67 × 10–7 9.84 × 10–7 2.6 × 10–9 8.2 × 

GST-T1+/+ Liver 1.84 × 10–4 5.64 × 10–6 1.56 × 10–5 2.60 × 10–5 1.0 × 10–9 2.9 × GST-T1+/+ Lung 2.06 × 10–3 3.31 × 10–7 8.55 × 10–7 1.34 × 10–6 6.8 × 10–10 1.8 × Mixed Liver 1.84 × 10–4 2.62 × 10–6 8.65 × 10–6 1.45 × 10–5 4.8 × 10–10 1.6 × 

Tissue-specific, scaling = 1.0 

Mixed Lung 2.06 × 10–3 1.81 × 10–7 5.67 × 10–7 9.84 × 10–7 3.7 × 10–10 1.2 × 

GST-T1+/+ Liver 3.03 × 10–2 1.53 × 10–7 4.87 × 10–7 9.20 × 10–7 4.6 × 10–9 1.5 × GST-T1+/+ Lung 6.80 × 10–2 1.53 × 10–7 4.87 × 10–7 9.20 × 10–7 1.0 × 10–8 3.3 × Mixed Liver 3.03 × 10–2 8.76 × 10–8 3.20 × 10–7 6.76 × 10–7 2.7 × 10–9 9.7 × 

Whole-body, allometric-scaled 

Mixed Lung 6.80 × 10–2 8.76 × 10–8 3.20 × 10–7 6.76 × 10–7 6.0 × 10–9 2.2 × 

aTissue specific dose units = mg dichloromethane metabolized via GST pathway per liter tissue (liver or lung, respectively, for liver and 

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 genotypes in current U.S. pop

/–, 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 huma

mouse BMDL10 (from Table 5-18) for the allometric-scaled and scaling = 1.0 risk factors, respectively. 

dMean, 95th, and 99th percentile of the human PBTK model-derived probability distribution of daily average internal dichloromethane d

chronic exposure to 1 μg/m3 (0.00029 ppm). 

eDerived by multiplying the dichloromethane tumor risk factor by the PBTK model-derived probabilistic internal doses from daily expo

8088 
287

8089 5.4.2.5. Cancer Inhalation Unit Risk 

 The recommended cancer IURs are 7 × 10–9(μg/m3

)

–1 and 5 × 10–9 (μg/m3

)

–1 8090 for the 

development of liver and lung cancer, respectively, based on the mean for the GST-T1+/+ 8091 

8092 population (the group with the greatest presumed sensitivity). These values are based on male 

8093 B6C3F1 mice, using a tissue-specific GST metabolism dose metric, with allometric scaling 

8094 (Table 5-19). Risk estimates were slightly higher for liver tumors and essentially equivalent for 

8095 lung tumors in males compared to females (Appendix F), so the estimates for males were 

8096 selected for the candidate values. 

8097 

8098 Consideration of combined risk (summing risk across tumors)

8099 With two significant tumor sites, focusing on the more sensitive response may 

8100 underestimate the overall cancer risk associated with exposure to this chemical. Following the 

8101 recommendations of the National Research Council (NRC, 1994) and the Guidelines for 

8102 Carcinogen Risk Assessment (U.S. EPA, 2005a), an upper bound on total risk was estimated in 

8103 order to gain some understanding of the total risk from multiple tumor sites in the selected data 

8104 set. Note that this estimate of overall risk describes the risk of developing either tumor type, not 

8105 just the risk of developing both simultaneously. 

8106 NRC (1994) stated that an approach based on counts of animals with one or more tumors 

8107 (or tumor-bearing animals) would tend to underestimate overall risk when tumor types occur 

8108 independently and that an approach based on combining the risk estimates from each separate 

8109 tumor type should be used. For dichloromethane, there is no reason to expect that the occurrence 

8110 of one tumor type depends on the incidence of the other, given the association of different dose 

8111 metrics with each tumor response. Therefore it appears reasonable to assume that the two tumor 

8112 types occur independently. However, simply summing upper limit risks may result in an 

8113 overestimation of overall of combined risk because of the statistical issues with respect to 

8114 summing variances of distributions. An additional challenge results from the use of different 

8115 internal dose metrics for different tumors, as is the case with the dose metrics based on tissue8116 specific metabolism. Statistical methods based on a common metric can not be used with the 

8117 tissue-specific metabolism metric used in these derivations. 

8118 An alternative approach is to derive an upper bound on the combined risk estimates by 

8119 summing central tendency risks and calculating a pooled SD by using BMD10 and BMDL10

8120 values for liver and lung tumors. The SD associated with the IUR for each tumor site is 

8121 calculated as the difference between 95th percentiles of the distribution for upper bound and 

8122 maximum likelihood estimate IURs (based on either female or male mouse tumor risk factors), 

8123 divided by 1.645 (the relevant t statistic, assuming normal distributions of summed quantities). 

8124 Variances for each tumor site are the squares of the SDs. Pooled variance and SD are defined as 

8125 the sum of variances for lung and liver tumors and the square root of that sum, respectively. 

8126 Finally, the upper bound on the combined lung and liver cancer risk is determined by multiplying 
288

8127 the cumulative SD by 1.645 and adding it to the summed central tendency IURs. The 

8128 calculations of these upper bound estimates for combined liver and lung tumor risks are shown in 

8129 Table 5-20. 

8130 Using this approach and the male mouse-derived risk factors, the combined human 

equivalent IUR values for both tumor types is 1 × 10

–8 (μg/m3

)

 –1 (rounded from 1.1 × 10–8 8131 ) in 

the most sensitive (GST-T1+/+ 8132 ) population. This is the recommended inhalation cancer unit risk 

8133 value to be used in deterministic risk assessments for chronic exposure to dichloromethane. The 

8134 corresponding value for a population with the frequency distribution of GST-T1 genotypes 

currently found in the U.S. population is 6 × 10–9 (μg/m3)

 –1 8135 . 
289

8136 

Table 5-20. Upper bound estimates of combined human IURs for liver and lung tumors resulting from 

exposure to 1 μg/m3 dichloromethane, based on liver-specific GST metabolism and whole body metabol

metrics, by population genotype 

 Internal dose 

metric and scaling 

factora

Population genotypeb Tumor site 

Upper 

bound IURc 

Central 

tendency 

IURd 

Variance of 

tissue-specific 

tumor riske 

Combined 

tumor risk 

SDf 

Liver 7.3 × 10–9 4.3 × 10–9 3.17 × 10–18 Lung 4.8 × 10–9 3.8 × 10–9 3.75 × 10–19 

Tissue-specific, allometric-scaled 

GST-T1+/+

Liver or lung 8.1 × 10–9 1.9 × 10–9 

Liver 3.4 × 10–9 2.0 × 10–9 6.84 × 10–19 Lung 2.6 × 10–9 2.1 × 10–9 1.12 × 10–19 

 Mixed 

Liver or lung 4.1 × 10–9 8.9 × 10–10 

Liver 1.0 × 10–9 6.2 × 10–10 6.48 × 10–20Lung 6.8 × 10–10 5.4 × 10–10 7.72 × 10–21

GST-T1+/+

Liver or lung 1.2 × 10–9 2.7 × 10–10 

Liver 4.8 × 10–10 2.9 × 10–10 1.40 × 10–20

Lung 3.7 × 10–10 2.9 × 10–10 2.31 × 10–21

Tissue-specific, scaling = 1.0 

Mixed 

Liver or lung 5.8 × 10–10 1.3 × 10–10 

Liver 4.6 × 10–9 2.8 × 10–9 1.29× 10–18

Lung 1.0 × 10–8 8.2 × 10–9 1.84 × 10–18

GST-T1+/+

Liver or lung 1.1 × 10–8 1.8 × 10–9 

Liver 2.7 × 10–9 1.6 × 10–9 4.23 × 10–19

Lung 6.0 × 10–9 4.7 × 10–9 6.04 × 10–19

Whole-body, allometric-scaled 

Mixed 

Liver or lung 6.3 × 10–9 1.0 × 10–9 

aTissue specific dose units = mg dichloromethane metabolized via GST pathway per liter tissue (liver or lung, respectively, for liv

lung 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 genotypes in current U.S

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 the t statistic, 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 the t statistic, 1.645, added to the sum of central tendency IURs. 

8137 
290

8138 5.4.2.6. Comparative Derivation Based on Rat Mammary Tumor Data 

8139 Mammary gland tumor data from male and female F344 rats following an inhalation 

8140 exposure to dichloromethane were considered in development of a comparative IUR for 

8141 dichloromethane (Mennear et al., 1988; NTP, 1986). In both the male and female rats, there 

8142 were significant increases in the incidence of adenomas, fibroadenomas, or fibromas in or near 

8143 the mammary gland. These were characterized as benign tumors in the NTP report (NTP, 1986). 

8144 Increased numbers of benign mammary tumors per animal in exposed groups were also seen in 

8145 two studies of Sprague-Dawley rats (Nitschke et al., 1988a; Burek et al., 1984). An oral 

8146 (gavage) study in Sprague-Dawley rats reported an increased incidence of malignant mammary 

8147 tumors, mainly adenocarcinomas (8, 6, and 18% in the control, 100, and 500 mg/kg dose groups, 

8148 respectively), but the increase was not statistically significant. Data were not provided to allow 

8149 an analysis that accounts for differing mortality rates (Maltoni et al., 1988). There are 

8150 considerably more uncertainties regarding the interpretation of these data with respect to 

8151 carcinogenic risk compared with the data pertaining to liver and lung tumors. The trends were 

8152 driven in large part by benign tumors; adenocarcinomas and carcinomas were seen only in the 

8153 females, with incidences of 1, 2, 2, and 0 in the 0, 1000, 2000 and 4000 ppm exposure groups, 

8154 respectively. There are little data to guide the choice of relevant dose metric, and the 

8155 genotoxicity and mechanistic studies have not included mammary tissue. For these reasons, the 

8156 analysis and the calculation of the comparative IUR based on rat mammary tumor data are 

presented in Appendix G. The IUR based on the female rat data was 1 × 10–7 (μg/m3

)

–1 8157 . 

8158 

8159 5.4.2.7. Alternative Based on Administered Concentration 

8160 Another comparison that can be made is with an alternative IUR based on liver and lung 

8161 tumors in mice, using the external concentrations of dichloromethane in the mouse studies as 

8162 converted to HECs, and then applying this using BMD modeling to obtain the BMDL10 and 

8163 resulting IUR. Mouse bioassay exposures were adjusted to HECs as follows: 

8164 • Adjusting to continuous exposure: External concentrationADJ = External 

8165 concentration × (6 hours/24 hours) × (5 days/7 days) 

• Concentrations in mg/m38166 = concentrations in ppm × 84.93/24.45. 

8167 • [Hb/g]A/[Hb/g]H = the ratio of blood:gas (air) partition coefficients in animals and 

8168 humans. Because the partition coefficient for mice (23.0) is higher than for humans 

8169 (9.7), a value of 1.0 was used, as per U.S. EPA (1994b) guidance. 

8170 

8171 Thus, 

8172 HECs = External concentrationADJ × [Hb/g]A/[Hb/g]H = External concentrationADJ × 1 

8173 
291

The HECs (mg/m38174 ) for the 0, 2,000, and 4,000 ppm exposure groups were 0, 1,241, and 

2,481 mg/m38175 , respectively. The BMD modeling and IURs derived from these values, in 

8176 conjunction with the liver and lung tumor data from Table 5-17 (NTP, 1986), are shown 

8177 in Table 5-21. The resulting IURs, based on the liver or lung tumors in the mouse, are 

8178 approximately one order of magnitude higher than the currently recommended value 

8179 obtained by using the mouse and human PBTK models. 

8180 

8181 

Table 5-21. Inhalation units risks based on human BMDL10 values using 

administered concentration for liver and lung tumors in B6C3F1 mice 

exposed by inhalation to dichloromethane for 2 years 

Sex, 

tumor type BMDS modela 

χ

2 

goodness of fit 

p-value BMD10

b BMDL10b 

Inhalation 

unit riskc 

(μg/m3)

–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 was fit to each of the four sets of mouse dose-response 

data shown in Table 5-17. The HEC used in these models for the 0, 2,000,and 4,000 ppm exposure groups were 

0, 1,241, and 2,481 mg/m3, 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. 

cBMD10 and BMDL10 refer to the BMD-model-predicted HECs (mg dichloromethane per cubic meter), and its 

95% lower confidence limit, associated with a 10% extra risk for the incidence of tumors. 

d

IUR, (risk/mg-m3) = 0.1/human BMDL10. 

Sources: Mennear et al. (1988); NTP (1986). 

8182 

8183 

8184 The difference between the administered concentration methodology and PBTK-based 

8185 approaches depends on two key differences: the use of a dichloromethane-metabolite dose8186 metric, rather than dichloromethane AUC, and the fact that the rate of dichloromethane 

8187 conversion to that metabolite is estimated in humans by using human data rather than default 

allometric scaling (BW0.75 8188 ). In addition, the administered concentration methodology does not 

8189 account in any way for the GST polymorphism and so might be considered as representing the 

8190 general/mixed-GST-genotype population rather than the +/+ subpopulation. 

8191 

8192 5.4.2.8. Previous IRIS Assessment: Cancer Inhalation Unit Risk

8193 The IUR in the previous IRIS assessment was determined from the combined incidence 

8194 of liver and lung adenomas and carcinomas in B6C3F1 mice exposed to dichloromethane for 

2 years by NTP (1986). A value of 4.7 × 10–7

 (μg/m3)

–1

8195 was derived by the application of a 

8196 modified version of the PBTK model of Andersen et al. (1987), which incorporated the 
292

8197 pharmacokinetics and metabolism of dichloromethane. Internal dose estimates, based on 

8198 dichloromethane metabolism via the GST pathway, were used and corrected for differences in 

8199 interspecies sensitivity by applying an interspecies scaling factor of 12.7, which was based on 

8200 dose equivalence adjusted to BW to the 2/3 power, to the human risks (Rhomberg, 1995; U.S. 

8201 EPA, 1987a). 

8202 5.4.2.8. Comparison of Cancer Inhalation Unit Risk Using Different Methodologies 

8203 In this assessment, cancer IURs derived by using different dose metrics and assumptions 

were examined, as summarized in Table 5-22. The recommended IUR value of 1 × 10–88204 

(μg/m3)

–1 8205 is based on a tissue-specific GST-internal dose metric, with allometric scaling, because 

8206 of the evidence for the involvement of highly reactive metabolites formed via the GST pathway. 

The value derived specifically for the GST-T1+/+ 8207 population is recommended to provide 

8208 protection for the population that is hypothesized to be most sensitive to the carcinogenic effect. 

The values based on the GST-T1+/+ 8209 group are approximately two to fivefold higher than those for 

8210 the full population, for all dose metrics used in this assessment. Within a genotype population, 

8211 the values of the IUR among the various dose metrics vary by about one to two orders of 

8212 magnitude. 
293

8213 

Table 5-22. Comparison of IURs derived by using various assumptions and metrics 

Populationa Dose metric Species, sex, Tumor type 

Scaling 

factor 

IU

(μg/m

GST-T1+/+ Tissue-specific GST-metabolism ratec Mouse, male Liver and lung 7.0 1.1 × GST-T1+/+ Tissue-specific GST-metabolism rate Mouse, male Liver 7.0 7.3 × 

GST-T1+/+ Tissue-specific GST-metabolism rate Mouse, male Lung 7.0 4.8 × GST-T1+/+ Tissue-specific GST-metabolism rate Mouse, male Liver and lung 1.0 1.6 × GST-T1+/+ Tissue-specific GST-metabolism rate Mouse, male Liver 1.0 1.0 × GST-T1+/+ Tissue-specific GST-metabolism rate Mouse, male Lung 1.0 6.8 × 

GST-T1+/+ Whole-body GST metabolism rate Mouse, male Liver and lung 7.0 1.4 × 

GST-T1+/+ Whole-body GST metabolism rate Mouse, male Liver 7.0 4.6 × GST-T1+/+ Whole-body GST metabolism rate Mouse, male Lung 7.0 1.0 × 

Mixed Tissue-specific GST-metabolism rate Mouse, male Liver and lung 7.0 5.5 × 

Mixed Tissue-specific GST-metabolism rate Mouse, male Liver 7.0 3.4 × 

Mixed Tissue-specific GST-metabolism rate Mouse, male Lung 7.0 2.6 × Mixed Tissue-specific GST-metabolism rate Mouse, male Liver and lung 1.0 7.9 × 

Mixed Tissue-specific GST-metabolism rate Mouse, male Liver 1.0 4.8 × Mixed Tissue-specific GST-metabolism rate Mouse, male Lung 1.0 3.7 × Mixed Whole-body GST metabolism rate Mouse, male Liver and lung 7.0 7.9 × 

Mixed Whole-body GST metabolism rate Mouse, male Liver 7.0 2.7 × Mixed Whole-body GST metabolism rate Mouse, male Lung 7.0 6.0 × 

 Administered concentration (HEC) Mouse, male Liver 3.7 × 

Administered concentration (HEC) Mouse, male Lung 9.2 × 1995 IRIS assessmentc Mouse, male Liver and lung 12.7 4.7 × 

aGST-T1+/+ = homozygous, full enzyme activity; Mixed = genotypes based on a population reflecting the estimated frequency of gen

U.S. population: 20% GST-T1–/–, 48% GST-T1+/–, and 32% GST-T1+/+ (Haber et al., 2002). 

bBased on mean value of the derived distributions Bolded value is the basis for the recommended IUR of 1× 10–8μg/m3 per mg/kg-day. 
294

8214 5.4.3. Differences Between Current Assessment and Previous IRIS PBTK-based 

8215 Assessment

8216 To better understand the changes in assessment risk predictions between previous EPA 

8217 evaluations and the current assessment, the differences in PBTK model parameters for the 

8218 B6C3F1 mouse were evaluated. Values that differed significantly between the model version 

8219 used previously and that of Marino et al. (2006), along with derived group parameters that lend 

8220 further insight, are shown in Table 5-23. 

8221 The tissue:air partition coefficients in Table 5-23 show that, while several of the blood:air 

8222 partition coefficients appear to differ significantly between the two models, the corresponding 

8223 blood:air partition coefficient does not, and it is the latter that can be more indicative of long8224 term equilibration between the tissue (tissue group) and air. Thus, the partition coefficients, the 

8225 ones that most significantly differ are the blood:air and liver:air partition coefficients that are, 

8226 respectively, 2.8- and 2.6-fold higher in the current version. The increased blood:air partition 

8227 coefficient results in a tendency for simulated blood concentrations to rise more quickly and 

8228 reach higher values, other parameters being equal. The significantly increased QCC and VPR 

8229 contribute even more to this difference, resulting in an even faster rise to steady state during 

8230 inhalation exposure simulations, though also more rapid delivery to the liver (decreasing blood8231 flow limitation of hepatic metabolism) and more rapid exhalation. The increased liver:air 

8232 partition coefficient leads to higher predicted liver concentrations (again, other parameters being 

8233 equal) and hence higher rates of metabolism. 

8234 For metabolism, a much reduced oxidative metabolism is seen, which at low doses 

8235 depends on Vmaxc/Km. The revised hepatic metabolism is over 40% lower and the total of lung + 

8236 liver metabolism is 50% lower than previously used. This lower rate of metabolism means that 

8237 far less of parent dichloromethane will be removed through metabolism, and hence predicted 

8238 blood concentrations will be higher still, relative to the impact of changes in partition coefficient,

8239 QCC, and VPR, as noted above. 

8240 The result of having higher predicted blood and liver dichloromethane concentrations is 

8241 that, while the GSH-pathway metabolic constant, kfC , is virtually the same for the mouse liver in 

8242 both cases, the much higher concentration of dichloromethane available will lead to a much 

8243 higher predicted rate of metabolism via this pathway. For the lung, since the lung:liver ratio 

8244 (A2) is 43% higher in the model of Marino et al. (2006), the relative increase will be even 

8245 greater. 

8246 Because the revised rate of GST metabolism in mice was estimated by using data with 

8247 CYP2E1 inhibited by a suicide inhibitor, there is considerable confidence in the relative rate of 

8248 metabolism by these two pathways, and the GST pathway in particular. The partition 

8249 coefficients used by Marino et al. (2006) are as measured by Clewell et al. (1993) and expected 

8250 to be at least as reliable as those used in the 1995 assessment. Considering that the revised 

8251 PBTK model does an excellent job of reproducing closed-chamber gas uptake data that were not 
295

8252 available for calibration of the 1987 model, as well blood concentrations after intravenous 

8253 injection, we can have fairly high confidence in its predictions. 

8254 The net result of these model changes is that, under mouse bioassay conditions, the 

8255 predicted dose metrics for liver and lung cancer, i.e., GST-mediated metabolism, are higher than 

8256 those obtained with the previous model, resulting in a lower risk estimated per unit of dose. 

8257 

Table 5-23. Comparison of key B6C3F1 mouse parameters differing 

between prior and current PBTK model application 

Parametera

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/kg0.74) 

VPR ventilation:perfusion ratio 

24.2 

1.45 

14.3 

1.0 

Metabolism parameters 

Vmaxc maximum CYP metabolic rate (mg/hour/kg0.7)

Km CYP affinity (mg/L) 

VmaxC/Km (L/hour/kg0.7) 

A1 ratio of lung Vmaxc to liver Vmaxc

Total lung + liver Vmaxc/Km

kfC first-order GST metabolic rate constant 

(kg0.3/hr) 

A2 ratio of lung kfC to liver kfC

Total lung + liver kfC

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% between versions. See Table 3-5 and associated text for details. 

8258 

8259 

8260 The other piece of the PBTK-based risk estimation is the human model. In updating the 

8261 parameter estimates for the human model (see Appendix B for details), the oxidative metabolism 

8262 Vmaxc/Km approximately doubled, which leads to lower predicted blood concentrations of 

8263 dichloromethane available for metabolism by GST. In addition, the liver GST was reduced by 

8264 almost 60%, and the lung:liver GST ratio decreased by almost fivefold, for a net change in lung 
296

8265 GST of over 90%. Given the larger human data set available to David et al. (2006) and the 

8266 sophisticated Bayesian analysis used to recalibrate the model, the expectation is that these values 

8267 are more reliable than the values used in the 1995 IRIS assessment. 

8268 Since actual rates of metabolism at a given exposure level also depend on respiration rate 

8269 and blood flows, these changes in metabolic parameters do not completely determine the relative 

8270 (predicted) dosimetry. But the difference in cancer risk predictions between the current and 

8271 previous assessments is primarily explained by the overall prediction of higher GST-mediated 

8272 dosimetry in the mouse (at bioassay conditions) and lower human GST metabolism (due in part 

8273 to greater predicted clearance of dichloromethane by oxidative metabolism). In addition to these 

8274 changes in PBTK parameters, the reduction of scaling factor from 12.7 to 7 is a significant factor 

8275 in the overall change from the previous assessments. 

8276 

8277 5.4.4. Application of Age-Dependent Adjustment Factors (ADAFs) 

8278 The available dichloromethane studies do not include an evaluation of early-life 

8279 susceptibility to dichloromethane cancer risk. In the absence of this type of data and if a 

8280 chemical follows a mutagenic mode of action for carcinogenicity, like dichloromethane, the 

8281 Supplemental Guidance for Assessing Susceptibility from Early-Life Exposure to Carcinogens

8282 (U.S. EPA, 2005b) recommends that ADAFs be applied to the cancer values. Since the OSF of 1 

× 10–3 (mg/kg-day)–1 and the IUR of 1 × 10–8 (μg/m3)

–1 8283 were calculated from adult 

8284 dichloromethane exposures, early-life cancer susceptibility has not been accounted for in these 

8285 values and ADAFs need to be applied in combination with exposure informations when 

8286 estimating cancer risks that include early-life exposures. Sample calculations that incorporate 

8287 ADAFs into the cancer risks are presented in subsequent sections. Additional examples of 

8288 evaluations of cancer risks incorporating early-life exposure are provided in section 6 of the 

8289 Supplemental Guidance (U.S. EPA, 2005b). 

8290 In the Supplemental Guidance (U.S. EPA, 2005b), ADAFs are established for three age 

8291 groups. An ADAF of 10 is applied for age groups less than 2 years, 3 is applied for ages 2 to 

8292 <16 years, and 1 is applied for 16 years and above (U.S. EPA, 2005b). The 10- and 3-fold 

8293 adjustments in cancer values are combined with age-specific exposure estimates, when early-life 

8294 exposure considerations need to be included in cancer risk estimates. The most current 

8295 information on usage of ADAFs can be found at www.epa.gov/cancerguidelines. For estimation 

8296 of risk, the Supplemental Guidance (U.S. EPA, 2005b) recommends obtaining and using age8297 specific values for exposure and cancer potency. In the absence of age-specific values cancer 

8298 potency values, as is true for dichloromethane, age-specific cancer values are estimated by using 

8299 the appropriate ADAFs. Using this process, a cancer risk is derived for each age group. The 

8300 risks are summed across the age groups to get the total cancer risk for the age-exposure period of 

8301 interest. 

8302 
297

8303 

8304 5.4.4.1. Application of ADAFs in Oral Exposure Scenarios

8305 Sample calculations incorporating the use of ADAFs are presented for three exposure 

8306 duration scenarios. These scenarios include full lifetime exposure (assuming a 70-year lifespan), 

8307 and two 30-year exposures from ages 0–30 and ages 20–50. A constant dichloromethane 

8308 exposure of 1 mg/kg-day was assumed for each scenario. 

8309 Table 5-24 lists the four factors (ADAFs, OSF, assumed exposure, and duration 

8310 adjustment) that are needed to calculate the partial cancer risk based on the early age-specific 

8311 group. The partial cancer risk for each age group is the product of the four factors in columns 2–

8312 5. Therefore, the partial cancer risk following daily dichloromethane oral exposure in the age 

group 0 to <2 years is the product of the values in columns 2–5 or 10 × (1 × 10–3 8313 ) × 1 × 2/70 = 

2.9 × 10–4 8314 . The partial risks that are listed in the last column of Table 5-24 are added together to 

8315 get the total risk. Thus, a 70-year (lifetime) risk estimate for continuous exposure to 1 mg/kgday dichloromethane is 1.7 × 10–3 8316 per mg/kg-day, which is adjusted for early-life susceptibility 

8317 and assumes a 70-year lifetime and constant exposure across age groups. 

8318 

Table 5-24. Application of ADAFs to dichloromethane 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 

8321 In calculating the cancer risk for a 30-year constant exposure to dichloromethane at an 

8322 exposure level of 1 mg/kg-day from ages 0–30, the duration adjustments would be 2/70, 14/70, 

and 14/70 and the partial risks for the three age groups would be 2.9 × 10–4, 6.0 × 10–4 8323 , and 2.0 × 

10–4, which would result in a total risk estimate of 1.1 × 10–3 8324 . 

8325 In calculating the cancer risk for a 30-year constant exposure to dichloromethane at an 

8326 exposure level of 1 mg/kg-day from ages 20–50, the duration adjustments would be 0/70, 0/70, 

and 30/70. The partial risks for the three groups are 0, 0, and 4.3 × 10–4 8327 , which would result in a 

total risk estimate of 4.3 × 10–4 8328 . 

8329 

8330 5.4.4.2. Application of ADAFs in Inhalation Exposure Scenarios 
298

8331 Sample calculations incorporating the use of ADAFs are presented for three exposure 

8332 duration scenarios involving inhalation exposure. These scenarios include full lifetime exposure 

8333 (assuming a 70-year lifespan) and two 30-year exposures from ages 0–30 and ages 20–50. A 

constant dichloromethane inhalation exposure of 1 μg/m3 8334 was assumed for each scenario. 

8335 Similar to the oral exposure scenarios presented in section 5.4.4.1, Table 5-25 lists the 

8336 four factors (ADAFs, unit risk, assumed exposure, and duration adjustment) that are needed to 

8337 calculate the partial cancer risk based on the early age-specific group. The partial cancer risk for 

8338 each age group is the product of the four factors in columns 2–5. Therefore, the partial cancer 

8339 risk following daily dichloromethane inhalation exposure in the age group 0 to <2 years is the 

product of the values in columns 2–5 or 10 × (1 × 10–8) × 1 × 2/70 = 2.9 × 10–9 8340 . The partial 

8341 risks that are listed in the last column of Table 5-25 are added together to get the total risk. Thus, 

a 70-year (lifetime) risk estimate for continuous exposure to 1 μg/m38342 dichloromethane is 1.8 × 

10–8 per μg/m38343 , which is adjusted for early-life susceptibility and assumes a 70-year lifetime and 

8344 constant exposure across age groups. 

8345 

Table 5-25. Application of ADAFs to dichloromethane 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 

8348 In calculating the cancer risk for a 30-year constant exposure to dichloromethane at a 

level of 1 μg/m38349 from ages 0–30, the duration adjustments would be 2/70, 14/70, and 14/70, and 

the partial risks for the three age groups are 2.9 × 10–9, 6.0 × 10–9, and 2.0 × 10–9 8350 . These partial 

risks result in a total risk estimate of 1.1 × 10–8 8351 . 

8352 In calculating the cancer risk for a 30-year constant exposure to dichloromethane at a 

level of 1 μg/m3

8353 from ages 20–50, the duration adjustments would be 0/70, 0/70, and 30/70, and 

the partial risks for the three age groups are 0, 0, and 4.3 × 10–9 8354 , resulting in a total risk estimate 

of 4.3 × 10–9 8355 . 

8356 

8357 5.4.5. Uncertainties in Cancer Risk Values 

8358 The derivation of cancer risk values often involves a number of uncertainties in the 

8359 extrapolation of dose-response data in animals to cancer risks in human populations. Several 

8360 types of uncertainty have been quantitatively integrated into the derivation of the recommended 
299

8361 OSFs and IURs for dichloromethane, while others are qualitatively considered. Table 5-26 and 

8362 the ensuing discussion summarize the principal uncertainties identified, their possible effects on 

8363 the cancer risk values, decisions made in the derivations, and justifications for the decisions. 

8364 

Table 5-26. Summary of uncertainty in the derivation 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 slightly 

higher 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 effects 
300

Table 5-26. Summary of uncertainty in the derivation 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 tumor 

responses 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 model 

parameters 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 

8367 Data selections for derivation of IUR and OSF 

8368 The database of animal bioassays identifies the liver and lung as the most sensitive target 

8369 organs for dichloromethane-induced tumor development. These effects demonstrate a dose8370 response relationship in mice exposed orally (liver only) or by inhalation (liver and lung). 

8371 Statistically significant increases in benign mammary gland tumors were observed in one study 

8372 of F344 rats exposed by inhalation to 2,000 or 4,000 ppm (Mennear et al., 1988; NTP, 1986), 

8373 and evidence for a tumorigenic mammary gland response in Sprague-Dawley rats was limited to 

8374 increased numbers of benign mammary tumors per animal at levels of 50–500 ppm (Nitschke et 

8375 al., 1988a) or 500–3,500 ppm (Burek et al., 1984). An oral (gavage) study in female Sprague8376 Dawley rats reported an increased incidence of malignant mammary tumors, mainly 

8377 adenocarcinomas (8, 6, and 18% in the control, 100, and 500 mg/kg dose groups, respectively), 
301

8378 but the increase was not statistically significant. Data were not provided to allow an analysis that 

8379 accounts for differing mortality rates (Maltoni et al., 1988). The toxicokinetic or mechanistic 

8380 events that might lead to mammary gland tumor development in rats are unknown, although 

8381 CYP2E1 (El-Rayes et al., 2003; Hellmold et al., 1998) and GST-T1 expression has been detected 

8382 in human mammary tissue (Lehmann and Wagner, 2008). Rare CNS tumors were observed in 

8383 one study in rats, a study spanning a relatively low range of exposures (0–500 ppm). These 

8384 cancers were not seen in two other studies (NTP, 1986; Burek et al., 1984) in rats, both involving 

8385 higher doses (1,000–4,000 ppm), or in a similar high-dose study (NTP, 1986) in mice. The 

8386 relative rarity of the tumors precludes the use of the low-dose exposure study in a quantitative 

8387 dose-response assessment. The in vivo genotoxicity and mechanistic data in rodents provide a 

8388 detailed sequence of steps from generation of reactive metabolites to mutagenic effects, such as 

8389 DNA-protein cross-links and DNA strand breaks. Further, the toxicokinetic pathways implicated 

8390 in production of the putative carcinogenic metabolites in animals also exist in humans. Thus, 

8391 there is high confidence that the dose-response data for liver and lung cancer in mice represents 

8392 the best data currently available for derivation of human cancer risks. A more complete 

8393 understanding of the carcinogenic potential of dichloromethane would be achieved by addressing 

8394 data gaps identified with respect to issues regarding potential risk and mechanisms relating to 

8395 brain cancer and mammary tumors. The available epidemiologic studies provide some evidence 

8396 of an association between dichloromethane and brain cancer (see Section 4.1.3.7.1). The 

8397 available epidemiologic studies do not provide an adequate basis for the evaluation of the role of 

8398 dichloromethane in breast cancer because there are currently no cohort studies with adequate 

8399 statistical power and no case-control studies with adequate exposure methodology to examine 

8400 this relationship (see section 4.1.3.7.6)

8401 

8402 Target organ 

8403 The liver and lung tumor incidence from chronic exposure biassays provide clear 

8404 evidence of the carcinogenic potential of dichloromethane exposure. The biassays are supported 

8405 by a substantial literature of genotoxicity and mechanistic studies (summarized in section 4.5). 

8406 The evidence for mammary gland tumors from dichloromethane exposure is based primarily on 

8407 observations of benign tumors in rats with inhalation exposure (NTP, 1986). Derivation of 

8408 cancer potenciy values based on these data are presented in Appendix G. The potential brain 

8409 cancer risk, suggested by a limited number of these relatively rare tumors in both animal and 

8410 human studies, is identified as a data gap which would benefit from additional research. 

8411 

8412 Extrapolation approach 

8413 A route-to-route extrapolation from the NTP (1986) inhalation mouse bioassay was used 

8414 to develop an oral cancer slope value. This value is inherently less certain than the values based 

8415 on oral exposure due to the influence of route of exposure on toxicokinetics. 
302

8416 

8417 Dose Metric 

8418 There is considerable data supporting the role of GST-related metabolism of 

8419 dichloromethane in carcinogenicity, as described in sections 4.5.1 and 4.7. Pretreatment of mice 

8420 with buthionine sulphoximine, a GSH depletor, caused a decrease, to levels seen in controls, in 

8421 the amount of DNA damage detected immediately after in vivo exposure in liver and lung tissue 

8422 (Graves et al., 1995). Although the results of Landi et al. (2003) indicate that GST activity is not 

8423 needed for the observation of DNA damage by the comet assay from some trihalomethanes (e.g., 

8424 bromodichloromethane), the results for dichloromethane were much weaker and of uncertain 

8425 significance. 

8426 

8427 Dose-response modeling 

8428 Because of the adequacy of the fit of the multistage model to the data, little modeling 

8429 uncertainty would be expected to be introduced by the choice of this model. Application of the 

8430 multistage model allowed for estimation of a point of departure in the lower region of exposure 

8431 for observable cancer effects. 

8432 

8433 Low-dose extrapolation

8434 The mode of action is a key consideration in determining how risks should be estimated 

8435 for low-dose exposure. The in vitro and in vivo genotoxicity data suggest that mutagenicity is 

8436 the most plausible mode of action, although key mutagenic events in the development of liver or 

8437 lung tumors have not been identified. No data are available that provide an adequate rationale 

8438 for choosing a nonlinear dose response in the low-dose region. Because a mutagenic mode of 

8439 action is most plausible, a linear-low-dose extrapolation approach was used to estimate OSFs and 

8440 IURs. 

8441 

8442 Interspecies extrapolation of dosimetry and risk 

8443 Target organ dosimetry for neoplastic mouse responses and estimation of equivalent 

8444 internal human doses were accomplished using PBTK models for dichloromethane in mice and 

8445 humans. Uncertainty in the ability of the PBTK models to estimate animal and human internal 

8446 doses from lifetime bioassay low-level environmental exposures may affect the confidence in the 

8447 cancer risk extrapolated from animal data. Uncertainties in the mouse and human model 

8448 parameter values were integrated quantitatively into parameter estimation by utilizing 

8449 hierarchical Bayesian methods to calibrate the models at the population level (David et al., 2006; 

8450 Marino et al., 2006). The use of Monte Carlo sampling to define human model parameter 

8451 distributions allowed for derivation of human distributions of dosimetry and cancer risk, 

8452 providing for bounds on the recommended risk values. 
303

8453 A detailed discussion of PBTK model structure (CYP rate equation) and parameter 

8454 uncertainties is provided in Section 5.3. While the structure and equations used in the existing 

8455 model have been described in multiple peer-review publications over the past two decades, there 

8456 are discrepancies between dichloromethane kinetics observed in vitro and the model parameters 

8457 obtained from in vivo data. However, an alternative (dual-binding-site) CYP metabolic equation 

8458 appears to resolve these discrepancies. Integration of the alternate rate equation into the PBTK 

8459 modeling, and then quantitative risk assessment, will likely require several years of further 

8460 research, and hence is beyond the scope of the current assessment. Since the GST activity in the 

8461 current model is within a factor of three of that measured in vitro (when both are evaluated on a 

8462 per gram of liver basis), the impact of that model uncertainty is also expected to be no more than 

8463 a factor of three. Sensitivity to the human PBTK parameter distributions was evaluated by 

8464 rescaling the parameters to the mean values obtained by David et al. (2006) for a specific data set 

8465 (DiVincenzo and Kaplan, 1981) for which the GST activity was close to a numerical average of 

8466 those obtained across individual data sets. When this was done, the upper bound estimates on 

8467 GST dosimetry (for low, fixed inhalation or oral exposures) in the GST-T1 +/+ subpopulation 

8468 increased by over an order of magnitude, as did the estimate of the mean activity for an 

8469 inhalation exposure, although the estimated mean GST activity for an oral exposure only 

8470 increased about two-fold. So while correspondence of the in vivo GST activity with that 

8471 measured in vitro suggests a lower degree of quantitative uncertainty, it is possible that revision 

8472 of the PBTK model could have a larger impact. The ultimate impact will depend on how 

8473 revisions effect model predictions for both the animal and the human. If the predicted GST 

8474 metabolism per unit exposure increases in both mice and humans by a similar factor, there will 

8475 be little impact on the risk estimate. But if the GST activity predicted in the mouse is decreased 

8476 by a factor of 3, while that in the human is increased by a factor of 3, for example, then the net 

8477 impact would be an increase of 9-fold in human risk estimates. 

8478 The PBTK animal models were utilized deterministically; i.e., the single-value parameter 

8479 estimates for the rat PBTK model were used for rat dosimetry simulations and the mean 

8480 parameter estimates from the Bayesian analysis of Marino et al. (2006) were used for the mouse 

8481 dosimetry simulations. To assess the effect of using point estimates of parameter values for 

8482 calculation of rodent dosimetry, a sensitivity analysis was performed to identify model 

8483 parameters most influential on the predictions of dose metrics used to estimate oral and 

8484 inhalation cancer risks. As was described in the RfD and RfC sensitivity analysis calculation, 

8485 this procedure used a univariate analysis in which the value of an individual model parameter 

8486 was perturbed by an amount (∆), in the forward and reverse direction (i.e., an increase and 

8487 decrease from the nominal value), and the change in the output variable was determined. Results 

8488 are for the effects of a perturbation of ±1% from the nominal value of each parameter on the 

8489 output values at the end of a minimum of 10,000 simulated hours. This time was chosen to 

8490 achieve a stable daily value of the dose metric, given that the simulated bioassay exposures did 
304

8491 not include weekend exposures. The exposure conditions represented the lowest bioassay 

8492 exposure, resulting in significant increases in the critical effect. For inhalation exposures in 

8493 mice, the blood:air partition coefficient, followed closely by the first-order GST-mediated 

8494 metabolism rate (kfC), had the greatest impact on the dose metric for liver cancer (mg 

8495 dichloromethane metabolized via GST pathway per liter liver per day) (Figure 5-17). For 

8496 drinking water exposures in mice, the kfC, followed by the CYP-mediated maximum reaction 

8497 velocity (Vmaxc), affected the liver cancer dose metric to the greatest extent (Figure 5-18). For 

8498 mice inhaling dichloromethane, the lung cancer dose metric (mg dichloromethane metabolized 

8499 via GST pathways per liter lung per day), like the liver cancer metric, was highly affected by the 

8500 kfC and the blood:air partition coefficient (Figure 5-19). However, since GST-mediated lung 

8501 metabolism is calculated as a constant fraction of the liver metabolism rate (A2 × kfC), the lung 

8502 cancer dose metric was most sensitive to the proportional yield of liver GST-mediated metabolic 

8503 activity attributed to the lung. The blood:air partition coefficient was experimentally determined, 

8504 lending high confidence to its value. Values for the three metabolic parameters were determined 

8505 by computational optimization against data sets not directly measuring dichloromethane or its 

8506 metabolites in the target/metabolizing tissues. It is uncertain how alternative values for these 

8507 three parameters would affect the estimation of animal BMDL10 values and, ultimately, the OSFs 

8508 and IURs. 

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

KFC

Parameter

Normalized sensitivity coefficient

8510 

8511 Figure 5-17. Sensitivity coefficients for long-term mass GST-mediated 

8512 metabolites per liver volume from a long-term average daily inhalation 

8513 concentration of 2000 ppm in mice. KFC = GST-mediated metabolism rate; A2 = 

8514 proportion of liver GST metabolism attributed to the lung; VMAXC = CYP-mediated 

8515 maximum rate of metabolism; PB = blood:air partition coefficient; VSC = slowly 

8516 perfused tissue volume; VLC = liver volume; VPR = Ventilation perfusion ratio ; 

8517 QCC = cardiac output constant. 
305

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

KFC

Parameter

Normalized sensitivity coefficient

8519 

8520 Figure 5-18. Sensitivity coefficients for long-term mass GST-mediated metabolites 

8521 per liver volume from a long-term average daily drinking water concentration of 

8522 500 mg/L in mice. KFC = GST-mediated metabolism rate; A2 = proportion of liver 

8523 GST metabolism attributed to the lung; KA = oral absorption rate from gut; VMAXC = 

8524 CYP-mediated maximum rate of metabolism; PB = blood:air partition coefficient; VSC 

8525 = slowly perfused tissue volume; VLC = liver volume; VPR = Ventilation perfusion 

8526 ratio ; QCC = cardiac output constant. 

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

KFC

Parameter

Normalized sensitivity coefficient

8529 

8530 Figure 5-19. Sensitivity coefficients for long-term mass GST-mediated metabolites 

8531 per lung volume from a long-term average daily inhalation concentration of 500 

8532 ppm in mice. KFC = GST-mediated metabolism rate; A2 = proportion of liver GST 

8533 metabolism attributed to the lung; VMAXC = CYP-mediated maximum rate of 

8534 metabolism; PB = blood:air partition coefficient; VSC = slowly perfused tissue volume; 

8535 VLC = liver volume; VPR = Ventilation perfusion ratio ; QCC = cardiac output 

8536 constant. 

8537 
306

8538 The comparison of the OSF derived from the oral exposure data and from the route–to8539 route extrapolation from the inhalation data provides a crude measure of the uncertainty in 

8540 recommending a human OSF based on the available rodent bioassay data. The oral cancer slope 

8541 factor based on route-to-route extrapolations from liver tumors in mice exposed by inhalation are 

8542 about an order of magnitude lower than those based on the liver tumor responses in mice 

8543 exposed via drinking water. This difference may be explained, at least partially, by the 

8544 heterogeneity of hepatic cell types within the sinusoid, resulting in regio-specific toxicity. Oral 

8545 exposure may result in a higher internal exposure of hepatocytes in the periportal region 

8546 (particularly those lining the portal vein, through which all gastrointestinal-absorbed 

8547 dichloromethane passes) than in the centrilobular region (Syracuse Research Corporation [SRC], 

8548 1989). Further, the metabolic capacity of hepatic cells is also regio-specific, with higher CYP 

8549 activity found in the centrilobular region compared to the periportal region. Thus, liver perfusion 

8550 via the systemic arterial circulation (through which inhaled dichloromethane would be 

8551 introduced) or portal drainage of the gastrointestinal tract may influence regio-specific 

8552 hepatotoxicity, resulting in the route-of-exposure effects on toxicity. The available PBTK 

8553 models do not have the capability to predict regio-specific disposition of dichloromethane in the 

8554 liver. 

8555 There is uncertainty as to whether the reactivity of the toxic dichloromethane metabolites 

8556 is sufficiently high enough to preclude systemic distribution. Therefore, alternative derivations 

8557 of cancer risk values were performed under the assumption that high reactivity leads to complete 

8558 clearance from the tissue in which the active metabolite is formed (scaling factor = 1.0). The 

8559 difference in scaling factor (7.0 for allometric scaling versus 1.0) results in a 7-fold decrease in 

8560 estimated cancer toxicity values. Using a whole-body GST metabolism dose metric, the 

8561 resulting OSF and IUR for liver and lung cancer were approximately five-fold higher than when 

8562 tissue-specific dose metrics were used (Table 5-16 and Table 5-22). The mechanistic data 

8563 support the notion that reactive metabolites produced in the target tissues are not well distributed 

8564 and produce deleterious effects in the metabolizing tissues soon after generation. Thus, there is 

8565 less uncertainty in the cancer risk values derived by using a tissue-specific GST metabolism dose 

8566 metric compared with those derived using a whole-body GST metabolism dose metric. 

8567 

8568 Sensitive human populations

8569 Possible sensitive populations include persons with altered CYP (e.g., obese individuals, 

8570 alcoholics, diabetics, and the very young) and GST (e.g., GST-T1 homozygous conjugators) 

8571 metabolic capacity. The PBTK model includes an estimate of the variability of CYP metabolism 

8572 (sixfold variation), within the general population but does not specifically address what could be 

8573 greater variation in these other groups. However, the known polymorphisms for GST-T1 

8574 expression were integrated into the human model. The distributions of human IUR values (from 

which the recommended [i.e., mean] values were taken) show that the 99th 8575 percentiles are 
307

8576 approximately 4-fold and 6-fold higher than means for liver and lung cancer. For the distribution 

of OSFs, the 99th 8577 percentile is approximately threefold higher than mean for liver cancer. 

8578 To further characterize the potential sensitivity of specific subpopulations, internal dose 

distributions for oral exposure to 1 mg/kg-day or inhalation exposure to 1 mg/m38579 were estimated 

8580 for 1-year-old children and 70-year-old men and women to compare with the broader population 

8581 results used to estimate cancer risks above. Since the recommended cancer risk estimate is based 

on the GST-T1+/+ 8582 subpopulation, this analysis was also restricted to that subpopulation, so that 

8583 only the factors of age and gender would differ. The impact of considering other GST-T1 groups 

can be seen where risk estimates for the GST-T1+/– 8584 and entire population mix are given above. 

8585 Specification of age- and gender-specific parameters are as described in Appendix B. This 

8586 sensitivity analysis is qualitatively similar to that described previously for the noncancer 

8587 assessments of dichloromethane, where the variability in human equivalent administered dose 

8588 and HEC values was estimated. 

For this analysis, however, consideration of exclusively GST-T1+/+ 8589 individuals will 

8590 clearly narrow any estimate of variability. This analysis will also differ from that for noncancer 

8591 effects in that the inverse of the former relationship is being considered (i.e., the variation in a 

8592 specific internal dose for a fixed exposure is being computed, whereas for the human equivalent 

8593 administered dose and HEC the variability in exposure levels corresponding to a fixed internal 

8594 dose are estimated. The results of this analysis for oral exposures are shown in Figure 5-20 and 

8595 Table 5-27 and for inhalation exposures in Figure 5-21 and Table 5-28. 

8596 For the oral exposure analysis, the distribution of internal doses shows little variation 

8597 among the different age/gender groups (Figure 5-20, Table 5-27). The cancer analysis is based 

8598 on a very low internal dose, where little enzymatic saturation is expected to occur, allowing for 

8599 efficient first-pass metabolism, which is independent of differences in respiration; differences 

8600 will be more significant at the higher doses analyzed for the noncancer human equivalent applied 

8601 dose. Thus, the consideration of only GST metabolism and the narrower range of metabolic rate 

8602 for that pathway in the +/+ population at low oral exposure rates results in minimal age/gender 

8603 sensitivity differences (the 7-year-old female is only 5% more sensitive from pharmacokinetic 

8604 factors than the general population). 

8605 For inhalation, an internal liver GST dose (mean value) about 2.5 times higher in the 

8606 child than the general population is predicted due to the higher inhalation rate. The results for 

8607 the liver GST dose for inhalation, Figure 5-21 and Table 5-28 indicate that the 70-year-old male 

8608 and female populations are only slightly shifted from the general population, while the 

8609 population for the 1-year-old child is a distinct upper tail of the general distribution. 

8610 
308

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 

8612 Figure 5-20. Histograms for a liver-specific dose of GST metabolism (mg 

8613 GST metabolites per liter liver per day) for the general population (0.5- to 

8614 80-year-old males and females) and specific age/gender groups, within the 

population of GST-T1+/+ 8615 genotypes, given a daily oral dose-rate of 1 mg/kg8616 day dichloromethane. 

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 95th percentile 99th percentile

All agesb 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. 

8619 
309

0

0.05

0.1

0.15

0.2

0.25

0 10 20 30 40

ng GST metabolites/L liver/day

Frequency 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 

8621 Figure 5-21. Histograms for liver-specific dose of GST metabolism (mg GST 

8622 metabolites per liter liver per day) for the general population (0.5- to 80-year8623 old males and females) and specific age/gender groups, within the population 

of GST-T1+/+ genotypes, given a continuous inhalation exposure to 1 mg/m3

8624 

8625 dichloromethane. 

8626 

8627 

Table 5-28. Statistical characteristics of human internal 

doses for 1 mg/m3 inhalation exposures in specific 

subpopulations

Internal dose (mg/L liver per day)a

Population Mean 95th percentile 99th percentile

All agesb 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/m3

 dichloromethane. 

b

0.5- to 80-year-old males and females. 

8628 

8629 

8630 

8631 
310

8632 6. MAJOR CONCLUSIONS IN THE CHARACTERIZATION OF HAZARD AND 

8633 DOSE RESPONSE 

8634 

8635 

8636 6.1. HUMAN HAZARD POTENTIAL 

8637 Dichloromethane (CASRN 75-09-2), also known as methylene chloride, is a colorless 

8638 liquid with a penetrating ether-like odor. It is produced by the direct reaction of methane with 

8639 chlorine at either high temperatures or low temperatures under catalytic or photolytic conditions. 

8640 The principal uses for dichloromethane have been in paint strippers and removers, as a propellant 

8641 in aerosols, in the manufacture of drugs, pharmaceuticals, film coatings, electronics, and 

8642 polyurethane foam, and as a metal-cleaning solvent. 

8643 Dichloromethane is rapidly absorbed through both oral administration and inhalation 

8644 exposure with a steady-state saturation occurring with inhalation. Results from studies of 

8645 animals show that, following absorption, dichloromethane is rapidly distributed throughout the 

8646 body and has been detected in all tissues that have been evaluated. Metabolism of 

8647 dichloromethane involves two primary pathways. Dichloromethane is metabolized to CO and to 

8648 a lesser extent CO2 in a CYP-dependent oxidative pathway (CYP2E1) that is predominant at low 

8649 exposure levels. The other major pathway for dichloromethane metabolism involves the 

8650 conjugation of dichloromethane to GSH, catalyzed by GST (GST-T1). This results in the 

8651 formation of a GSH conjugate that is eventually metabolized to CO2. The conjugation of 

8652 dichloromethane to GSH results in the formation of two reactive intermediates that have been 

8653 hypothesized to be involved in dichloromethane carcinogenicity, S-(chloromethyl)glutathione 

8654 and formaldehyde. Formation of formaldehyde leads to several covalent modifications of 

8655 cellular macromolecules, including DNA-protein cross-links (Casanova et al., 1996) and RNA8656 formaldehyde adducts (Casanova et al., 1997). Evidence is also available that 

8657 S-(chloromethyl)glutathione can result in both DNA SSBs and DNA mutations, presumably 

8658 through DNA alkylation (Green, 1997; Graves and Green, 1996; Graves et al., 1996, 1994a; 

8659 Hashmi et al., 1994). However, DNA reaction products (e.g., DNA adducts) produced by S8660 (chloromethyl)glutathione have not been found, possibly due to potential instability of these 

8661 compounds (Watanabe et al., 2004; Hashmi et al., 1994). 

8662 Information on noncancer effects in humans exposed orally to dichloromethane are 

8663 restricted to case reports of neurological impairment (general CNS depression), liver and kidney 

8664 effects (as severe as organ failure), and gastrointestinal irritation in individuals who ingested 

8665 amounts ranging from about 25 to 300 mL (Chang et al., 1999; Hughes and Tracey, 1993). The 

8666 animal toxicity database identifies hepatic effects (hepatic vacuolation, nonneoplastic liver foci) 

8667 as the critical dose-dependent noncancer endpoint associated with oral exposure to 

8668 dichloromethane. The most frequently observed liver effect was hepatocyte vacuolation, seen 

8669 with drinking water exposure (90 days) in F344 rats at ≥166 mg/kg-day and B6C3F1 mice at 
311

8670 586 mg/kg-day (Kirschman et al., 1986) and with gavage exposure (14 days) in CD-1 mice at 

8671 333 mg/kg-day (Condie et al., 1983). Hepatocyte degeneration or necrosis was observed in 

8672 female F344 rats exposed in drinking water for 90 days to 1,469 mg/kg-day (Kirschman et al., 

8673 1986) and in female F344 rats exposed by gavage for 14 days to 337 mg/kg-day (Berman et al., 

8674 1995). In the chronic-duration (104-week) study, liver effects were described as nonneoplastic 

8675 foci in F344 rats exposed to drinking water doses between 50 and 250 mg/kg-day (Serota et al., 

8676 1986a). In the reproductive oral administration studies, no significant effect on reproductive 

8677 function or parameter was observed in rats up to 225 mg/kg-day (General Electric Co., 1976) or 

8678 in mice up to 500 mg/kg-day (Raje et al., 1988). The NOAEL and LOAEL for altered 

8679 neurological functions in female F344 rats were 101 and 337 mg/kg-day (as reported by Moser et 

8680 al., 1995). 

8681 Acute inhalation exposure of humans to dichloromethane has been associated with 

8682 cardiovascular impairments due to decreased oxygen availability from COHb formation and 

8683 neurological impairment from interaction of dichloromethane with nervous system membranes 

8684 (Bos et al., 2006; ACGIH, 2001; ATSDR, 2000; Cherry et al., 1983; Putz et al., 1979; Gamberale 

8685 et al., 1975; Winneke, 1974). Relatively little is known about the long-term neurological effects 

8686 of chronic exposures, although there are studies that provide some evidence of an increased 

8687 prevalence of neurological symptoms among workers with average exposures of 75–100 ppm 

8688 (Cherry et al., 1981) and long-term effects on some neurological measures (i.e., possible 

8689 detriments in attention and reaction time in complex tasks) in workers whose past exposures 

8690 were in the 100–200 ppm range (Lash et al., 1991). These studies are limited by the relatively 

8691 small sample sizes and low power for detecting statistically significant results for these 

8692 endpoints. 

8693 Following repeated inhalation to dichloromethane, the liver is the most sensitive target 

8694 for noncancer toxicity in rats and mice. Lifetime exposure was associated with hepatocyte 

8695 vacuolation and necrosis in F344 rats exposed to 1,000 ppm 6 hours/day (Mennear et al., 1988; 

8696 NTP, 1986), hepatocyte vacuolation in Sprague-Dawley rats exposed to 500 ppm 6 hours/day 

8697 (Nitschke et al., 1988a; Burek et al., 1984), and hepatocyte degeneration in B6C3F1 mice 

8698 exposed to 2,000 ppm 6 hours/day (lower concentrations were not tested in mice) (Mennear et 

8699 al., 1988; NTP, 1986). Other effects observed include renal tubular degenerations in F344 rats 

8700 and B6C3F1 mice at 2,000 ppm, testicular atrophy in B6C3F1 mice at 4,000 ppm, and ovarian 

8701 atrophy in B6C3F1 mice at 2,000 ppm. 

8702 Other studies with inhalation exposure to dichloromethane revealed no significant effects 

8703 on reproductive performance in rats (up to 1,500 ppm) (Nitschke et al., 1988b), although some 

8704 evidence of a decrease in fertility index was seen in male mice exposed to 150 and 200 ppm 

8705 (Raje et al., 1988), and no adverse effects on fetal development of mice or rats exposed up to 

8706 1,250 ppm were seen by Schwetz et al. (1975). Decreases in fetal BW and changes in behavioral 

8707 habituation were observed in Long-Evans rats exposed to 4,500 ppm during the gestational 
312

8708 period (Bornschein et al., 1980; Hardin and Manson, 1980). Exposure-related noncancer effects 

8709 on the lungs consisted of foreign-body pneumonia in rats exposed to 8,400 ppm 6 hours/day for 

8710 13 weeks (NTP, 1986), Clara cell vacuolation in mice exposed to 4,000 ppm 6 hours/day for 

8711 13 weeks (Foster et al., 1992), and pulmonary congestion in guinea pigs exposed to 5,000 ppm 

8712 7 hours/day for 6 months (Heppel et al., 1944). Several neurological mediated parameters 

8713 including decreased activity (Kjellstrand et al., 1985; Weinstein et al., 1972; Heppel and Neal, 

8714 1944), impairment of learning and memory (Alexeef and Kilgore, 1983), and changes in 

8715 responses to sensory stimuli (Rebert et al., 1989) are reported from acute and short-term 

8716 dichloromethane exposure. Evidence of a localized immunosuppressive effect in the lung, 

8717 resulting from inhalation dichloromethane exposure, was seen in an acute exposure (3 hours, 

8718 100 ppm) study in CD-1 mice (Aranyi et al., 1986). 

8719 Numerous in vitro studies have demonstrated mutagenic and genotoxic effects 

8720 associated with dichloromethane exposure. For example, bacterial assays, yeast, and fungi 

8721 provide evidence that the mutagenic action of dichloromethane in bacterial systems is enhanced 

8722 by metabolic activation (e.g., Dillon et al., 1992; Jongen et al., 1982; Gocke et al., 1981). 

8723 Positive results from assays of DNA damage with in vitro mammalian systems provide support 

8724 that dichloromethane genotoxicity is linked to metabolism by GST enzymes (Graves et al., 1996, 

8725 1995, 1994b). Consistent evidence for several genotoxic endpoints in target tissues (liver and 

8726 lung) in mice following in vivo exposure to dichloromethane provides supporting evidence that 

8727 GST-pathway metabolites are key actors in the mutagenic and carcinogenic mode of action for 

8728 dichloromethane. Pretreatment of mice with buthionine sulphoximine, a GSH depletor, caused a 

8729 decrease to levels seen in controls in the amount of DNA damage detected immediately after in 

8730 vivo exposure in liver and lung tissue, indicating GSH involvement in the genotoxic process 

8731 (Graves et al., 1995). DNA damage (detected by the comet assay) was also reported in liver and 

8732 lung tissues from male CD-1 mice sacrificed 4 hours after administration of a single oral dose of 

8733 1,720 mg/kg of dichloromethane (Sasaki et al., 1998). In this study, DNA damage in lung and 

8734 liver was not detected 3 hours after dose administration, and no DNA damage occurred at either 

8735 time point in several other tissues in which a carcinogenic response was not seen in chronic 

8736 animal cancer bioassays (e.g., stomach, kidney, bone marrow). The weight of evidence from 

8737 these studies suggests that dichloromethane is carcinogenic by a mutagenic mode of action. 

8738 Dichloromethane is “likely to be carcinogenic in humans” by the inhalation and oral 

8739 routes of exposure under the Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a). 

8740 Results from several 2-year bioassays provide adequate evidence of the carcinogenicity of 

8741 dichloromethane in mice and rats exposed by inhalation, as well as adequate data to describe 

8742 dose-response relationships. Oral exposure to dichloromethane produced statistically significant 

8743 increases in hepatocellular adenomas and carcinomas in male B6C3F1 mice (Serota et al., 1986b; 

8744 Hazelton Laboratories, 1983). Inhalation exposure to concentrations of 2,000 or 4,000 ppm 

8745 dichloromethane produced increased incidences of lung and liver tumors in B6C3F1 mice 
313

8746 (Maronpot et al., 1995; Foley et al., 1993; Kari et al., 1993; Mennear et al., 1988; NTP, 1986). 

8747 Significantly increased incidences of benign mammary tumors (adenomas or fibroadenomas) 

8748 were observed in male and female F344/N rats exposed by inhalation to 2,000 or 4,000 ppm 

8749 (Mennear et al., 1988; NTP, 1986). A statistically significant increased incidence of brain or 

8750 CNS tumors has not been observed in any of the animal cancer bioassays, but a 2-year study 

8751 using relatively low exposure levels (0, 50, 200, and 500 ppm) in Sprague-Dawley rats observed 

8752 a total of six astrocytoma or glioma (mixed glial cell) tumors in the exposed groups (Nitschke et 

8753 al., 1988a). These tumors are exceedingly rare in rats, and there are few examples of statistically 

8754 significant trends in animal bioassays (Sills et al., 1999). 

8755 

8756 6.2. DOSE-RESPONSE 

8757 6.2.1. Oral RfD 

8758 The available oral toxicity data for animals identify hepatic effects (hepatic vacuolation, 

8759 nonneoplastic liver foci) as the most sensitive noncancer endpoint associated with chronic oral 

8760 exposure to dichloromethane. The 104-week drinking-water study in F344 rats (Serota et al., 

8761 1986a) was selected as the principal study for RfD derivation because the study provided a 

8762 sensitive endpoint (nonneoplastic liver foci) and used lower doses in comparison to other chronic 

8763 oral administration studies. In this study, four doses (6, 52, 125, and 235 mg/kg-day in males; 6, 

8764 58, 136, and 263 mg/kg-day in females) were used. A NOAEL of 6 mg/kg-day in males and 

8765 females and a LOAEL of 52 (male) and 58 (female) mg/kg-day for nonneoplastic alterations of 

8766 liver foci was identified. 

 An RfD of 7 × 10–3 8767 mg/kg-day is recommended for use in humans. The RfD derivation 

8768 process involved first fitting all available dichotomous models in BMDS version 2.0 to the 

8769 incidence data for male rats. The male data were used because a greater sensitivity was seen in 

8770 males compared with females in this study. A dose metric of average daily mass of 

8771 dichloromethane metabolized via the CYP pathway per unit volume of liver was derived from an 

8772 EPA-modified rat PBTK model (see Appendix C). This metric was chosen because there are no 

8773 data to support the role of a specific metabolite in the development of the noncancer liver lesions 

8774 seen in oral and inhalation exposure studies and the CYP-metabolism dose metric was 

8775 determined to be most consistent with the data. Then, the lower 95% confidence limit on the 

8776 dose associated with a 10% risk for liver lesions (BMDL10) was derived, based on the best fitting 

8777 model (in terms of the value of the AIC and examination of model fit and residuals). Because 

8778 the metric is a rate of metabolism, rather than the concentration of putative toxic metabolites, and 

8779 the clearance of these metabolites may be slower per volume tissue in the human compared with 

8780 the rat, this rodent internal dose metric for noncancer effects was adjusted by dividing by a 

8781 pharmacokinetic scaling factor to obtain a human equivalent internal BMDL10. This BMDL10

8782 was then converted to the human equivalent dose by using a human PBTK model (adapted from 

8783 David et al., 2006; see Appendix B) that utilizes Monte Carlo sampling techniques to provide a 
314

8784 distribution of human equivalent doses. The first percentile of the distribution of human 

8785 equivalent doses was chosen to include the most sensitive population, while staying within 

8786 bounds of what is considered computationally stable. The first percentile human equivalent 

8787 administered dose was used as a point of departure and was divided by a composite UF of 30 (3 

[100.5] to account for uncertainty about interspecies toxicodynamic equivalence, 3 [100.5 8788 ] to 

account for uncertainty about toxicodynamic variability in humans, and 3 [100.5 8789 ] for database 

deficiencies) to arrive at an RfD of 7 × 10–3 8790 mg/kg-day. 

 Use of the mean value (3.6 × 10–1 8791 mg/kg-day) of the human equivalent administered dose 

distribution instead of the 1st percentile, with an additional UF of 3 (100.5 8792 ) to account for human 

toxicokinetic variability, would yield a candidate RfD of 4 × 10–3 8793 , which is relatively similar to 

the recommended RfD of 1 × 10–3 8794 . 

8795 

8796 6.2.2. Inhalation RfC 

8797 The liver is the most sensitive target for noncancer toxicity in rats and mice, following 

8798 repeated inhalation exposure to dichloromethane. Nonneoplastic liver lesions (specifically, 

8799 hepatic vacuolation) in rats are the critical noncancer effect from chronic dichloromethane 

8800 inhalation in animals. Inhalation bioassays with Sprague-Dawley rats identified the lowest 

8801 inhalation LOAEL for nonneoplastic liver lesions in the database: 500 ppm (6 hours/day, 

8802 5 days/week for 2 years) (Nitschke et al., 1988a; Burek et al., 1984). Nitscke et al. (1988a) 

8803 identified a NOAEL of 200 ppm for hepatocyte vacuolation in female rats. Because the Nitschke 

8804 et al. (1988a) study more adequately covers the range spanning the BMR compared with the 

8805 study by Burek et al. (1984), the former study was selected as the principal study for derivation 

8806 of a chronic inhalation RfC. 

 An RfC of 0.2 mg/m38807 is derived based on the observed critical effect in the principal 

8808 study. As was described above for the RfD, the RfC derivation process was based on a dose 

8809 metric of average daily mass of dichloromethane metabolized via the CYP pathway per unit 

8810 volume of liver. This metric was derived from an EPA-modified rat PBTK model (see Appendix 

8811 C). Then, the lower 95% confidence limit on the dose associated with a 10% risk for liver 

8812 lesions (BMDL10) was derived, based on the best fitting model in terms of the value of the AIC 

8813 and examination of model fit and residuals. Because the metric is a rate of metabolism, rather 

8814 than the concentration of putative toxic metabolites, and the clearance of these metabolites may 

8815 be slower per volume tissue in the human compared with the rat, this rodent internal dose metric 

8816 for noncancer effects was adjusted by dividing by a pharmacokinetic scaling factor to obtain a 

8817 human-equivalent internal BMDL10. This BMDL10 was then converted to the HEC by using a 

8818 human PBTK model (adapted from David et al., 2006; see Appendix B) that utilizes Monte 

8819 Carlo sampling techniques to provide a distribution of HECs. 

8820 The first percentile HEC was used as a point of departure. This percentile was chosen 

8821 because it included the most sensitive population while staying within bounds of what is 
315

8822 considered computationally stable. This point of departure was divided by a composite UF of 

100 (3 [100.5] to account for uncertainty about interspecies toxicodynamic equivalence, 3 [100.5 8823 ] 

8824 to account for uncertainty about toxicodynamic variability in humans, and 10 for database 

deficiencies) to arrive at an RfC of 0.2 mg/m38825 . 

 Use of the mean value (47.36 mg/m3)of the HEC distribution instead of the 1st 8826 percentile, 

with an additional UF of 3 (100.5 8827 ) to account for human toxicokinetic variability would yield a 

candidate RfC identical to the recommended value of 0.2 mg/m38828 . In addition, two comparison 

values derived from occupational studies produced values of 3.5 mg/m3 8829 (Cherry et al., 1983) and 

0.55 mg/m38830 (Lash et al., 1991). The animal-derived candidate RfC is preferable to the human8831 derived candidate RfC because of the uncertainties about the exposure durations for the workers 

8832 in the Cherry et al. (1983) study and uncertainties regarding the exposures and effect sizes in 

8833 Lash et al. (1991) and because the RfC based on the rat data is more health protective. 

8834 

8835 6.2.3. Uncertainties in Reference Dose and Reference Concentration Values 

8836 One data uncertainty identified is the potential for neurodevelopmental effects. Animal 

8837 bioassays have not identified gross or microscopic effects on neural tissues from long-term 

8838 exposures or single (Schwetz et al., 1975) or multigenerational (Nitschke et al., 1988b) 

8839 developmental toxicity studies. However, behavioral changes were observed in pups born to rats 

8840 exposed to high levels (4,500 ppm) of dichloromethane (Bornschein et al., 1980; Hardin and 

8841 Manson, 1980); 4,500 ppm was the only dose used in this study. Thus uncertainty exists as to 

8842 the development of neurological effects from lower gestational exposures in animals, or in 

8843 humans. In addition, immunotoxicity data revealed an additional area of data uncertainty 

8844 specifically with respect to inhalation exposure. Data from Aranyi et al. (1986) demonstrated 

8845 evidence of immunosuppression, following a single 100 ppm dichloromethane exposure for 3 

8846 hours in CD-1 mice. The weight of evidence for nonneoplastic effects in humans and animals 

8847 suggests that the development of liver lesions is the most sensitive effect, with a UF applied 

8848 because of the lack of neurodevelopmental studies and, for the RfC, the uncertainty regardingthe 

8849 lack of a low dose developmental study. 

8850 The extrapolation of internal dichloromethane dosimetry from nonneoplastic rat 

8851 responses to human risk was accomplished by using PBTK models for dichloromethane in rats 

8852 and humans. Uncertainties in rat and human dosimetry used for RfD and RfC derivation can 

8853 arise from uncertainties in the PBTK models to accurately simulate the toxicokinetics of 

8854 dichloromethane for animals under bioassay conditions and humans experiencing relatively low, 

8855 chronic environmental exposures. Further, the dose metric used in the models is the rate of 

8856 metabolism to a putative toxic metabolite, rather than the concentration (average or area under 

8857 the concentration curve of the metabolite), so the model specifically fails to account for rodent8858 human differences in clearance or removal of the toxic metabolite. A scaling factor, based on 

8859 BW ratios, was used to account for this difference. 
316

8860 Uncertainties in the human population model were quantitatively accounted for by 

8861 utilizing hierarchical Bayesian calibration methods during model development (David et al., 

8862 2006; Marino et al., 2006). The rat model was modified and utilized in a deterministic manner. 

8863 Data were not available to perform a hierarchical Bayesian calibration in the rat, but 

8864 uncertainties in the rat model predictions were assessed qualitatively. For both oral and 

8865 inhalation exposures, the liver volume, followed closely by the volume of slowly perfused 

8866 tissues, had the greatest impact on the internal dose of mg dichloromethane metabolized via CYP 

8867 pathway per liter tissue per day. This was due to the fact that the dose metric is a tissue-specific 

8868 concentration, the majority of CYP metabolism is attributed to the liver, and changes in liver 

8869 volume have a greater impact on the total CYP metabolism than either of the individual Vmax

8870 values. There is high confidence in the values used for volume of liver and slowly perfused 

8871 tissues in the rat, as these are well studied (Brown et al., 1997). Therefore, the uncertainties 

8872 associated with use of the rat PBTK model should not markedly affect the values of the RfD and 

8873 RfC. 

8874 An additional uncertainty inherent in this process, however, is the lack of knowledge 

8875 concerning the most relevant dose metric (e.g., a specific metabolite) within the context of the 

8876 development of the noncancer liver effects. This basic research question represents a data gap, 

8877 and this uncertainty is not addressed quantitatively or qualitatively in the assessment. 

8878 The effect of dichloromethane on human populations that are sensitive due to 

8879 pharmacokinetic differences was addressed quantitatively by using a human probabilistic PBTK 

8880 model to generate distributions of human exposures likely to occur given a specified internal 

8881 BMDL10. The model and resulting distributions take into account the known differences in 

8882 human physiology and metabolic capability with regard to dichloromethane dosimetry. The first 

8883 percentile values of the distributions of human equivalent doses (Table 5-3) and HECs (Table 5-

8884 7) served as points of departure for candidate RfDs and RfCs, respectively, to protect 

8885 toxicokinetically sensitive individuals. No data are available regarding toxicodynamic 

8886 differences within a human population. Therefore, a UF of 3 for possible differences in human 

8887 toxicodynamic responses is intended to be protective for sensitive individuals. 

8888 

8889 6.2.4. Oral Cancer Slope Factor 

The recommended oral cancer slope factor for dichloromethane is 1 × 10–3 (mg/kg-day)–1 8890 , 

8891 which is based on liver tumor responses in male B6C3F1 mice exposed to dichloromethane in 

8892 drinking water for 2 years (Serota et al., 1986b; Hazelton Laboratories, 1983). This value was 

8893 derived by using a tissue-specific GST metabolism dose metric with allometric scaling to 

8894 account for uncertainty regarding the reactivity and clearance of the metabolite(s) involved in the 

8895 carcinogenic response. 

8896 There was only one adequate oral exposure cancer bioassay (Serota et al., 1986a, b; 

8897 Hazelton Laboratories, 1983) evaluating the carcinogenic potential of orally administered 
317

8898 dichloromethane in F344 rats and B6C3F1 mice. Significant increases in incidence of liver

8899 adenomas and carcinomas were observed in male (trend p-value = 0.058) but not female B6C3F1

8900 mice (Serota et al., 1986b; Hazelton Laboratories, 1983). In F344 rats (Serota et al., 1986a), no 

8901 increased incidence of liver tumors was seen in male rats, and the pattern in female rats was 

8902 characterized by a jagged stepped pattern of increasing incidence of hepatocellular carcinoma or 

8903 neoplastic nodules; a similar pattern, but based on more sparse data, was seen when limited to 

8904 hepatocellular carcinomas. Statistically significant increases in tumor incidences were observed 

8905 in the 50 and 250 mg/kg-day groups (incidence rates of 10 and 14%, respectively) but not in the 

8906 125 mg/kg-day group (incidence rate of 3%). Incidence was also increased (10%) in a group 

8907 exposed for 78 weeks followed by a 26-week period of no exposure. The derivation of the oral 

8908 cancer slope factor is based on the male mice data because of their greater sensitivity to liver 

8909 cancer compared with female rats. 

8910 A modified mouse PBTK model of Marino et al. (2006) was used to approximate the 

8911 internal dose of daily dichloromethane (mg) metabolized via the GST pathway per unit volume 

8912 of liver from the daily oral administered doses. This approach was taken based on evidence that 

8913 GST-pathway metabolites produced from dichloromethane are primarily responsible for 

8914 dichloromethane carcinogenicity in mouse liver. The multistage dose-response model (BMDS 

8915 version 2.0) was used to fit the mouse liver tumor incidence and PBTK model-derived internal 

8916 dose data and derive a mouse internal BMD and BMDL associated with 10% extra risk 

8917 (BMDL10). The human BMDL10 was derived by multiplying the mouse BMDL10 by allometric 

scaling factor (BWhuman/BWmouse)

0.25 8918 ≈ 7). Linear extrapolation from the internal human 

8919 BMDL10 (0.1/BMDL10) was used to derive oral risk factors for liver tumors based on tumor 

8920 responses in male mice. The linear low-dose extrapolation approach for agents with a mutagenic 

8921 mode of action was selected because GST-metabolism of dichloromethane is expected to occur 

8922 at and below exposures producing the mouse BMDL10, even though CYP2E1 metabolism is 

8923 expected to be unsaturated and to represent the predominant metabolic pathway in the liver. 

8924 Currently, there are no data from chronic oral cancer bioassays in mice providing support for a 

8925 nonlinear dose-response relationship. 

8926 Probability distributions of human oral cancer slope factors were derived by using a 

8927 human PBTK model (adapted from David et al. [2006]; see Appendix B). The cancer reference 

8928 values (OSF and IUR) were derived for a sensitive population: a population composed entirely 

of carriers of the GST-T1+/+ 8929 homozygous genotype (that is, the group that would be expected to 

8930 be most sensitive to the carcinogenic effects of dichloromethane). In addition, cancer values 

8931 derived for a population reflecting the estimated frequency of GST-T1 genotypes in the current 

U.S. population (20% GST-T1–/–, 48% GST-T1+/–, and 32% GST-T1+/+ 8932 ) were presented. All 

8933 simulations also included a distribution of CYP activity based on data from Lipscomb et al. 

8934 (2003). The mean OSF based on liver tumors in mice exposed to dichloromethane in drinking 

water, 1 × 10–3 (mg/kg-day)–1 8935 , based on what is assumed to be the most sensitive of the 
318

populations, the GST-T1+/+ 8936 group, is the recommended OSF to be used in deterministic risk 

8937 assessments for chronic oral exposures to dichloromethane. 

8938 An OSF derived from the liver tumor data in the Serota et al. (1986b) study, using 

8939 administered dose dosimetry rather than PBTK modeling, is approximately one order of 

magnitude higher than the current recommended value of 1 × 10–3 8940 (per mg/kg-day). There is 

8941 approximately one to two orders of magnitude difference among the values based on different 

8942 dose metrics, scaling factors, and populations (Table 6-1). 

 The recommended OSF of 1 × 10–3 8943 (per mg/kg-day) is based on a tissue-specific GST 

8944 internal dose metric with allometric scaling. Although the involvement of the GST pathway in 

8945 carcinogenic response has been established, some uncertainty remains as to the metabolite(s) 

8946 involved and the rate at which those metabolites are cleared. The value derived specifically for 

the GST-T1+/+ 8947 population is recommended to provide protection for the population that is 

8948 hypothesized to be most sensitive to the carcinogenic effect. Application of ADAFs to the 

8949 cancer OSF is recommended in combination with appropriate exposure data when assessing 

8950 risks associated with early-life exposure (see section 5.4.4 for more details). 8951 
319

8952 

Table 6-1. Comparison of OSFs derived by using various assumptions and metrics, based on liver tu

in male mice 

Populationa Dose metric 

Species, 

sex Tumor 

Scaling 

factor 

Mean OSF 

(mg/kg-day)–

GST-T1+/+ Tissue-specific GST-metabolism rateb Mouse, male Liver 7.0 1.4 × 10–3 

 Tissue-specific GST-metabolism rate Mouse, male Liver 1.0 2.0 × 10–4 

 Whole-body GST metabolism rate Mouse, male Liver 7.0 8.1 × 10–4  Route-to-route extrapolation, tissue-specific metabolism Mouse, male Liver 7.0 1.0 × 10–4  Route-to-route extrapolation, tissue-specific metabolism Mouse, male Liver 1.0 1.5 × 10–5  Route-to-route extrapolation, whole-body metabolism Mouse, male Liver 7.0 5.8 × 10–5 

Mixed Tissue-specific GST-metabolism rateb Mouse, male Liver 7.0 8.0 × 10–4 

 Tissue-specific GST-metabolism rate Mouse, male Liver 1.0 1.2 × 10–4 

 Whole-body GST metabolism rate Mouse, male Liver 7.0 4.6 × 10–4  Route-to-route extrapolation, tissue-specific metabolism Mouse, male Liver 7.0 5.8 × 10–5  Route-to-route extrapolation, tissue-specific metabolism Mouse, male Liver 1.0 8.3 × 10–6  Route-to-route extrapolation, whole-body metabolism Mouse, male Liver 7.0 3.3 × 10–5  Applied dose (human equivalent dose) Mouse, male Liver 1.0 × 10–2 

1995 IRIS assessment Mouse, male Liver 7.5 × 10–3

aGST-T1+/+ = homozygous, full enzyme activity; Mixed = genotypes based on a population reflecting the estimated frequency of genotyp

U.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 × 10–3 per mg/kg-day. 
320

8953 

8954 6.2.5. Cancer Inhalation Unit Risk 

The recommended cancer IUR is 1 × 10–8 (μg/m3)

–1 8955 for the development of liver and lung 

8956 cancers, based on data from male B6C3F1 mice, using a tissue-specific GST metabolism dose 

8957 metric. Data for liver and lung tumors in male and female B6C3F1 mice, following exposure to 

8958 airborne dichloromethane, were used to develop IURs for dichloromethane (Mennear et al., 

8959 1988; NTP, 1986). This study was selected as the principal study to derive an IUR for 

8960 dichloromethane because of the completeness of the data, adequate sample size, and clear dose 

8961 response. In the NTP (1986) study, significant increases in incidence of liver and lung adenomas 

8962 and carcinomas were observed in both sexes of B6C3F1 mice exposed 6 hours/day, 5 days/week 

8963 for 2 years. 

8964 The PBTK model of Marino et al. (2006) for dichloromethane in the mouse was used to 

8965 calculate long-term daily average internal liver doses. The selected internal dose metrics for 

8966 liver tumors and lung tumors were long-term average daily mass of dichloromethane 

8967 metabolized via the GST pathway per unit volume of liver and lung, respectively. This approach 

8968 was taken based on evidence that GST-pathway metabolites produced from dichloromethane are 

8969 primarily responsible for dichloromethane carcinogenicity in mouse liver. The multistage dose8970 response model (BMDS version 2.0) was used to fit the mouse liver tumor incidence and PBTK 

8971 model-derived internal dose data and derive a mouse internal BMD and BMDL associated with 

8972 10% extra risk (BMDL10). The human BMDL10 was derived by multiplying the mouse BMDL10

by allometric scaling factor (BWhuman/BWmouse)

0.25 8973 ≈ 7). A linear extrapolation approach using 

8974 the internal human BMDL10 for liver and lung tumors was used to calculate human tumor risk 

8975 factors by dividing the BMR of 0.1 by the human BMDL for each tumor type. Currently, there 

8976 are no data from chronic inhalation cancer bioassays in mice or rats providing support for a 

8977 nonlinear dose-response relationship. 

8978 The human PBTK model (adapted from David et al. [2006]; see Appendix B) provided 

8979 distributions of human internal dose metrics of daily mass of dichloromethane metabolized via 

8980 the GST pathway per unit volume of liver and lung resulting from chronic inhalation exposure to 

a unit concentration of 1 μg/m3

8981 dichloromethane (0.00029 ppm). As with the OSF, the cancer 

8982 IUR was derived for a sensitive population: a population composed entirely of carriers of the 

8983 GST-T1 homozygous positive genotype (that is, the group that would be expected to be most 

8984 sensitive to the carcinogenic effects of dichloromethane). In addition, cancer values derived for 

8985 a population reflecting the estimated frequency of GST-T1 genotypes in the current U.S. 

population (20% GST-T1–/–, 48% GST-T1+/–, and 32% GST-T1+/+ 8986 ) were also presented. The 

8987 distributions of IURs for liver or lung tumors were generated by multiplying the human tumor 

8988 risk factor for each tumor type and sex by the distribution of internal doses from chronic 

exposure to 1 μg/m3

8989 dichloromethane. A procedure to combine risks for liver and lung tumors 

8990 using different dose metrics for the different tumors (i.e., liver-specific and lung-specific 
321

8991 metabolism for liver and lung tumors, respectively), was used to derive the recommended IUR of 

1 × 10–8 (μg/m3)

–1 8992 based on what is assumed to be the most sensitive of the populations, the 

GST-T1+/+ 8993 group. 

 The current recommended IUR value of 1 × 10–8 (μg/m3)

–1

8994 is approximately 1.5 orders 

of magnitude lower than the previous IRIS value of 4.7 × 10–7 (μg/m3)

–1

8995 and similar to the 

occupational exposure-based risk value of 4.17 × 10

–8

 (μg/m3)

–1

8996 promulgated by OSHA (1997) 

and derived from an estimated risk of 3.62 × 10–38997 for a lifetime occupational inhalation exposure 

8998 of 25 ppm. The current use of the updated mouse PBTK model, with blood and tissue 

8999 equilibrium partition coefficients and metabolic parameters updated with MCMC calibration, 

9000 resulted in approximately three- and four-fold increases in the values of internal liver and lung 

9001 dose metrics, respectively, associated with the dichloromethane exposure concentrations, 

9002 compared with estimates from the model used in the previous IRIS assessment. For a unit 

9003 inhalation exposure, the mean internal lung GST dose predicted for the entire population 

9004 predicted by the MCMC updated human PBTK model is approximately thirteen-fold lower 

9005 compared with the human PBTK model used in the U.S. EPA (1995) assessment. The mean 

9006 internal liver GST dose, however, is approximately the same as (only 16% higher than) that 

9007 obtained with the previous PBTK parameters. For unit oral exposures, the mean internal liver 

9008 GST dose predicted by the MCM updated model is about 80% of that predicted using the 

9009 previous parameters, while the mean whole-body GST dose is predicted to be about 50% of that 

9010 predicted using the previous parameters.

9011 An IUR derived from the liver tumor data of the NTP (1986) study using applied 

concentration dosimetry rather than PBTK modeling, 3.7 × 10–7 (μg/m3)

–1 9012 , is approximately one 

order of magnitude higher than the currently recommended value of 1 × 10–8 (μg/m3

)

–1 9013 (Table 

9014 6-2). There is approximately one- to two- orders of magnitude difference among the values 

9015 based on different dose metrics, scaling factors, and populations. 

The recommended IUR value of 1 × 10–8 (μg/m3

)

–1 9016 is based on a tissue-specific GST9017 internal dose metric with allometric scaling. Although the involvement of the GST pathway in 

9018 carcinogenic response has been established, some uncertainty remains as to the metabolite(s) 

9019 involved and the rate at which those metabolites are cleared. The value derived specifically for 

the GST-T1+/+ 9020 population is recommended to provide protection for the population that is 

9021 hypothesized to be most sensitive to the carcinogenic effect. Application of ADAFs to the 

9022 cancer IUR is recommended when assessing risks associated with early-life exposure (see 

9023 section 5.4.4 for more details). 
322

9024 9025 

Table 6-2. Comparison of IURs derived by using various assumptions and metrics 

Populationa Dose metric Species, sex, Tumor type 

Scaling 

factor 

IURb

(μg/m3)–

GST-T1+/+ Tissue-specific GST-metabolism ratec Mouse, male Liver and lung 7.0 1.1 × 10–GST-T1+/+ Tissue-specific GST-metabolism rate Mouse, male Liver 7.0 7.3 × 10–

GST-T1+/+ Tissue-specific GST-metabolism rate Mouse, male Lung 7.0 4.8 × 10–GST-T1+/+ Tissue-specific GST-metabolism rate Mouse, male Liver and lung 1.0 1.6 × 10–GST-T1+/+ Tissue-specific GST-metabolism rate Mouse, male Liver 1.0 1.0 × 10–GST-T1+/+ Tissue-specific GST-metabolism rate Mouse, male Lung 1.0 6.8 × 10–

GST-T1+/+ Whole-body GST metabolism rate Mouse, male Liver and lung 7.0 1.4 × 10–

GST-T1+/+ Whole-body GST metabolism rate Mouse, male Liver 7.0 4.6 × 10–GST-T1+/+ Whole-body GST metabolism rate Mouse, male Lung 7.0 1.0 × 10–

Mixed Tissue-specific GST-metabolism rate Mouse, male Liver and lung 7.0 5.5 × 10–

Mixed Tissue-specific GST-metabolism rate Mouse, male Liver 7.0 3.4 × 10–

Mixed Tissue-specific GST-metabolism rate Mouse, male Lung 7.0 2.6 × 10–Mixed Tissue-specific GST-metabolism rate Mouse, male Liver and lung 1.0 7.9 × 10–

Mixed Tissue-specific GST-metabolism rate Mouse, male Liver 1.0 4.8 × 10–Mixed Tissue-specific GST-metabolism rate Mouse, male Lung 1.0 3.7 × 10–Mixed Whole-body GST metabolism rate Mouse, male Liver and lung 7.0 7.9 × 10–

Mixed Whole-body GST metabolism rate Mouse, male Liver 7.0 2.7 × 10–Mixed Whole-body GST metabolism rate Mouse, male Lung 7.0 6.0 × 10–

 Administered concentration (HEC) Mouse, male Liver 3.7 × 10–

Administered concentration (HEC) Mouse, male Lung 9.2 × 10–1995 IRIS assessmentc Mouse, male Liver and lung 12.7 4.7 × 10–

aGST-T1+/+ = homozygous, full enzyme activity; Mixed = genotypes based on a population reflecting the estimated frequency of genotyp

population: 20% GST-T1–/–, 48% GST-T1+/–, and 32% GST-T1+/+ (Haber et al., 2002). 

bBased on mean value of the derived distributions Bolded value is the basis for the recommended IUR of 1 × 10–8 μg/m3 per mg/kg-day. 

9026 
323

9027 6.2.6. Uncertainties in Cancer Risk Values 

9028 The database of animal bioassays identifies the liver and lung as the most sensitive target 

9029 organs for dichloromethane-induced tumor development, and there is high confidence that the 

9030 dose-response data for liver and lung cancer in mice represent the best available data for 

9031 derivation of human cancer risks. A dose-response relationship was seen with respect to liver 

9032 cancer in mice exposed orally and with respect to liver and lung cancer in mice exposed by 

9033 inhalation. Statistically significant increases in benign mammary gland tumors were observed in 

9034 one study of F344 rats exposed by inhalation to 2,000 or 4,000 ppm (Mennear et al., 1988; NTP, 

9035 1986); evidence for a tumorigenic mammary gland response in Sprague-Dawley rats was limited 

9036 to increased numbers of benign mammary tumors per animal at levels of 50–500 ppm (Nitschke 

9037 et al., 1988a) or 500–3,500 ppm (Burek et al., 1984). An oral (gavage) study in female Sprague9038 Dawley rats reported an increased incidence of malignant mammary tumors, mainly 

9039 adenocarcinomas (8, 6, and 18% in the control, 100, and 500 mg/kg dose groups, respectively), 

9040 but the increase was not statistically significant. Data were not provided to allow an analysis that 

9041 accounts for differing mortality rates (Maltoni et al., 1988). The toxicokinetic or mechanistic 

9042 events that might lead to mammary gland tumor development in rats are unknown, although 

9043 CYP2E1 (El-Rayes et al., 2003; Hellmold et al., 1998) and GST-T1 expression has been detected 

9044 in human mammary tissue (Lehmann and Wagner, 2008). Rare CNS tumors were observed in 

9045 one study in rats, a study spanning a relatively low range of exposures (0–500 ppm) (Nitschke et 

9046 al., 1988a). These cancers were not seen in two other studies in rats, both involving higher doses 

9047 (1,000–4,000 ppm) (NTP, 1986; Burek et al., 1984), or in a similar high-dose study in mice 

9048 (NTP, 1986). The relative rarity of the tumors precludes the use of the low-dose exposure study 

9049 (Nitschke et al., 1988a) in a quantitative dose-response assessment. The available epidemiologic 

9050 studies provide some evidence of an association between dichloromethane and brain cancer. The 

9051 available epidemiologic studies do not provide an adequate basis for the evaluation of the role of 

9052 dichloromethane in breast cancer because there are currently no cohort studies with adequate 

9053 statistical power and no case-control studies with adequate exposure methodology to examine 

9054 this relationship. 

9055 There is uncertainty as to whether the reactivity of the toxic dichloromethane metabolites 

9056 is sufficiently high enough to preclude systemic distribution. Therefore, alternative derivations 

9057 of cancer risk values were performed under the assumption that high reactivity leads to complete 

9058 clearance from the tissue in which the active metabolite is formed (scaling factor = 1.0). The 

9059 difference in scaling factor (7.0 for allometric scaling versus 1.0) results in a 7-fold decrease in 

9060 estimated cancer toxicity values. Using a whole-body GST metabolism dose metric, the 

9061 resulting OSF and IUR for liver and lung cancer were approximately five-fold higher than when 

9062 tissue-specific dose metrics were used (Table 5-16 and Table 5-22). The mechanistic data 

9063 support the notion that reactive metabolites produced in the target tissues are not well distributed 

9064 and produce deleterious effects in the metabolizing tissues soon after generation. Thus, there is 
324

9065 less uncertainty in the cancer risk values derived by using a tissue-specific GST metabolism dose 

9066 metric compared with those derived using a whole-body GST metabolism dose metric. 

9067 Uncertainty in the ability of the PBTK models to estimate animal and human internal 

9068 doses from lifetime bioassay low-level environmental exposures may affect the confidence in the 

9069 cancer risk extrapolated from animal data. Uncertainties in the mouse and human model 

9070 parameter values were integrated quantitatively into parameter estimation by utilizing 

9071 hierarchical Bayesian methods to calibrate the models at the population level (David et al., 2006; 

9072 Marino et al., 2006). The use of Monte Carlo sampling to define human model parameter 

9073 distributions allowed for derivation of human distributions of dosimetry and cancer risk, 

9074 providing for bounds on the recommended risk values. However, the PBTK animal models were 

9075 utilized deterministically, and a sensitivity analysis was performed to identify model parameters 

9076 most influential on the predictions of dose metrics used to estimate oral and inhalation cancer 

9077 risks. For inhalation exposures in mice, the blood:air partition coefficient, followed closely by 

9078 the first-order GST-mediated metabolism rate, had the greatest impact on the dose metric for 

9079 liver cancer (mg dichloromethane metabolized via GST pathway per liter liver per day). For 

9080 drinking water exposures in mice, the first-order GST-mediated metabolism rate, followed by the 

9081 CYP-mediated maximum reaction velocity (Vmaxc) affected the liver cancer dose metric to the 

9082 greatest extent. For mice inhaling dichloromethane, the lung cancer dose metric (mg 

9083 dichloromethane metabolized via GST pathways per liter lung per day), like the liver cancer 

9084 metric, was highly affected by the first-order GST-mediated metabolism rate and the blood:air 

9085 partition coefficient. However, the lung cancer dose metric was most sensitive to the 

9086 proportional yield of liver GST-mediated metabolic activity attributed to the lung. The blood:air 

9087 partition coefficient was experimentally determined, lending high confidence to its value. In 

9088 contrast, values for the three metabolic parameters were determined by computational 

9089 optimization against data sets not directly measuring dichloromethane or its metabolites in the 

9090 target/metabolizing tissues. It is uncertain how alternative values for these three parameters 

9091 would affect the estimation of animal BMDL10 values and, ultimately, IURs and OSFs. In 

9092 addition, specific uncertainty remains concerning the human PBTK parameter distributions. In 

9093 addition, while the structure and equations used in the existing model have been described 

9094 extensively in peer-reviewed publications, uncertainty remains concerning the model structure, 

9095 and specifically the potential of an alternative (dual-binding-site) CYP metabolic rate equation 

9096 for dichloromethane Integration of the alternate rate equation into the PBTK modeling, and then 

9097 quantitative risk assessment, will likely require several years of further research, and hence is 

9098 beyond the scope of the current assessment. 
325

7. REFERENCES 

ACGIH (American Conference of Governmental Industrial Hygienists). (2001) Dichloromethane. In: 

Documentation of the threshold limit values and biological exposure indices. 7th edition. Cincinnati, OH: 

American Conference of Governmental Industrial Hygienists. 

Alexeef, GV; Kilgore, WW. (1983) Learning impairment in mice following acute exposure to 

dichloromethane and carbon tetrachloride. J Toxicol Environ Health 11:569–581. 

Allen, J; Kligerman, A; Campbell, J; et al. (1990) Cytogenetic analyses of mice exposed to 

dichloromethane. Environ Mol Mutagen 15:221–228. 

Anders, MW; Sunram, JM. (1982) Transplacental passage of dichloromethane and carbon monoxide. 

Toxicol Lett 12:231–234. 

Andersen, ME; Clewell, HJ, III; Gargas, ML; et al. (1987) Physiologically based pharmacokinetics and the 

risk assessment process for methylene chloride. Toxicol Appl Pharmacol 87:185–205. 

Andersen, ME; Clewell, HJ, III; Gargas, ML; et al. (1991) Physiologically based pharmacokinetic 

modeling with dichloromethane, its metabolite, carbon monoxide, and blood carboxyhemoglobin in rats 

and humans. Toxicol Appl Pharmacol 108:14–27. 

Andrae, U; Wolff, T. (1983) Dichloromethane is not genotoxic in isolated rat hepatocytes. Arch Toxicol 

52:287–290. 

Angelo, MJ; Pritchard, AB; Hawkins, DR; et al. (1986a) The pharmacokinetics of dichloromethane. I. 

Disposition in B6C3F1 mice following intravenous and oral administration. Food Chem Toxicol 

24(9):965–974. 

Angelo, MJ; Pritchard, AB; Hawkins, DR; et al. (1986b) The pharmacokinetics of dichloromethane. II. 

Disposition in Fischer 344 rats following intravenous and oral administration. Food Chem Toxicol 

24(9):975–980. 

Anthony, T. (1979) Methylene chloride. In: Kirk, RE; Othmer, DF; eds. Kirk-Othmer encyclopedia of 

chemical technology. 3rd edition. New York, NY: John Wiley & Sons; pp. 686–693. 

Anundi, H; Lind, ML; Friis, L; et al. (1993) High exposures to organic solvents among graffiti removers. 

Int Arch Occup Environ Health 65:247–251. 

Aranyi, C; O'Shea, WJ; Graham, JA; et al. (1986) The effects of inhalation of organic chemical air 

contaminants on murine lung host defenses. Fundam Appl Toxicol 6:713–720. 

Arcus-Arth, A; Blaisdell, RJ. (2007) Statistical distributions of daily breathing rates for narrow age groups 

of infants and children. Risk Anal 27:97–110. 

Astrand, I; Ovrum, P; Carlsson, A. (1975) Exposure to methylene chloride. I. Its concentration in alveolar 

air and blood during rest and exercise and its metabolism. Scand J Work Environ Health 1:78–84. 

Atkinson, R. (1989) Kinetics and mechanisms of the gas-phase reactions of the hydroxyl radical with 

organic compounds. J Phys Chem Ref Data 1:63. 

ATSDR (Agency for Toxic Substances and Disease Registry). (2000) Toxicological profile for methylene 

chloride. Public Health Service, U.S. Department of Health and Human Services, Atlanta, GA. Available

online at http://www.atsdr.cdc.gov/toxpro2.html. 
326

Bakinson, MA; Jones, RD. (1985) Gassings due to methylene chloride, xylene, toluene, and styrene 

reported to Her Majesty’s factory inspectorate 1961–80. Br J Ind Med 42:184–190. 

Bell, BP; Franks, P; Hildreth, N; et al. (1991) Methylene chloride exposure and birth weight in Monroe

County, New York. Environ Res 55:31–39. 

Berman, E; Schlicht, M; Moser, VC; et al. (1995) A multidisciplinary approach to toxicological screening: 

I. Systemic toxicity. J Toxicol Environ Health 45:27–143. 

Bernauer, U; Vieth, B; Ellrich, R; et al. (2000) CYP2E1 expression in bone marrow and its intra- and 

interspecies variability: approaches for a more reliable extrapolation from one species to another in the risk 

assessment of chemicals. Arch Toxicol 73:618–624. 

Blair, A; Hartge, P; Stewart, P; et al. (1998) Mortality and cancer incidence of aircraft maintenance 

workers exposed to trichloroethylene and other organic solvents and chemicals: extended follow up. Occup 

Environ Med 55:161–171. 

Blocki, FA; Logan, MSP; Baoli, C; et al. (1994) Reaction of rat liver glutathione s-transferase and bacterial 

dichloromethane dehalogenase with dihalomethanes. J Biol Chem 269(12):8826–8830. 

Bogaards, JJP; van Ommen, B; van Bladeren, PJ. (1993) Interindividual differences in the in vitro 

conjugation of methylene chloride with glutathione by cytosolic glutathione S-transferase in 22 human liver 

samples. Biochem Pharmacol 45(10):2166–2169. 

Bornschein, RL; Hastings, L; Manson, JM. (1980) Behavioral toxicity in the offspring of rats following

maternal exposure to dichloromethane. Toxicol Appl Pharmacol 52:29–37. 

Bos, PMJ; Zeilmaker, MJ; van Eijkeren, JCH. (2006) Application of physiologically based 

pharmacokinetic modeling in setting acute exposure guideline levels for methylene chloride. Toxicol Sci 

91(2):576–585. 

Boublik, T; Fried, V; Hala, E. (1984) The vapor pressures of pure substances: selected values of the 

temperature dependence of the vapour pressures of some pure substances in the normal and low pressure 

region. Vol. 17. 2nd revised edition. Amsterdam, Netherlands: Elsevier; p. 42. 

Bowen, SE; Batis, JC; Paez-Martinez, N; Cruz, SL. (2006) The last decade of solvent research in animal

models of abuse: mechanistic and behavioral studies. Neurotoxicol Teratol 28(6): 636-647.. 

Breslow, NE; Day, NE. (1987) Statistical methods in cancer research. Volume II. The design and analysis 

of cohort studies. IARC Sci Publ. 82:1–406. 

Briving, C; Hambergerm, A; Kjellstrand, P; et al. (1986) Chronic effects of dichloromethane on amino 

acids, glutathione and phosphoethanolamine in gerbil brain. Scand J Work Environ Health 12:216–220. 

Brown, RP; Delp, MD; Lindstedt, SL; et al. (1997) Physiological parameter values for physiologicallybased pharmacokinetic models. Toxicol Ind Health 13(4):407–484. 

Brown-Woodman, PDC; Hayes, LC; Huq, F; et al. (1998) In vitro assessment of the effect of halogenated 

hydrocarbons: chloroform, dichloromethane, and dibromoethane on embryonic development of the rat. 

Teratology 57:321–333. 

Bruhn, C; Bröckmoller, J; Kerb, R; et al. (1998) Concordance between enzyme activity and genotype of 

glutathione S-transferase theta (GST-T1). Biochem Pharmacol 56:1189–1193. 
327

Brzezinski, MR; Boutelet-Bochan, H; Person, RE; et al. (1999) Catalytic activity and quantitation of 

cytochrome P-450 2E1 in prenatal human brain. J Pharmacol Exp Ther 289:1648–1653. 

Burek, JD; Nitschke, KD; Bell, TJ; et al. (1984) Methylene chloride: a two-year inhalation toxicity and 

oncogenicity study in rats and hamsters. Fundam Appl Toxicol 4:30–47. 

Cagiano, R; Ancona, D; Cassano, T; et al. (1998) Effects of prenatal exposure to low concentrations of

carbon monoxide on sexual behaviour and mesolimbic dopaminergic function in rat offspring. Br J 

Pharmacol 125(4):909–915. 

Callen, DF; Wolf, CR; Philpot, RM. (1980) Cytochrome P-450 mediated genetic activity and cytotoxicity 

of seven halogenated aliphatic hydrocarbons in Saccharomyces cerevisiae. Mutat Res 77:55–63. 

Candrian, U; You, M; Goodrow, T; Maronpot, RR; Reynolds, SH; Anderson, MW (1991). Activation of 

protooncogenes in spontaneously occurring non-liver tumors from C57BL/6 x C3H F1 mice. Cancer Res 

51:1148-1153. 

Cantor, KP; Stewart, PA; Brinton, LA; et al. (1995) Occupational exposures and female breast cancer 

mortality in the United States. J Occup Environ Med 37:336–348. 

Carlsson, A; Hultengren, M. (1975) Exposure to methylene chloride. III. Metabolism of 14C-labelled 

methylene chloride in rat. Scand J Work Environ Health 1:104–108. 

Casanova, M; Deyo, DF; Heck, Hd'A. (1992) Dichloromethane (methylene chloride): metabolism to 

formaldehyde and formation of DNA-protein cross links in B6C3F1 mice and Syrian golden hamsters. 

Toxicol Appl Pharmacol 114:162–165. 

Casanova, M; Conolly, RB; Heck, Hd'A. (1996) DNA-protein cross-links (DPX) and cell proliferation in 

B6C3F1 mice but not Syrian golden hamsters exposed to dichloromethane: pharmacokinetics and risk 

assessment with DPX as dosimeter. Fundam Appl Toxicol 31:103–106. 

Casanova, M; Bell, DA; Heck, Hd'A. (1997) Dichloromethane metabolism to formaldehyde and reaction of 

formaldehyde with nucleic acids in hepatocytes of rodents and humans with and without glutathione 

S-transferase T1 and M1 genes. Fundam Appl Toxicol 37:168–180. 

Chang, Y-L; Yang, CC; Deng, JF; et al. (1999) Diverse manifestations of oral methylene dichloride 

poisoning: report of 6 cases. Clin Toxicol 37(4):497–504. 

Cherry, N; Venables, H; Waldron, HA; et al. (1981) Some observations on workers exposed to methylene 

chloride. Br J Ind Med 38:351–355. 

Cherry, N; Venables, H; Waldron, HA. (1983) The acute behavioural effects of solvent exposure. J Soc 

Occup Med 33:13–18. 

Clewell, HJ, III; Gearhart, JM; Andersen, ME. (1993) Analysis of the metabolism of methylene chloride in 

the B6C3F1 mouse and its implications for human carcinogenic risk. Submitted to Occupational Safety and 

Health Administration, U.S. Department of Labor, Washington, DC. Docket # H-071, Exhibit #96. 

Clewell, HJ; Gentry, PR; Covington, TR; et al. (2004) Evaluation of the potential impact of age- and 

gender-specific pharmacokinetic differences on tissue dosimetry. Toxicol Sci 79:381–393. 

CMR (Chemical Marketing Reporter). (1973) Chemical profile on methylene chloride. Chemical 

Marketing Reporter, October 22, 1973. 
328

CMR (Chemical Marketing Reporter). (1979) Chemical profile on methylene chloride. Chemical 

Marketing Reporter, August 6, 1979. 

CMR (Chemical Marketing Reporter). (1982) Chemical profile on methylene chloride. Chemical 

Marketing Reporter, July 12, 1982. 

CMR (Chemical Marketing Reporter). (1997) Chemical profile on methylene chloride. Chemical 

Marketing Reporter, November 24, 1997. 

CMR (Chemical Marketing Reporter). (2000) Chemical profile on methylene chloride. Chemical 

Marketing Reporter, October 9, 2000. 

Cocco, P; Heineman, EF; Dosemeci, M. (1999) Occupational risk factors for cancer of the central nervous 

system (CNS) among US women. Am J Ind Med 36:70–74. 

Cohen J. (1987) Statistical power analysis for the behavioral sciences. Hillsdale NJ: Lawrence Erlbaum 

Associates; p. 36–37. 

Condie, LW; Smallwood, CL; Laurie, RD. (1983) Comparative renal and hepatotoxicity of halomethanes: 

bromodichloromethane, bromoform, chloroform, dibromochloromethane, and methylene chloride. Drug 

Chem Toxicol 6(6):563–578. 

Crebelli, R; Benigni, R; Franekic, J; et al. (1988) Induction of chromosome malsegregation by halogenated 

organic solvents in Aspergillus nidulans: unspecific or specific mechanism? Mutat Res 201:401–411. 

Crebelli, R; Carere, A; Leopardi, P; et al. (1999) Evaluation of 10 aliphatic halogenated hydrocarbons in 

the mouse bone marrow micronucleus test. Mutagenesis 14(2):207–215. 

Dankovic, DA; Bailer, AJ. (1994) The impact of exercise and intersubject variability on dose estimates for 

dichloromethane derived from a physiologically based pharmacokinetic model. Fundam Appl Toxicol 

22:20–25. 

David, RM; Clewell, HJ; Gentry, PR; et al. (2006) Revised assessment of cancer risk to dichloromethane II. 

Application of probabilistic methods to cancer risk determinations. Regul Toxicol Pharmacol 45:55–65.

De Salvia, MA; Cagiano, R; Carratù, MR; et al. (1995) Irreversible impairment of active avoidance 

behavior in rats prenatally exposed to mild concentrations of carbon monoxide. Psychopharmacology 

(Berl) 122(1):66–71. 

Dearfield, KL; Moore, MM. (2005) Use of genetic toxicology information for risk assessment. Environ 

Mol Mutagen 46(4):236–245. 

DeMarini, DM; Shelton, ML; Warren, SH; et al. (1997) Glutathione S-transferase-mediated induction of 

GC→AT transitions by halomethanes in Salmonella. Environ Mol Mutagen 30:440–447. 

Devereux, TR; Foley, JF; Maronpot, RR; et al. (1993) Ras proto-oncogene activation in liver and lung 

tumors from B6C3F1 mice exposed chronically to methylene chloride. Carcinogenesis 14(5):795–801. 

Dillon, D; Edwards, I; Combes, R; et al. (1992) The role of glutathione in the bacterial mutagenicity of 

vapour phase dichloromethane. Environ Mol Mutagen 20:211–217. 

Dinse GE; Lagakos SW. (1982) Nonparametric estimation of lifetime and disease onset distributions from

incomplete observations. Biometrics. 38(4):921–932. 
329

DiVincenzo, GD; Kaplan, CJ. (1981) Uptake, metabolism, and elimination of methylene chloride vapor by 

humans. Toxicol Appl Pharmacol 59:130–140. 

DiVincenzo, GD; Yanno, FJ; Astill, BD. (1971) The gas chromatographic analysis of methylene chloride in 

breath, blood, and urine. Am Ind Hyg Assoc J 32:387–391. 

DiVincenzo, GD; Yanno, FJ; Astill, BD. (1972) Human and canine exposures to methylene chloride vapor. 

Am Ind Hyg Assoc J 33:125–134. 

Doherty, AT; Ellard, S; Parry, EM; et al. (1996) An investigation into the activation and deactivation of 

chlorinated hydrocarbons to genotoxins in metabolically competent human cells. Mutagenesis 11(3):247–

274. 

Dosemeci, M; Cocco, P; Chow, WH. (1999) Gender differences in risk of renal cell carcinoma and 

occupational exposures to chlorinated aliphatic hydrocarbons. Am J Ind Med 36:54–59. 

Dumas, S; Parent, ME; Siemiatychi, J. (2000) Rectal cancer and occupational risk factors: a hypothesisgenerating, exposure-based case-control study. Int J Cancer 87:874–879. 

Eastmond, DA; Hartwig, A; Anderson , D. et al. (2009). Mutagenicity testing for chemical risk assessment: 

update of WHO/IPCS harmonized scheme. Mutagenesis 24:341-349. 

Ehrlich, R. (1980) Interaction between environmental pollutants and respiratory infections. Environ Health 

Perspect 35:89–100 

El-Masri, HA; Bell, DA; Portier, CJ. (1999) Effects of glutathione transferase theta polymorphism on the 

risk estimates of dichloromethane to humans. Toxicol Appl Pharmacol 158:221–230. 

El-Rayes, BF; Ali, S; Heilbrun, LK; et al. (2003). Cytochrome P450 and glutathione transferase expression 

in human breast cancer. Clin Cancer Res 9(5):1705–1709. 

Engström, J; Bjurström, R. (1977) Exposure to methylene chloride. Content in subcutaneous adipose tissue. 

Scand J Work Environ Health 3:215–224. 

Estill, CF; Spencer, AB. (1996) Case study: control of methylene chloride exposures during furniture 

stripping. Am Ind Hyg Assoc J 57:43–49. 

Fechner, G; Ortmann, C; Du Chesne, A; et al. (2001) Fatal intoxication due to excessive dichloromethane 

inhalation. Forensic Sci Int 122:69–72. 

Fechter, LD (1987) Neurotoxicity of prenatal carbon monoxide exposure. Res Rep Health Eff Inst. 12: 3-

22. 

Fodor, GG; Prajsnar, D; Schlipkoter, H-W. (1973) Endogenous CO formation by incorporated halogenated 

hydrocarbons of the methane series. Staub Reinhalt Luft 33:260–261. 

Foley, JF; Tuck, PD; Ton, T-VT; et al. (1993) Inhalation exposure to a hepatocarcinogenic concentration of 

methylene chloride does not induce sustained replicative DNA synthesis in hepatocytes of female B6C3F1

mice. Carcinogenesis 14(5)811–817. 

Forster, HV; Graff, S; Hake, CL; et al. (1974) Pulmonary-hematologic studies on humans during exposure

to methylene chloride. Prepared by the Department of Environmental Medicine, Medical College of 

Wisconsin, Milwaukee, WI, for the National Institute for Occupational Safety and Health, Centers for 

Disease Control and Prevention, U.S. Department of Health and Human Services, Cincinnati, OH. 

Available from the National Technical Information Service, Springfield, VA; PB82-151697. 
330

Foster, JR; Green, T; Smith, LL; et al. (1992) Methylene chloride-an inhalation study to investigate 

pathological and biochemical events occurring in the lungs of mice over an exposure period of 90 days. 

Fundam Appl Toxicol 18:376–388. 

Foster, JR; Green, T; Smith, LL; et al. (1994) Methylene chloride: an inhalation study to investigate 

toxicity in the mouse lung using morphological, biochemical, and Clara cell culture techniques. 

Toxicology 91:221–234. 

Friedlander, BR; Hearne, T; Hall, S. (1978) Epidemiologic investigation of employees chronically exposed 

to methylene chloride. J Occup Med 20:657–666. 

Fujimoto, K; Arakawa, S; Watanabe, T; et al. (2007) Generation and functional characterization of mice

with a disrupted glutathione S-transferase, theta 1 gene. Drug Metab Dispos. 35:2196-2202. 

Fuxe, K; Andersson, K; Hansson, T; et al. (1984) Central catecholamine neurons and exposure to 

dichloromethane. Selective changes in amine levels and turnover in tel- and diencephalic and Na nerve 

terminal systems and in the secretion of anterior pituitary hormone in the male rat. Toxicol 29:293–305. 

Gamberale, F; Annwall, G; Hultengren, M (1975) Exposure to methylene chloride. II. Psychological 

function. Scand J Work Environ Health 1:95–103. 

Gargas, ML; Clewell, HJ; Andersen, ME. (1986) Metabolism of inhaled dihalomethanes in vivo: 

differentiation of kinetic constants for two independent pathways. Toxicol Appl Pharmacol 82:211–223.

Garrett, NE; Lewtas, J. (1983) Cellular toxicity in Chinese hamster ovary cells culture. I. Analysis of 

cytotoxicity endpoints for twenty-nine priority pollutants. Environ Res 32:455–465. 

Garte, S; Gaspari, L; Alexandrie, AK; et al. (2001) Metabolic gene polymorphism frequencies in control

populations. Cancer Epidemiol Biomarkers Prev 10:1239–1248. 

General Electric Co. (1976) Dichloromethane and ninety day oral toxicity study in rats. Prepared by the 

International Research and Development Corporation, Mattawan, MI for the Plastics Tech Department, 

General Electric Co., Pittsfield, MA. Submitted under TSCA Section 8D; EPA Document No. 86-

878210707; NTIS No. OTS0205887. 

Gibbs, GW. (1992) The mortality of workers employed at a cellulose acetate and triacetate fibers plant in 

Cumberland, Maryland: a “1970” cohort followed 1970–1989 [final report]. Prepared by Safety Health 

Environment International Consultants Corporation, Winterburn, Alberta, Canada, for the Hoechst 

Celanese Corporation, Somerville, NJ. 

Gibbs, GW; Amsel, J; Soden, K. (1996) A cohort mortality study of cellulose triacetate-fiber workers 

exposed to methylene chloride. J Occup Environ Med 38(7):693–697. 

Giustino, A; Cagiano, R; Carratù, MR; et al. (1999) Prenatal exposure to low concentrations of carbon 

monoxide alters habituation and non-spatial working memory in rat offspring. Brain Res 844(1–2):201–

205. 

Glatzel, W; Tietze, K; Gutewort, R; et al. (1987) Interaction of dichloromethane and ethanol in rats: 

toxicokinetics and nerve conduction velocity. Alcoholism: Clin Exp Res 11:450–455. 

Gocke, E; King, MT; Eckhardt, K; et al. (1981) Mutagenicity of cosmetic ingredients licensed by the 

European communities. Mutat Res 90:91–109. 

Gomez, MR. (1996) Exposure determinants needed to improve the assessment of exposure. Am J Ind Med 

29:569–570. 
331

Gomez, MR; Cocco, P; Dosemeci, M; et al. (1994) Occupational exposure to chlorinated aliphatic 

hydrocarbons: job exposure matrix. Am J Ind Med 26:171–183. 

Goodman, DG; Maronpot RR; Newbene PM, Popp JA; Squire RA. (1994) Proliferative and selected other 

lesions in the liver of rats. G1-5. In: Guides for Toxologic Pathology. STP/ARP/AFIP, Washington DC:1-

24. 

Goulle, JP; Lacroix, C; Vaz, E; et al. (1999) Fatal case of dichloromethane poisoning. J Anal Toxicol

23:380–383. 

Graves, RJ; Green, T. (1996) Mouse liver glutathione S-transferase mediated metabolism of methylene 

chloride to a mutagen in the CHO/HPRT assay. Mutat Res 367:143–150. 

Graves, RJ; Callander, RD; Green, T. (1994a) The role of formaldehyde and S-chloromethylglutathione in

the bacterial mutagenicity of methylene chloride. Mutat Res 320:235–243. 

Graves, RJ; Coutts, C; Eyton-Jones, H; et al. (1994b) Relationship between hepatic DNA damage and 

methylene chloride-induced hepatocarcinogenicity in B6C3F1 mice. Carcinogenesis 15(5):991–996. 

Graves, RJ; Coutts, C; Green, T. (1995) Methylene chloride-induced DNA damage: an interspecies 

comparison. Carcinogenesis 16(8):1919–1926. 

Graves, RJ; Trueman, P; Jones, S; et al. (1996) DNA sequence analysis of methylene chloride-induced 

HPRT mutations in Chinese hamster ovary cells: comparison with the mutation spectrum obtained for 

1,2-dibromoethane and formaldehyde. Mutagenesis 11(3):229–233. 

Green, T. (1983) The metabolic activation of dichloromethane and chlorofluoromethane in a bacterial 

mutation assay using Salmonella typhimurium. Mutat Res 118:277–288. 

Green, T. (1989) A biological data base for methylene chloride risk assessment. In: Travis, CC; ed. 

Biologically based methods for cancer risk assessment. New York, NY: Plenum Press, p. 289–300. 

Green, T. (1997) Methylene chloride induced mouse liver and lung tumours: an overview of the role of 

mechanistic studies in human safety assessment. Hum Exp Toxicol 16:3–13. 

Guengerich, FP. (1997) Mechanisms of mutagenicity of DNA adducts derived from alkyl and vinyl halides. 

Jpn J Toxicol Environ Health 43:69–82. 

Haber, LT; Maier, A; Gentry, PR; et al. (2002) Genetic polymorphisms in assessing interindividual 

variability in delivered dose. Reg Toxicol Pharmacol 35:177–197. 

Hall, AH; Rumack, BH. (1990) Methylene chloride exposure in furniture-stripping shops: ventilation and

respirator use practices. J Occup Med 32:33–37. 

Hallier, E; Schröder, KR; Asmuth, K; et al. (1994) Metabolism of dichloromethane (methylene chloride) to 

formaldehyde in human erythrocytes: influence of polymorphism of glutathione transferase theta (GST T1-

1). Arch Toxicol 68:423–427. 

Hansch, C; Leo, A; Hoekman, D. (1995) Exploring QSAR. Hydrophobic, electronic, and steric constants. 

ACS professional reference book. Washington, DC: American Chemical Society; p. 3. 

Hardie, DWF. (1964) Methylene chloride. In: Kirk-Othmer encyclopedia of chemical technology. 2nd 

edition. New York, NY: John Wiley & Sons; pp. 111–119. 
332

Hardin, BD; Manson, J. (1980) Absence of dichloromethane teratogenicity with inhalation exposure in rats. 

Toxicol Appl Pharmacol 52:22–28. 

Hashmi, M; Dechert, S; Dekant, W; et al. (1994) Bioactivation of [13C]dichloromethane in mouse, rat, and 

human liver cytosol: 13C nuclear magnetic resonance spectroscopic studies. Chem Res Toxicol 7:291–296. 

Haufroid, V; Ligocka, D; Buysschaert, M; et al. (2003) Cytochrome P4502E1 (CYP2E1) expression in 

peripheral blood lymphocytes: evaluation in hepatitis C and diabetes. Eur J Clin Pharmacol 59:29–33. 

Haun, CC; Harris, ES; Darmer, KI, Jr. (1971) Continuous animal exposure to methylene chloride. In: 

Proceedings of the 2nd conference of environmental toxicology; August 31–September 2; Fairborn, OH; 

Paper No. 10; AMRL-TR-71-120. Aerospace Medical Research Laboratory, Wright-Patterson Air Force 

Base, OH; pp. 125–135. Available from the National Technical Information Service, Springfield, VA; 

AD751432 (individual paper); AD746660 (entire proceedings). 

Haun, CC; Vernot, EH; Darmer, KI, Jr; et al. (1972) Continuous animal exposure to low levels of 

dichloromethane. Proceedings of the 3rd conference of environmental toxicology; October 25–27; 

Fairborn, OH; Paper No. 12; AMRL-TR-72-120. Aerospace Medical Research Laboratory, WrightPatterson Air Force Base, OH; pp. 199–208. Available from the National Technical Information Service, 

Springfield, VA; AD773766. 

Hazelton Laboratories. (1983) 24-Month oncogenicity study of methylene chloride in mice [final report]. 

Performed by Hazelton Laboratories America, Inc., Vienna, VA for the National Coffee Association, New 

York, NY; Submitted under TSCA Section 4; EPA Document No. 45-8303005; NTIS No. OTS0206132. 

Hearne, FT; Pifer, JW. (1999) Mortality study of two overlapping cohorts of photographic film base 

manufacturing employees exposed to methylene chloride. J Occup Environ Med 41(12):1154–1169. 

Hearne, FT; Grose, F; Pifer, JW; et al. (1987) Methylene chloride mortality study: dose-response 

characterization and animal model comparison. J Occup Med 29(3):217–228. 

Hearne, FT; Pifer, JW; Grose, F. (1990) Absence of adverse mortality effects in workers exposed to 

methylene chloride: an update. J Occup Med 32(3):234–240. 

Hegi, ME; Söderkvist, P; Foley, JF; et al. (1993) Characterization of p53 mutations in methylene chlorideinduced lung tumors from B6C3F1 mice. Carcinogenesis 14(5):803–810. 

Heineman, EF; Cocco, P; Gomez, MR; et al. (1994) Occupational exposure to chlorinated aliphatic 

hydrocarbons and risk of astrocytic brain cancer. Am J Ind Med 26:155–169. 

Heineman, EF; Gomez, MR; Dosemeci, M; et al. (1996) Methylene chloride and brain cancer: interpreting 

a new study in light of existing literature. Am J Ind Med 30:506–507. 

Hellmold, H; Rylander, T; Magnusson, M; et al. (1998) Characterization of cytochrome P450 enzymes in 

human breast tissue from reduction mammaplasties. J Clin Endocrinol Metab 83(3):886–895. 

Heppel, LA; Neal, PA. (1944) Toxicology of dichloromethane (methylene chloride). II. Its effect upon 

running activity in the male rat. J Ind Hyg Toxicol 26(1):17–21. 

Heppel, LA; Neal, PA; Perrin, TL; et al. (1944) Toxicology of dichloromethane (methylene chloride). J

Ind Hyg Toxicol 26:8–16. 

Herr, DW; Boyes, WK. (1997) A comparison of the acute neuroactive effects of dichloromethane, 1,3-

dichloropropane, and 1,2-dichlorobenzene on rat flash evokes potentials (FEPs). Fundam Appl Toxicol 

35:31–48. 
333

Hines, RN. (2007) Ontogeny of human hepatic cytochromes P450. J Biochem Mol Toxicol 21:169–75. 

Holbrook, MT. (2003) Methylene chloride. In: Kirk-Othmer encyclopedia of chemical technology. New 

York, NY: John Wiley & Sons. Available online at 

http://www.mrw.interscience.wiley.com/kirk/kirk_search_fs.html. (subscription required). 

Horvath, AL. (1982) Halogenated hydrocarbons: solubility-miscibility with water. New York, NY: Marcel

Dekker, Inc., p. 543. 

Hu, Y; Kabbler, SL; Tennant, AH et al. (2006) Induction of DNA-protein crosslinks by dichloromethane in 

a V79 cell line transfected with the murine glutathione-S-transferase theta 1 gene. Mutation Res 607:231-

239. 

Hughes, NJ; Tracey, JA. (1993) A case of methylene chloride (nitromors) poisoning, effects on 

carboxyhaemoglobin levels. Hum Exp Toxicol 12:159–160. 

IARC (International Agency for Research on Cancer). (1999) Dichloromethane. IARC monographs on the 

evaluation of carcinogenic risk of chemicals to humans. Vol. 71. Re-evaluation of some organic chemicals, 

hydrazine and hydrogen peroxide. Lyon, France: International Agency for Research on Cancer, pp. 251–

315. Available online at http://inchem.org/documents/iarc/vol71/004-dichloromethane.html. 

Infante-Rivard, C; Siemiatycki, J; Lakhani, R; et al. (2005) Maternal exposure to occupational solvents and 

childhood leukemia. Environ Health Perspect 113:787–792. 

Ingelman-Sundberg, M. (2004) Human drug metabolising cytochrome P450 enzymes: properties and 

polymorphisms. Naunyn-Schmiedeberg’s Arch Pharmacol 369:89–104. 

Jacubovich, RM; Landau, D; Dayan, YB; et al. (2005) Facial nerve palsy after acute exposure to 

dichloromethane. Am J Ind Med 48:389–392. 

Johnsrud, EK; Koukouritaki, SB; Divakaran, K; et al. (2003) Human hepatic CYP2E1 expression during 

development. J Pharmacol Exp Ther 307:402–407. 

Jongen, WMF; Alink, GM; Koeman, JH. (1978) Mutagenic effect of dichloromethane on Salmonella 

typhimurium. Mutat Res 56:245–248. 

Jongen, WMF; Lohman, PHM; Kottenhagen, MJ; et al. (1981) Mutagenicity testing of dichloromethane in 

short-term mammalian test systems. Mutat Res 81:203–213. 

Jongen, WMF; Harmsen, EGM; Alink, GM; et al. (1982) The effect of glutathione conjugation and 

microsomal oxidation on the mutagenicity of dichloromethane in S. typhimurium. Mutat Res 95:183–189. 

Jonsson, F; Johanson, G. (2001) A Bayesian analysis of the influence of GST-T1 polymorphism on the 

cancer risk estimate for dichloromethane. Toxicol Appl Pharmacol 174:99–112. 

Jonsson, F; Johanson, G. (2003) The Bayesian population approach to physiological toxicokinetictoxicodynamic models―an example using the MCSim software. Toxicol Lett 138:143–150. 

Jonsson, F; Bois, F; Johanson, G. (2001) Physiologically based pharmacokinetic modeling of inhalation 

exposure of humans to dichloromethane during moderate to heavy exercise. Toxicol Sci 59:209–218. 

Juronen, E; Tasa, G; Uusküla, M;, Pooga, M; Mikelsaar, AV (1996). Purification, characterization and 

tissue distribution of human class theta glutathione S-transferase T1-1. Biochem Mol Biol Int. 39:21-9 
334

Kanada, M; Miyagawa, M; Sato, M; et al. (1994) Neurochemical profile of effects of 28 neurotoxic 

chemicals on the central nervous system in rats (1). Effects of oral administration on brain contents of 

biogenic amines and metabolites. Ind Health 32:145–164. 

Kanno, J; Foley, JF; Kari, F; et al. (1993) Effect of methylene chloride inhalation of replicative DNA 

synthesis in the lungs of female B6C3F1 mice. Environ Health Perspect 101(Suppl. 5):271–276. 

Kari, FW; Foley, JF; Seilkop, SK; et al. (1993) Effect of varying exposure regimens on methylene chlorideinduced lung and liver tumors in female B6C3F1 mice. Carcinogenesis 14(5):819–826. 

Karlsson, JE; Rosengren, LE; Kjellstrand, P; et al. (1987) Effects of low-dose inhalation of three 

chlorinated aliphatic organic solvents on deoxyribonucleic acid in gerbil brain. Scand J Work Environ

Health 13:453–458. 

Kayser, MF; Vuilleumier, S. (2001) Dehalogenation of Dichloromethane by Dichloromethane 

Dehalogenase/Glutathione S-Transferase Leads to Formation of DNA Adducts. J Bacteriology 183:5209-

5212. 

Kelly, M. (1988) Case reports of individuals with oligospermia and methylene chloride exposures. Reprod 

Toxicol 2:13–17. 

Kernan, GJ; Ji, BT; Dosemeci, M; et al. (1999) Occupational risk factors for pancreatic cancer: a casecontrol study based on death certificates from 24 US states. Am J Ind Med 36:260–270. 

Kim, SK; Kim YC. (1996) Effect of a single administration of benzene, toluene or m-xylene on 

carboxyhaemoglobin elevation and metabolism of dichloromethane in rats. J Appl Toxicol 16:437–444. 

Kim, RB; O’Shea, D; Wilkinson, GR. (1995) Interindividual variability of chlorzoxazone 6-hydroxylation

in men and women and its relationship to CYP2E1 genetic polymorphisms. Clin Pharmacol Ther 

57(6):645–655. 

Kim, NY; Park, SW; Suh, JK. (1996) Two fatal cases of dichloromethane or chloroform poisoning. J 

Forensic Sci 41:527–529. 

Kirschman, JC; Brown, NM; Coots, RH; et al. (1986) Review of investigations of dichloromethane 

metabolism and subchronic oral toxicity as the basis for the design of chronic oral studies in rats and mice. 

Food Chem Toxicol 24(9):943–949. 

Kitchin, KT; Brown, JL. (1989) Biochemical effects of three carcinogenic chlorinated methanes in rat liver. 

Teratog Carcinog Mutagen 9:61–69. 

Kjellstrand, P; Holmquist, B; Jonsson, I; et al. (1985) Effects of organic solvents on motor activity in mice. 

Toxicol 35:35–46. 

Kolodner, K; Cameron, L; Gittlesohn, A; et al. (1990) Morbidity study of occupational exposure to 

methylene chloride using a computerized surveillance system. Center of Occupational Health and 

Environmental Health, Johns Hopkins School of Hygiene and Public Health, Baltimore, MD. Submitted to 

TSCA under Section 8D; EPA Document No. 86-900000421; NTIS No. OTS0522984. 

Korzekwa, KR; Krishnamachary, N; Shou, M; et al. (1998) Evaluation of atypical cytochrome P450 

kinetics with two-substrate models: evidence that multiple substrates can simultaneously bind to 

cytochrome P450 active sites Biochemistry 37:4137-4147. 
335

Kramers, PGN; Mout, HCA; Bissumbhar, B; et al. (1991) Inhalation exposure in Drosophila mutagenesis 

assays: experiments with aliphatic halogenated hydrocarbons, with emphasis on the genetic activity profile 

of 1,2-dichloroethane. Mutat Res 252:17–33. 

Landi, S; Naccarati, A; Ross, MK; et al. (2003) Induction of DNA strand breaks by trihalomethanes in 

primary human lung epithelial cells. Mutat Res 538:41–50. 

Landry, TD; Ramsey, JC; McKenna, MJ. (1983) Pulmonary physiology and inhalation dosimetry in rats: 

development of a method and two examples. Toxicol Appl Pharmacol 71:72–83. 

Lanes, SF; Cohen, A; Rothman, KJ; et al. (1990) Mortality of cellulose fiber production workers. Scand J 

Work Environ Health 16:247–251. 

Lanes, SF; Rothman, KJ; Dreyer, NA; et al. (1993) Mortality update of cellulose fiber production workers. 

Scand J Work Environ Health 19:426–428. 

Lash, AA; Becker, CE; So, Y; et al. (1991) Neurotoxic effects of methylene chloride: are they long lasting 

in humans? Br J Ind Med 48:418–426. 

Lefevre, PA; Ashby, J. (1989) Evaluation of dichloromethane as an inducer of DNA synthesis in the 

B6C3F1 mouse liver. Carcinogenesis 10(6):1067–1072. 

Lehmann, L; Wagner, J. (2008) Gene expression of 17beta-estradiol-metabolizing isozymes: comparison of

normal human mammary gland to normal human liver and to cultured human breast adenocarcinoma cells. 

Adv Exp Med Biol 617:617–624. 

Lehnebach, A; Kuhn, C; Pankow, D. (1995) Dichloromethane as an indicator of cytochrome c oxidase in 

different tissues of rats. Arch Toxicol 69:180–184. 

Leighton, DT, Jr; Calo, JM. (1981) Distribution coefficients of chlorinated hydrocarbons in dilute air-water 

systems for groundwater contamination applications. J Chem Eng 26:382–385. 

Leikin, JB; Kaufman, D; Lipscomb, JW; et al. (1990) Methylene chloride: report of five exposures and two 

deaths. Am J Emerg Med 8:534–537. 

Leitao, MM, Jr; White, P; Cracchiolo, B. (2008). Cervical cancer in patients infected with the human 

immunodeficiency virus. Cancer 112:2683–2689. 

Leuschner, F; Neumann, B-W; Hübscher, F. (1984) Report on subacute toxicological studies with 

dichloromethane in rats and dogs by inhalation. Arzneim Forsch/Drug Res 34(12):1772–1774. 

Lewis, RJ, Sr; ed. (1997) Hawley's condensed chemical dictionary. 13th edition. New York, NY: John 

Wiley & Sons, Inc., p. 736. 

Lide, DR; ed. (2000) CRC handbook of chemistry and physics. 81st edition. Boca Raton, FL: CRC Press; 

pp. 3–206. 

Lipscomb, JC; Garrett, CM; Snawder, JE. (1997) Cytochrome P450-dependent metabolism of 

trichloroethylene: interindividual differences in humans. Toxicol Appl Pharmacol 142:311–318. 

Lipscomb, JC; Teuschler, LK; Swartout, J; et al. (2003) The impact of cytochrome P450 2E1-dependent 

metabolic variance on a risk-relevant pharmacokinetic outcome in humans. Risk Anal 23(6):1221–1238. 

Lowenfels, AB; Maisonneuve, P. (2005) Risk factors for pancreatic cancer. J Cell Biochem 95:649–656. 
336

Lucas, D; Ferrara, R; Gonzalez, E; et al. (1999) Chlorzoxazone, a selective probe for phenotyping CYP2E1 

in humans. Pharmacogenetics 9:377–388. 

Lucas, D; Ferrara, R; Gozales, E; et al. (2001) Cytochrome CYP2E1 phenotyping and genotyping in the 

evaluation of health risks from exposure to polluted environments. Toxicol Lett 124:71–81. 

Madle, S; Dean, SW; Andrae, U; et al. (1994) Recommendations for the performance of UDS tests in vitro 

and in vivo. Mutation Res 312:263-285. 

Mahmud, M; Kales, SN. (1999) Methylene chloride poisoning in a cabinet worker. Environ Health 

Perspect 107:769–772. 

Mainwaring, GW; Williams, SM; Foster, JR; et al. (1996) The distribution of Theta-class glutathione Stransferase in the liver and lung of mouse, rat and human. Biochem J 318:297–303. 

Mainwaring, GW; Foster, JR; Green, T. (1998) Nuclear and cellular immunolocalization of theta-class 

glutathione S-transferase GST-T1-1 in the liver and lung of the mouse. Biochem J 329:431–432. 

Maltoni, C; Cotti, G; Perino, G. (1988) Long-term carcinogenicity bioassays on methylene chloride 

administered by ingestion to Sprague-Dawley rats and Swiss mice and by inhalation to Sprague-Dawley 

rats. Ann NY Acad Sci 534:352–366. 

Manno, M; Rugge, M; Cocheo, V. (1992) Double fatal inhalation of dichloromethane. Hum Exp Toxicol 

11:540–545. 

Marino, DJ; Clewell, HJ; Gentry, PR; et al. (2006) Revised assessment of cancer risk to dichloromethane: 

part I Bayesian PBPK and dose-response modeling in mice. Regul Toxicol Pharmacol 45:44–54. 

Maronpot, RR; Devereux, TR; Hegi, M; et al. (1995) Hepatic and pulmonary carcinogenicity of methylene 

chloride in mice: a search for mechanisms. Toxicol 102:73–81. 

Marriott, HM; Dockrell, DH. (2007) The role of the macrophage in lung disease mediated by bacteria. Exp 

Lung Res 33:493–505. 

Marsch, GA; Mundkowski, R; Morris, BKJ; et al. (2001) Characterization of Nucleoside and DNA 

Adducts Formed by S-(1-Acetoxymethyl)glutathione and Implications for Dihalomethane-Glutathione 

Conjugates Chem Res Toxicol 14:600-608. 

Marsch, GA; Botta, S; Martin, MV; et al. (2004) Formation and mass spectrometric analysis of DNA and 

nucleoside adducts by S-(1-acetoxymethyl)glutathione and by glutathione S-transferase-mediated 

activation of dihalomethanes. Chem Res Toxicol 17:45–54. 

Mathews, JM; Raymer, JH; Etheridge, AS; Velez, GR; Bucher, JR (1997) Do endogenous volatile organic 

chemicals measured in breath reflect and maintain CYP2E1 levels in vivo? Toxicol Appl Pharmacol 

146:255-260. 

Mattsson, JL; Albee, RR; Eisenbrandt, DL. (1990) Neurotoxicologic evaluation of rats after 13 weeks of

inhalation exposure to dichloromethane or carbon monoxide. Pharmacol Biochem Behav 36:671–681. 

McKenna, MJ; Zempel, JA. (1981) The dose-dependent metabolism of [14C]methylene chloride following 

oral administration to rats. Food Cosmet Toxicol 19:73–78. 
337

McKenna, MJ; Zempel, JA; Braun, WH. (1982) The pharmacokinetics of inhaled methylene chloride in 

rats. Toxicol Appl Pharmacol 65:1–10. 

Mendoza-Cantú, A; Castorena-Torres, F; et al. (2004) Genotype and allele frequencies of polymorphic 

cytochromes P450 CYP1A2 and CYP2E1 in Mexicans. Cell Biochem Funct 22:29–34. 

Mennear, JH; McConnell, EE; Huff, JE; et al. (1988) Inhalation and carcinogenesis studies of methylene

chloride (dichloromethane) in F344/n rats and B6C3F1 mice. Ann NY Acad Sci 534:343–351. 

Meyer, DJ; Coles, B; Pemble, SE; et al. (1991) Theta, a new class of glutathione transferases purified from 

rat and man. Biochem J 274:409–414. 

Miksys, S; Tyndale, RF. (2004) The unique regulation of brain cytochrome P450 2 (CYP2) family enzymes 

by drugs and genetics. Drug Metab Rev 36(2):313–333. 

Mirsalis, JC; Tyson, CK; Steinmetz, KL; et al. (1989) Measurement of unscheduled DNA synthesis and sphase synthesis in rodent hepatocytes following in vivo treatment: testing of 24 compounds. Environ Mol 

Mutagen 14:155–164. 

Moser, VC; Cheek, BM; MacPhail, RC. (1995) A multidisciplinary approach to toxicological screening: 

III. Neurobehavioral toxicity. J Toxicol Environ Health 45:173–210. 

Narotsky, MG; Kavlock, RJ. (1995) A multidisciplinary approach to toxicological screening: II. 

Developmental toxicity. J Toxicol Environ Health 45:145–171. 

Nelson, HH; Wiencke, JK; Christiani, DC; et al. (1995) Ethnic differences in the prevalence of the 

homozygous deleted genotype of glutathione S-transferase theta. Carcinogenesis 16(5):1243–1245. 

NIOSH (National Institute of Occupational Safety and Health). (1986) Methylene chloride. Current 

intelligence bulletin 46. National Institute of Occupational Safety and Health, Centers for Disease Control 

and Prevention, U.S. Department of Health and Human Services, Cincinnati, OH; DHHS (NIOSH) 

Publication No. 86-114. Available online at: http://www.cdc.gov/niosh/86114_46.html (accessed June 29,

2006). 

Nishimura, M; Yaguti, H; Yoshitsugu, H. (2003) Tissue distribution of mRNA expression of human 

cytochrome P450 isoforms assessed by high-sensitivity real-time reverse transcription PCR. Yakugaku 

Zasshi 123(5):369–375. 

Nitschke, KD; Burek, JD; Bell, TJ; et al. (1988a) Methylene chloride: a 2-year inhalation toxicity and

oncogenicity study in rats. Fundam Appl Toxicol 11:48–59. 

Nitschke, KD; Eisenbrandt, DL; Lomax, LG; et al. (1988b) Methylene chloride: two-generation inhalation

reproductive study in rats. Fundam Appl Toxicol 11:60–67. 

NLM (National Library of Medicine). (2003) Dichloromethane. HSDB (Hazardous Substances Data 

Bank). National Library of Medicine, National Institutes of Health, U.S. Department of Health and Human 

Services, Bethesda, Maryland. Available online at http://toxnet.nlm.nih.gov/. 

Norman, WC, III; Boggs, P. (1996) Flawed estimates of methylene chloride exposures. Am J Ind Med 

30:504–509. 

NRC (National Research Council). (1983) Risk assessment in the federal government: managing the 

process. Washington, DC: National Academy Press. 
338

NRC (National Research Council). (1994) Science and judgment in risk assessment. Washington, DC: 

National Academy Press. 

NTP (National Toxicology Program). (1986) Toxicology and carcinogenesis studies of dichloromethane 

(methylene chloride) (CAS No. 75-09-2) in F344/N rats and B6C3F1 mice (inhalation studies). Public 

Health Service, U.S. Department of Health and Human Services; NTP TR 306. Available from the National 

Institute of Environmental Health Sciences, Research Triangle Park, NC. Available online at 

http://ntp.niehs.nih.gov/ntp/htdocs/LT_rpts/tr306.pdf 

Oda, Y; Yamazaki, H; Thier, R; et al. (1996) A new Salmonella typhimurium NM5004 strain expressing rat 

glutathione S-transferase 5-5: use in detection of genotoxicity of dihaloalkanes using an SOS/umu test 

system. Carcinogenesis 17:297–302. 

Ognenovski, VM; Marder, W; Somers EC; et al. (2004) Increased incidence of cervical intraepithelial 

neoplasia in women with systemic lupus erythematosus treated with intravenous cyclophosphamide. J 

Rheumatol 31:1763–1767. 

Oh, SJ; Kim, SK; Kim, YC. (2002) Role of glutathione in metabolic degradation of dichloromethane in 

rats. Toxicol Lett 129:107–114. 

Ojajärvi, A; Partanen, T; Ahlbom, A; et al. (2001) Risk of pancreatic cancer in workers exposed to 

chlorinated hydrocarbon solvents and related compounds: a meta-analysis. Am J Epidemiol 153:841–850. 

O’Neil, MJ; Smith, A; Heckelman, PE; et al. (2001) Methylene chloride. The Merck index: an 

encyclopedia of chemicals, drugs, and biologicals. Whitehouse Station, NJ: Merck & Co., Inc.; p. 1082.

Osterman-Golkar, S; Hussain, S; Walles, S; et al. (1983) Chemical reactivity and mutagenicity of some 

dihalomethanes. Chem Biol Interact 46:121–130. 

OSHA (Occupational Safety and Health Administration). (1997) Occupational exposure to methylene 

chloride. Federal Register 63(7):1383–1658. 

Ott, MG; Skory, LK; Holder, BB; et al. (1983a) Health evaluation of employees occupationally exposed to 

7methylene chloride: general study design and environmental considerations. Scand J Work Environ 

Health 9(Suppl. 1):1–7. 

Ott, MG; Skory, LK; Holder, BB; et al. (1983b) Health evaluation of employees occupationally exposed to 

methylene chloride: mortality. Scand J Work Environ Health 9(Suppl. 1):8–16. 

Ott, MG; Skory, LK; Holder, BB; et al. (1983c) Health evaluation of employees occupationally exposed to 

methylene chloride: clinical laboratory evaluation. Scand J Work Environ Health 9(Suppl. 1):17–25. 

Ott, MG; Skory, LK; Holder, BB; et al. (1983d) Health evaluation of employees occupationally exposed to 

methylene chloride: twenty-four hour electrocardiographic monitoring. Scand J Work Environ Health 

9(Suppl. 1):26–30. 

Ott, MG; Skory, LK; Holder, BB; et al. (1983e) Health evaluation of employees occupationally exposed to 

methylene chloride: metabolism data and oxygen half-saturation pressures. Scand J Work Environ Health

9(Suppl. 1):31–38. 

Ott, MG; Carlo, GL; Steinberg, S; et al. (1985) Mortality among employees engaged in chemical 

manufacturing and related activities. Am J Epidemiol 122:311–322. 

Pankow, D. (1988) Enhancement of dichloromethane-induced carboxyhemoglobinemia by isoniazid 

pretreatment. Biomed Biochem Act 3:293–295. 
339

Pankow, D; Hoffmann, P. (1989) Dichloromethane metabolism to carbon monoxide can be induced by 

isoniazid, acetone and fasting. Arch Toxicol Suppl 13:302–303. 

Pankow, D; Jagielki, S. (1993) Effect of methanol or modifications of the hepatic glutathione concentration 

on the metabolism of dichloromethane to carbon monoxide in rats. Hum Exp Toxicol 12:227–231. 

Pankow, D; Kretschmer, S; Weise, M. (1991a) Effect of pyrazole on dichloromethane metabolism to 

carbon monoxide. Arch Toxicol Suppl 14:246–248. 

Pankow, D; Matschiner, F; Weigmann, H-J. (1991b) Influence of aromatic hydrocarbons on the 

metabolism of dichloromethane to carbon monoxide in rats. Toxicol 68:89–100. 

Pegram, RA; Andersen, ME; Warren, SH; et al. (1997) Glutathione S-transferase-mediated mutagenicity of

trihalomethanes in Salmonella typhimurium: contrasting results with bromodichloromethane and

chloroform. Toxicol Appl Pharmacol 144:183–188. 

Perocco, P; Prodi, G. (1981) DNA damage by haloalkanes in human lymphocytes cultured in vitro. Cancer

Lett 13:213–218. 

Putz, VR; Johnson BL; Setzer, JV. (1979) A comparative study of the effects of carbon monoxide and 

methylene chloride on human performance. J Environ Pathol Toxicol 2:97–112. 

Quondamatteo, F; Schulz, TG; Bunzel, N; et al. (1998) Immunohistochemical localization of glutathione 

S-transferase-T1 in murine kidney, liver, and lung. Histochem Cell Biol 110:417–423. 

Raimondi, S; Paracchini, V; Autrup, H; et al. (2006) Human genome epidemiology (HuGE) review. Metaand pooled analysis of GST-T1 and lung cancer: a HuGE-GSEC review. Am J Epidemiol 164(11):1027–

1042. 

Raje, R; Basso, M; Tolen, T; et al. (1988) Evaluation of in vivo mutagenicity of low-dose methylene 

chloride in mice. J Am Coll Toxicol 7(5):699–703. 

Ramsey, JR; Andersen, ME. (1984) A physiologically-based description of the inhalation pharmacokinetics 

of styrene in rats and humans. Toxicol Appl Pharmacol 73:159–175. 

Raphael, M; Nadiras, P; Flacke-Vordos, N. (2002) Acute methylene chloride intoxication―a case report on 

domestic poisoning. Eur J Emerg Med 9:57–59. 

Rebert, CS; Matteucci, MJ; Pryor, GT. (1989) Acute effects of inhaled dichloromethane on the EEG and 

sensory-evoked potentials of Fischer-344 rats. Pharmacol Biochem Behav 34:619–629. 

Reitz, RH. (1991) Estimating the risk of human cancer associated with exposure to methylene chloride. 

Ann Ist Super Sanita 27(4):609–614. 

Reitz, RH; McDougal, JN; Himmelstein, MW; et al. (1988a) In vitro metabolism of methylene chloride in 

human and animal tissues: use in physiologically based pharmacokinetic models. Toxicol Appl Pharmacol

97:230–246. 

Reitz, RH; Mendrala, AL; Park, CN; et al. (1988b) Incorporation of in vitro enzyme data into the 

physiologically-based pharmacokinetic (PB-PK) model for methylene chloride: implications for risk 

assessment. Toxicol Lett 43:97–116. 

Reitz, RH; Mendrala, AL; Guengerich, FP. (1989) In vitro metabolism of methylene chloride in human and

animal tissues: use in physiologically based pharmacokinetic models. Toxicol Appl Pharmacol 97:230–

246. 
340

Reitz, RH; Hays, SM; Gargas, ML. (1997) Addressing priority data needs for methylene chloride with 

physiologically based pharmacokinetic modeling. Prepared for the Agency for Toxic Substances and 

Disease Registry, Public Health Service, U.S. Department of Health and Human Services, Atlanta, GA, on

behalf of the Halogenated Solvents Industry Alliance (HSIA), Arlington, VA. 

Rhomberg, L. (1995) Use of quantitative modeling in methylene chloride risk assessment. Toxicology 

102:95–114. 

Riley, EC; Fassett, DW; Sutton, WL. (1966) Methylene chloride vapor in expired air of human subjects. 

Am Ind Hyg Assoc J 27:341–348. 

Rioux, JP; Myers, RA. (1988) Methylene chloride poisoning: a paradigmatic review. J Emerg Med 6:227–

238. 

Rodkey, FL; Collison, HA. (1977) Effect of dihalogenated methanes on the in vivo production of carbon 

monoxide and methane by rats. Toxicol Appl Pharmacol 40:39–47. 

Rodriguez-Arnaiz, R. (1998) Biotransformation of several structurally related 2B compounds to reactive

metabolites in the somatic w/w+ assay of Drosophila melanogaster. Environ Mol Mutagen 31(4):390–401. 

Roldán-Arjona, T; Pueyo, C. (1993) Mutagenic and lethal effects of halogenated methanes in the Ara test 

of Salmonella typhimurium: quantitative relationship with chemical reactivity. Mutagenesis 8(2):127–131. 

Rosengren, LE; Kjellstrand, P; Aurell, A; et al. (1986) Irreversible effects of dichloromethane on the brain 

after long term exposure: a quantitative study of DNA and the glial cell marker proteins S-100 and GFA. 

Br J Ind Med 43:291–299. 

Roth, RP; Drew, RT; Lo, RJ; et al. (1975) Dichloromethane inhalation, carboxyhemoglobin concentrations, 

and drug metabolizing enzymes in rabbits. Toxicol Appl Pharmacol 33:427–437. 

Rothman, KJ; Greenland, S. (1998) Precision and validity in epidemiologic studies. In: Modern 

epidemiology. 2nd edition. Philadelphia, PA: Lippincott-Raven Publishers; pp. 115–134. 

Sakai, T; Morita, Y; Wakui, C. (2002) Biological monitoring of workers exposed to dichloromethane using 

head-space gas chromatography. J Chromat B 778:245–250. 

Sasaki, YF; Saga, A; Akasaka, M; et al. (1998) Detection of in vivo genotoxicity of haloalkanes and 

haloalkenes carcinogenic to rodents by the alkaline single cell gel electrophoresis (comet) assay in multiple 

mouse organs. Mutat Res 419:13–20. 

Savolainen, H; Pfäffli, P; Tengén M; et al. (1977) Biochemical and behavioural effects of inhalation 

exposure to tetrachloroethylene and dichloromethane. J Neuropathol Exp Neurol 36:941–949. 

Savolainen, H; Kurppa, K; Pfäffli, P; et al. (1981) Dose-related effects of dichloromethane in rat brain in 

short-term inhalations exposure. Chem-Biol Inter 34:315–322. 

Schwetz, BA; Leong, BKJ; Gehring, PJ. (1975) The effect of maternally inhaled trichloroethylene, 

perchloroethylene, methyl chloroform, and methylene chloride on embryonal and fetal development in 

mice and rats. Toxicol Appl Pharmacol 32:84–96. 

Searles, J; McPhail, HA. (1949) Methylene chloride, CH2Cl2. In: Kirk, RE; Othmer, DF; eds. 

Encyclopedia of chemical technology. New York, NY: Interscience Encyclopedia, Inc.; pp. 747–751. 

Selgrade, MK; Gilmour, MI. (2006) Immunotoxicology of inhaled compounds–assessing risks of local 

immune suppression and hypersensitivity. J Toxicol Environ Health A 69:827–844. 
341

Serota, DG; Thakur, AK; Ulland, BM; et al. (1986a) A two-year drinking water study of dichloromethane 

in rodents. I. Rats. Food Chem Toxicol 24(9):951–958. 

Serota, DG; Thakur, AK; Ulland, BM; et al. (1986b) A two-year drinking water study of dichloromethane 

in rodents. II. Mice. Food Chem Toxicol 24(9):959–963. 

Sheldon, T; Richardson, CR; Elliott, BM. (1987) Inactivity of methylene chloride in the mouse bone 

marrow micronucleus assay. Mutagenesis 2(1):57–59.

Sherratt, PJ; Pulford, DJ; Harrison, DJ; et al. (1997) Evidence that human class Theta glutathione Stransferase T1-1 can catalyse the activation of dichloromethane, a liver and lung carcinogen in the mouse. 

Biochem J 326:837–846. 

Sherratt, PJ; Williams, S; Foster, J; et al. (2002) Direct comparison of the nature of mouse and human GST 

T1-1 and the implications on dichloromethane carcinogenicity. Toxicol Appl Pharmacol 179:89–97. 

Shimada, T; Yamazaki, H; Mimura, M; et al. (1994) Interindividual variations in human liver cytochrome

P-450 enzymes involved in the oxidation of drugs, carcinogens and toxic chemicals: studies with liver 

microsomes of 30 Japanese and 30 Caucasians. J Pharmacol Exp Ther 270(1):414–423. 

Shusterman, D; Quinlan, P; Lowengart, R; et al. (1990) Methylene chloride intoxication in a furniture 

refinisher. A comparison of exposure estimates utilizing workplace air sampling and blood 

carboxyhemoglobin measurements. J Occup Med 32:451–454. 

Sills, RC; Hailey, JR; Neal, J; et al. (1999) Examination of low-incidence brain tumor responses in F344 

rats following chemical exposures in National Toxicology Program carcinogenicity studies. Toxicol Pathol 

27(5):589–599. 

Simula, TP; Glancey, MJ; Wolf, CR. (1993) Human glutathione S-transferase-expressing Salmonella 

typhimurium tester strains to study the activation/detoxification of mutagenic compounds: studies with 

halogenated compounds, aromatic amines and aflatoxin B1. Carcinogenesis 14(7):1371–1376. 

Slikker, W, Jr; Andersen, ME; Bogdanffy, MS; et al. (2004) Dose-dependent transition in mechanisms of 

toxicity: case studies. Toxicol Appl Pharmacol 201:226–294. 

Soden, KJ. (1993) An evaluation of chronic methylene chloride exposure. J Occup Med 35(3):282–286. 

Soden, KJ; Marras, G; Amsel, J. (1996) Carboxyhemoglobin levels in methylene chloride-exposed 

employees. J Occup Environ Med 38(4):367–371. 

Spirtas, R; Stewart, PA; Lee, JS; et al. (1991) Retrospective cohort mortality study of workers at an aircraft 

maintenance facility. I. Epidemiological results. Br J Ind Med 48:515–530. 

SRC (Syracuse Research Corporation). (1989) A review of in vitro test methodology for assessment of 

hepatotoxicity with a view to application to chemical mixtures. Prepared for U.S. EPA, Environmental 

Criteria and Assessment Office, U.S. Environmental Protection Agency; SRC TR-89-205. 

Stephens, EA; Taylor, JA; Kaplan, N; et al. (1994) Ethnic variation in the CYP2E1 gene: polymorphism 

analysis of 695 African-Americans, European-Americans and Taiwanese. Pharmacogenetics 4:185–192. 

Stewart, RD; Fisher, TN; Hosko, MJ; et al. (1972a) Experimental human exposure to methylene chloride. 

Arch Environ Health 25:342–348. 

Stewart, RD; Fisher, TN; Hosko, MJ; et al. (1972b) Carboxyhemoglobin elevation after exposure to 

dichloromethane. Science 176:295–296. 
342

Stewart, PA; Lee, JS; Marano, DE; et al. (1991) Retrospective cohort mortality study of workers at an 

aircraft maintenance facility. II Exposures and their assessment. Br J Ind Med 48:531–537. 

Stott, WT; McKenna, MJ. (1984) The comparative absorption and excretion of chemical vapors by the 

upper, lower, and intact respiratory tract of rats. Fundam Appl Toxicol 4:594–602. 

Sweeney, LM; Gargas, ML; Strother, DE; Kedderis, GL. (2003) Physiologically based pharmacokinetic 

model parameter estimation and sensitivity and variability analyses for acrylonitrile disposition in humans. 

Toxicol Sci. 71:27-40. 

Sweeney, LM; Kirman, CR; Morgott, DA; et al. (2004) Estimation of interindividual variation in oxidative 

metabolism of dichloromethane in human volunteers. Toxicol Lett 154:201–216. 

Takeshita, H; Mogi, K; Yasuda, T; et al. (2000) Postmortem absorption of dichloromethane: a case study

and animal experiments. Int J Legal Med 114:96–100. 

Taskinen, H; Lindbohm, ML; Hemminki, K. (1986) Spontaneous abortions among women working in the 

pharmaceutical industry. Br J Ind Med 43:199–205. 

Tay, P; Tan, KT; Sam, CT. (1995) Fatal gassing due to methylene chloride―a case report. Singapore Med 

J 36:444–445. 

Teschke, K; Olshan, AF; Daniels, JL; et al. (2002) Occupational exposure assessment in case-control 

studies: opportunities for improvement. Occup Environ Med 59:575–593. 

Thier, R; Foest, U; Deutschmann, S; et al. (1991) Distribution of methylene chloride in human blood. Arch 

Toxicol Suppl 14:254–258. 

Thier, R; Taylor, JB; Pemble, SE; et al. (1993) Expression of mammalian glutathione S-transferase 5-5 in 

Salmonella typhimurium TA1535 leads to base-pair mutations upon exposure to dihalomethanes. Proc Natl 

Acad Sci USA 90:8576–8580. 

Thier, R; Wiebel, FA; Hinkel, A; et al. (1998) Species differences in the glutathione transferase GST-T1-1 

activity towards the model substrates methyl chloride and dichloromethane in the liver and kidney. Arch 

Toxicol 72:622–629. 

Thilagar, AK; Kumaroo, V. (1983) Induction of chromosome damage by methylene chloride in CHO cells. 

Mutat Res 116:361–367. 

Thilagar, AK; Back, AM; Kirby, PE; et al. (1984) Evaluation of dichloromethane in short term in vitro 

genetic toxicity assays. Environ Mutagen 6:418–419. 

Thomas, AA; Pinkerton, MK; Warden, JA. (1972) Effects of low level dichloromethane exposure on the 

spontaneous activity of mice. Proceedings of the 3rd conference of environmental toxicology; October 25–

27; Fairborn, OH; Paper No. 14; AMRL-TR-72-130. Aerospace Medical Research Laboratory, WrightPatterson Air Force Base, OH; pp. 223–238. Available from the National Technical Information Service, 

Springfield, VA; AD773766. 

Tomenson, JA; Bonner, SM; Heijne, CG; et al. (1997) Mortality of workers exposed to methylene chloride

employed at a plant producing cellulose triacetate film base. Occup Environ Med 54:470–476. 

Treluyer, JM; Cheron, G; Sonnier M; et al. (1996) Cytochrome P-450 expression in sudden infant death 

syndrome. Biochem Pharmacol 52:497–504. 
343

Trueman, RW; Ashby, J. (1987) Lack of UDS activity in the livers of mice and rats exposed to 

dichloromethane. Environ Mol Mutagen 10:189–195. 

U.S. Coast Guard. (1999) Dichloromethane. CHRIS: hazardous chemical data. U.S. Coast Guard, 

Department of Transportation, Washington, DC. Available online at 

http://www.chrismanual.com/findform.htm. 

U.S. EPA (Environmental Protection Agency). (1986a) Guidelines for the health risk assessment of 

chemical mixtures. Federal Register 51(185):34014–34025. Available online at 

http://www.epa.gov/ncea/raf/rafguid.htm. 

U.S. EPA (Environmental Protection Agency). (1986b) Guidelines for mutagenicity risk assessment. 

Federal Register 51(185):34006–34012. Available online at http://www.epa.gov/ncea/raf/rafguid.htm. 

U.S. EPA (Environmental Protection Agency). (1987a) Update to the health assessment document and 

addendum for dichloromethane (methylene chloride): pharmacokinetics, mechanism of action and 

epidemiology [review draft]. Office of Health and Environmental Assessment, Office of Research and 

Development, Washington, DC; EPA/600/8-87/030A. Available from the National Technical Information 

Service, Springfield, VA, PB87228565, and online at

http://nepis.epa.gov/Exe/ZyPURL.cgi?Dockey=30001GFH.txt. 

U.S. EPA (Environmental Protection Agency). (1987b) Technical analysis of new methods and data 

regarding dichloromethane hazard assessments [review draft]. Office of Health and Environmental 

Assessment, Office of Research and Development, Washington, DC; EPA/600/8-87/029A. Available from 

the National Technical Information Service, Springfield, VA, PB87-228557, and online at 

http://nepis.epa.gov/Exe/ZyPURL.cgi?Dockey=30001GAN.txt. 

U.S. EPA (Environmental Protection Agency). (1988a) Recommendations for and documentation of 

biological values for use in risk assessment. Environmental Criteria and Assessment Office, Office of

Health and Environmental Assessment, Cincinnati, OH; EPA/600/6-87/008. Available from the National 

Technical Information Service, Springfield, VA, PB88-179874/AS, and online at 

http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=34855. 

U.S. EPA (Environmental Protection Agency). (1988b) The impact of pharmacokinetics on the risk 

assessment of dichloromethane. Office of Health and Environmental Assessment, Office of Research and 

Development, Washington, DC; EPA/600/D-88/219. Available from National Technical Information 

Service, Springfield, VA; PB89-173249. 

U.S. EPA (Environmental Protection Agency). (1991) Guidelines for developmental toxicity risk 

assessment. Federal Register 56(234):63798–63826. Available online at 

http://www.epa.gov/ncea/raf/rafguid.htm. 

U.S. EPA (Environmental Protection Agency). (1994a) Interim policy for particle size and limit 

concentration issues in inhalation toxicity: notice of availability. Federal Register 59(206):53799. 

Available online at http://www.epa.gov/EPA-PEST/1994/October/Day-26/pr-11.html. 

U.S. EPA (Environmental Protection Agency). (1994b) Methods for derivation of inhalation reference 

concentrations and application of inhalation dosimetry. Environmental Criteria and Assessment Office,

Office of Health and Environmental Assessment, Cincinnati, OH; EPA/600/8-90/066F. Available from the 

National Technical Information Service, Springfield, VA, PB2000-500023, and online at 

http://cfpub.epa.gov/ncea/raf/recordisplay.cfm?deid=71993. 

U.S. EPA (Environmental Protection Agency). (1995) Use of the benchmark dose approach in health risk 

assessment. Risk Assessment Forum, Washington, DC; EPA/630/R-94/007. Available from the National 
344

Technical Information Service, Springfield, VA, PB95-213765, and online at 

http://cfpub.epa.gov/ncea/raf/raf_pubtitles.cfm?detype=document&excCol=archive. 

U.S. EPA (Environmental Protection Agency). (1996) Guidelines for reproductive toxicity risk assessment. 

Federal Register 61(212):56274–56322. Available online at http://www.epa.gov/ncea/raf/rafguid.htm. 

U.S. EPA (Environmental Protection Agency). (1998) Guidelines for neurotoxicity risk assessment. 

Federal Register 63(93):26926–26954. Available online at http://www.epa.gov/ncea/raf/rafguid.htm. 

U.S. EPA (Environmental Protection Agency). (2000a) Science policy council handbook: risk 

characterization. Office of Science Policy, Office of Research and Development, Washington, DC; 

EPA/100-B-00-002. Available online at http://www.epa.gov/OSA/spc/pdfs/prhandbk.pdf. 

U.S. EPA (Environmental Protection Agency). (2000b) Benchmark dose technical guidance document 

[external review draft]. Risk Assessment Forum, Washington, DC; EPA/630/R-00/001. Available online at

http://cfpub.epa.gov/ncea/cfm/ 

nceapublication.cfm?ActType=PublicationTopics&detype=DOCUMENT&subject=BENCHMARK+DOS

E&subjtype=TITLE&excCol=Archive. 

U.S. EPA (Environmental Protection Agency). (2000c) Supplementary guidance for conducting health risk 

assessment of chemical mixtures. Risk Assessment Forum, Washington, DC; EPA/630/R-00/002. 

Available online at http://cfpub.epa.gov/ncea/raf/chem_mix.cfm. 

U.S. EPA (Environmental Protection Agency). (2000d) Toxicological review of vinyl chloride. Integrated 

Risk Information System (IRIS), National Center for Environmental Assessment, Washington, DC; 

EPA/635/R-00/004. Available online at http://www.epa.gov/iris. 

U.S. EPA (Environmental Protection Agency). (2002) A review of the reference dose concentration and 

reference concentration processes. Risk Assessment Forum, Washington, DC; EPA/630/P-02/002F. 

Available online at http://cfpub.epa.gov/ncea/raf/raf_pubtitles.cfm?detype=document&excCol=archive. 

U.S. EPA (Environmental Protection Agency). (2005a) Guidelines for carcinogen risk assessment. Federal 

Register 70(66):17765–18717. Available online at http://www.epa.gov/cancerguidelines. 

U.S. EPA (Environmental Protection Agency). (2005b) Supplemental guidance for assessing susceptibility

from early-life exposure to carcinogens. Risk Assessment Forum, Washington, DC; EPA/630/R-03/003F. 

Available online at http://www.epa.gov/cancerguidelines. 

U.S. EPA (Environmental Protection Agency). (2006a) Science policy council handbook: peer review. 3rd

edition. Office of Science Policy, Office of Research and Development, Washington, DC; EPA/100/B06/002. Available online at http://www.epa.gov/OSA/spc/2peerrev.htm. 

U.S. EPA (Environmental Protection Agency). (2006b) A framework for assessing health risk of 

environmental exposures to children. National Center for Environmental Assessment, Washington, DC; 

EPA/600/R-05/093F. Available online at http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=158363. 

Vieira, I; Sonnier, M; Cresteil, T. (1996) Developmental expression of CYP2E1 in human liver. 

Hypermethylation control of gene expression during the neonatal period. Eur J Biochem 238:476–483. 

Warbrick, EV; Kilgour, JD; Dearman, RJ; et al. (2003) Inhalation exposure to methylene chloride does not 

induce systemic immunotoxicity in rats. J Toxicol Environ Health A 66:1207–1219. 

Warholm, M; Alexandrie, AK; Högberg, J; et al. (1994) Polymorphic distribution of glutathione transferase 

activity with methyl chloride in human blood. Pharmacogenetics 4:307–311. 
345

Watanabe K, Liberman RG, Skipper PL, Tannenbaum SR, Guengerich FP. (2007). Analysis of DNA 

adducts formed in vivo in rats and mice from 1,2-dibromoethane, 1,2-dichloroethane, dibromomethane, and 

dichloromethane using HPLC/accelerator mass spectrometry and relevance to risk estimates. Chem Res 

Toxicol. 20:1594-600. 

Weinstein, RS; Boyd, DD; Back, KC. (1972) Effects of continuous inhalation of dichloromethane in the 

mouse: morphologic and functional observations. Toxicol Appl Pharmacol 23:660–679. 

Wells, VE; Schrader, SM; McCamon, CS; et al. (1989) Letter to the editor. Reprod Toxicol 3:281–282. 

Westbrook-Collins, B; Allen, JW; Sharief, Y; et al. (1990) Further evidence that dichloromethane does not 

induce chromosome damage. J Appl Toxicol 10(2):79–81. 

Winneke, G. (1974) Behavioral effects of methylene chloride and carbon monoxide as assessed by sensory

and psychomotor performance. In: Xintaras, C; Johnson, BL; de Groot, I; eds. Behavioral toxicology: early 

detection of occupational hazards. Washington, DC: National Institute for Occupational Safety and Health, 

Center for Disease Control, Public Health Service, U.S. Department of Health, Education and Welfare; pp. 

130–144. 

Wirkner, K; Damme, B; Peolchen, W; et al. (1997) Effect of long-term ethanol pretreatment on the 

metabolism of dichloromethane to carbon monoxide in rats. Toxicol Appl Pharmacol 143:83–88. 

Withey, JR; Karpinski, K. (1985) The fetal distribution of some aliphatic chlorinated hydrocarbons in the 

rat after vapor phase exposure. Biol Res Pregnancy Perinatol 6(2):79–88. 

Zarrabeitia, MT; Ortega, C; Altuzarra, E; et al. (2001) Accidental dichloromethane fatality: a case report. J 

Forensic Sci 46:726–727 

Zeiger; E. (1990) Mutagenicity of 42 chemicals in Salmonella. Environ Mol Mutagen 16:32–54. 

Zielenska, M; Ahmed, A; Pienkowska, M; et al. (1993) Mutational specificities of environmental 

carcinogens in the lacI gene of Escherichia coli. VI. Analysis of methylene chloride-induced mutational 

distribution in Uvr+ and UvrB- strains. Carcinogenesis 14(5):789–794.