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TOXICOLOGICAL REVIEW
OF
ACRYLONITRILE
(CAS No. 107-13-1)
In Support of Summary Information on the
Integrated Risk Information System (IRIS)
June 2011
NOTICE
This document is an External Review draft. This information is distributed solely for the
purpose of pre-dissemination peer review under applicable information quality guidelines. It has
not been formally disseminated by EPA. It does not represent and should not be construed to
represent any Agency determination or policy. It is being circulated for review of its technical
accuracy and science policy implications.
U.S. Environmental Protection Agency
Washington, DC
-------
DISCLAIMER
This document is a preliminary draft for review purposes only. This information is
distributed solely for the purpose of pre-dissemination peer review under applicable information
quality guidelines. It has not been formally disseminated by EPA. It does not represent and
should not be construed to represent any Agency determination or policy. Mention of trade
names or commercial products does not constitute endorsement or recommendation for use.
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CONTENTS —TOXICOLOGICAL REVIEW OF ACRYLONITRILE
(CAS No. 107-13-1)
LIST OF TABLES vii
LIST OF FIGURES xiii
LIST OF ABBREVIATIONS AND ACRONYMS xiv
FOREWORD xviii
AUTHORS, CONTRIBUTORS, AND REVIEWERS xix
1. INTRODUCTION 1
2. CHEMICAL AND PHYSICAL INFORMATION 3
3. TOXICOKINETICS 5
3.1. ABSORPTION 5
3.1.1. Studies in Humans 5
3.1.2. Studies in Animals 5
3.2. DISTRIBUTION 7
3.3. METABOLISM 13
3.3.1. Oxidation of AN to CEO 15
3.3.2. Interact! on of AN with GSH 23
3.3.3. Covalent Binding of AN and Its Metabolites to Subcellular Macromolecules 28
3.4. ELIMINATION 34
3.4.1. Studies in Humans 34
3.4.2. Studies in Animals 34
3.4.2.1. Exhalation 34
3.4.2.2. Fecal Excretion 34
3.4.2.3. Urinary Excretion 35
3.5. PHYSIOLOGICALLY BASED PHARMACOKINETIC MODELS 40
4. HAZARD IDENTIFICATION 52
4.1. STUDIES IN HUMANS—EPIDEMIOLOGY AND CASE REPORTS 52
4.1.1. Oral Exposure 52
4.1.2. Inhalation Exposure 52
4.1.2.1. Acute Exposure 52
4.1.2.2. Chronic Exposure 54
4.1.3. Dermal Exposure 113
4.1.3.1. Acute Exposure 113
4.1.3.2. Chronic Exposure 114
4.1.4. Ocular Exposure 115
4.2. SUBCHRONIC AND CHRONIC STUDIES AND CANCER BIOASSAYS IN
ANIMALS—ORAL AND INHALATION 115
4.2.1. Oral Exposure 115
4.2.1.1. Subchronic Studies 115
4.2.1.2. Chronic Studies 122
4.2.2. Inhalation Exposure 149
4.2.2.1. Subchronic Studies 149
4.2.2.2. Chronic Studies 150
4.3. REPRODUCTIVE/DEVELOPMENTAL STUDIES—ORAL AND INHALATION.. 158
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4.3.1. Studies in Humans 158
4.3.2. Studies in Animals 162
4.3.2.1. Oral Studies 162
4.3.2.2. Inhalation Exposure 170
4.3.2.3. Intraperitoneal Administration 177
4.3.3. In Vitro Studies 179
4.4. OTHER DURATION- OR ENDPOINT-SPECIFIC STUDIES 180
4.4.1. Acute Toxicity Data 180
4.4.1.1. Effects of AN on the GI Tract 184
4.4.1.2. Effects of AN on the Kidney 186
4.4.1.3. Effects of AN on the Adrenal Gland 186
4.4.1.4. Effects of AN on Neurological Endpoints 187
4.4.1.5. Effects of AN on Hearing 188
4.4.2. Immunological Effects of AN 191
4.5. MECHANISTIC DATA AND OTHER STUDIES IN SUPPORT OF THE MODE
OF ACTION 196
4.5.1. Mode-of-Action Studies 196
4.5.1.1. Noncancer Endpoints 196
4.5.1.2. Cancer Effects 210
4.5.2. Genotoxicity Studies 222
4.5.2.1. Studies in Humans 222
4.5.2.2. In Vivo Tests in Mammals 225
4.5.2.3. Short-term Tests: Bacteria, Fungi, Drosophila, Others 228
4.5.2.4. Mammalian Cell Short-term Tests 233
4.6. SYNTHESIS OF MAJOR NONCANCER EFFECTS 254
4.6.1. Oral 254
4.6.2. Inhalation 258
4.6.3. Mode-of-Action Information 262
4.6.3.1. GI Effects 263
4.6.3.2. Neurological Effects 264
4.6.3.3. Reproductive/Developmental Effects 266
4.6.3.4. Hematological Effects 267
4.6.3.5. Immunological Effects 267
4.6.3.6. Covalent Binding to Sulfhydryl Groups 268
4.7. EVALUATION OF CARCINOGENICITY 268
4.7.1. Summary of Overall Weight of Evidence 268
4.7.2. Synthesis of Human, Animal, and Other Supporting Evidence 269
4.7.3. Mode-of-Action Information 272
4.7.3.1. Hypothesized Mode of Action for Brain Tumors: Mutagenic Mode-of-
Action 272
4.7.3.2. Experimental Support for the Hypothesized Mode of Action 273
4.7.3.3. Other Possible Modes of Action 283
4.7.3.4. Possible Modes of Action for Forestomach Tumors 292
4.7.3.5. Possible Modes of Action for Other Tumors - Hepatoma, Mammary
Gland, Lung, Intestinal, Tongue, Zymbal Gland, and Harderian Gland
Tumors 293
4.8. SUSCEPTIBLE POPULATIONS AND LIFE STAGES 294
4.8.1. Possible Childhood Susceptibility 294
4.8.2. Possible Geriatric Susceptibility 295
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4.8.3. Possible Gender Differences 296
4.8.4. Genetic Polymorphisms 299
4.8.4.1. CYP450 299
4.8.4.2. Glutathione S-transferases 300
4.8.4.3. EH 301
5. DOSE-RESPONSE ASSESSMENTS 302
5.1. ORAL REFERENCE DOSE (RfD) 302
5.1.1. Choice of Principal Study and Critical Effect 302
5.1.2. Methods of Analysis—Including Models 303
5.1.2.1. PBPK Modeling 305
5.1.2.2. BMD Modeling 306
5.1.3. RfD Derivation—Including Application of Uncertainty Factors (UFs) 311
5.1.4. Data Array for Oral Noncancer Endpoints 313
5.1.5. Previous RfD Assessment 315
5.2. INHALATION REFERENCE CONCENTRATION (RfC) 315
5.2.1. Choice of Principal Study and Critical Effect 315
5.2.2. Methods of Analysis 318
5.2.3. RfC Derivation—Including Application of Uncertainty Factors (UFs) 319
5.2.4. Data Array for Inhalation Noncancer Endpoints 320
5.2.5. Previous RfC Assessment 321
5.3. UNCERTAINTIES IN THE ORAL REFERENCE DOSE AND INHALATION
REFERENCE CONCENTRATION 321
5.4. CANCER ASSESSMENT 323
5.4.1. Choice of Study/Data—with Rationale and Justification 323
5.4.2. Dose-Response Data 325
5.4.2.1. Human Occupational Data 325
5.4.2.2. Rat Oral Data 325
5.4.2.3. Rat Inhalation Data 329
5.4.3. Dose-Response Modeling 330
5.4.3.1. Human Occupational Data 330
5.4.3.2. Rat Oral Data 331
5.4.3.3. Rat Inhalation Data 338
5.4.4. Oral Slope Factor and Inhalation Unit Risk 341
5.4.4.1. OralCSFs 341
5.4.4.2. Inhalation Unit Risk 343
5.4.4.3. Application of Age-Dependent Adjustment Factors (ADAFs) 345
5.4.5. Uncertainties in Cancer Risk Values 348
5.4.5.1. Oral Cancer Assessment 348
5.4.5.2. Inhalation Cancer Assessment 354
5.4.6. Previous Cancer Assessment 358
6. MAJOR CONCLUSIONS IN THE CHARACTERIZATION OF HAZARD AND
DOSE-RESPONSE 359
6.1. HUMAN HAZARD POTENTIAL 359
6.2. DOSE-RESPONSE 361
6.2.1. Oral RfD 361
6.2.2. Inhalation RfC 362
6.2.3. Oral Slope Factor 364
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6.2.4. Cancer Inhalation Unit Risk 365
7. REFERENCES 367
APPENDIX A. SUMMARY OF EXTERNAL PEER REVIEW AND PUBLIC
COMMENTS AND DISPOSITION A-l
APPENDIX B. BENCHMARK DOSE CALCULATIONS B-l
APPENDIX B-l. NONCANCER ORAL DOSE-RESPONSE ASSESSMENT (RfD):
BENCHMARK DOSE MODELING RESULTS EMPLOYING THE INCIDENCE
OF FORESTOMACH LESIONS (HYPERPLASIA AND HYPERKERATOSIS)
IN MALE AND FEMALE SPRAGUE-DAWLEY RATS, F344 RATS, AND
B6C3Fi MICE CHRONICALLY EXPOSED ORALLY TO AN FOR 2 YEARS B-l
APPENDIX B-2. NONCANCER INHALATION DOSE-RESPONSE ASSESSMENT
(RfC): BMD MODELING RESULTS EMPLOYING THE INCIDENCE DATA
FOR NONNEOPLASTIC NASAL LESIONS IN RATS EXPOSED TO AN BY
INHALATION FOR 2 YEARS (TABLES B-6 THROUGH B-9) B-35
APPENDIX B-3. CANCER ORAL DOSE-RESPONSE ASSESSMENT: BMD DOSE
MODELING RESULTS FOR TUMOR INCIDENCE DATA FROM RATS
CHRONICALLY EXPOSED TO AN IN DRINKING WATER B-42
APPENDIX B-4. CANCER INHALATION DOSE-RESPONSE ASSESSMENT: BMD
MODELING RESULTS FOR TUMOR INCIDENCE DATA FROM RATS
CHRONICALLY EXPOSED TO AN VIA INHALATION B-149
APPENDIX B-5. ANALYSIS TO ASSESS COMBINING TUMOR INCIDENCE
DATA FROM TWO CANCER BIO AS SAYS EMPLOYING SPRAGUE-
DAWLEY RATS B-183
APPENDIX B-6. ESTIMATION OF COMPOSITE CANCER RISK FROM
EXPOSURE TO AN BY COMBINING RISK ESTIMATES ACROSS MULTIPLE
TUMOR SITES B-189
APPENDIX B-7. STATISTICAL ANALYSIS OF BLAIR ETAL. (1998) B-l93
APPENDIX C. PBPK MODEL DESCRIPTIONS AND SOURCE CODE C-l
APPENDIX D. UNCERTAINTIES ASSOCIATED WITH CHEMICAL-SPECIFIC
PARAMETERS EMPLOYED IN THE PBPK MODEL FOR AN DOSIMETRY
IN HUMANS D-l
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LIST OF TABLES
Table 3-1. Recovery of radioactivity from male Sprague-Dawley rats exposed to 5 or 100
ppm [1-14C]-AN for 6 hours via inhalation [[[ 6
Table 3-2. Recovery of radioactivity after a single gavage dose of 0.1 or 10 mg/kg [1-14C]-
AN to male Sprague-Dawley rats [[[ 6
Table 3-3. Percentage recovery of radioactivity in tissues of male Wistar rats following a
single oral dose of radiolabeled AN [[[ 9
Table 3-4. Apparent kinetic parameters of CEO formation from AN in B6C3F1 mice,
F344rats, and humans [[[ 20
Table 3-5. Apparent kinetic parameters for glutathione conjugation of AN and CEO at pH
6.5 [[[ 26
Table 3-6. Glutathione conjugation of AN and CEO with or without microsomal or
cytosolic GST from rat, mouse, and human liver preparations ................................. 26
Table 3-7. Urinary excretion of thioethers derived from AN [[[ 38
Table 4-1. Clinical signs in 144 subjects accidentally exposed to AN ......................................... 53
Table 4-2. Distribution of select incidence and mortalities among wage workers and all
workers at an AN plant in South Carolina [[[ 56
Table 4-3. Distribution of select incidence and mortalities among wage workers and all
workers at an AN plant in South Carolina (updated follow-up) ................................ 59
Table 4-4. Distribution of select incidence and mortalities among wage and salary workers
at an AN plant in Virginia [[[ 60
Table 4-5. Distribution of select incidences and mortalities among exposed workers in two
AN plants in South Carolina and Virginia [[[ 63
Table 4-6. Distribution of select mortalities among exposed workers in two AN plants in
South Carolina and Virginia (updated follow-up) [[[ 64
Table 4-7. Crude and adjusted hazard ratio estimates for cumulative exposure and adjusted
hazard ratio estimates for exposure intensity among workers in two AN plants
in South Carolina and Virginia (updated follow-up) ................................................. 65
Table 4-8. Adjusted hazard ratio estimated for select cancer mortality by lagged
cumulative exposure for 100-ppm-year increases in cumulative exposure
among workers in two AN plants in South Carolina and Virginia (updated
follow-up) [[[ 66
Table 4-9. Distribution of select mortalities among AN-exposed and unexposed workers in
the Netherlands [[[ 73
Table 4-10. Lung cancer mortality among AN-exposed workers in the Netherlands,
stratified by cumulative dose and latency [[[ 73
Table 4-11. Distribution of select mortalities among AN-exposed and unexposed workers
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Table 4-17. Derived SMRs for all-cancer mortality and AN exposure in major cohort
studies 95
Table 4-18. Derived associations between lung cancer and AN exposure in major cohort
and case-control studies 97
Table 4-19. Epidemiology studies of general symptoms, clinical chemistry, and
neurological outcomes among cohorts of workers exposed to AN 103
Table 4-20. Industrial AN exposure, levels of AN and thiocyanate in urine, and prevalence
of physical signs of adverse effects in workers exposed to AN at six acrylic
fiber factories in Japan 106
Table 4-21. Effect on SCV in male Sprague-Dawley rats exposed to AN via gavage for 12
weeks 120
Table 4-22. Incidence of nonneoplastic lesions in Sprague-Dawley rats exposed to AN in
drinking water for 2 years 125
Table 4-23. Selected tumor incidences in response to AN administered to Sprague-Dawley
rats in drinking water for up to 2 years 126
Table 4-24. Histopathologic location of astrocytomas in the CNS of male Sprague-Dawley
rats administered AN in drinking water for 2 years 127
Table 4-25. Incidence of nonneoplastic lesions in Sprague-Dawley rats exposed to AN in
drinking water for 2 years 129
Table 4-26. Selected tumor incidences in Sprague-Dawley rats exposed to AN in drinking
water for up to 2 years 131
Table 4-27. Incidence of nonneoplastic lesions in Sprague-Dawley rats exposed to AN by
gavage for 20 months 133
Table 4-28. Cumulative incidence of tumors in response to AN administered to Sprague-
Dawley rats by gavage for up to 2 years 134
Table 4-29. Incidences of nontumorous lesions in F344 rats exposed to AN in drinking
water for 2 years 137
Table 4-30. Selected tumor incidences in F344 rats exposed to AN in drinking water for 2
years 139
Table 4-31. Incidence and severity of nonneoplastic lesions in B6C3Fi mice exposed by
gavage to AN for 2 years 141
Table 4-32. Incidences of selected neoplastic lesions in B6C3Fi mice exposed by gavage
to AN for 2 years 142
Table 4-33. Incidence of tumors in female Sprague-Dawley rats exposed to AN in drinking
water for up to 46 weeks 146
Table 4-34. Summary of chronic oral toxicity studies of AN: noncancer effects in rats and
mice 146
Table 4-35. Summary of chronic oral toxicity studies of AN: cancer effects in rats and
mice 148
Table 4-36. Effect on SCV in male Sprague-Dawley rats exposed to AN via inhalation for
24 weeks 150
Table 4-37. Incidence of histopathological lesions of the nasal turbinates in Sprague-
Dawley rats exposed to AN via inhalation for 2 years 152
Table 4-38. Incidence of dose-related noncancerous histopathological lesions in Sprague-
Dawley rats exposed to AN via inhalation for 2 years 154
Table 4-39. Cumulative incidence of tumors in Sprague-Dawley rats exposed to AN via
inhalation for up to 2 years 155
Table 4-40. Comparison of carcinogenic effects of chronic exposure to AN at 60 ppm
starting either in utero or in adulthood, in Sprague-Dawley rats 157
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Table 4-41. Epidemiology studies of reproductive and developmental outcomes among
cohorts of workers exposed to AN 158
Table 4-42. Incidence of fetal abnormalities among litters of Sprague-Dawley rats
following maternal exposure to AN on GDs 6-15 164
Table 4-43. Morphological alterations in GD 12 fetuses of Sprague-Dawley rats exposed
to lOOmg/kgANonGD 10 165
Table 4-44. Group-specific reproductive indices in three generations of Sprague-Dawley
rats receiving AN in drinking water 168
Table 4-45. Group-specific pup weights in three generations of Sprague-Dawley rats
receiving AN in drinking water 169
Table 4-46. Incidence of fetal malformations among litters of Sprague-Dawley rats
exposed to AN by inhalation 171
Table 4-49. Effects of AN on organ weight, clinical chemistry, and biochemical
parameters when administered to male Wistar rats via inhalation 183
Table 4-50. Time course of the effect of AN administration on [3H]-thymidine uptake into
mouse splenocytes under the influence of different mitogens in vitro 194
Table 4-51. Summary of immunotoxicity studies of AN 195
Table 4-52. Effect of AN on RBC metabolic intermediates following a single oral dose 197
Table 4-53. Detection of N7-(2-oxoethyl)guanine after i.p. administration of 50 mg/kg AN
or CEO to male F344 rats 212
Table 4-54. Summary of studies on the mutagenicity/genotoxicity of AN 239
Table 4-55. Formation of 8-oxodG in DNA from tissues of male Sprague-Dawley and
F344 rats exposed to AN in drinking water for 21 days 251
Table 4-56. Summary of studies on the mutagenicity or genotoxicity resulting from
oxidative stress of AN 253
Table 4-57. Noncancer effects in animals in chronic oral exposure studies 254
Table 4-58. Noncancer effects observed in epidemiology studies among cohorts of workers
exposed to AN 259
Table 4-59. Noncancer effects observed in chronic inhalation exposure studies in animals 260
Table 4-59. 8-OxodG in brain DNA and brain tumor incidence in male F344 rats exposed
to AN in drinking water 288
Table 4-60. 8-OxodG in brain DNA and brain tumor incidence in male Sprague-Dawley
rats exposed to AN in drinking water 288
Table 5-1. Incidences of forestomach lesions (hyperplasia or hyperkeratosis) in Sprague-
Dawley and F344 rats exposed to AN in drinking water for 2 years 304
Table 5-2. Incidences of forestomach lesions (hyperplasia or hyperkeratosis) in male and
female B6C3Fi mice administered AN via gavage for 2 years 305
Table 5-3. Candidate RfDs based on BMD modeling of the incidence of forestomach
lesions (hyperplasia or hyperkeratosis) in male and female Sprague-Dawley
rats exposed to AN in drinking water for 2 years 308
Table 5-4. Candidate RfDs based on BMD modeling of the incidence of forestomach
lesions (hyperplasia or hyperkeratosis) in male and female F344 rats exposed
to AN in drinking water for 2 years 309
Table 5-5. Candidate RfDs based on BMD modeling of the incidence of forestomach
lesions (hyperplasia or hyperkeratosis) in male and female B6C3Fi mice
exposed to AN via gavage for 2 years 310
Table 5-6. Results of dose-response analyses of incidence data for selected nasal lesions in
male and female Sprague-Dawley rats exposed by inhalation to AN for 2 years ...319
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Table 5-8. Incidence of CNS tumors in Sprague-Dawley and F344 rats exposed to AN in
drinking water for 2 years 326
Table 5-9. Incidence of mammary gland tumors in F344 and Sprague-Dawley rats exposed
to AN in drinking water for 2 years 327
Table 5-10. Tumor incidences Sprague-Dawley and F344 in rats exposed to AN in
drinking water for 2 years 328
Table 5-11. Incidences of intestinal, CNS, Zymbal gland, tongue, and mammary gland
tumors in Sprague-Dawley rats exposed to AN via inhalation for 2 years 330
Table 5-12. Four different dose metrics, two external and two internal, based on doses
employed in studies of Sprague-Dawley and F344 rats exposed to AN in
drinking water for 2 years 332
Table 5-13. BMD modeling results using tumor incidence data from male and female
Sprague-Dawley and F344 rat studies in which animals were exposed to AN in
drinking water for 2 years 333
Table 5-14. Site-specific oral CSFs for AN based on BMD modeling of tumor incidence
data in rats and predicted CEO levels in blood (AUC/24 hours) of rats and
humans assuming episodic exposure to AN 335
Table 5-15. Site-specific oral CSFs for AN based on BMD modeling of tumor incidence
data in rats and predicted AN levels in blood (AUC/24 hours) of rats and
humans assuming episodic exposure to AN 336
Table 5-16. Site-specific oral CSFs for AN based on BMD modeling of tumor incidence
data in rats and BW scaling to the % power to convert from rat to human
administered doses 337
Table 5-17. Two different dose metrics, one external and one internal, based on
administered air concentrations of AN employed in a 2-year bioassay in
Sprague-Dawley rats 338
Table 5-18. BMD modeling results using tumor incidence data from male and female
Sprague-Dawley rats exposed to AN via inhalation for 2 years and CEO
concentration in blood predicted from an EPA-modified PBPK model 339
Table 5-19. Site-specific lURs for AN based on BMD modeling of tumor incidence data in
Sprague-Dawley rats and PBPK modeling of CEO levels in blood (AUC/24
hours) of rats and humans 340
Table 5-20. Site-specific lURs for AN based on BMD modeling of tumor incidence data in
Sprague-Dawley rats exposed to AN via inhalation 340
Table 5-21. Estimated human equivalent oral CSFs for composite risk based on tumor
incidence in AN-exposed rats and predicted CEO-AUC levels in blood 342
Table 5-22. Estimated human equivalent composite lURs based on tumor incidence in
AN-exposed rats and predicted CEO-AUC levels in blood 344
Table 5-23. Application of ADAFs to AN cancer risk following a lifetime (70-year) oral
exposure 346
Table 5-24. Application of ADAFs to AN cancer risk following a lifetime (70-year)
inhalation exposure 347
Table 5-25. Summary of uncertainty in the AN oral cancer risk assessment 349
Table 5-26. Summary of uncertainty in the AN inhalation cancer risk assessment 355
Table B-l. Incidences of forestomach lesions (hyperplasia or hyperkeratosis) in Sprague-
Dawley and F344 rats exposed to AN in drinking water for 2 years B-l
Table B-2. Incidences of forestomach lesions (hyperplasia or hyperkeratosis) in male and
female B6C3Fi mice administered AN via gavage for 2 years B-3
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Table B-3. Summary of the BMD modeling results based on the incidence of forestomach
lesions (hyperplasia or hyperkeratosis) in male and female Sprague-Dawley
rats exposed to AN in drinking water for 2 years B-4
Table B-4. Summary of the BMD modeling results based on the incidence of forestomach
lesions (hyperplasia or hyperkeratosis) in male and female F344 rats exposed
to AN in drinking water for 2 years B-17
Table B-5. Summary of the BMD modeling results based on the incidence of forestomach
lesions (hyperplasia or hyperkeratosis) in male and female B6C3Fi mice
exposed to AN via gavage for 2 years B-30
Table B-6. Incidence data for selected nasal lesions in Sprague-Dawley rats exposed by
inhalation to AN for 2 years B-3 5
Table B-7. A summary of BMDS (version 1.3.2) modeling results based on incidence of
hyperplasia of mucus-secreting cells in male Sprague-Dawley rats exposed to
AN via inhalation for 2 years B-36
Table B-8. A summary of BMDS (version 1.3.2) modeling results based on incidence of
flattening of respiratory epithelium in female Sprague-Dawley rats exposed to
AN via inhalation for 2 years B-39
Table B-9. Incidence of forestomach (nonglandular) tumors in Sprague-Dawley rats
exposed to AN in drinking water for 2 years B-43
Table B-10. Summary of BMD modeling results based on incidence of forestomach
(nonglandular) tumors in Sprague-Dawley rats exposed to AN in drinking
water for 2 years B-43
Table B-l 1. Incidence of CNS tumors in Sprague-Dawley rats exposed to AN in drinking
water for 2 years B-56
Table B-12. Summary of BMD modeling results based on incidence of CNS tumors in
Sprague-Dawley rats exposed to AN in drinking water for 2 years B-56
Table B-13. Incidence of Zymbal gland tumors in Sprague-Dawley rats exposed to AN in
drinking water for 2 years B-69
Table B-14. Summary of BMD modeling results based on incidence of Zymbal gland
tumors in Sprague-Dawley rats exposed to AN in drinking water for 2 years B-69
Table B-15. Incidence of tongue tumors in Sprague-Dawley rats exposed to AN in
drinking water for 2 years B-82
Table B-16. Summary of BMD modeling results based on incidence of tongue tumors in
Sprague-Dawley rats exposed to AN in drinking water for 2 years B-82
Table B-17. Incidence of mammary gland tumors in Sprague-Dawley rats exposed to AN
in drinking water for 2 years B-95
Table B-18. Summary of BMD modeling results based on incidence of mammary gland
tumors in Sprague-Dawley rats exposed to AN in drinking water for 2 years B-95
Table B-19. Incidence of forestomach (nonglandular) tumors in F344 rats exposed to AN
in drinking water for 2 years B-l02
Table B-20. Summary of BMD modeling results based on incidence of forestomach
(nonglandular) tumors in F344 rats exposed to AN in drinking water for 2
years B-103
Table B-21. Incidence of CNS tumors in F344 rats exposed to AN in drinking water for 2
years B-l 16
Table B-22. Summary of BMD modeling results based on incidence of CNS tumors in
F344 rats exposed to AN in drinking water for 2 years B-l 16
Table B-23. Incidence of Zymbal gland tumors in F344 rats exposed to AN in drinking
water for 2 years B-129
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Table B-24. Summary of BMD modeling results based on incidence of Zymbal gland
tumors inF344 rats exposed to AN in drinking water for 2 years B-129
Table B-25. Incidence of mammary gland tumors in F344 rats exposed to AN in drinking
water for 2 years B-142
Table B-26. Summary of BMD modeling results based on incidence of mammary gland
tumors inF344 rats exposed to AN in drinking water for 2 years B-142
Table B-27. Incidence of intestinal tumors in Sprague-Dawley rats exposed to AN in air
for 2 years B-150
Table B-28. Summary of BMD modeling results based on incidence of intestinal tumors in
Sprague-Dawley rats exposed to AN in air for 2 years B-150
Table B-29. Incidence of CNS tumors in Sprague-Dawley rats exposed to AN in air for 2
years B-155
Table B-30. Summary of BMD modeling results based on incidence of CNS tumors in
Sprague-Dawley rats exposed to AN in air for 2 years B-155
Table B-31. Incidence of Zymbal gland tumors in Sprague-Dawley rats exposed to AN in
air for 2 years B-164
Table B-32. Summary of BMD modeling results based on incidence of Zymbal's gland
tumors in Sprague-Dawley rats exposed to AN in air for 2 years B-164
Table B-33. Incidence of tongue tumors in Sprague-Dawley rats exposed to AN in air for 2
years B-173
Table B-34. Summary of BMD modeling results based on incidence of tongue tumors in
Sprague-Dawley rats exposed to AN in air for 2 years B-173
Table B-3 5. Incidence of mammary gland tumors in Sprague-Dawley rats exposed to AN
in air for 2 years B-178
Table B-36. Summary of BMD modeling results based on incidence of mammary gland
tumors in Sprague-Dawley rats exposed to AN in air for 2 years B-178
Table B-37. Summary of PODs for composite cancer risk associated with episodic oral
exposure to AN, using CEO-AUC levels in blood as dose metric and multiple
tumor incidence data in rats B-191
Table B-38. Summary of PODs for composite cancer risk associated with inhalation
exposure to AN, based on multiple tumor incidence data in rats and CEO-AUC
levels in blood B-191
Table B-39. Estimated human oral CSFs for AN based on multiple tumor incidence data in
rats and CEO-AUC level sin blood B-192
Table B-40. Estimated human lURs for AN based on multiple tumor incidence data in rats
and CEO-AUC levels in blood B-192
Table C-l. Rat and human PBPK model parameter values C-2
TableD-l. Tissue:bloodPCs D-2
Table D-2. Impact of method of estimation of PCs on the peak AN and CEO model
predictions D-3
Table D-3. PBPK mass balance predictions for an 8-hour human exposure to 2 ppm AN D-6
Table D-4. Sensitivity of AN and CEO metrics for continuous inhalation exposure D-9
Table D-5. Sensitivity of AN and CEO metrics for continuous oral exposure D-10
Table D-6. Estimated CVs for AN and CEO metrics D-12
Table D-7. Parameter contributions to overall CVs for inhalation exposure D-13
Table D-8. Parameter contributions to overall CVs for oral exposure D-14
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LIST OF FIGURES
Figure 2-1. Chemical structure of AN 3
Figure 3-1. Scheme for the metabolic transformation of AN 14
Figure 3-2. Structure of the PBPK model for AN and CEO 40
Figure 3-3. Intravenous exposure, dosimetry, and model fits 45
Figure 3-4a. Inhalation exposure, CEO concentrations 46
Figure 3-4b. Inhalation exposure, AN concentrations 47
Figure 3-5a. Oral exposure, CEO concentrations 48
Figure 3-5b. Oral exposure, AN concentrations 49
Figure 3-5c. Oral exposure, urinary excretion, andHb binding 50
Figure 5-1. Exposure-response array for noncancer endpoints across target organs
following oral exposure to AN in animals 314
Figure 5-2. Comparison of composite oral CSFs derived from tumor incidence data in four
different sex/strain/species of rats exposed chronically to AN. For each
sex/strain/species combination, two different dose metrics were employed: (1)
CEO concentration in blood, and (2) human equivalent administered dose 350
Figure 5-3. Comparison of composite lURs derived from: (1) tumor incidence data in
male and female Sprague-Dawley rats exposed chronically to AN, and (2) lung
cancer mortality in humans exposed to AN occupationally. In deriving the
animal-based lURs, two different dose metrics were employed: (1) predicted
CEO concentration in blood, and (2) human equivalent administered AN
concentration in air 356
Figure C-1. Human inhalation exposure level vs. internal AN concentration C-5
Figure C-2. Human inhalation exposure level vs. internal CEO concentrations C-6
Figure C-3. Human oral exposure level vs. internal AN concentration for continuous
exposure C-8
Figure C-4. Human oral exposure level vs. internal CEO concentration for continuous
exposure C-8
Figure C-5. Human oral exposure level vs. internal AN concentration for episodic
exposure C-9
Figure C-6. Human oral exposure level vs. internal CEO concentration for episodic
exposure C-9
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LIST OF ABBREVIATIONS AND ACRONYMS
ABS
ABT
ADAF
ADUR
AIC
ALT
AMAP
AN
ASAP
AST
ATP
ATPase
ATSDR
AUC
BAL
BMC
BMCL
BMD
BMDL
BMDS
BMR
BrdU
BSO
BUN
BW
CA
CAIII
CAP
CASRN
CEMA
CEO
CEVal
CF
CHL
CHO
CI
CNS
con-A
CSF
CV
CYP450
DEM
DEX
dGMP
DMSO
DNA
AN butadiene styrene
1 -aminobenzotriazole
age-dependent adjustment factor
AN-derived undialyzable radioactivity
Akaike's Information Criterion
alanine aminotransferase
amplitude of the motor action potential
acrylonitrile
amplitude of the sensory action potential
aspartate aminotransferase
adenosine triphosphate
adenosine triphosphatase
Agency for Toxic Substances and Disease Registry
area under the curve
bronchoalveolar lavage
benchmark concentration
95% lower bound of the BMC
benchmark dose
95% lower confidence limit of benchmark dose
benchmark dose software
benchmark response
bromodeoxyuridine
buthionine sulfoximine
blood urea nitrogen
body weight
chromosomal aberration
carbonic anhydrase III
compound action potential
Chemical Abstracts Service Registry Number
2-cyanoethyl mercapturic acid
2-cyanoethylene oxide
N-(2-cyanoethyl)valine
correction factor
Chinese hamster lung
Chinese hamster ovary
confidence interval
central nervous system
concanavalin-A
cancer slope factor
coeffi ci ent of vari ati on
cytochrome P450
diethylmaleate
dexamethasone
deoxyguanosine-5'-monophosphate
dimethyl sulfoxide
deoxyribonucleic acid
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DPOAE
DTK
EC
ECG
EH
ERK
EROD
FISH
FSH
GAPDH
GD
GEI
GI
GJIC
GSH
GSSG
GST
GSTM
y-GTP
Hb
HbCO
HbO
HEC
HED
HMPA
HPLC
hprt
IC50
IPCS
IRIS
IUR
i.v.
KCN
LC50
LD50
LDH
LEC
LH
LOAEL
LPS
MCMC
MCV
MDA
mEH
MEK
MEL
MetHb
distortion product otoacoustic emission
delayed-type hypersensitivity
exposure concentration
electrocardiogram
epoxide hydrolase
extracellular signal-regulated kinase
ethoxyresorufin-O-deethylase
fluorescence in situ hybridization
follicle stimulating hormone
glyceraldehyde-3-phosphate dehydrogenase
gestation day
gastric erosion severity index
gastrointestinal
gap junction intercellular communications
reduced glutathione
glutathione disulfide
glutathi one- S -transferase
GST of the |i subclass
y-glutamyl transpeptidase
hemoglobin
carboxyhemoglobin
oxyhemoglobin
human equivalent concentration
human equivalent dose
hexamethylphosphoramide
high performance liquid chromatography
hypoxanthine guanine phosphoribosyl transferase
median inhibitory concentration
immunoglobulin
intraperitoneal
International Programme on Chemical Safety
Integrated Risk Information System
inhalation unit risk
intravenous
potassium cyanide
median lethal concentration
median lethal dose
lactate dehydrogenase
95% lower bound of exposure concentration
luteinizing hormone
lowest-observed-adverse-effect level
lipopolysaccharide
Markov chain Monte Carlo
motor conduction velocity
malondialdehyde
microsomal epoxide hydrolase
mitogen-activated/ERK-activating kinase
melatonin
methemoglobin
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MN
MNU
MP
4MP
NAC
NADP(H)
NAS
NCI
NCTB
NHA
NIHL
NIOSH
NMR
NOAEL
NRC
NTP
OBN
OHC
OR
OTC
8-oxodG
PB
PEN
PBPK
PC
PCR
PH
PHA
PK
PMA
PND
POD
PRL
RBC
RfC
RfD
RLC
RNA
ROS
RR
s.c.
SCE
SCV
SD
SDH
SEER
SGPT
SHE
micronucleus, micronuclei
methylnitrosourea
microsomal protein
4-methylpyrazole
N-acetylcysteine
nicotinamide adenine dinucleotide phosphate (reduced)
National Academy of Sciences
National Cancer Institute
Neurobehavioral Core Test Battery
normal human astrocyte
noise-induced hearing loss
National Institute for Occupational Safety and Health
nuclear magnetic resonance
no-observed-adverse-effect level
National Research Council
National Toxicology Program
octave band of noise
outer hair cell
odds ratio
L-2-oxothiazolidine-4-carboxylic acid
8-oxodeoxyguanosine, also referred to as 8-oxo-7,8-dihydro-
2'deoxyguanosine and 8-hydroxy-2'-deoxyguanosine
phenobarbital
phenyl -N-terti ary-butylnitrone
physiologically based pharmacokinetic
partition coefficient
polymerase chain reaction
phorone
phytohemagglutinin
protein kinase
phorbol 12-myristate 13-acetate
postnatal day
point of departure
prolactine
red blood cell
reference concentration
reference dose
rat liver cell
ribonucleic acid
reactive oxygen species
relative risk
subcutaneous
sister chromatid exchange
sensory conduction velocity
standard deviation
sorbitol dehydrogenase
Surveillance, Epidemiology and End Results
serum glutamate pyruvate transaminase
Syrian hamster embryo
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SIR
SMR
SOD
SRBC
STS
TAU
TEARS
TCPO
6-TG
TNF-a
TPO
TRX
TWA
UCL
UDS
UF
U.S. EPA
v/v
WBC
WHO
WT
standardized incidence ratio
standardized mortality ratio
superoxide dismutase
sheep red blood cell
sodium thiosulfate
taurine
thiobarbituric acid-reactive substances
1,1,1 -trichloropropane-2,3 -oxide
6-thioguanine
tumor necrosis factor
3,3,3-trichloropropylene oxide
trolox
time-weighted average
upper confidence limit
unscheduled DNA synthesis
uncertainty factor
U.S. Environmental Protection Agency
volume/volume
white blood cell
World Health Organization
wild-type
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FOREWORD
The purpose of this Toxicological Review is to provide scientific support and rationale
for the hazard and dose-response assessment in IRIS pertaining to chronic exposure to
acrylonitrile. It is not intended to be a comprehensive treatise on the chemical or toxicological
nature of acrylonitrile.
The intent of Section 6, Major Conclusions in the Characterization of Hazard and Dose
Response, is to present the major conclusions reached in the derivation of the reference dose,
reference concentration and cancer assessment, where applicable, and to characterize the overall
confidence in the quantitative and qualitative aspects of hazard and dose response by addressing
the quality of data and related uncertainties. The discussion is intended to convey the limitations
of the assessment and to aid and guide the risk assessor in the ensuing steps of the risk
assessment process.
For other general information about this assessment or other questions relating to IRIS,
the reader is referred to EPA's IRIS Hotline at (202) 566-1676 (phone), (202) 566-1749 (fax), or
hotline.iris@epa.gov (email address).
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AUTHORS, CONTRIBUTORS, AND REVIEWERS
CHEMICAL MANAGER/AUTHOR
Ambuja Bale, Ph.D., DABT
National Center for Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
Washington, DC
AUTHORS
Diana Wong, Ph.D., DABT
Amanda Persad, Ph.D.
Ted Berner, MS
Karen Hogan, MS
Leonid Kopylev, Ph.D.
Paul Schlosser, Ph.D.
John Fox, Ph.D.
April Luke, MS
Rosemarie Hakim, Ph.D.
National Center for Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
Washington, DC
Roglio Tornero-Valez, Ph.D.
Office of Research and Development
U.S. Environmental Protection Agency
Washington, DC
CONTRACTOR SUPPORT
Lutz W.D. Weber, Ph.D., DABT (First Draft)
George Holdsworth, Ph.D. (First Draft)
Janusz Z. Byczkowski, Ph.D., D.Sc., DABT (First Draft)
Donna Cragle, Ph.D. (First Draft)
Virginia H. Sublet, Ph.D. (First Draft)
Oak Ridge Institute for Science and Education
Oak Ridge Associated Universities
Oak Ridge, TN
Margaret Fransen, Ph.D. (Second Draft)
Michael Lumpkin, Ph.D. (Second Draft)
Jennifer Rhoades, B.S. (Second Draft)
Peter McClure, Ph.D., DABT (Second Draft)
Environmental Science Center
SRC, Inc.
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North Syracuse, NY
REVIEWERS
This document has been provided for review to EPA scientists and interagency reviewers
from other federal agencies and White House offices.
INTERNAL EPA REVIEWERS
Dave Bayliss, Ph.D.
Lynn Flowers, Ph.D., DABT
Kate Guyton, Ph.D., DABT
Jennifer Jinot
Cheryl Scott
Lawrence Valcovic, Ph.D.
John Whalan
National Center for Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
Washington, DC
Rob DeWoskin, Ph.D.
Channa Keshava, Ph.D.
National Center for Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
Research Triangle Park, NC
John Lipscomb, Ph.D.
National Center for Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
Cincinnati, OH
Jerry Blancato, Ph.D.
NERL
U.S. Environmental Protection Agency
Research Triangle Park, NC
James Allen, Ph.D.
Anthony DeAngelo, Ph.D.
Karl Jensen, Ph.D.
Robert MacPhail, Ph.D.
Jeff Ross, Ph.D.
NHEERL
Office of Research and Development
U.S. Environmental Protection Agency
Research Triangle Park, NC
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1. INTRODUCTION
This document presents background information and justification for the Integrated Risk
Information System (IRIS) Summary of the hazard and dose-response assessment of acrylonitrile
(AN). IRIS Summaries may include oral reference dose (RfD) and inhalation reference
concentration (RfC) values for chronic and other exposure durations, and a carcinogenicity
assessment.
The RfD and RfC, if derived, provide quantitative information for use in risk assessments
for health effects known or assumed to be produced through a nonlinear (presumed threshold)
mode of action (MO A). The RfD (expressed in units of mg/kg-day) is defined as an estimate
(with uncertainty spanning perhaps an order of magnitude) of a daily exposure to the human
population (including sensitive subgroups) that is likely to be without an appreciable risk of
deleterious effects during a lifetime. The inhalation RfC (expressed in units of mg/m3) is
analogous to the oral RfD, but provides a continuous inhalation exposure estimate. The
inhalation RfC considers toxic effects for both the respiratory system (portal of entry) and for
effects peripheral to the respiratory system (extrarespiratory or systemic effects). Reference
values are generally derived for chronic exposures (up to a lifetime), but may also be derived for
acute (<24 hours), short-term (>24 hours up to 30 days), and subchronic (>30 days up to 10% of
lifetime) exposure durations, all of which are derived based on an assumption of continuous
exposure throughout the duration specified. Unless specified otherwise, the RfD and RfC are
derived for chronic exposure duration.
The carcinogenicity assessment provides information on the carcinogenic hazard
potential of the substance in question and quantitative estimates of risk from oral and inhalation
exposure may be derived. The information includes a weight-of-evidence judgment of the
likelihood that the agent is a human carcinogen and the conditions under which the carcinogenic
effects may be expressed. Quantitative risk estimates may be derived from the application of a
low-dose extrapolation procedure. If derived, the oral slope factor is a plausible upper bound on
the estimate of risk per mg/kg-day of oral exposure. Similarly, an inhalation unit risk is a
plausible upper bound on the estimate of risk per ug/m3 air breathed.
Development of the hazard identification and dose-response assessments for AN has
followed the general guidelines for risk assessment as set forth by the National Research Council
(NRC, 1983). U.S. Environmental Protection Agency (U.S. EPA) Guidelines and Risk
Assessment Forum Technical Panel Reports that may have been used in the development of this
assessment include the following: Guidelines for the Health Risk Assessment of Chemical
Mixtures (U.S. EPA, 1986a), Guidelines for Mutagenicity Risk Assessment (U.S. EPA, 1986b),
Recommendations for and Documentation of Biological Values for Use in Risk Assessment (U.S.
EPA, 1988), Guidelines for Developmental Toxicity Risk Assessment (U.S. EPA, 1991), Interim
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Policy for Particle Size and Limit Concentration Issues in Inhalation Toxicity (U.S. EPA,
1994a), Methods for Derivation of Inhalation Reference Concentrations and Application of
Inhalation Dosimetry (U.S. EPA, 1994b), Use of the Benchmark Dose Approach in Health Risk
Assessment (U.S. EPA, 1995), Guidelines for Reproductive Toxicity Risk Assessment (U.S. EPA,
1996), Guidelines for Neurotoxicity Risk Assessment (U.S. EPA, 1998a), Science Policy Council
Handbook: Risk Characterization (U.S. EPA, 2000a), Benchmark Dose Technical Guidance
Document (U.S. EPA, 2000b), Supplementary Guidance for Conducting Health Risk Assessment
of Chemical Mixtures (U.S. EPA, 2000c), A Review of the Reference Dose and Reference
Concentration Processes (U.S. EPA, 2002), Guidelines for Carcinogen Risk Assessment (U.S.
EPA, 2005a), Supplemental Guidance for Assessing Susceptibility from Early-Life Exposure to
Carcinogens (U.S. EPA, 2005b), Science Policy Council Handbook: Peer Review (U.S. EPA,
2006a), and A Framework for Assessing Health Risks of Environmental Exposures to Children
(U.S. EPA, 2006b).
The literature search strategy employed for this compound was based on the Chemical
Abstracts Service Registry Number (CASRN) and at least one common name. Any pertinent
scientific information submitted by the public to the IRIS Submission Desk was also considered
in the development of this document. The relevant literature was reviewed through February
2011.
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2. CHEMICAL AND PHYSICAL INFORMATION
AN (CASRN 107-13-1) is a three-carbon alkene carrying a nitrile substituent group as
part of the terminal carbon atom (carbon 1) (Figure 2-1). Synonyms for the compound include
vinyl cyanide, propenenitrile, and cyanoethylene, and there are a variety of trade names. AN is a
colorless, flammable, and volatile liquid with a weakly pungent onion- or garlic-like odor. Some
physical and chemical properties are shown below (NLM, 2003; IPCS, 2002; ATSDR, 1990).
H
H
Figure 2-1. Chemical structure of AN.
Formula: CsHsN
Molecular weight: 53.06
Melting point: -83 °C
Boiling point: 77.4°C
Density: 0.806 g/mL (at 20°C)
LogKow: -0.92
LogKoc: -0.07
Vapor pressure: 100 mmHg at 22.8°C
Henry's law constant: 8.8 x 10"5 atm-m3/mol
Conversion factors: 1 ppm = 2.17 mg/m3
1 mg/m3 = 0.46 ppm
AN is a commercially important chemical with a wide range of uses in the chemical
industry. It is used in the production of acrylic and modacrylic fibers, plastics (AN butadiene
styrene [ABS] and AN-styrene resins), and nitrile rubbers and as an intermediate in the
production of other important chemicals, such as adiponitrile and acrylamide. AN is used in the
plastics industry in the formation of surface coatings and adhesives. It is a chemical intermediate
in the synthesis of antioxidants, pharmaceuticals, and dyes and, in general, for processes
requiring the introduction of a cyanoethyl group into a molecule (NLM, 2003). AN is also used
in clinical practice in the form of dialysis tubing (Mulvihill et al., 1992). AN was used
occasionally as a fumigant insecticide for stored grain. A measure of the commercial importance
of AN may be judged by the amount produced in a given year. The Agency for Toxic
Substances and Disease Registry (ATSDR, 1990) reports that 1,1 12,754 metric tons of AN were
produced in the United States in 1987. Production had increased to 1,455,735 metric tons by
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1995 (ACS, 2003). Exposure of the general public to AN can potentially occur through
migration of residual monomer in polymeric products via contact with food or water. Exposure
to airborne AN is possible among members of the general population living in the vicinity of
emission sources such as acrylic fiber or chemical manufacturing plants or waste sites (ATSDR,
1990). In addition, smokers are expected to be exposed to AN, which has been detected in
cigarette smoke at levels of 3.2-15 mg per cigarette (IARC, 1999).
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3. TOXICOKINETICS
Although only limited information on the toxicokinetics of AN in exposed humans is
available, a substantial body of evidence has accumulated on the absorption, distribution,
metabolism, and excretion of AN in experimental animals. AN is rapidly and nearly completely
absorbed, widely distributed among the tissues, and biochemically transformed into several
discrete metabolites that are excreted in the urine and, to a much lesser extent, in feces and
expired air. Two primary metabolic processes appear to be involved for AN: (1) interaction
with reduced glutathione (GSH) and (2) cytochrome P450 (CYP450) 2E1-mediated formation of
2-cyanoethylene oxide (CEO). Each product of these processes can undergo further metabolic
transformations. Important data elements that have contributed to the current understanding of
the toxicokinetics of AN are summarized in the following sections.
3.1. ABSORPTION
3.1.1. Studies in Humans
Jakubowski et al. (1987) administered AN via inhalation to six male volunteers for 8
hours at concentrations of either 5 or 10 mg/m3 from a chamber. Lung ventilation and retention
of AN in the lungs of these individuals were measured by determining the concentrations of AN
in the inhaled and expired air. A respiratory retention of 52% was estimated, based on 90
minutes to 8 hours of observation.
3.1.2. Studies in Animals
Most of the available information on the absorption of AN has come from studies in
experimental animals. Young et al. (1977) carried out a series of experiments in which
[1-14C]-AN was administered to male Sprague-Dawley rats via the oral, inhalation, or
intravenous (i.v.) route. Semi quantitative evidence of extensive absorption of AN has come
from an inhalation experiment of Young et al. (1977) in which four rats/group were exposed
nose only to 5 or 100 ppm AN vapor for 6 hours. Following the exposure, the animals were kept
in metabolism cages and excreta were collected for 220 hours. Total recovered doses estimated
from total recovered radioactivity for the low and high exposure levels were 0.7 and 10.2 mg/kg
[1-14C]-AN, respectively. Substantial amounts of the recovered dose were found in urine, feces,
and tissues, which indicated that AN has the capacity to be absorbed via the inhalation route
(Table 3-1).
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Table 3-1. Recovery of radioactivity from male Sprague-Dawley rats
exposed to 5 or 100 ppm [1-14C]-AN for 6 hours via inhalation
Site of recovery
Urine
Feces
14C02
Body
Cage wash
Total dose (mg/kg)
5 ppm
100 ppm
Percentage of recovered dose (mean ± standard deviation [SD])
68.50 ±9.38
3.94 ±0.97
6.07 ±1.58
18.53 ±4.68
2.95 ±3.95
0.7
82.17 ±4.21
3. 15 ±0.82
2.60 ±0.83
11.24 ±2.85
0.85 ±0.58
10.2
Source: Young etal. (1977).
In the gavage experiments, animals were kept in metabolic cages after a single dose of
either 0.1 or 10 mg/kg [1-14C]-AN (vehicle not stated), and the fate of the radiolabel was
monitored for 72 hours. As shown in Table 3-2, only about 5% of the administered radiolabel
was recovered in the feces after 72 hours. By contrast, most of the administered radiolabel was
recovered in the urine, carcass, and skin, with smaller fractions of the administered radiolabel
expired as 14CC>2. These data suggest that at least 95% of administered AN was absorbed.
Table 3-2. Recovery of radioactivity after a single gavage dose of 0.1 or
10 mg/kg [1-14C]-AN to male Sprague-Dawley rats
Recovery site
Urine
Feces
Expired CO2
Carcass
Skin
Total recovery
0.1 mg/kg
10 mg/kg
Percent of dose
34.22 ±6.26
5.36 ±1.43
4.56 ±1.82
24.24 ± 5.02
12.78 ±1.17
81.2
66.68 ±10.6
5.22 ±1.17
3. 93 ±1.79
16.04 ±1.87
10.57 ±4.55
102.4
Source: Young etal. (1977).
Ahmed et al. (1983) administered a single oral dose of 46.5 mg/kg AN to male Sprague-
Dawley rats in distilled water, using 50 uQ/kg of either [2,3-14C]- or [1-14C]-AN as a tracer.
Only 8-10% of administered radioactivity from the [2,3-14C]-AN dose had been excreted in
feces 72 hours after dosing, compared with 2% from orally administered [1-14C]-AN during the
same period. Thus, AN was readily absorbed at the gastrointestinal (GI) tract. The time course
of radioactivity released from feces showed a major spike of [2,3-14C]-AN-derived radiolabel
after 72 hours. The delayed release suggests that at least a portion of this released radioactivity
may have resulted from biliary elimination of absorbed AN. The major peak of released
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radiolabel from [1-14C]-AN was at 24 hours postexposure, possibly indicative of a different
metabolic fate for the nitrile portion of the molecule.
Kedderis et al. (1993a) administered [2,3-14C]-AN orally to male F344 rats (0.09-
28.8 mg/kg) and male B6C3Fi mice (0.09-10 mg/kg). Three to five percent of the administered
dose was recovered in the feces of rats after 72 hours. Between 2 and 8% of the dose was
recovered in the feces of male B6C3Fi mice. These data indicate the near-complete absorption
of the compound when administered via the oral route.
3.2. DISTRIBUTION
The toxicokinetic experiments of Young et al. (1977) provide data on the deposition of
radiolabel when male Sprague-Dawley rats were exposed to [1-14C]-AN via inhalation, gavage,
or i.v. injection. As shown in Table 3-1, an average of 11.24% of total recovered dose
(10.2 mg/kg [1-14C]-AN) was obtained in the tissues when a group of four rats was exposed to
100 ppm of [1-14C]-AN for 6 hours via inhalation, and excreta were collected for 220 hours.
Another group exposed to 5 ppm [1-14C]-AN had an average of 18.53% of the recovered
radiolabel (0.7 mg/kg [1-14C]-AN) deposited in the tissues. When the fate of radiolabel
administered by gavage was monitored, combining the recoveries of the administered
radioactivity in carcass and skin gave a value of 37% for tissue deposition at the lower AN
concentration (0.1 mg/kg) vs. 27% at the higher concentration (10 mg/kg) (Table 3-2), indicating
possible metabolic saturation at the higher dose. The percentage of expired CC>2 was lower for
the high-dose group than the low-dose group (3.93 vs. 4.56%), providing support for possible
metabolic saturation.
Young et al. (1977) examined the distribution of the radiolabel among the major organs
and tissues after oral and i.v. administration of [1-14C]-AN. Radioactivity was detected in
several tissues, including lungs, liver, kidneys, stomach, intestines, skeletal muscle, heart, spleen,
brain, thymus, testes, skin, carcass, and blood cells. Of these, the stomach, red blood cells
(RBCs), and skin appeared to be the most important deposition sites for radiolabeled AN or its
metabolites, regardless of the route of administration, dose level, or time. Radioactivity found in
the stomach was highest after either oral or i.v. route and was not due to unabsorbed AN since
similar results were obtained with either i.v. or oral dosing.
In a follow-up time course experiment, Young et al. (1977) administered 10 mg/kg
[1-14C]-AN intravenously via the tail vein to three male Sprague-Dawley rats and examined the
distribution of radioactivity to the stomach and adrenal glands. The level of radioactivity in the
stomach and its contents increased from 5 minutes to 24 hours postinjection. On the other hand,
radioactivity was high in the adrenal gland at 5 minutes but decreased 13-fold in 24 hours.
Radioactivity was observed in the bile of a single bile-duct-cannulated rat, indicating biliary
excretion. Maximal radioactivity in bile was observed after 15 minutes but declined at later time
points.
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When both sexes of adult Sprague-Dawley rats (including some pregnant animals) and
two cynomolgus monkeys were exposed orally (26 mg/kg) or intravenously (13 mg/kg) to single
doses of [1-14C]-AN, whole-body autoradiography from 20 minutes after injection primarily
showed radioactivity in the bile, intestinal contents, and urine (Sandberg and Slanina, 1980).
Other organs showing accumulation of isotope in the autoradiogram were blood, liver, kidney,
lung, and adrenal cortex, in which the activity declined slowly during 24 hours. There also was
uptake of the label in the stomach mucosa and hair follicles. Distribution of radioactivity in fetal
tissue following i.v. and oral administration to pregnant rats was uniform and showed a low
uptake compared to maternal tissue. An exception was found in the eye lens, in which the
radioactivity exceeded that of the maternal blood (Sandberg and Slanina, 1980).
Jacob and Ahmed (2003a) used whole-body autoradiography to examine the distribution
of 11.5 mg/kg [2-14C]-AN administered orally or intravenously to male F344 rats 5 minutes and
8, 24, and 48 hours postexposure. Levels of radioactivity per gram of tissue were highest
5 minutes after oral dosing for stomach lumen and mucosa, small intestine lumen and mucosa,
liver, nasal mucosa, spleen, and kidney; other tissues (including the lung, brain, spinal cord,
thyroid, and testis) had peak levels at 8 hours. Covalently bound radioactivity was detected
48 hours later in stomach mucosa, blood, and hair follicles. Five minutes following i.v.
administration, the highest levels of radioactivity per gram of tissue were detected in the lung,
liver, spleen, small intestine lumen, kidney, epididymis, and adrenal gland. At 24 hours, tissues
that showed peak levels included bone marrow, brain, lacrimal gland, and testis. Tissues with
the highest level at 24 hours included the lung, liver, and bone marrow. The levels of covalent
bound radioactivity (nCi/g) were higher 48 hours after i.v. exposure compared with oral
exposure, with bound radioactivity retained in liver, spleen, bone marrow, lung, kidney, and
adipose tissue ranging from 2.5 to 23 times higher following i.v. exposure. The only organs that
retained higher levels of covalently bound radioactivity 48 hours following oral exposure were
the stomach mucosa (3 x) and heart blood (5.7x). The total radioactive dose retained in animals
after i.v. and oral exposures were 70 and 38%, respectively. Jacob and Ahmed (2003a)
concluded that the metabolism and distribution of AN is greatly influenced by the portals of
entry, with a higher amount of AN metabolized and excreted following oral exposure compared
with exposure by i.v. injection. Rapid delivery of AN after i.v. treatment resulted in fast
conjugation and/or covalent interaction of the parent compound with biological molecules,
resulting in minimal metabolism and excretion in urine or feces.
Sapota (1982) administered 40 mg/kg AN in saline, containing either 40 uCi/kg
[1,2-14C]-AN or [1-14C]-AN to male Wistar rats, either via gavage or intraperitoneally. Tissue
distribution of radioactivity as measured by liquid scintillation counting after intraperitoneal
(i.p.) and oral administration showed that the highest specific radioactivity in tissue (nCi/g) was
found in RBCs, liver, and kidneys. Tissue-wide recovery of the radiolabel from either
[1,2-14C]-AN or [1-14C]-AN at 2, 8, and 24 hours after a single oral dose is shown in Table 3-3.
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Statistically significant differences were observed in the distribution of radioactivity from the
two forms of labeled AN in RBCs, plasma, liver, and kidney at 8 and 24 hours after
administration. More rapid loss of tissue radioactivity in the liver, kidneys, and brain was also
observed after oral administration of [1-14C]-AN than [1,2-14C]-AN. These results suggested
different pathways for disposition and biotransformation of the cyano and vinyl moieties of the
AN molecule.
Table 3-3. Percentage recovery of radioactivity in tissues of male Wistar
rats following a single oral dose of radiolabeled AN
Target organ/
tissue
RBCs
Plasma
Liver
Kidney
Spleen
Lung
Brain
Total
Recovered radioactivity (percent of dose in tissue)
Distribution from [1-14C]-AN
2h
5.36
2.63
6.13
1.17
0.22
0.36
0.25
16.1
8h
4.82a
4.00a
1.21a
0.27a
0.10
0.25
0.12a
10.8
24 h
5.45
0.50a
1.00a
0.15a
0.08
0.25
0.09
7.5
Distribution from [1,2-14C]-AN
2h
5.31
1.93
7.00
0.82
0.14
0.30
0.24
15.7
8h
7.27a
1.92a
5.98a
0.77a
0.17
0.27
0.25a
16.6
24 h
6.72
1.70a
2.67a
0.30a
0.10
0.11
0.12
11.7
a/> < 0.05; significant difference between [14CN]-AN and [1,2-14C]-AN at the same time point.
Source: Sapota (1982).
In an in vivo study on the interaction of orally administered 46.5 mg/kg [1-14C]-AN or
5 mg/kg K14CN with rat blood, Farooqui and Ahmed (1982) reported that up to 94% of 14C from
AN in RBCs was covalently bound to cytoplasmic and membrane proteins. On the other hand,
90% of the radioactivity from K14CN in erythrocytes was bound to the heme fraction of
hemoglobin (Hb), indicating that CN~ liberated from potassium cyanide (KCN) interacted with
heme. In addition, distribution of 14C from erythrocytes of rats treated with [1-14C]-AN showed
that more than 40% of total radioactivity was localized in membrane residue, 20-35% in the
globin fraction, and 11-25% in the heme fraction. In contrast, 70% of 14C from K14CN in red
cells was localized in the heme fraction, 14-25% in globin, and 5-10% in cell membrane. The
study authors concluded that KCN interacted with rat blood mainly through liberation of CIST,
which was bound to heme. Since AN was found to be mainly covalently bound to cell
membranes, AN might cause damage to RBCs by mechanisms other than the release of CIST.
In male Wistar rats that received 40 mg/kg AN (about half the median lethal dose [LD50])
by gavage in water, peak blood levels of AN (2 ug/mL) were detected 1.5 hours after dosing
(Shibata et al., 2004). By 2.5 hours after dosing, blood levels of AN had dropped to less than
0.2 ug/mL. Ten hours after dosing, AN was still detectable in blood at 50 ng/mL.
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Ahmed et al. (1983, 1982) studied the distribution of AN administered to male Sprague-
Dawley rats by gavage with a dose equivalent to one-half the LDso (46.5 mg/kg). In the first
experiment (Ahmed et al., 1982), the dose additionally contained 50 uCi/kg [1-14C]-AN. In the
second experiment (Ahmed et al., 1983), both [2,3-14C]-AN (both carbons in the vinyl moiety
were radiolabeled) and [1-14C]-AN were studied in rats administered an oral dose of 46.5 mg/kg.
Radioactivity from both forms of labeled AN or its metabolites initially was sequestered
mainly in the stomach and stomach content, followed by the rest of the GI tract, including small
and large intestines. The GI tract contained the highest levels of radioactivity up to 72 hours
before beginning to decline, suggesting that AN or its metabolites were re-secreted in the
stomach (Ahmed et al., 1983). In addition, radioactivity became widely distributed in all tissues
within 1-6 hours after dosing, with liver, kidney, and blood showing higher radioactivity than
muscle, fat, and bone (Ahmed et al., 1983, 1982). Heart, spleen, brain, and thymus showed
maximum concentrations between 3 and 6 hours. By 24 hours after administration, the levels of
radioactivity found in liver, kidney, and lung began to decline, resulting in 10- to several
100-fold reductions from peak concentrations over the course of the 10-day experiment.
However, in several tissues, the decline of radioactivity was much lower: at 10 days postdosing,
radiolabel concentrations had declined only 2.4-fold in skin, 2.9-fold in blood, 3.8-fold in spleen,
and 4.9-fold in eyes, as compared with peak levels.
Two differently labeled AN preparations were administered to rats to elucidate potential
differences in distribution and metabolism between the cyano and vinyl groups. One important
finding of these studies was that radioactivity in blood was predominantly in the RBCs,
especially for radioactivity from [2,3-14C]-AN. Radioactivity from [1-14C]-AN in plasma was
higher than that from [2,3-14C]-AN. For radioactivity from [1-14C]-AN in RBCs, 40% was
localized in membrane residue, 20-35% in the globin fraction, and 11-25% in the heme fraction.
In contrast, 50% of radioactivity from [2,3-14C]-AN was in the membrane fraction, 45% in the
globin fraction, and only a trace amount in the heme fraction.
In addition, compared to [1-14C]-AN administered to animals, the percentage of covalent
binding of [2,3-14C]-AN to proteins was significantly higher even 72 hours after dosing.
Subcellular distribution of radioactivity from [2,3-14C]-AN was also different from that derived
from [1-14C]-AN. For [2,3-14C]-AN, the cytosol fraction attained the lowest covalent protein
binding in tissues. The percentage of covalently bound radioactivity in tissues relative to the
total increased four- to fivefold over that of the 1-hour level. At 72 hours after administration,
the highest bound radioactivity was in the mitochondrial fractions of kidney, spleen, lung, and
heart. However, in liver, the microsomal fraction contained the highest radioactivity (Ahmed et
al., 1983).
For [1-14C]-AN, covalent binding to macromolecules in tissues remained unchanged over
time. Cytosol contained the highest levels of total radioactivity in the six tissues (liver, kidney,
spleen, brain, lung, and heart) selected for the study of subcellular distribution. Twenty to 40%
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of total radioactivity was bound to nuclear, mitochondrial, or microsomal fractions. Only 6-14%
of total radioactivity was bound to cytosol over 6 hours (Ahmed et al., 1982).
Farooqui and Ahmed (1983a) demonstrated the irreversible binding of radiolabel from
[2,3-14C]-AN to proteins and nucleic acids in vivo. AN was administered by gavage as a bolus
dose of 46.5 mg/kg (one-half LDso) in distilled water to male Sprague-Dawley rats (3-4/group).
Proteins were extracted by chloroform-isoamyl alcohol-phenol, and ribonucleic acid (RNA) and
deoxyribonucleic acid (DNA) were isolated by hydroxyapatite chromatography. Protein binding
at 1 hour after dosing was highest in spleen and stomach, followed by liver, brain, and kidney.
Protein binding in spleen and stomach declined after 1 hour, whereas binding in liver, kidney,
and brain increased. At 6 hours after dosing, protein binding fell to lower values in spleen and
stomach but increased in other tissues. Binding plateaued in all tissues between 6 and 48 hours,
with levels in spleen > liver > stomach > kidney > brain. Binding to RNA was highest in liver,
stomach, and brain, with liver attaining a maximum by 6 hours and stomach and brain by
24 hours. The subsequent decline until 48 hours was also slow. DNA binding did not reach a
maximum until 24 hours after dosing, with levels in brain > stomach > liver. Again, the decline
of DNA-bound radioactivity during the following 24 hours was slow. Binding of AN to DNA in
brain, stomach, and liver was 56, 45, and 5 umol AN per mol DNA, respectively, at 24 hours.
The study authors also calculated a covalent binding index for DNA, defined as the ratio of
(umol AN bound per mol DNA) to (mmol AN applied per kg body weight [BW]). The values
for brain, stomach, and liver were 65, 52, and 6, respectively.
Silver et al. (1987) examined the distribution of 100 mg [1-14C]-AN in female Sprague-
Dawley rats after i.v. injection to investigate why this procedure induced acute hemorrhagic
necrosis of the adrenal gland 2 hours after administration and why this damage was more
prominent with i.v. injection than with oral administration. Total radioactivity was found to be
highest in the blood, liver, kidney, duodenum, and adrenals 15-90 minutes following i.v.
injection of [1-14C]-AN . (This result was largely in agreement with the whole-body
autoradiographic findings of Sandberg and Slanina [1980] in rats and monkeys.) Total radiolabel
in blood increased over this time period, whereas the total radiolabel in other organs remained
constant or decreased with time. The level of covalently bound radiolabel in the adrenals was
lower than that observed in blood, liver, kidney, forestomach, and glandular stomach.
Silver et al. (1987) also administered the same dose of [1-14C]-AN in water by gavage to
female Sprague-Dawley rats (four/group) and investigated the distribution of radiolabel up to
24 hours after dosing. Total radioactivity was highest during the first 8 hours after dosing in the
blood, GI tract, liver, and kidney. At 24 hours after dosing, the highest total radioactivity was
found in the blood, forestomach, and glandular stomach. The highest level of covalently bound
radioactivity was found in these same tissues during the first 8 hours after dosing and remained
highest in the blood and forestomach at 24 hours after dosing. The study authors concluded that
their observations would not support a role of covalent binding in the hemorrhagic effect of AN
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on the adrenals. Rather, the initial high concentrations of radiolabel from AN might play a role
in the action of AN on the adrenal gland. However, it should be noted that in both studies, the
cyano carbon, not the vinyl carbons, was labeled.
The rapidity with which AN was biotransformed and distributed to the brain as its
epoxide metabolite, CEO, was reported in two studies (Kedderis et al., 1993b; Roberts et al.,
1991). Roberts et al. (1991) administered 4 mg/kg AN to male F344 rats and B6C3Fi mice
(three/group) and measured CEO levels in blood at 0.5, 1, 4, or 24 hours after dosing. Higher
levels of CEO were found in rat blood than in mouse blood. In addition, CEO was cleared from
mouse blood in 4 hours, but was cleared in rat blood in 24 hours. The dose dependence of CEO
concentrations in blood was also evaluated. Blood CEO was measured 0.5 hours after oral
dosing in F344 rats given either 0, 1, 4, 10, or 30 mg/kg AN and in B6C3Fi mice given either 0,
1, 4, 8, or 10 mg/kg AN. Blood CEO concentrations increased with dose in rats and mice but at
higher concentrations in rats at the same doses.
In the first experiment by Kedderis et al. (1993b), three male F344 rats and three male
B6C3Fi mice were administered 10 mg/kg AN in water by gavage. The rats were sacrificed
10 minutes after dosing, while the mice were sacrificed 5 minutes after dosing. CEO
concentrations from blood and brains of rats and mice were measured. Higher CEO
concentrations were found in the blood and brains of rats than in mice (13% higher in blood and
23% higher in brain). In addition, CEO concentration in rat blood 10 minutes after oral
administration was about twice the concentration previously reported by Roberts et al. (1991) at
30 minutes after oral dosing of 4 mg/kg AN to three male F344 rats. On the other hand, CEO
concentration in mice 5 minutes after oral administration was about 10 times higher than that
reported at 30 minutes after oral dosing of 4 mg/kg AN to three male B6C3Fi mice (Roberts et
al., 1991). These results suggested that CEO was rapidly cleared in both rats and mice and that
the clearance of CEO in mice was more rapid than in rats.
Kedderis et al. (1993b) also administered 3 mg/kg [2,3-14C]-CEO orally to F344 rats and
B6C3Fi mice to determine the tissue distribution of radioactivity from labeled CEO after 2 and
24 hours. Radioactivity from labeled CEO was widely distributed in major organs of rats and
mice 2 hours after administration, with the highest level of radioactivity found in the stomach
and intestines of rats and mice. However, radioactivity detected in the stomach and intestines of
mice was only about 15 and 40%, respectively, of that detected in rats, suggesting that mice
absorbed CEO more rapidly than rats (Kedderis et al., 1993b). By 24 hours, radioactivity
decreased by 71-90% in all tissues, including the brain, liver, and lung. Stomach and intestines
continued to retain the highest level of radioactivity, probably due to covalent binding of CEO to
macromolecules in these organs.
Burka et al. (1994) monitored the tissue distribution of radiolabel derived from
[2-14C]-AN after oral dosing at 0.87 mmol/kg (46 mg/kg) to untreated, phenobarbital
(PB)-pretreated, or SKF 525A pretreated male F344 rats (three/group). PB induces a number of
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CYP450 isozymes, including CYP2B1 and CYP2B2, whereas SKF 525A is a general inhibitor
of CYP450. After 24 hours, about 10% of the administered dose in untreated rats was present in
the blood with a further 4% sequestered in the tissues. In PB-pretreated rats, a 40% increase in
AN-derived radioactivity was found in both the liver and glandular stomach when compared
with rats treated with AN alone, with no changes in other tissues. However, AN-derived
radioactivity in most tissues of SKF 525A pretreated rats were up to 278% higher, suggesting the
involvement of CYP450 metabolism in the disposition of AN and/or its metabolites. (AN reacts
more rapidly with tissue nucleophiles than CEO; hence, decreasing its oxidative metabolism to
CEO would increase tissue binding of radiolabel.) Because PB pretreatment had little effect on
tissue distribution, the isoforms of CYP450 induced by PB are probably not the ones involved in
the metabolism of AN. It is known that AN is metabolized by CYP2E1, not CYP2B1 or
CYP2B2, to CEO (see Section 3.3). PB might increase AN-derived radioactivity in the liver and
stomach by inducing other enzymes, such as nicotinamide adenine dinucleotide phosphate
(NADPH)-cytochrome CYP450 reductase.
Ahmed et al. (1996a) monitored the tissue distribution of AN-derived radioactivity in
F344 rats (four/group) up to 48 hours following i.v. injection of 11.5 mg/kg of [2-14C]-AN
(50 uCi/kg). The study authors used whole-body autoradiography to chart a time course of tissue
deposition and obtained the highest levels of activity in lung (998 nCi/mg), intestinal contents
(752 nCi/mg), liver (713 nCi/mg), and spleen (539 nCi/mg) 5 minutes after dosing. Other tissues
with high radiolabel at this time point were the kidney (283 nCi/mg), epididymis (266 nCi/mg),
adrenal gland (241 nCi/mg), intestinal mucosa (245 nCi/mg), heart-blood (166 nCi/mg), bone-
marrow (178 nCi/mg), thyroid (121 nCi/mg), adipose tissue (169 nCi/mg), and lacrimal gland
(122 nCi/mg), while the brain, spinal cord, and testis had the lowest levels of radioactivity. At
8 hours after dosing, the contents of the large intestine, especially the cecum, had the highest
level of radioactivity (852 nCi/mg). Radioactivity in brain (92 nCi/mg), lacrimal gland
(294 nCi/mg), and thyroid (211 nCi/mg) peaked at 24 hours after dosing, while radioactivity
level in bone marrow (698 nCi/mg) peaked at 48 hours. Covalent bound radioactivity, as
determined after acid-extraction techniques on freeze-dried sections, was observed in the spleen,
liver, bone marrow, and lung.
3.3. METABOLISM
A proposed scheme for the metabolic pathways of AN in mammals is shown in
Figure 3-1. The scheme has been developed as a result of studies on the identification of urinary
metabolites following acute exposure (Fennell and Sumner, 1994; Kedderis et al., 1993a; Fennell
et al., 1991; Turner et al., 1989; Muller et al., 1987; Tardiff et al., 1987; Gut et al., 1985;
Kopecky et al., 1980; Langvardt et al., 1980), measurements of the in vivo modulation of
AN-induced toxicological and biochemical changes by enzyme inhibitors or inducers (Wang et
al., 2002; Sumner et al., 1999; Burka et al., 1994; Kedderis et al., 1993c; Pilon et al., 1988a, b;
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Ghanayem and Ahmed, 1986; Ahmed and Abreu, 1981; Abreu and Ahmed, 1980), determination
of the subcellular distribution of metabolic products of AN (Nerland et al., 2001; Ahmed et al.
1996a), and analysis of metabolites formed in vitro by subcellular fractions of liver, lung, and
kidney in response to AN administration (Mostafa et al., 1999; Kedderis et al., 1995; Kedderis
and Batra, 1993;, Roberts et al., 1991, 1989; Hogy, 1986; Geiger et al., 1983; Ahmed and Abreu,
1981; Guengerich et al., 1981; Abreu and Ahmed, 1980).
RSCH2CH2CN—
HOOC CH2—S CH2CH2CN giutathon,
S-(2-cyanoethyl)thioacetic acid
GS CH2CH2CN
S-(2-cyanoethyl)glutathione
=CH— C=
Acrylonitrile
S-transferase
CYP2E1
Ac-N-Cys S CH2CH2CN
N-acetyl-S-(2-cyanoethyl)cysteine
H
NC
rhodanese
CH2 OH + CN -SCN
CHO Glycol Cyanide Thiocyanate
epoxide/ aldehyde
— / hydrolase
CH2 CH—CN
2-Cyanoethylene
oxide GS—CHCN
I CH2OH
Ac-N-Cys^S—CHCN
N-acetyl-S-(l-cyano-2-hydroxyethyl)cysteine
i^r^n
-X CH2OH
rhodi
rnodanese
Cyanide
- SCN
Thiocyanate
CH2
Cyanoethanol
CH2
Cyanoacetic acid
OH
GS—CH2—CHCN
Cyanohydrin
GS—CH2CHO - Ac-N-Cys S CH2CH2OH
S-2-oxoethylglutathione N-acetyl-S-(2-hydroxyethyl)cysteine
Cys S—CH2COOH ^^-
S-(carboxymethyl)cysteine HOOC CH2 S CH2COOH
I Thionyldiacetic acid
Ac-N-Cys -^-S CH2COOH
N-acetyl-S-(carboxymethyl)cysteine
Source: National Toxicology Program (NTP) (2001).
Figure 3-1. Scheme for the metabolic transformation of AN.
Primary components of the scheme are formation of CEO by the action of mixed function
oxidases (predominantly CYP2E1), detoxification of AN by interaction with GSH, and covalent
binding of AN reactive metabolite to other biological macromolecules.
In rats, CYP2E1 is present in the liver and widespread among other tissues. This isoform
can be induced in the tongue, esophagus, forestomach squamous epithelia (Shimizu et al., 1990),
and intestinal mucosa (Subramanian and Ahmed, 1995). It is present, albeit at levels 10-
20 times lower than in liver, in such tissues as kidney and lung. It is also present in the brain
(Geng and Strobel, 1993; Sohda et al., 1993). The formation of CEO has important implications
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for the toxicity of AN, because the intermediate has been proposed as the principal carcinogenic
metabolite of AN. However, CEO can undergo a number of further transformations. These
include the interaction with GSH to form a series of cysteine or N-acetyl cysteine derivatives and
the production of cyanide via the action of epoxide hydrolase (EH). Subsequent detoxification
of cyanide to thiocyanate is thought to occur under the action of rhodanese (Kopecky et al.,
1980).
3.3.1. Oxidation of AN to CEO
Evidence for the oxidation of AN to CEO and its subsequent transformations came from
a number of studies. Abreu and Ahmed (1980) studied the in vitro conversion of AN to cyanide
in subcellular fractions of liver from Sprague-Dawley rats (also reported in Ahmed and Abreu
[1981]). The metabolic activity was localized in the microsomal fraction and required NADPH,
MgCb, and oxygen for maximal activity. Determination of the kinetic parameters (Km and Vmax)
of the transformation of AN to cyanide pointed to a higher affinity and faster rate of product
formation in microsomes from rats pretreated with CYP450-inducing agents, such as
Aroclor 1254 and PB. (Six AN concentrations ranging from 10 to 300 mM were used for each
preparation.) The Km values calculated for the PB and Aroclor 1254 preparations were 54.8 and
40.9 mM, respectively, and were lower than the control (190 mM). Pretreatment of rats or
addition to incubation mixtures with agents that inhibit CYP450 activity, such as SKF 525A or
cobalt chloride, reduced the amount of cyanide formed by rat liver microsomes.
Abreu and Ahmed (1980) studied the effect on cyanide formation when 1,1,1-trichloro-
propane-2,3-oxide (TCPO), a specific inhibitor of EH, was added to the microsomal incubation
mixtures. Production of cyanide was dose-dependently reduced to 19% of control levels at
TCPO concentration of 1 x 10"2 M. Abreu and Ahmed (1980) also tested the effect of sulfhydryl
compounds on microsomal metabolism of AN, as measured by the rate of cyanide formation.
Only cysteamine decreased cyanide formation; other sulfhydryl compounds, including GSH and
cysteine, enhanced the rate of cyanide formation from AN.
Abreu and Ahmed (1980) suggested that probably more than one step was involved in the
enzymatic conversion of AN to cyanide. The study authors proposed the initial product of AN
oxidation to be CEO, which could then undergo a number of alternative transformations, one of
which would be non-enzymatic conversion to cyanide. That another transformation might
involve EH was indicated by the decrease in cyanide formation following administration of the
inhibitor TCPO. Because cyanide production was enhanced in the presence of sulfhydryl
compounds, such as GSH, chemical interaction of CEO with GSH could lead to formation of
cyanohydrin. Rearrangement of this cyanohydrin to an aldehyde could result in the release of
cyanide. Cysteamine might diminish cyanide formation due to its inhibition of CYP450-
dependent metabolism (Buckpitt et al., 1979).
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Geiger et al. (1983) studied the conversion of AN to its metabolic products in isolated rat
hepatocytes and demonstrated the formation of CEO and its hydrolysis to cyanide, which itself
was transformed and detected as thiocyanate.
In vitro incubation of AN with liver microsomes isolated from male F344 rats pretreated
with inhibitors or inducers of specific members of the CYP450 family of mixed function
oxidases indicated that CYP2E1 is the major catalyst in the oxidation of AN to CEO (Kedderis et
al., 1993c). The rate (Vmax) at which rat liver microsomes oxidized AN to CEO was increased
more than fivefold from Vmax of 366 pmol CEO/minute-mg for untreated rats following acetone
pretreatment, although Km was increased from 11 to 19 uM. Because acetone is a potent inducer
of CYP2E1, the data suggest that this isoform is a primary catalyst of AN epoxidation in rats.
Treatment with p-naphthoflavone to induce CYP1 Al and CYP1A2 or with dexamethasone
(DEX) to induce the CYP3 A enzymes increased Vmax only less than twofold, but Km was
increased by 3.5 and 5.2-fold, respectively, for AN epoxidation in these two pretreatment
systems. Treatment with PB to induce CYP2B1 and CYP2B2 slightly decreased the Vmax but
increased the Km in microsomes from rats. These studies demonstrated that other forms of
CYP450 (CYP2B1, CYP2B2, and the 3 A enzymes) can oxidize AN but with specific activities
much lower than CYP2E1.
The effect of a number of CYP450 inhibitors on epoxidation of 1.2 mM AN by rat
hepatic microsomes was investigated (Kedderis et al., 1993c). Neither SKF 525A nor
metyrapone were effective inhibitors, retaining 87% of control activity. After treatment of rats
with DEX or PB, SKF 525 A became a more effective inhibitor, retaining 45 and 47% of control
activity. Metyrapone also became a more effective inhibitor of epoxidation of AN after DEX
treatment. The CYP450 ligand, 1-phenylimidazole, was a potent inhibitor of AN epoxidation
(4% of control activity). Chlorzoxazone (27%), ethanol (42%), and diethyldithiocarbamate
(17%) also inhibited this pathway. The changes in the degree of inhibition of epoxidation of AN
following DEX and PB treatments could be interpreted as multiple CYP450 enzymes from rat
hepatic microsomes were capable of oxidizing AN.
Antibodies to CYP2E1 (sheep or goat anti-rabbit CYP2E1) inhibited more than 85% of
AN epoxidation in liver microsomes from untreated or acetone-treated rats but only 40 and 60%
inhibition following DEX and PB treatment, respectively (Kedderis et al., 1993c), suggesting
that CYP450 enzymes other than CYP2E1 might participate in the epoxidation. However, it
should be noted that the AN concentration in these in vitro studies was high (1 mM). Forms of
CYP450 enzymes other than CYP2E1 might have been recruited to AN metabolism.
Kedderis et al. (1993c) also investigated the kinetics of the epoxidation of 1.2 mM AN by
using human hepatic microsomes. Km and Vmax values for oxidation of AN by liver microsomes
from six uninduced individuals ranged from 12 to 18 uM and from 129 to 315 pmol/minute-mg,
respectively. Antibodies to CYP2E1 produced 58-70% inhibition of AN epoxidation catalyzed
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by the six human liver microsomal preparations. This suggests that, while CYP2E1 was the
major catalyst of AN epoxidation in humans, other isoforms of CYP450 may also be involved.
Sumner et al. (1999) investigated the role of CYP450 in the metabolism of AN in mice.
Three male wild-type (WT) mice and three to four male CYP450 2El-null mice were treated
orally with either 2.5 mg or 10 mg/kg of [1,2,3-13C] AN. Urinary metabolites in samples
collected over 24-h were characterized using 13C nuclear magnetic resonance (NMR). In WT
mice, urinary metabolites of CEO predominated, with metabolites derived from GSH
conjugation at the 3-carbon of CEO accounting for 67-71%. Metabolites from GSH conjugation
at the 2-carbon accounted for about 13%. Metabolites from direct GSH conjugation with the
parent compound, AN, accounted for 15-21%. In the urine of CYP2El-null mice, however,
only metabolites from direct GSH conjugation were detected. Sumner et al. (1999) interpreted
their data as indicating that CYP2E1 may be the only CYP450 involved in the metabolism of AN
in mice.
Subramanian and Ahmed (1995), attempting to characterize the specific intestinal
toxicity of AN, incubated microsomes isolated from male Sprague-Dawley rat intestinal mucosa
in vitro with AN in the presence of NADPH. AN metabolism to cyanide was enhanced by the
addition of sulfhydryl compounds such as GSH, cysteine, and D-penicillamine. AN metabolism
to cyanide was also enhanced following the induction of microsomal proteins (MPs) by treating
rats with PB (inducer of CYP2B1), p-naphthoflavone (inducer of CYP1A1), and
4-methylpyrazole (inducer of CYP2E1). AN metabolism to cyanide was inhibited to 8 and 20%
of control, respectively, when dimethyl sulfoxide (DMSO) or ethanol (competitive inhibitors of
CYP2E1) was added to the incubation mixtures.
Subramanian and Ahmed (1995) showed that the intestinal CYP450 isoform had a high
affinity for AN, with a Km of 1.1 uM and a Vmax of 1,250 pmol/mg protein/minute. Addition of
DMSO in varying concentrations (final 30 mM) increased the Kmof the reaction to 10 uM, but
Vmax remained unchanged. Since DMSO is a specific substrate and competitive inhibitor for
CYP2E1, these studies indicated that CYP2E1 was the main CYP450 isoform that bio-activates
AN in the intestine. In addition, anti-P450 3a immunoglobulin (Ig)G (which cross-reacts with rat
CYP2E1) caused a concentration-dependent inhibition of the metabolism of AN to cyanide in the
ethanol-induced intestinal microsomes (Subramanian and Ahmed, 1995). These results showed
that CYP2E1 was the main intestinal mucosa enzyme metabolizing AN to cyanide.
Similarly, Abdel-Aziz et al. (1997) demonstrated the metabolism of AN to cyanide when
AN was incubated in vitro with NADPH and a microsomal fraction prepared from Sprague-
Dawley rat testis. The Vmax of this reaction was 65 pmol CNVmg protein/minute, and the Km
was 88.6 uM AN. Addition of SKF 525A or benzimidazole (competitive inhibitors of CYP450)
to the incubation mixture inhibited the formation of cyanide, whereas microsomes obtained from
PB-treated rats increased activation of AN to cyanide. Thus, AN was metabolized in rat testis
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via CYP450 mixed function oxidase. Addition of GSH, L-cysteine, D-penicillamine, or
2-mercaptoethanol also enhanced the release of cyanide from AN.
The capability of rat kidney to metabolize AN to cyanide was demonstrated by Mostafa
et al. (1999) in an in vitro study that investigated the mechanism by which AN caused renal
toxicity. In renal subcellular fractions from Sprague-Dawley rats, the metabolism of AN to
cyanide was highest in the microsomal fraction. An NADPH-generating system in the presence
of magnesium ions was required for maximal activity. The Vmax of this reaction was 118 pmol
CNVmg protein/minute, and the Km was 160 jiM AN. Metabolism of AN to cyanide was
increased when microsomes were obtained from PB-, ethanol-, 4-methylpyrazole-, and
3-methylcholanthrene-treated rats. On the other hand, addition of SKF 525A or benzimidazole
to the incubation mixture inhibited AN metabolism. These data suggested that AN was
metabolized in the kidney via a CYP450-dependent mixed function oxidase system. Addition of
GSH, L-cysteine, cysteamine, D-penicillamine, or 2-mercaptoethanol to the incubation mixture
enhanced AN metabolism.
Ahmed and Patel (1981) carried out a series of single-dose gavage experiments on male
Sprague-Dawley rats and male Swiss mice at fractions of the LDso values of AN and KCN. (The
LDso of AN is 93 mg/kg in rats and 27 mg/kg in mice; the LDso of KCN is 10 mg/kg in rats and
8.5 mg/kg in mice.) Cyanide was measured and detected in blood, liver, kidney, and brain of
both rats and mice 1 hour after administration in a dose-dependent manner. However, cyanide
concentrations from metabolism of one LD50 AN in blood and tissues of rats were significantly
lower than those produced from one LDso of KCN. On the other hand, comparable
concentrations of cyanide in blood and tissues were observed after one LDso AN or KCN was
administered to mice. Blood and liver contained higher amounts of cyanide per unit volume than
kidney and brain (the other two organs evaluated).
Observed signs of toxicity were also different in rats and mice administered an LD50 of
AN. Rats developed severe cholinomimetic signs including salivation, lacrimation, diarrhea,
wheezing on expiration, and peripheral vasodilatation within 10 minutes after administration of
93 mg/kg AN. These signs were not observed in rats treated with KCN. Severe central nervous
system (CNS) effects such as depression, convulsions, and asphyxia were observed in rats 10-
20 minutes after treatment with 10 mg/kg KCN (LDso). These CNS signs of cyanide toxicity
were observed in AN-treated rats 2-3 hours after dosing. No physiological adverse effects were
observed in rats receiving 0.25 LDso AN. Mild salivation, diarrhea, and vasodilation were
observed after one-half LD50 AN in rats. However, in mice treated with equitoxic dose (LD50
27 mg/kg) AN, CNS signs identical to those observed after KCN was administered were
observed. These results demonstrated species differences in the toxicity and metabolism of AN.
Moreover, Ahmed and Patel (1981) showed that pretreatment of rats with Aroclor 1254
and PB increased AN metabolism to cyanide in rats, and pretreatment of rats with CoCb or SKF
525 A decreased blood cyanide concentrations. These results showed that the AN
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transformations demonstrated by Abreu and Ahmed (1980) in vitro also could take place in vivo.
In addition, increased metabolism of AN to cyanide would increase CNS effects. However,
acute AN toxicity was also manifested as cholinomimetic signs, which were not from cyanide.
Shibata et al. (2004) employed headspace gas chromatography to simultaneously measure
blood levels of AN and its metabolite, hydrogen cyanide, following oral administration to rats.
Plasma and urinary thiocyanate concentrations were also measured by the colorimetric method.
Male Wistar rats that received 40 mg/kg AN (about half the LDso) by gavage in water showed
toxic signs such as tachycardia 1 hour later. Peak blood levels of AN (2 ug/mL) and cyanide
(0.7 ug/mL) were detected 1.5 hours after dosing. By 2.5 hours after dosing, blood levels of AN
had dropped to less than 0.2 ug/mL and blood levels of cyanide decreased to 0.1 ug/mL; at that
time, thiocyanate was detected in plasma (20 ug/mL). Plasma thiocyanate concentrations rose
over time, peaking at 5 hours (31.3 ug/mL). At the same time, excretion of thiocyanate in urine
began to increase significantly. Ten hours after dosing, AN was still detectable in plasma at
50 ng/mL, but cyanide had decreased to a background level of about 5 ng/mL. The cumulative
urinary elimination of thiocyanate gradually increased, and at 10 hours, about 1.2 mg thiocyanate
was excreted into the urine. This amount was calculated to be 7% of the total administered AN.
Urinary AN level was not measured.
The capacity for formation of CEO from AN has been demonstrated in F344 rat liver
microsomes, lung microsomes, and isolated lung cells. The rate of CEO formation in rat lung
was cell specific, with the Clara cell-enriched fraction having a rate of CEO formation 7 times
greater than other cell fractions (Roberts et al., 1989). The overall rate of CEO formation was
about 15 times greater in the livers than the lungs (Roberts et al., 1989).
Roberts et al. (1991) provided data on the kinetics of CEO formation in liver and lung
microsomes isolated from male F344 rats, B6C3Fi mice, and humans (Table 3-4). While CEO
was produced in vitro by lung and liver microsomes in both rats and mice, the metabolite was
produced at a greater rate in liver compared with lung and in mice vs. rats. These data
potentially implicated the liver as the primary site of CEO formation after oral challenge with
AN but suggested differences in the kinetics of CEO formation between species. The rate of
CEO formation in microsomes isolated from human livers was comparable to that of F344 rats,
but about 4 times lower than that of B6C3Fi mice. The average rate of CEO formation in liver
microsome samples from six human donors was 501 ±112 pmol/minute-mg protein. (The
almost eightfold variation in enzymatic activity among these human samples appeared to
correlate with the amount of CYP450 in each preparation.) However, after oral administration of
AN, the concentration of CEO in mouse blood was about one-third that in rat blood at all doses
and time points tested. Thus, blood CEO concentration did not correlate with rate of microsomal
CEO formation, suggesting that species differences in detoxification of CEO might play a role in
determining CEO concentrations in blood. The single human lung microsome sample tested in
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the study formed 0.55 pmol CEO/minute-mg protein, which was much lower than that for rat
liver or lung microsomes.
Table 3-4. Apparent kinetic parameters of CEO formation from AN in
B6C3F1 mice, F344 rats, and humans
Tissue
Liver
Lung
Species
Mouse
Rat
Human
Mouse
Rat
Human
Vmax (pmol/min-mg protein)
2,801a
667a
501b
570a
45a
0.55C
Km (uM)a
67
52
Not available
1,229
1,854
Not available
aValues are the mean of eight replicates.
bValue is the mean of six human donors.
°Value is from one human donor.
Source: Roberts etal. (1991).
Guengerich et al. (1981) used a reconstituted enzyme system containing purified rat liver
CYP450 and NADPH-P450 reductase and a NADPH-generating system to oxidize AN in vitro
to a metabolite that they identified colorimetrically as CEO. The extent of CEO accumulation
was decreased by the addition of purified rat liver EH to the incubation medium. When 0.5 mM
CEO was incubated with 30 ug/mL purified EH, the rate that CEO was hydrolyzed was
5.5 nmol/minute. The rate of disappearance of CEO (due to nonenzymatic hydrolysis) was
1.7 nmol/minute in the absence of EH or in the presence of inactivated EH. HCN was released at
a rate of 1.5 nmol/minute during the hydrolysis of CEO by EH. (The rate of HCN release was
0.2 nmol/minute in the absence of EH.) The study authors suggested that the reason the HCN
release was not stoichiometric with epoxide disappearance might be due to a finite level of
cyanohydrin existing in solution. A Km of 0.8 mM and a Vmax of 300 nmol/minute-mg based on
disappearance of CEO was estimated for the hydrolysis of CEO by purified rat liver microsomal
EH. The half-life of CEO in 0.1 M potassium phosphate was estimated to be about 2 hours at
37°C (Guengerich et al., 1981).
Kopecky et al. (1980) also investigated the role of EH and the generation of hydrogen
cyanide from AN metabolism. AN was incubated in vitro with liver microsomes isolated from
female Wistar rats, with and without cofactors (NADP, Mg2+, etc.) for 60 minutes. After the
incubation, the mixtures were adjusted to either pH 1.8 (acidic processing) or 6.3 (alkaline
processing). Cyanide release, in the presence of cofactors, was found to be fourfold higher under
alkaline processing condition than acidic processing condition. The role of EH in AN
metabolism in rats was supported by the increase in cyanide release after alkaline processing,
indicating the existence of a cyanohydrin intermediate (glycolaldehyde cyanohydrin) in the
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biotransformation of AN. (Cyanohydrins generally decompose spontaneously to hydrogen
cyanide and a carbonyl compound at pH higher than 7.) Moreover, when 3,3,3-trichloro-
propylene oxide (TPO), a potent inhibitor of EH, was added to the incubation mixture, the
conversion of AN to cyanide was decreased by 70%. This result also provided evidence for the
participation of the cyanohydrin in AN metabolism because the hydration of CEO to
glycolaldehyde cyanohydrin was significantly inhibited by TPO.
Kedderis and Batra (1993) compared the rates of CEO hydrolysis, enhanced by liver
cytosol and microsomes from rats, mice, and humans, to the background rate of non-enzymatic
hydrolysis displayed by the chemical. [2,3-14C]-CEO was incubated at pH 7.3 and 37°C for
5 minutes with liver cytosol or microsomes. [2,3-14C]-CEO and its hydrolysis products were
separated by high performance liquid chromatography (HPLC). The identity of the hydrolysis
products could not be determined and did not correspond to aldehydes. Human hepatic
microsomes enhanced the formation of hydrolysis products of CEO, whereas both human hepatic
cytosol and liver cytosol and microsomes from F344 rats and B6C3Fi mice had no effect on
hydrolysis product formation from CEO. The study authors concluded that rodent hepatic
microsomal and cytosolic EHs were not active toward CEO, contrary to conclusions developed
by Guengerich et al. (1981) on rat purified microsomal EH and the conclusions by Kopecky et al.
(1980).
One possible explanation that EH activity was not observed by Kedderis and Batra
(1993) in microsomes from rats and mice was that their enzymatic reactions had not been
optimized. No cofactors were used in the incubation mixture, and the incubation duration was
only for 5 minutes. Both Kopecky et al. (1980) and Guengerich et al. (1981) used an NADPH-
generating system in their incubation mixture.
Kedderis and Batra (1993) also showed that the heat-labile human EH activity was
inhibited by the specific inhibitor, 1,1,1-trichloropropene oxide (median inhibitory concentration
[ICso] of 23 uM), indicating that EH was the catalyst in the hydrolysis of CEO. The half-life of
CEO in sodium phosphate buffer (pH 7.3), as estimated from hydrolysis by human liver
microsomes, was 99 minutes. Estimated Km using liver microsomes from six individuals ranged
from 0.6 to 3.2 mM. Vmax ranged from 8.3 to 18.8 nmol hydrolysis products/minute-mg protein.
The affinity of the human liver microsomal EH for CEO was relatively low, suggesting that the
contribution of the hydrolysis pathway to the clearance of CEO would be small at low substrate
concentrations. Increase in microsomal hydrolysis was observed after treatment of mice and rats
with PB or acetone, suggesting that CEO hydrolysis is inducible. However, treatment of rats and
mice with AN did not induce hepatic EH activity towards CEO (Kedderis and Batra, 1993).
Studies that demonstrated that EH is present in rats and humans are also available.
Immunoblot analysis of MPs was used by de Waziers et al. (1990) to measure EH in different
organs and tissues of rats and humans. They reported that EH occurred in rat liver microsomes
at 165 ug/mg protein and in human liver microsomes at 170 ug/mg protein. Therefore, the
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concentrations of EH in MP are similar in rats and humans. Guengerich et al. (1979) also
reported that multiple forms of EH exist in rats and humans.
EH activity was demonstrated in mice in a recent study, contrary to the conclusion by
Kedderis and Batra (1993). El Hadri et al. (2005) demonstrated that microsomal EH was present
in WT mice, which metabolized AN administered by gavage to cyanide in a dose- and time-
dependent manner. Blood cyanide levels in microsomal EH-null mice treated with a gavage
dose of 0.047-0.38 mmol/kg AN were lower than levels in similarly treated WT mice. Blood
cyanide level was also largely abolished in CYP2El-null mice and in WT mice pretreated with a
nonselective CYP inhibitor, 1-aminobenzotriazole (ABT), confirming that CYP2E1 was the key
enzyme for the epoxidation of AN and the subsequent formation of cyanide. AN-treated
CYP2E1- and mEH-null mice showed less severe symptoms of cyanide poisoning (labored
breathing, lethargy, and trembling) than similarly treated WT mice (El Hadri et al., 2005).
Significantly higher levels of AN-derived blood cyanide levels were observed in male mice than
in female mice, suggesting gender-related differences in toxicity. Western blot analysis also
demonstrated that expression of soluble EH was greater in male than female mice.
Detection and identification of urinary metabolites of AN from male F344 rats or B6C3Fi
mice exposed orally to [1,2,3- 13C]-AN (10 or 30 mg/kg for rats, 10 mg/kg for mice), using
13C NMR spectroscopy, also offered more information of its possible metabolic interactions
(Fennel and Sumner, 1994; Fennel et al., 1991). As detailed in Section 3.4.2.3, some of the
hypothetical metabolites of AN shown in Figure 3-1 were detected in the urine of animals
exposed to AN. A major urinary metabolite in rats was N-acetyl-S-(2-cyanoethyl)cysteine, from
conjugation of AN with GSH. Other metabolites, formed following oxidation of AN to CEO and
subsequent conjugation to GSH, were identified as N-acetyl-S-(2-hydroxyethyl)cysteine,
thiodiglycolic acid, S-carboxylmethylcysteine, and thionyldiacetic acid, all derived from addition
of GSH to the 3-position of CEO. Thiocyanate was detected in urine as a metabolite of released
cyanide. Moreover, N-acetyl-S-(l-cyano-2-hydroxyethyl)cysteine was formed after addition of
GSH to the 2-position of CEO (Fennel and Sumner, 1994). These metabolites were also found in
mouse urine.
Species differences in the extent of AN metabolism via oxidation to CEO, and
subsequent conjugation of CEO with GSH, may exist. Fennell et al. (1991) and Fennell and
Sumner (1994) noted differences in the relative abundance of these urinary metabolites in mice
compared with rats. After oral administration of 10 mg/kg AN to mice, 80% of the urinary
metabolites were derived from CEO, most notably thiodiglycolic acid and S-(carboxymethyl)-
cysteine. By contrast, these metabolites made up only 60% of metabolites in the urine from rats
administered orally with 10 or 30 mg/kg AN. This difference indicated that more CEO was
produced in the mouse than in rats. In addition, the ratio of metabolites derived from glutathione
conjugation of CEO at the 2- and 3-positions determined the amount of cyanide released, since
cyanide is released from CEO metabolites conjugated at the 3-position. This ratio was 0.43 in
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rats and 0.21 in mice, indicating that a greater percentage of the CEO produced in mice was
metabolized to release cyanide. Thus, mice were likely to be exposed to a higher cyanide level
produced from CEO, possibly accounting for the greater acute toxicity of AN in the mouse
(Fennel and Sumner, 1994).
Wang et al. (2002) confirmed the central role of CYP2E1 in the metabolism of AN to
cyanide via CEO. Male WT and CYP2El-null mice were dosed by gavage with 0, 2.5, 10, 20, or
40 mg/kg AN, and cyanide was measured in blood and tissues. Expression of CYP2E1 and EH
was monitored concurrently using Western blot techniques. Cyanide concentrations in blood and
tissues of AN-treated WT mice increased dose dependently but remained at background levels in
CYP2El-null mice or control WT mice. Results from Western blots showed CYP2E1 to be well
expressed in the liver, kidney, and lung of WT mice and not detected in tissues of CYP2El-null
mice. EH was equally expressed in both WT and CYP2E1 mice, supporting the hypothesis that
CYP2E1-mediated oxidation of AN is an early step in the metabolism of AN to cyanide. The
role of cyanide in the acute toxicity of AN was confirmed by the lack of acute symptoms of AN
toxicity in CYP2El-null compared with WT mice. Pretreatment of WT mice with a universal
CYP450 inhibitor, ABT, likewise blocked cyanide formation and abolished the symptoms of
acute toxicity. Wang et al. (2002) concluded that the metabolism of AN to CEO was exclusively
catalyzed by CYP2E1.
3.3.2. Interaction of AN with GSH
Young et al. (1977) suggested that since the toxicity of AN was due to the parent
compound (AN) or its oxidative metabolites, the cyanoethylation of sulfhydryl-containing
compounds, such as GSH or cysteine, by AN represented a detoxification mechanism. This
metabolic pathway was shown indeed to play a role in the detoxification of AN (Ghanayem and
Ahmed, 1986; Ghanayem et al., 1985; Appel et al., 1981). The toxicity of AN would be
expected to increase in severity as the GSH level becomes depleted (Benz et al., 1997a).
Kopecky et al. (1980) administered 0.75 mmol/kg AN or 0.5 mmol/kg [1-14C]-AN to
female Wistar rats by different routes (oral, i.p., subcutaneous [s.c.], and i.v.) and measured
radioactivity and thiocyanate excreted in the urine. "Non-thiocyanate" metabolites excreted in
urine constituted about two-thirds of the administered dose. Paper chromatography of the
metabolites in urine identified AN mercapturic acid as the key metabolite. Kopecky et al. (1980)
proposed that there were at least two pathways for AN metabolism. The minor route was
oxidative metabolism to cyanide, which was further metabolized to thiocyanate and other "non-
thiocyanate" metabolites. The major route was conjugation with glutathione, catalyzed by
glutathione S-alkenetransferases, to N-acetyl-S-(2-cyanoethyl)cysteine.
Further support for this proposition came from Geiger et al. (1983), who studied AN
metabolism in isolated F344 rat hepatocytes. GSH levels and AN-protein binding were
measured after incubating rat hepatocytes with [1-14C]- or [2,3-14C]-AN. GSH-adduct levels
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were determined by chromatographic procedures of aliquots of the trichloroacetic acid
supernatant. Exposure to AN at 5 or 10 mM resulted in decrease of GSH levels to 15-20% of
controls within 10 minutes. The primary radiolabeled product was S-(2-cyanoethyl)glutathione,
and not S-(2-oxoethyl)GSH (the compound formed by reaction of CEO with GSH in the
presence of purified GSH transferase).
Indirect evidence for the involvement of GSH in the metabolism of AN was provided by
Langvardt et al. (1980, 1979), who used gas chromatography-mass spectrometry and gas
chromatography-infrared spectroscopy to identify urinary components in male Sprague-Dawley
rats 16 hours after exposure to [1-14C]- or [2,3-14C]-labeled AN by gavage. They identified two
major components: the first was N-acetyl-S-(2-cyanoethyl)cysteine, which they assumed to be a
product of AN conjugation with GSH, and the other was thiocyanate. A third metabolite,
N-acetyl-S-(2-cyanoethyl)cysteine, was tentatively identified and was proposed by Langvardt et
al. (1980) to have resulted from the action of GSH on the epoxide intermediate. The authors
speculated that the detoxification of AN likely involved conjugation with GSH, and the toxicity
of AN was likely affected by the status of GSH pools in target tissues, since a rapid and dose-
dependent decrease in GSH stores in the liver, lungs, kidney, and adrenals was observed by
Szabo et al. (1977) after i.v. injection of 1-15 mg/100 g AN to Sprague-Dawley rats. A sharp
decrease in cerebral GSH concentrations occurred between 5 and 15 mg/100 g AN and correlated
with the occurrence of mortality. On the other hand, oral dosing of 0.002-0.05% AN to rats for
21 days resulted in up to 25% increase in hepatic GSH and might represent a rebound
phenomenon (Szabo et al., 1977).
Benz et al. (1997a) studied the time and dose dependence of the depletion of tissue GSH
and tissue cyanide and the covalent binding to tissue after s.c. injection of 0, 20, 50 (LDio),
80 (LDso), or 115 mg/kg (LDgo) AN to male Sprague-Dawley rats. GSH levels in liver were the
most sensitive marker of AN exposure and were depleted by 50% at 20 mg/kg, a dose without
overt toxicity. At 50 mg/kg, the threshold dose for overt toxicity, GSH was depleted by>85%
and followed by a rapid recovery of 60% at 4 hours. Liver GSH was depleted almost completely
within 30 minutes when rats were injected with 80 mg/kg AN. The depletion was sustained
through 120 minutes and followed by 40% recovery through the end of study period of 4 hours.
Blood and brain GSH were more resistant to the GSH depleting effects of AN and were depleted
less extensively in a dose-dependent manner as the doses were in the toxic range. (The highest
dose of 115 mg/kg depleted only 40% brain GSH at 2 hours.) In addition, brain GSH levels
showed little capacity for recovery during the study period, unlike liver and kidney. Glandular
and forestomach GSH were also dose-dependently depleted by AN treatment and were unable to
recover within the study period.
GSH depletion was accompanied by a dose-dependent increase of cyanide in the blood
and brain during the first 60 minutes. At the lowest dose of 20 mg/kg, blood and brain cyanide
declined after 60 minutes. At the higher doses, blood and brain cyanide continued to increase to
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120 minutes and then declined. Covalent binding of AN to tissue protein increased in all tissues
rapidly during the first 30 minutes at all doses. At 20 mg/kg, covalent binding reached a plateau
at 30 minutes. At the three higher doses, covalent binding continued to increase after 30 minutes
and reached a plateau level by 2-4 hours. Benz et al. (1997a) concluded that when liver GSH
was depleted, detoxification of AN was terminated. Acute AN toxicity became apparent, and a
sustained increase in covalent binding to tissue protein was observed.
Further experiments by Ahmed and coworkers (Ahmed et al., 1983, 1982; Ghanayem and
Ahmed, 1982) confirmed the involvement of hepatic GSH in AN metabolism. When bile was
collected from male Sprague-Dawley rats given a single oral dose of 46.5 mg/kg AN containing
12 uCi/kg [1-14C]-AN, four metabolites were isolated and characterized in biliary extracts at 6
hours after treatment; the two main metabolites, glutathione conjugates of AN, were S-
cyanoethyl glutathione andN-acetyl-S-(2-cyanoethyl)cysteine (Ghanayem and Ahmed, 1982).
Pretreatment of rats with diethyl malate (a glutathione-depleting agent) significantly decreased or
abolished all of the metabolites in the bile. The study authors proposed GSH conjugation as a
major pathway of AN metabolism, probably catalyzed by glutathione transferases. The product,
S-cyanoethyl glutathione, is further metabolized to N-acetyl-S-(2-cyanoethyl)cysteine. Nearly
27% of the administered dose was excreted in the bile after 6 hours (Ahmed et al., 1982).
Kedderis et al. (1995) compared the kinetics of AN and CEO interaction with GSH in
vitro by measuring the formation of conjugates when [2,3-14C]-labeled AN or CEO were
incubated for a very short time (20 seconds for mice, 30 seconds for rats) with [glycine-
2-3H]-GSH in the presence of microsomal and cytosolic subcellular fractions of human, rat, or
mouse liver. Because of the rapid non-enzymatic reaction of AN and CEO with GSH at pH 7.3,
the steady-state kinetics of GSH conjugation were determined at pH 6.5. HPLC-mass
spectrometry was used to separate and identify the conjugates, which included S-(2-cyanoethyl)-
glutathione from AN; and S-(l-cyano-2-hydroxyethyl)glutathione and S-(2-cyano-2-hydroxy-
ethyl)-glutathione from CEO. The apparent kinetic parameters for the conjugation reactions
(Vmaxapp and Km app) were estimated by fitting the Michaelis-Menten equation to the data, giving
estimates of Vmaxapp for the conjugation reactions catalyzed by mouse cytosolic enzymes that
were four to six times greater than those for rat cytosolic enzymes at pH 6.5 (Table 3-5). These
data suggest that mouse liver cytosolic glutathione-S-transferase (GST) has a greater capacity for
conjugating AN and CEO than do rat liver enzymes. Initial velocity studies were also carried out
with microsomes and cytosol from human liver, providing data to suggest that GSH conjugation
of AN with human liver cytosol was broadly similar to that of rodent liver cytosol (Table 3-6).
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Table 3-5. Apparent kinetic parameters for glutathione conjugation of AN
and CEO at pH 6.5
Fixed substrate
AN
CEO
Species
Rat
Mouse
Rat
Mouse
Vmaxapp (umol/min-mg)a
3.32 ±0.29
21.68 ± 1.75
2.00 ± 0.20
8.34 ±0.75
Kmapp(GSH)(mM)a
17.47 ±3.60
23.39 ±3.99
9.36 ±2.27
15. 12 ±2.84
aValues are the means ± SDs of three determinations.
Source: Kedderis et al. (1995).
Table 3-6. Glutathione conjugation of AN and CEO with or without
microsomal or cytosolic GST from rat, mouse, and human liver
preparations
Species"
Rat
Mouse
Human
Rat
Mouse
Human
Substrate
AN
CEO
Nanomoles of product
Non-enzymatic
25.5 ±2.0
15.5 ±1.3
Microsomal
32.6±1.9b
31.3±1.2b
19.5-24.5C
16.6 ±1.1
18.9±1.0b
11.2-14.3C
Cytosolic
36.1±3.0b
36.6±1.4b
25.2-34.4c
25.2±l.lb
25.9±3.7b
12.6-14.5C
"Rodent values are means ± SDs of three determinations.
bStatistically significantly greater than the non-enzymatic reaction (p < 0.05) as calculated by the authors.
°Range of values for subcellular fractions prepared from six human subjects.
Source: Kedderis et al. (1995).
In the same report, Kedderis et al. (1995) described the determination of the initial
velocities of non-enzymatic vs. enzymatic GSH conjugation of AN and CEO at pH 7.3 in the
absence or presence of GST (Table 3-6). Both substrates reacted rapidly with GSH in a non-
enzymatic reaction, and addition of hepatic microsomes or cytosol from rats and mice
statistically significantly enhanced the rate of product formation from AN (p < 0.05). Initial
velocities indicated that AN conjugated more effectively with GSH than did CEO. Hepatic
cytosol enhanced the rate of product formation from CEO to a greater extent (up to 52% higher)
than did microsomes, suggesting that rodent cytosolic GST is more active toward CEO than
microsomal GST. Under physiological conditions (pH 7.3), addition of liver cytosol from four
out of six individuals similarly enhanced GSH conjugation with AN but not CEO. Human liver
microsomes did not enhance the velocity of CEO or AN conjugation with GSH, suggesting that
human microsomal GST forms do not catalyze this reaction (Table 3-6).
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Similarly, Guengerich et al. (1981) studied the reaction of AN and CEO with glutathione
in vitro. The pseudo first-order rate constant for disappearance of GSH at pH 7.7 (37°C) was
0.28/minute, when 0.5 mM was mixed with 0.1 M AN. The rate constant for the disappearance
of GSH in the presence of 0.1 mM CEO was 0.1 I/minute.
Guengerich et al. (1981) also incubated [1-14C]-AN or -CEO with GSH at pH 6.5 in the
presence of cytosolic fraction from rat liver, human liver, or rat brain to determine GST activity.
Rat liver cytosol showed GSH S-transferase activity towards AN, with a Km of 33 mM and a
Vmax of 57 nmol/minute-mg protein. Human liver and rat brain cytosolic fraction had no activity
towards AN. Rat liver also showed activity towards CEO, with activity in rat brain more than an
order of magnitude lower.
The importance of sulfhydryl compounds in the detoxification of AN was shown in a
number of studies, particularly by the demonstration that treatment with exogenous sulfhydryl
compounds could protect the organism from the harmful effects of AN. Appel et al. (1981)
reported that sulfhydryl compounds such as cysteine were effective antidotes for both orally and
intraperitoneally administered lethal doses of AN in rats. N-acetyl-cysteine was ineffective
when given intraperitoneally and less effective than cysteine when administered orally. The
GSH-depleting effect of a single s.c. dose of AN administered to male Sprague-Dawley rats and
the subsequent GI bleeding and gastric mucosal necrosis could be blocked by pretreatment with
sulfhydryl-containing agents such as L-cysteine or cysteamine (Ghanayem and Ahmed, 1986;
Ghanayem et al., 1985).
Benz et al. (1990) studied the effectiveness of D- or L-cysteine and N-acetyl-D-cysteine
or N-acetyl-L-cysteine in the detoxification of acutely administered AN by determining the s.c.
LDso of AN in male Sprague-Dawley rats either administered AN alone or in combination with
individual antidotes. The LDso of AN alone was determined to be 74.7 mg/kg. The LDso of AN
when combined with other antidotes ranged from 93.3 mg/kg (with N-acetyl-D-cysteine) to
151.4 mg/kg (with L-cysteine). The antidote protective index [(LDso AN with antidote)/(LD5o
AN alone)] ranged from 1.25 for N-acetyl-D-cysteine to 2.03 for L-cysteine. Thus, N-acetyl-D-
cysteine was less effective than other antidotes in reducing acute lethality of AN. Measurement
of urinary N-acetyl-S-cyanoethyl cysteine, which was derived from conjugation with GSH
pathway, following s.c. injection of 50 mg/kg AN alone or AN plus an antidote, indicated that
none of the antidotes significantly increased the excretion of this metabolite.
Blood cyanide levels were also measured in rats at 0.5, 1, 2, 4, and 6 hours following s.c.
injection of 50 mg/kg AN or AN plus an antidote. Benz et al. (1990) showed that all of the
antidotes, except N-acetyl-D-cysteine, lowered blood cyanide levels. Since N-acetyl-D-cysteine
was the least effective antidote, the antidotal effectiveness of these cysteine enantiomers was
related to their cyanide detoxification mechanism. Discussing these findings, Borak (1992)
pointed out that, because both D- and L-cysteine provided equivalent protection against AN
poisoning but only L-cysteine could be incorporated into GSH, the antidotal effects of these
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compounds may be unrelated to GSH repletion. In fact, since the potency of each antidote was
proportional to their ability to lower cyanide levels, Borak (1992) suggested that their effects
may be due to the ability of cysteine derivatives to serve as sulfur donors for the detoxification of
cyanide via rhodanese-mediated transformation of cyanide to thiocyanate. The protection
provided by these antidotes for cyanide poisoning from AN exposure, however, does not
necessarily extend to other forms of AN-induced toxicity.
3.3.3. Covalent Binding of AN and Its Metabolites to Subcellular Macromolecules
When isolated hepatocytes from male F344 rats were incubated with 1 mM [2,3-14C]-AN
for 2 hours (Hogy, 1986; Geiger et al., 1983), a radiolabeled protein adduct was formed that
could be characterized after the removal of residual AN and other low molecular weight products
by dialysis. Analysis of the hydrolyzed non-dialyzable fraction indicated that about 75% of the
radioactivity was contained in a single major peak that was chromatographically identified as
S-(2-carboxyethyl)cysteine, the hydrolysis product of S-(2-cyanoethyl)cysteine, indicating direct
cyanoethylation of cysteinyl residues by AN. When isolated hepatocytes were incubated with
2 mM [2,3-14C]-AN for 60 minutes, Geiger et al. (1983) estimated the level of irreversible
binding to protein to be 8.1 nmol/mg protein, a level of alkylation corresponding to modification
of 1 in every 900 amino acid residues or roughly 1 of every 20 cysteine groups. In contrast to
AN-protein binding, detection of binding of [2,3-14C]-AN-derived radiolabel to DNA and RNA
in hepatocytes was limited by the low level of radioactivity that could be incorporated into
[14C]-AN because of polymerization and microsynthesis problems and by the high level of
protein binding. Therefore, hepatocytes were incubated with 2 mM [2,3-14C]-AN for 60 minutes
in the presence of extracellular calf thymus DNA (0.5 mg/mL) (Hogy, 1986; Geiger et al., 1983).
RNA and both intracellular and extracellular DNA were isolated. The isolated RNA contained
53 pmol AN metabolites per mg RNA of which 9 pmol/mg RNA could be attributed to the
contaminating protein. The isolated extracellular DNA contained 47 pmol AN adducts per mg
DNA, of which 30 pmol/mg DNA could be attributed to protein. A portion of incubated
extracellular DNA was further purified; alkylation of extracellular DNA was not observed at a
detection limit at 3.5 x 105 bases (Geiger et al., 1983).
Hogy and Guengerich (1986) treated an F344 rat intraperitoneally with 0.6 mg/kg
[2,3-14C]-CEO to assess macromolecular binding in liver and brain 1 hour after treatment. DNA
and RNA in liver and brain were isolated, and the amounts were estimated. Bound radioactivity
was estimated by liquid scintillation counting. Covalent binding of CEO-derived radioactivity
was detected in both liver and brain protein at rates of 1.1 and 1.0 alkylations per 106 amino
acids, respectively. However, no covalent binding to DNA or RNA could be detected at the
level of 0.3 alkylations per 106 bases in liver and brain.
In another experiment by Hogy and Guengerich (1986), liver and brain DNA were
isolated from three rats 2 hours after treatment with 50 mg/kg AN i.p. or 6 mg/kg CEO i.p.
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Using thin layer chromatography with fluorescent plates, the study authors were able to detect
N7-(2-oxoethyl)guanine at 0.014 and 0.032 alkylations per 106 DNA bases in liver for rats treated
with CEO and AN, respectively. This DNA adduct was apparently derived from CEO,
consistent with its recovery after AN and CEO treatment. The apparent lineage of this DNA
adduct was different from the protein adduct, S-(2-cyanoethyl)cysteine, described previously,
which was derived from AN.
N7-(2-oxoethyl)guanine was only at the level of detection in rat brain DNA. In addition,
l,N6-ethenoadenosine or l,N6-ethenodeoxyadenosine were not detected in liver DNA, using
HPLC with fluorescence detector in the analyses of these two adducts. Binding of AN and/or
CEO to DNA is also discussed in Section 4.5.1.2.1.
As discussed in Section 3.3.2, when 0, 20, 50, 80, or 115 mg/kg [2,3-14C]-AN was
injected subcutaneously into male Sprague-Dawley rats, GSH became almost completely
depleted (>95%) in liver at 80 mg/kg within 30 minutes, while blood and brain GSH were more
resistant to the depleting effect of AN. Brain and blood GSH were not affected at 20 mg/kg.
The amount of cyanide in blood and brain increased dose dependently in the first hour after
dosing (Benz et al., 1997a). Covalent binding to tissue proteins increased in a dose-dependent
fashion during the first 30 minutes at all doses, with binding to blood proteins being 3-4 times
greater than in any other tissue. Benz et al. (1997a) suggested that GSH depletion in liver was
related to AN toxicity and covalent binding.
The effect of GSH depletion on the irreversible binding of AN to tissue macromolecules
has been studied in male F344 rats exposed to 4 mg/kg [2,3-14C]-AN either by inhalation (Pilon
et al., 1988a) or gavage (Pilon et al., 1988b). Binding of radiolabel to tissue macromolecules
was evaluated in control rats or rats depleted of GSH by an i.p. injection of phorone (PH)/
buthionine sulfoximine (BSO) about 30 minutes prior to AN exposure. GSH contents in control
rats were as follows: liver (17.3 umol/g), kidney (4.5 umol/g), lung (3.1 umol/g), stomach
(5.3 umol/g), brain (3.9 umol/g), and blood (4.2 umol/g). A significant depletion of GSH was
produced in liver (43%), kidney (42%), and lung (22%) after PH/BSO treatment. No significant
depletion of GSH was observed in blood, brain, or stomach 30 minutes after a combined
PH/BSO treatment.
In the inhalation studies (Pilon et al., 1988a), three rats were exposed to initial AN
concentrations of 0, 25, 50, 100, 500, or 750 ppm in a closed-circuit inhalation chamber for
240 minutes. AN was not replenished during the exposure, and the decrease in chamber AN
concentration was monitored by taking samples every 10 minutes during the exposure. Uptake
of AN vapor by control rats showed two distinct phases: an initial, rapid phase that lasted about
60 minutes, followed by a slower phase. An uptake rate of 4.82 mg/kg-hour was estimated for a
concentration of 100 ppm using the uptake curve for the rapid phase. In GSH-depleted rats, the
mortality rate was higher. The rate of AN uptake was increased in the rapid phase but decreased
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at 500 and 750 ppm. In the slow phase, uptake was similar to that in control rats for
concentrations below 200 ppm, but was elevated at 500 and 750 ppm.
Radioactivity irreversibly associated with tissue macromolecules was measured in control
rats 1, 2, 4, 6, 12, or 24 hours after the AN dose. In GSH-depleted rats, radioactivity was
measured 1, 6, or 24 hours postexposure. In most tissues, the concentration of AN-derived
undialyzable radioactivity (ADUR) reached a maximum in
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higher recovered radioactivity was found in stomach, lung, and blood of GSH-depleted rats than
of control rats 1 hour after dosing, while the radioactivity in kidney was lower in GSH-depleted
rats (Pilon et al., 1988b).
The concentration of ADUR in brain reached a maximum in <1 hour for both control and
GSH-depleted rats and remained constant for 24 hours. In GSH-depleted rats, 60-80% increase
in ADUR was measured in brain at 1, 6, and 24 hours after dosing. The highest ADUR level was
found in stomach and increased with time in control rats. ADUR levels in liver and kidney of
control rats was characterized by an increase over the first 6 hours and a decrease between 5 and
24 hours. GSH-depletion resulted in an increase in ADUR levels in liver, lung, kidney, stomach,
blood, and brain between 6 and 24 hours after the dose.
GSH depletion also caused a significant increase in ADUR levels in total nucleic acid
(DNA + RNA) in both brain and stomach (one and a half- and threefold, respectively) 6 hours
after the dose. No change was found in liver. ADUR associated with DNA was detected in
stomach tissue of control rats only. Pilon et al. (1988b) suggested that ADUR levels reflected
the relative concentration of covalently bound radioactivity in control and GSH-depleted rats and
that the reaction of AN with protein and other macromolecules was responsible for the rapid
increase in ADUR at <1 hour. The slower increase in ADUR in metabolically competent organs
(liver, kidney, and lung) of control rats and all organs of GSH-depleted rats might represent the
binding of CEO to macromolecules. Urinary excretion of thiocyanate, a final metabolite from
the epoxide pathway of AN metabolism, was increased twofold in GSH-depleted rats. Since
urinary thiocyanate is indicative of CEO formation, Pilon et al. (1988a) interpreted their results
as indicating that more CEO was formed after GSH-depletion.
Farooqui and Ahmed (1983a) also reported covalent binding of [2,3-14C]-AN to protein,
DNA, and RNA of tissues of male Sprague-Dawley rats treated with a single oral dose of
46.5 mg/kg. DNA from tissue homogenate was isolated by extraction with chloroform/isoamyl
alcohol/phenol and application of the aqueous extract to hydroxyapatite chromatography. DNA
alkylation was higher in brain and stomach than that in the liver, with highest levels of covalent
binding in the brain. The covalent binding indices for the liver, stomach, and brain at 24 hours
after dosing were 5.9, 51.9, and 65.3, respectively.
Ahmed et al. (1992a) demonstrated the covalent binding of radiolabel from [2,3-14C]-AN
to testicular DNA after a single gavage of radiolabeled AN (46.5 mg/kg) to male Sprague-
Dawley rats. In a time course study, bound activity was shown to be greatest after 30 minutes
(8.93 ± 0.80 umol AN bound per mol nucleotide). Using an identical experimental protocol,
Ahmed et al. (1992b) demonstrated the capacity of AN to bind covalently to DNA in the lung.
Binding was associated with a 28-41% decrease in replicative DNA synthesis at time points up
to 24 hours after dosing.
Jacob and Ahmed (2003a) used whole-body autoradiography to examine the distribution
of [2-14C]-AN administered orally or intravenously to male F344 rats. Two days after oral
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dosing, covalently bound radioactivity was retained at higher levels in the gastric mucosa, blood,
and hair follicles. After i.v. injection, covalently bound radioactivity was retained in liver,
spleen, bone marrow, adipose tissue, and lung. The amount of radioactivity associated with
covalent binding was lower for oral dosing than for i.v. injection, which was consistent with the
higher recovery of radioactivity excreted following oral dosing compared with injection (see
Section 3.4.2.1).
Covalent binding of [2,3-14C]-AN to tissue protein and globin was also studied in male
Sprague-Dawley rats after a single s.c. injection of 1.2-115 mg/kg (Benz et al., 1997b).
Covalent binding to tissue protein reached completion in 1-4 hours and was linear in the low
dose range (1.2-50 mg/kg), with the relative order (in descending order) as follows: blood >
kidney = liver > forestomach = brain > glandular stomach > muscle. Covalent binding to globin
followed a similar dose-response curve. Benz et al. (1997b) also measured an N-(2-cyanoethyl)-
valine (CEVal) adduct of globin at this dose range. This adduct was formed by reaction of AN
with the NH2-terminal residue of globin (Osterman-Golkar et al., 1994) and represented only
0.2% of total AN binding to globin. However, regression of tissue protein binding vs. globin
total covalent binding or globin CEVal adduct indicated that both globin biomarkers could be
used as surrogates for the amount of AN bound to tissue protein.
Using a similar dosing regimen, Nerland et al. (2001) employed sodium dodecyl sulfate-
polyacrylamide gel electrophoresis to separate labeled proteins isolated from subcellular
fractions of liver from treated rats. Binding of AN was found to be associated preferentially with
GST of the [j, subclass (GSTM). Within this subclass, GSTM1 was labeled about seven times
more strongly than GSTM2, while, from the a-subclass, only GSTA3 was labeled (at about
1/35 the strength of GSTM1). No label was associated with GSTA1 or GSTA2. The site of
binding was identified as exclusively cysteine 86. Since this particular cysteine residue in
rGSTMl appeared to have been targeted specifically, the study authors hypothesized that high
reactivity at cysteine 86 was due to its potential interaction with histidine residue at position 84,
which would lower the pka of cysteine 86, increasing reactivity towards sulfhydryl reagents.
These data would suggest that tissue proteins containing cysteine residues with an abnormally
low pka value would be likely targets for AN. In an in vivo experimental approach, Nerland et
al. (2003) demonstrated that subcutaneously administered AN preferentially bound to the
cysteine 186 residue of carbonic anhydrase III (CAIII) in rat liver.
A considerable body of evidence demonstrated the ability of AN to bind to intercellular
proteins, in particular to Hb. Osterman-Golkar et al. (1994) reported a method for quantifying an
N-terminal cyanoethyl-valine adduct, CEVal, the product of reaction between AN and the
N-terminal valine of Hb. The method was based on the N-alkyl Edman procedure involving the
derivatization of globin with pentafluorophenyl isothiocyanate and gas chromatography-mass
spectrometry analysis. Osterman-Golkar et al. (1994) showed that the method was applicable to
experimental animals exposed to AN in drinking water and to humans exposed to AN in tobacco
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smoke. Hb from smokers (10-20 cigarettes/day) with a daily intake of 0.5-5 ug AN per kg BW
contained about 90 pmol CEVal/g, whereas adduct levels in the Hb of nonsmokers were below
the detection limit of about 20 pmol/g Hb. This utility was confirmed by Tavares et al. (1996)
for occupationally exposed workers and for smokers.
A subsequent study by Bergmark (1997) showed that the Hb of smokers contained
adducts of ethylene oxide and acrylamide as well as AN. In nonsmokers, the CEVal Hb adduct
of AN was below the detection limit of <2 pmol/g of globin. In the 10 smokers studied, the
levels of this adduct ranged from 25 to 178 pmol/g (mean 106 pmol/g) and correlated with the
number of cigarettes smoked per day (correlation coefficient = 0.94).
Fennell et al. (2000) determined CEVal from blood samples of 16 nonsmokers and
32 smokers and reported that CEVal Hb adducts increased with increased cigarette smoking.
The estimated CEVal level from smoking was 170 fmol per mg Hb per pack-day. Two
participants in a smoking cessation program showed a gradual reduction of CEVal levels (Perez
et al., 1999). Thus, the use of the Hb adduct CEVal may have utility as a biomarker to assess
low-level exposure to AN (in the region of 50 ppb), even in a complex mixture of toxicants,
although smoking would be a confounding variable. Fennell et al. (2000) also reported that the
null genotypes for GSTM1 or GSTT1 had little effect on CEVal levels when compared to active
genotypes.
Borba et al. (1996) measured CEVal as a marker of AN exposure in occupationally
exposed workers in an acrylic fiber factory. The values for CEVal among the subjects were 8.5-
70.5 pmol/g globin in controls, 635.2-4,603.5 pmol/g globin for continuous polymerization
workers, and 93.9-4,746 pmol/g globin for maintenance workers. These findings pointed to the
ready formation of AN adducts with Hb in an occupational setting.
CEVal Hb adduct was also used as a follow-up dose monitor after accidental exposure of
four cleaning workers to AN in an AN-containing tank wagon (Bader and Wrbitzky, 2006). On
day 25 after exposure, CEVal adduct levels in Hb ranged from 640 (blood sample partly
hemolyzed) to 2,020 pmol/g globin for the three smokers and was 566 pmol/g globin for the
nonsmoker, indicating residual AN adducts from the accidental exposure. On day 175, CEVal
adduct levels were 81-276 pmol/g globin for the smokers and 2 pmol/g globin for the nonsmoker
and represented background CEVal of the study participants according to their smoking status.
For both the smokers and nonsmoker, the adduct concentrations in blood declined linearly with
time. Linear regression analysis of the data estimated a total elimination interval of 148 days,
longer than the standard lifespan of 126 days of erythrocytes. Linear regression analysis also
allowed estimation of the initial adduct levels of the workers on the day of the accident and
estimation of the exposed AN concentrations based on the correlation from the former German
exposure equivalent that 3 ppm AN yields 17,200 pmol/g globin.
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3.4. ELIMINATION
3.4.1. Studies in Humans
In an inhalation study of six male volunteers exposed to 7.8-10 mg/m3 AN for 8 hours
(Jakubowski et al., 1987), 44-58% (mean = 52%) of the inhaled AN was absorbed, of which
about 22% of absorbed AN was metabolized and excreted in the urine over 31 hours from the
start of exposure as N-acetyl-S-(2-cyanoethyl)-L-cysteine (2-cyanoethyl mercapturic acid
[CEMA]). Elimination followed first-order kinetics, with a half-life of about 8 hours. The
authors concluded that individual urinary CEMA levels were not a useful measure for exposure
to AN.
3.4.2. Studies in Animals
3.4.2.1. Exhalation
In a study by Young et al. (1977), male Sprague-Dawley rats were exposed to [1-14C]-AN
via inhalation (5 or 100 ppm) or a single oral dose (0.1 or 10 mg/kg). Exhalation of AN as CC>2
within 6 hours after dosing decreased with increasing dose, from 6.1 to 2.6% of the dose
following inhalation exposure and from 4.6 to 3.9% after gavage (see Tables 3-1 and 3-2).
Ahmed et al. (1983) administered a single oral dose of 46.5 mg/kg AN to male Sprague-
Dawley rats in distilled water, using 50 uCi/kg of either [2,3-14C]- or [1-14C]-AN as tracer. The
recovery of total dose in expired 14CO2 varied from 2% for [2,3-14C]-AN to 12% for [1-14C]-AN
24 hours after dosing. Burka et al. (1994) administered 0.87 mmol/kg (46.2 mg/kg) [2-14C]-AN
by gavage to male F344 rats. About 2% of the dose was expired as volatile organic components,
predominantly unchanged AN, while 11% was liberated as 14CC>2 24 hours after dosing.
Jacob and Ahmed (2003a) compared excretion of 11.5 mg/kg [2-14C]-AN administered
orally or intravenously to male F344 rats. In the 48 hours after oral dosing, 61% of the
radioactive dose was excreted, with 4% in exhaled CC>2, 4% in urine, and 53% in feces.
Following i.v. administration, 30% of the dose was eliminated over 48 hours, with 2% in expired
air, 8% in urine, and 21% in feces. In the 8 hours after oral dosing, 3% of the radioactive dose
was in exhaled CC>2, 2% in urine, and 48% in feces. Following i.v. administration, 2% of the
dose was in exhaled CC>2, 3% in urine, and 0.2% in feces. Jacob and Ahmed (2003a) concluded
that these results indicated a significant difference in biological fate of AN following i.v. or oral
treatment.
3.4.2.2. Fecal Excretion
Young et al. (1977) investigated the fecal excretion of AN in male Sprague-Dawley rats.
This route of excretion showed little dependence on the route of administration, amounting to 3-
4% of the dose following inhalation exposure and about 5% following gavage within 6 hours of
dosing. Young et al. (1977) also investigated the possibility of biliary excretion with subsequent
reabsorption from the intestines in one male rat. A cannula was inserted between the common
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bile duct and the duodenum for sampling of bile, and [14C]-AN was given intravenously. Based
on the concentrations of radioactivity in bile, urine, RBCs, and plasma as a function of time, the
rate of elimination from the bile was found to be faster than from other fluids during the first 6
hours after dosage. The initial half-life for excretion of radioactivity via bile was estimated to be
about 15 minutes. One unidentified metabolite was excreted in the bile and underwent
enterohepatic circulation to the GI tract, contributing partially to 5% of the dose excreted in
feces.
Kedderis et al. (1993a) estimated that 3-5% of AN doses were excreted in feces of F344
rats (0.09-28.8 mg/kg orally), while 2-8% of the dose were recovered from the feces of B6C3Fi
mice (0.09-10 mg/kg orally). In either species, the differences in fecal excretion were not
related to the administered dose. Farooqui and Ahmed (1982) administered a single oral dose of
46.5 mg/kg [1-14C]-AN to male Sprague-Dawley rats, the highest percentage of 14C excreted in
feces (2%) occurred between 12 and 24 hours after dosing. At the end of 10 days, 2.5% of the
dose was excreted in feces. In another study, male Sprague-Dawley rats were given a single
gavage dose of 46.5 mg/kg [1-14C]-AN; four radioactive peaks were identified in biliary extracts
at 6 hours after treatment. The two major metabolites in bile were GSH conjugates of AN:
S-cyanoethyl glutathione and N-acetyl-S-(2-cyanoethyl)cysteine (Ghanayem and Ahmed, 1982).
Nearly 27% of the dose appeared in the bile after 6 hours (Ghanayem and Ahmed, 1982; Ahmed
et al., 1982). The GI tract contained the highest level of radioactivity up to 72 hours, suggesting
resecretion of AN metabolites to the stomach or binding of metabolites to the stomach mucosa
(Ahmed et al., 1982).
3.4.2.3. Urinary Excretion
A substantial number of studies demonstrated the rapid urinary elimination of AN or its
metabolites when AN was administered to experimental animals via the oral or inhalation routes
(Burka et al., 1994; Fennell and Sumner, 1994; Kedderis et al., 1993a; Ahmed et al., 1983, 1982;
Young et al., 1977). For example, when Young et al. (1977) exposed male Sprague-Dawley rats
to radiolabeled AN via inhalation (5 or 100 ppm) or a single oral dose (0.1 or 10 mg/kg), most of
the radiolabel was recovered in the urine, with much lower proportions of the initial dose in feces
or expired air (see Table 3-1 and 3-2). Urinary excretion increased with dose, from 69 to 82% of
the dose following inhalation exposure and from 34 to 67% after gavage (Tables 3-1 and 3-2).
Excreted dose in urine was mostly unidentified metabolites.
In another study, Sapota (1982) administered 40 mg/kg AN, containing either 40 uCi/kg
[1,2-14C]- or [1-14C]-AN, in saline to male Wistar rats, either via gavage or intraperitoneally. In
parallel to the depletion of radiolabel in tissues at 24 hours postexposure, 82-93% of the dose
was eliminated from the body in the urine, with 3-7% exhaled unchanged in the breath in
24 hours, independent of the route of administration and the position of the radiolabel. However,
when Farooqui and Ahmed (1982) administered an oral dose of 46.5 mg/kg [1-14C]-AN to male
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Sprague-Dawley rats, 40% of the radioactivity was excreted in urine over 24 hours after dosing.
At the end of 10 days, 61% of the total dose was excreted in urine. Similarly, when Ahmed et al.
(1983) gave a single oral dose of 46.5 mg/kg AN to male Sprague-Dawley rats in distilled water,
using 50 uCi/kg of either [2,3-14C]- or [1-14C]-AN as tracer, 40% of the radioactivity from
[1-14C]-AN but 60% of [2,3-14C]-AN-derived radiolabel were detected in the urine in the initial
24 hours. Neither of the two studies provided details that might explain the discrepancy.
However, the observed discrepancy might indicate the impact of different strains used in the
studies.
The single gavage dose experiment of Burka et al. (1994), in which male F344 rats were
placed in metabolic cages after receiving 0.87 mmol/kg (46.2 mg/kg) [2-14C]-AN, resulted in
67% of the load being voided to the urine after 24 hours. This proportion of the recovered load
was similar to the 55-56% value obtained when male F344 rats or B6C3Fi mice were orally
exposed to [1,2,3-13C]-AN (Fennell and Sumner, 1994). However, Ahmed et al. (1996a)
recovered only about 4% of the counts in 48-hour urine samples when F344 rats were injected
intravenously with 11.5 mg/kg [2-14C]-AN. About 27% of the load was eliminated via all routes
combined. In this study, much higher levels of tissue binding were observed than in other
studies.
Identification of the urinary metabolites derived from AN metabolism was attempted in
several studies. Langvardt et al. (1980) administered [2,3-14C]- and [1-14C]-AN orally to male
Sprague-Dawley rats, thereby specifically tracing the fate of the vinyl and cyano groups of AN.
Two main urinary metabolites, thiocyanate and N-acetyl-S-(2-cyanoethyl)cysteine, were
identified. Thiocyanate, formed from the epoxide metabolite CEO, was the predominant urinary
metabolite following oral dosing with 30 mg/kg of [1-14C]-AN and accounting for 54% of the
injected radioactivity within 16 hours after dosing. On the other hand, thiocyanate only
accounted for 1% of the urinary radioactivity after dosing with [2,3-14C]-AN. The second
metabolite, N-acetyl-S-(2-cyanoethyl)cysteine, a mercapturic acid derived from direct
conjugation between AN and GSH, constituted 18% of the radiolabel derived from [1-14C]-AN
and 28% from [2,3-14C]-AN. Another metabolite, tentatively identified as N-acetyl-3-carboxy-
5-cyanotetrahydro-l,4-2H-thiazine, constituted 19% of the label from [1-14C]-AN and 35% from
[2,3-14C]-AN. Evidently, this metabolite was formed from conjugation of CEO with GSH.
Langvardt et al. (1980) found another four minor metabolites that were not identified
structurally.
Tardiff et al. (1987) monitored the urinary metabolites of AN-treated male Sprague-
Dawley rats 24 hours after i.v. or i.p. single doses (0.6, 3, or 15 mg/kg) or inhalation exposure to
4, 20, and 100 ppm for 6 hours. Three major metabolites were measured: thiocyanate, N-acetyl-
S-(2-hydroxyethyl)cysteine, and N-acetyl-S-(2-cyanoethyl)cysteine, that were also detected by
Langvardt et al. (1980). These three excreted metabolites represented 50-54% of the
administered dose in urine within 24 hours, independent of the dose. Tardiff et al. (1987) found
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that following i.p. or i.v. administration of AN, N-acetyl-S-(2-hydroxyethyl)cysteine and
thiocyanate each represented between 5 and 10% of the urinary metabolites; the major
metabolite was N-acetyl-S-(2-cyanoethyl)cysteine, representing 74-78% of the urinary
metabolite when AN was administered intraperitoneally or intravenously. However, following
inhalation exposure of AN, only 8% of the dose was excreted as N-acetyl-S-(2-cyanoethyl)-
cysteine, and the major urinary metabolite was thiocyanate at 20 and 100 ppm. Moreover,
N-acetyl-S-(2-hydroxyethyl)cysteine was excreted in larger amounts than N-acetyl-S-(2-cyano-
ethyl)cysteine. These results also showed that the percentage of dose excreted as urinary
thiocyanate increased with dose when AN was administered by inhalation. The study authors
concluded that the route of administration of AN had an important influence on the pattern of
metabolic excretion.
Shibata et al. (2004) investigated the urinary excretion of thiocyanate in male Wistar rats
that received 40 mg/kg AN (about half the LDso) by gavage in water. Urinary excretion of
thiocyanate became measurable at the time of peak plasma thiocyanate levels, 5 hours after
dosing. Excretion of urinary thiocyanate gradually increased so that at 10 hours after dosing,
about 1.2 mg thiocyanate (7% of administered dose) had been excreted into urine.
Gut et al. (1981) administered [1-14C]-AN to male Wistar rats (dose not given) by the
oral, i.v., i.p., and s.c. routes. Total excretion of radioactivity after 48 hours was reported to be
close to 100% following oral administration but 75-84% following the other routes of
administration. The patterns of urinary radioactivity elimination were also different: after
parenteral administrations, elimination of radioactivity was highest within the first 4 hours after
dosing, with much smaller amounts for the remaining 44 hours. After oral dosing, between
6 and 8% of the dose was eliminated in urine for the time periods 4, 8, and 12 hours after dosing
with much lower amounts thereafter. However, the major difference was found to be urinary
thiocyanate elimination: 23% during the 48 hours after oral dosing and 4% following i.p., 4.6%
following s.c., and 1.2% following i.v. administration.
In another study (Gut et al., 1985), male Wistar rats were exposed via inhalation to 57,
125, or 271 mg/m3 AN for 12 hours. A constant ratio between thiocyanate and the sum of
thioether compounds (AN mercapturic acids) in urine was found throughout the three doses.
Average total amounts of thioethers excreted in urine during 12 hours of exposure were 24, 63,
or 83 umol/kg for 57, 125, or 271 mg/m3 AN, respectively. Average total amounts of
thiocyanate excreted were 14, 24, and 49 umol/kg for 57, 125, and 271 mg/m3 AN, respectively.
The ratio of thioethers to thiocyanate was about 2:4 during the 12 hours of exposure and was
similar to that from oral exposure to AN. Thus, thioethers, not thiocyanate, were the major
urinary metabolites following inhalation and oral exposure. These results were different from
those reported by Langvardt et al. (1980) and Tardiff et al. (1987).
Miiller et al. (1987) quantified four urinary metabolites and unchanged AN following
inhalation exposure of male Wistar rats to 1-100 ppm of AN for 8 hours (thiocyanate was not
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measured). At 24 hours postexposure, N-acetyl-S-(2-cyanoethyl)cysteine was the primary
urinary metabolite, followed by N-acetyl-S-(2-hydroxyethyl)cysteine, thiodiglycolic acid (also
known as thiodiacetic acid), S-carboxyethyl-L-cysteine, and unchanged AN. The excretion
pattern of AN and its metabolites was dependent on the inhalation exposure concentrations. The
study authors proposed cyanoethyl mercapturic acid was the most sensitive indicator metabolite
of AN exposure at levels of 5 ppm.
Fennell et al. (1991) also measured GSH-derived metabolites, but not thiocyanate, in the
urine of male F344 rats (10 or 30 mg/kg) and male B6C3Fi mice (10 mg/kg), following oral
administration of [1,2,3-13C]-AN. The results of this study are shown in Table 3-7. N-acetyl-
S-(2-cyanoethyl)cysteine was formed by conjugation of AN with GSH, whereas the other
metabolites were from reaction of CEO with GSH. The results support the finding that rats and
mice differ in the way they metabolize AN, and mice evidently metabolize more AN through the
CYP2E1-mediated formation of CEO.
Table 3-7. Urinary excretion of thioethers derived from AN
Metabolite (see Figure 3-2)
N-acetyl-S-(2-cyanoethyl)cysteine
N-acetyl-S-(2-hydroxyethyl)cysteine
N-acetyl-S-(l-cyano-2-hydroxyethyl)cysteine
N-acetyl-S-(carboxymethyl)cysteine and thiodiacetic acid
Thionyldiacetic acid
Rat (30 mg/kg)
Mouse (10 mg/kg)
Percent of total metabolites
42.8 ±4.8
26.7 ±1.8
17.4 ±2.2
7.4 ±2.9
5.7 ±0.21
20.5 ±2.0
22.3 ±1.2
13.9 ±3.2
43.2 ±3.5
Not detected
Source: Fennell etal. (1991).
Kedderis et al. (1993a) studied the dose dependence of the urinary excretion of AN
metabolites in male F344 rats and male B6C3Fi mice. In rats during the 72 hours following oral
doses of 0.09-28.8 mg/kg [2,3-14C]-AN, 73-99% of the dose was excreted in urine, while 3-5%
was found in feces. In mice receiving 0.09-10 mg/kg [2,3-14C]-AN, 83-94% of the dose was
excreted in urine and 2-8% in feces. Excretion of radioactivity by both routes was not dose
dependent in either species.
The position of the radiolabel did not allow detection of thiocyanate. Radiochromato-
grams of urine from rats or mice identified five major peaks, two of which contained more than
one compound. Following administration of [2,3-14C]-CEO, two of the five peaks were not
found, indicating that those peaks were derived from direct conjugation of AN without metabolic
activation to CEO. The sum of the percent of total radioactivity from CEO conjugate-derived
peaks was higher than that from the AN conjugate-derived peaks in both rats and mice. None of
the metabolites appeared to be glucuronides. Kedderis et al. (1993a) detected four of the five
metabolites shown in Table 3-7, but could not identify thionyldiacetic acid. In addition,
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S-(2-cyanoethyl)-thioacetic acid was detected in mouse urine only, likely a degradation product
of N-acetyl-S-(2-cyanoethyl)cysteine.
The excretion of metabolites derived from CEO was approximately linear with the AN
dose in both rats and mice. However, the urinary excretion of N-acetyl-S-(2-cyanoethyl)cysteine
increased nonlinearly with increasing dose of AN, an effect that was much more pronounced in
rats than in mice. This probably indicated the presence of a competing pathway, namely,
epoxidation of AN, with the conjugation of AN with GSH. The fraction of the total dose
recovered as metabolite from CEO was 0.5 in rats and 0.67 in mice. The study authors estimated
that the ratio of AN epoxidation to GSH conjugation ranged from 4.8 to 1.3 as the AN dose
increased in rats and from 5.7 to 2.7 as the dose increased in mice. Kedderis et al. (1993a) also
pointed specifically to the species differences detected in this study—a roughly 10-fold higher
excretion of thiodiglycolic acid (thiodiacetic acid) in mice as compared with rats and measurable
excretion of S-(2-cyanoethyl)thioacetic acid in mice. This urinary metabolite could not be
detected in rats at all.
Sumner et al. (1999) treated three male WT and four male CYP2El-null mice
(C57BL/GN x Svl29) orally to 0, 2.5, or 10 mg/kg [1,2,3-13C]-AN and used NMR spectroscopy
to characterize AN metabolites in urine samples collected over 24 hours. WT mice excreted
metabolites derived from CEO (80-85% of total excreted) and from direct GSH conjugation with
AN (15-21% of total excreted), with the largest percentage of metabolites from conjugation of
GSH with the 3-carbon of CEO. CYP2El-null mice displayed only metabolites derived from
direct GSH conjugation with AN in their urine following administration of 2.5 or 10 mg/kg
[1,2,3-13C]-AN. This confirmed the role of CYP2Elin the oxidation of AN to CEO and its
subsequent transformation to a range of other products. Since CYP2El-null mice did not excrete
metabolites that would be produced by oxidation by other CYP450s, CYP2E1 may be the only
CYP450 enzyme involved in the metabolism of AN. In addition, CYP2El-null mice excreted
about the same percentage of administered dose as the WT mice, indicating CYP2El-null mice
compensated for the CYP2E1 deficiency by producing more metabolites from direct conjugation
of AN with GSH.
Taken together, the animal data indicate that AN can be exhaled as parent compound at a
low percentage of the administered dose, probably increasing at high doses (e.g., 10 mg/kg).
Fecal excretion amounts to about 5% of a given dose. The biliary pathway leading to fecal
excretion has not been well characterized. To a small extent, similar to fecal excretion, AN is
metabolized completely and exhaled as CO2. Both pathways appear to be quite independent of
the administered dose. Urinary excretion of AN metabolites has been well characterized but is
not without contradictory findings. There appears to be little doubt that mice metabolize more
AN via the CYP2E1-mediated oxidative pathway than do rats. To what extent this pathway is
likely to be overcome by large doses, or what contribution the GSH conjugation makes with
varying doses of AN and different species exposed, is not known.
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3.5. PHYSIOLOGICALLY BASED PHARMACOKINETIC MODELS
A physiologically based pharmacokinetic (PBPK) model for AN was previously
developed in rats (Kedderis et al., 1996; Gargas et al., 1995) and extended to describe the
dosimetry of both AN and CEO in humans (Sweeney et al., 2003). The PBPK model structure
consists of two parallel modules, one for AN and one for CEO, interlinked by the rate of
oxidative metabolism of AN to CEO in the liver, essentially as described by Gargas et al. (1995).
Each module consists of seven dynamic tissue compartments representing the lung, slowly
perfused tissues, fat, well-perfused tissues, brain, stomach, and liver (Figure 3-2). All perfusion-
limited tissue compartments are linked through blood flow, following an anatomically accurate,
typical, physiologically based description (Andersen, 1991).
Inhalation
CEO
Drinking
water
Metabolism
Source: Sweeney et al. (2003).
Figure 3-2. Structure of the PBPK model for AN and CEO.
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Because AN and CEO are retained by the tissue in each compartment according to their
tissue/blood partition coefficients (PCs) (which were measured in vitro), the concentrations of
both chemicals in venous blood (leaving the tissue) are lower than those in arterial blood during
the equilibration phase (except CEO in the liver). Therefore, the rate of change in the amounts
of both chemicals in each tissue compartment is given by the difference between concentration in
blood entering and exiting the tissue multiplied by the blood flow. Simple differential equations
for each compartment, corrected for the non-enzymatic conjugation with GSH, are integrated
over time, giving the amounts of AN and CEO present in the tissue (Kedderis et al., 1996) (see
equation below). Therefore, knowing (from the literature) the actual volume of each tissue,
concentrations of AN and CEO in each tissue can be calculated over time.
dA/dT = QJCA-CVHKso x CVt x GSH) x V,
where
Qi = blood flow rate to the target tissue (i)
CA = concentration of AN in arterial blood
CVi = concentration of AN in venous blood
KSO = rate constant for the conjugation of AN with glutathione (GSH)
Vi = volume of the target tissue (i)
For the lung compartment, with two mass inputs (mixed venous blood and inhaled air)
and two outputs (arterial blood and exhaled air), the amount of either chemical in alveolar air is
in equilibrium with the amount in lung blood at the steady state. Thus, concentrations of AN and
CEO in arterial blood can be calculated from simple mass balance equations. Such calculations
take into account the alveolar ventilation rate and the rate of blood flow through the lung, a
parameter made equal to cardiac output (both known from the literature) and corrected for
binding to Hb and blood sulfhydryls.
For the liver compartment, with mass input from blood and two outputs (venous blood
and metabolism, excluding biliary excretion that was not considered), the chemical mass transfer
is given by the difference between concentrations in portal and venous blood multiplied by
hepatic blood flow and corrected for: (1) metabolism of AN to yield CEO (calculated from the
Michaelis-Menten equation, subtracted from the mass of AN, and added to the mass of CEO),
(2) enzymatic hydrolysis of CEO (also described by the Michaelis-Menten equation), and (3) the
first-order conjugation with GSH. A simplified scheme of the mass flow in the PBPK model for
AN and CEO is shown in Figure 3-2 (Sweeney et al., 2003).
The model was initially calibrated in rats for three routes of AN administration—oral, i.v.
(bolus), and inhalation (Kedderis et al., 1996)—by manually adjusting the metabolic parameters
Vmax and Km for AN oxidation, first-order constants for AN- and CEO-GSH conjugation, and
first-order constant for absorption of AN from the GI tract, guided by the approximate statistical
likelihood calculation of SimuSolv. Absorption through the skin, which has been estimated at
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about 1% of that through the lung (Rogaczewska, 1975), was not considered in this model.
Chemical-specific modeling parameters were either: (1) measured in vitro in rats (PC [Teo et al.,
1994]; macromolecular interaction constants [Gargas et al., 1995]), (2) fitted to the experimental
data by the model (metabolism parameters: Vmax and Km), or (3) estimated from the literature
(CEO elimination constants).
The model assumed that rats do not have EH activity, based on data from Kedderis and
Batra (1993). As discussed previously in Section 3.3.1, this assumption is probably incorrect,
based on data from de Waziers et al. (1990), Guengerich et al. (1981), and Kopecky et al. (1980).
This PBPK model did not scale allometrically EH activity to humans, whose liver microsomes
display EH activity toward CEO (Kedderis and Batra, 1993). Instead, Kedderis (1997) estimated
this activity in humans in vivo based on the ratio of EH to CYP450 activity in subcellular hepatic
fractions multiplied by the CYP450 activity in vivo. In addition, the human blood-to-air PC for
AN was determined experimentally (Kedderis and Held, 1998). Because no in vivo
pharmacokinetic data were available to validate the human model, human in vitro data were
scaled to experimental data obtained from rats by using a "parallelogram approach."
The model by Kedderis et al. (1996) overestimates the oral exposure of venous blood
CEO concentration by 3- to 10-fold. This is a systematic bias, which suggests that the model
parameters need to be reevaluated or that the model is not capturing an important kinetic process
for CEO clearance. In particular, for all tissues except the liver, the measured rise in CEO
concentration for the first 10-15 minutes after oral and i.v. exposures is much slower than
predicted by the model, and the model continues to overpredict the oral data for all measured
time points, with the overprediction in the brain being the worst. Kedderis et al. (1996)
suggested that the overestimation of CEO concentration in blood at early time points is "most
likely due to a large intrahepatic first-pass effect." This hypothesis has not been further
evaluated, but, even if correct, the model structure has not been revised to simulate this
phenomenon. The model continues to overestimate CEO levels in blood (albeit to a lesser
degree) at later time points when the hypothesized first-pass effect would be less of a
contributing factor.
It should also be noted that Kedderis et al. (1996) did not compare their model
predictions to data on the urinary elimination of AN-GSH and CEO-GSH conjugates and on total
activity bound to Hb, as was done in a previous version of the model (Gargas et al., 1995).
While fits to those data may have informed the parameters reported by Kedderis et al. (1996),
this is clearly not the case in this instance. Also, the blood and tissue time-course data shown by
Kedderis et al. (1996) do not include all of the points shown in the previous publication.
Since Kedderis and colleagues (1996) did not include EH activity in their rat model (but
other evidence strongly indicates significant CEO hydrolysis by EH), did not perform a more
global, numerical optimization of their parameters, may not have included the urinary and Hb
data, and the subsequent model fits consistently overpredicted CEO pharmacokinetics in blood
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and brain to a large extent, the model parameterization was revised to improve the model
characterization of the rat pharmacokinetic data and provide a more sound footing for human-rat
dosimetry comparisons and hence risk extrapolation, at least making the results numerically
reproducible. Guengerich et al. (1981) measured CEO hydrolysis with purified rat EH and
obtained a Vmax of 0.3 umol/minute-mg protein and a Km of 800 uM. Since the highest blood
levels observed for CEO were ~2 uM (0.1 ug/mL), to a good approximation hydrolysis can then
be described using a first-order rate constant of 0.3/800 = 3.75 x 10"4 L/minute-mg EH protein.
The EH content of rat liver was measured by de Waziers et al. (1990), who found 0.165 mg
EH/mg MP and a value of 40 mg MP/g liver from Ploemen et al. (1997) can be applied.
Finally, the liver fraction for the rat as used in the model 40 g/kg BW was applied,
assuming a standard 0.25 kg rat. The rate constant for a standard rat would then be kEH =
(3.75 x 10"4 L/minute-mg EH) x (0.165 mg EH/mg MP) x (40 mg MP/g liver) x (40 g liver/kg
BW) x (0.25 kg BW) x (60 minutes/hour) = 1.49 L/hour. Because kEH is effectively the ratio,
Vmax/Km, for EH, and a Vmax is expected to scale with BW across species while Km is not
expected to scale with BW, kEH is expected to scale in the same way as a Vmax. In particular,
since kEH is a total activity (rather than an activity per unit volume of liver), it is expected to
scale as BW0'7. Therefore, as for Vmax values in general, one derives a scaled constant, kEHc =
kEH/(0.25 kg)0'7 = 3.92 L/h-kg0'7. Based on similar in vitro (Krause et al., 1997) and in vivo
(Kemper et al., 2001) EH activity data for butadiene mono-epoxide, it is suspected that the in
vivo enzymatic hydrolysis of CEO is also similar to in vitro values, further supporting this direct
extrapolation.
Because of the close dose spacing used in the protein kinase (PK) studies and the limited
data on metabolite disposition, not all of the remaining parameters were readily identifiable from
the in vivo PK data. Therefore, a value of Km for AN oxidation to CEO from in vitro
experiments with rat liver microsomes from Roberts et al. (1991) will be used: 52 uM x 0.05306
mg AN/umol = 2.76 mg/L.
While the model predictions of AN and CEO blood levels after oral exposures tended to
either match or exceed the measured levels, the model predictions of the Hb-binding data were
below the measured levels, indicating that the corresponding binding constants that had been
determined in vitro were too low. Since the data are only for total binding to Hb (AN plus
CEO), the ratio of the AN: CEO binding constants from in vitro was held constant, but the
absolute value of the constants was allowed to vary.
Given the values for Km and the ksnc obtained above, the parameters varied during model
fitting were Vmax (AN oxidation), kFc (GST activity towards AN), kFc2 (GST activity towards
CEO), kA (oral absorption), and kn (AN-Hb binding constant, with the CEO-Hb binding
constant, km, varied in direct proportion, so that kn/km remained equal to the ratio of the values
measured in vitro). An approximate log likelihood function as described by Cole et al. (2001)
was used as a measure of goodness of fit. During initial optimizations, it was assumed that the
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heteroscedasticity values for AN and CEO were distinct but that the same value held for each
compound among all three tissues sampled (brain, liver, and blood). However, examination of
the standard deviations (SDs) noted in the data of Kedderis et al. (1996) and Gargas et al. (1995)
indicated that measurement error was approximately proportional to the signal strength, so the
heteroscedasticities were set to two.
Initial attempts to fit the model to the entire data set by varying other model parameters
were not satisfactory. While the results are not shown here in detail, a possible reason that the
model could not entirely fit the data is that it assumes constant GSH levels in each tissue, while
previous studies (e.g., Benz et al., 1997a) have shown significant GSH depletion at dose levels in
the range of those used in the PK studies. Since GSH depletion will be least significant at the
lowest doses used, and the most interest is in the low-dose range for extrapolation, it was decided
to reestimate the model parameters by using only the lowest concentration pharmacokinetic data:
the oral doses at or below 3 mg/kg, i.v. dose of 3.4 mg/kg, and inhalation concentration of
186 ppm. Also, of the data used to estimate parameters, it was determined that the last AN blood
measurement from both the 3 mg/kg oral data and the 3.4 mg/kg i.v. data appeared to be outliers
since they were well above the otherwise log-linear clearance curve defined by the preceding
data, so they too were not used in the estimation. (Model simulations of all exposure
concentrations and doses compared with the data are shown in figures in this section, but only
the data for the exposure values specified two sentences above were used for fitting.)
The following figures show the model fits obtained with the parameters of Kedderis et al.
(1996) as compared with EPA's revised parameters. (All parameter values are listed in Table C-
1 of Appendix C.) The term "fits" is used here to describe the closeness of model simulations to
the data, recognizing that only a subset of the data, as described in the preceding paragraph and
indicated in the figure legends, was actually used in parameter estimation.) Overall, the revised
model fits proved to be almost identical to the original model fits. Results are nearly identical
for most of the i.v., inhalation, and oral (Figures 3-3 to 3-5a) blood- and tissue-time-course data.
The revised fits to the CEO blood- and tissue-time-course data after oral dosing are slightly
worse than in the original model (Figure 3-6b), but then the fits to the urine and especially the
Hb-binding data (Figure 3-6c) are considerably better than those obtained with the parameters of
Kedderis et al. (1996). Since only a few of the fits were degraded slightly, while others were
improved, the revised model parameters are considered to represent the overall data set at least as
well as the original. The fact that the revised parameters are specifically defined by the low-dose
data, which are closest to the range where the model will be applied and are obtained using an
objective criteria and numerically reproducible methods (the approximate likelihood function
with numerical optimization), gives further support for their use.
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Kedderis etal. 1996
Revised parameters
CEO in blood
+ 3.4 mg/kg data
3.4 mg/kg fit
O 47 mg/kg data
47 mg/kg fit
•* 55 mg/kg data
55 mg/kg fit
x 84 mg/kg data
84 mg/kg fit
AN in blood
0.3 0.6
0.9
1.2 0
Time (h)
0.3
0.6
0.9
1.2
Note: Only 3.4 mg/kg data were used to estimate revised parameters, with circled
point excluded.
Figure 3-3. Intravenous exposure, dosimetry, and model fits.
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Kedderisetal. 1996
0.1
0.03
0.01
0.1
c
I 0.01
0>
u
c
o
O
0.001
0.03
0.01
Revised parameters
V
13 [
j i
CEO in liver
+ 186 ppm data
186 ppm fit
O 254 ppm data
254 ppm fit
* 291 ppm data
291 ppm fit
n c
3.2
3.4 2.8
Time (h)
3
3.2
3.4
Note: Only 186 ppm data were used to estimate revised parameters.
Figure 3-4a. Inhalation exposure, CEO concentrations.
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Kedderisetal. 1996
Revised parameters
V
AN in blood
186ppmdata
186ppmfit
o 254 ppm data
254 ppm fit
291 ppm data
291 ppm fit
AN in liver
AN in brain
3.4 2.8
Time (h)
Note: Only 186 ppm data were used to estimate revised parameters.
Figure 3-4b. Inhalation exposure, AN concentrations.
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Kedderis etal. 1996
Revised parameters
0.1
0.01
0.001
0.1
Rf
£0.01
Q>
(J
c
o
O
0.001
0.1
0.01
0.001
0
1 V '
•/ "^ " " ' • • :
/ 0* * + ^
* °0
' ' .* '
.•'/
'/ — _
.(/ i X
/ +
_L * °9
+ + + CEO in blood] + + ' +
* CEOinbloodf* ;
o+ :
i i i
:o+ ;
1 1 1
CEO in liver
^Q°°^rtr- ;
+ 3 mg/kg data
3t-r-lf-lf\^f-1 f if
rng/Kg TIT
o 10 mg/kg data
10 mg/kg fit
+ 30 mg/kg data
on rnn'l 'ri fit
1 • .jU rng/Kg IIL
S "~ " ~~ •— —
/ -•
0 :
r"°"^oo- ;
f ;
^--
/ ;
CEO in brain :
0^
:
J oo
=
0.2
0.4
o.e o
Time (h)
0.2
Note: Only 3 mg/kg data were used to estimate revised parameters.
Figure 3-5a. Oral exposure, CEO concentrations.
0.4
0.6
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Kedderis et al. 1996
Revised parameters
ou
10
3
1
0.3
10
3*
"5i o
£
c
o
1 1
-u
0)
£0.3
o
O
10
3
1
0.3
n 1
*
•y '*""---..
*•••..
;/oO--^ '••••..
- d ° ~-^. '••*
•I °
]/ 4^^.
+
*/•* ••-••••-....
4- •••••.. :
: 0$-^ '" '•
or o--- . +
=/ °
: +±
:f/^r— -
v/
jl + AN in blood + ;
J . . :
.
n
AN in liver
- r^
I . "X + 3mg/kgdata
'Ox.
I *-T. * "3 K-p-i^yiy-^ fit
. ' O v. o my/KCI Til
j 010 mg/kg data
I o 10 mg/kg fit
* 30 mg/kg data
1 • • "3 ^ ^-^-|^^ylx^^ fit
j ou mg/Kg rn
lii
.
0
f -Q-.___
!° ^^--
1 0
t
1
\ o :
/
i
i
J
0 AN in brain °
r*-^
t ° ^
i o
!
- 1
1
I
I
o
/ -" "~Cr - -• •
/ °
/ °
/
f
f
J
I :
/
I
0
0.2
0.4
0.6 0
Time (h)
0.2
Note: Only 3 mg/kg data were used to estimate revised parameters.
Figure 3-5b. Oral exposure, AN concentrations.
0.4
0.6
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Kedderis et al. 1996
•o 6
t- ,-,
o 3
o ^
E
<1.6
1.2
0.8
0.4
0
Revised parameters
AN-GSH in urine
o
CEO-GSH in urine
o
o
0
10
20 30 0
Dose (rng/kg)
10
20
30
Note: points are data; lines are model fits or simulations. Only data for doses
<2 mg/kg were used to estimate revised parameters.
Figure 3-5c. Oral exposure, urinary excretion, and Hb binding.
Since the human metabolic parameters were extrapolated from in vitro measurements by
using the relationship between in vitro measured and in vivo estimated values from the rat
(Sweeney et al., 2003), it is appropriate to update the human metabolic parameters and model in
parallel with EPA's update of those in the rat. For EH, Sweeney et al. (2003) used the in vivo:in
vitro relationship for the CYP450-mediated metabolism, since there had been no parallel
relationship for hydrolysis. Because in the revised model the rat EH is extrapolated from in vitro
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to in vivo based only on enzyme content, microsomal content, liver fraction, and BW, a parallel
approach for the human would be to do the same, rather than using a correction factor based on a
different class of enzymes. In the case of humans, Kedderis and Batra (1993) measured EH
activity in vitro by using hepatic microsomes, so that use of the amount of enzyme per mg MP
was not needed in the calculation. Kedderis and Batra (1993) determined a Vmax and Km EH-
mediated hydrolysis of CEO using liver samples from six individual humans. The lowest
estimated Km in the group was 600 uM, so again, the metabolism was described as first order,
using the ratio of Vmax/Km. The ratio of Vmax/Km was first calculated for each individual since
Vmax and Km tend to be statistically correlated due to the way they are estimated, and an average
value for the ratio was then determined to be 7.02 x 10"6 L/minute-mg MP. The following were
then applied: the value of 56.9 mg MP per g liver from Lipscomb et al. (2003), the liver fraction
of 25.7 g/kg BW, and the standard value of 70 kg for a human. The rate constant for a standard
human is then kEH = (7.02 x 10"6 L/minute-mg MP) x (56.9 mg MP/g liver) x (25.7 g liver per kg
BW) x (70 kg BW) x (60 minutes/hour) = 43.1 L/hour, assuming it also scales as BW0'7,
kEH/(0.70 kg)0'7 = 2.20 L/hour-kgOJ. Recall that the value for rats was estimated to be
3.92L/hour-kg°'7.
The original PBPK model for humans was assessed for its sensitivity to changes in key
input parameters, and the expected variability in CEO concentrations in humans under different
AN exposure scenarios was estimated (Sweeney et al., 2003). In addition to updating the CEO
hydrolysis rate constant for the human model (and using a first-order equation for that reaction)
as previously discussed, the ratio of Vmax for the oxidation step as estimated in vivo vs. measured
in vitro for the rat was used to estimate the human in vivo Vmaxc, and the enzymatic GSH
conjugation rate constants for AN (kpc) was likewise estimated from the rat estimated in vivo vs.
measured in vitro ratio (kFc2 was unchanged during the reestimation). Since the rat values for
Vmaxc and kpc were revised, and these values were each one leg of the metabolic extrapolation
"parallelogram," the human in vivo values should be accordingly varied. The result is that the
ratio of revised/original human values for each of these constants is simply equal to the
respective revised/original in vivo values for the rat constants; the revised human values are
Vmaxc = 22. 1 mg/hour-kg0'7 and kpc = 77 kg°3/hour. Finally, since the original model
extrapolation assumed that the oral absorption constant, kA, and the Hb-binding constants, kH and
km, were the same in humans as in rats, these assumptions were retained, updating the human
parameters appropriately.
Appendix C provides model source codes written in acslXtreme (AEgis Technologies,
Huntsville, AL) and Matlab (The Mathworks, Inc., Natick, MA) that were used to model
AN/CEO pharmacokinetics with the PBPK models of Kedderis et al. (1996), as revised, and
Sweeney et al. (2003) in this evaluation.
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4. HAZARD IDENTIFICATION
4.1. STUDIES IN HUMANS—EPIDEMIOLOGY AND CASE REPORTS
4.1.1. Oral Exposure
No studies were identified that addressed the exposure of human beings to AN via the
oral route.
4.1.2. Inhalation Exposure
4.1.2.1. Acute Exposure
In a study on six male volunteers, a single 8-hour inhalation exposure of 5-10 mg/m3 AN
produced no subjective symptoms, such as headache, nausea, or general weakness (Jakubowski
etal., 1987).
A report by Chen et al. (1999) collated 144 acute AN poisoning cases that had occurred
between 1977 and 1994 in China. Each case involved a brief workplace accidental exposure to a
high concentration of AN. There were few reliable data on the levels of exposure in these cases,
although transient concentrations of AN were thought to range from 40 to 560 mg/m3 (18-
258 ppm) for 60 cases and may have been over 1,000 mg/m3 for the remaining 84 cases. All but
9 of the 144 subjects were males, ranging in age from 18 to 53 years old. Forty-two of the cases
were considered to have resulted in severe acute AN poisoning, while the rest fell into the mild
acute category.
Table 4-1 summarizes the incidence of symptoms and signs of toxicity that were evident
in the 144 cases. Other changes in monitored biochemical or physiological parameters included
transient increases in peripheral white blood cell (WBC) count to greater than 10 x 109/L in
66 cases and a number of apparent fluctuations in clinical chemistry parameters. Out of
120 subjects whose liver function was monitored, seven subjects showed abnormal increases in
alanine aminotransferase (ALT) (slight), aspartate aminotransferase (AST), and cholylglycine
(i.e., glycocholic acid). Up to 23 times the normal serum concentration of the latter compound
was detected. The seven subjects with abnormal liver functions were poisoned at low
concentrations in air (40-79 mg/m3 [18-36 ppm]), but for a comparatively longer duration (36
hours). Blood levels of GSH were depressed up to 50% in the severe cases, and urine
thiocyanate was increased up to fourfold.
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Table 4-1. Clinical signs in 144 subjects accidentally exposed to AN
Symptoms
Dizziness
Headache
Feebleness
Sore throat
Chest tightness
Cough
Dyspnea
Nausea
Vomiting
Abdominal pain
Numbness of the limbs
Fainting
Convulsion
Cases (%)
144 (100)
144 (100)
144 (100)
87 (60)
144 (100)
16(11)
118(82)
133 (92)
95 (66)
97 (68)
50 (40)
104 (72)
46 (32)
Symptoms
Congestion of pharynx
Hoarseness
Pallor
Profuse diaphoresis
Rough breathing sound
Rapid heart rate
ECG abnormalities3
High blood pressure
Low blood pressure
Liver tenderness
Hepatomegaly and splenomegaly
Coma
Hyperactive knee jerk
Cases (%)
105 (73)
13(9)
108 (75)
95 (66)
18(13)
36 (25)
15 (10)
20(13)
5(4)
9(7)
7(5)
7(5)
137 (95)
aAll 15 instances of electrocardiogram (ECG) abnormalities occurred in cases of severe poisoning.
Source: Chen etal. (1999).
These changes disappeared after poisoned subjects were removed from the accident site
and underwent treatment that included amyl nitrite via inhalation and an i.v. injection of 3%
sodium nitrite followed by 50% sodium thiosulfate (STS). Other treatments included infusion
with glucose, adenosine triphosphate (ATP), coenzyme A, and DEX along with oxygen
inhalation to prevent the development of cerebral edema and to protect brain and liver cells.
In a study that was possibly reflective of combined inhalation and dermal exposure to
AN, Wilson et al. (1948) reported that workers exposed to AN vapors at 16-100 ppm (35-
217 mg/m3) for periods of 20-45 minutes while involved in cleaning operations in polymerizer
facilities complained of dull headache, fullness in the chest, irritation of all mucous membranes
(eyes, nose, and throat), and a feeling of apprehension and nervous irritability. Some workers
also complained of intolerable itching of the skin, but had no clinically demonstrable dermatitis.
Workers who had direct skin contact with AN displayed a sequence of symptoms, including skin
irritation and erythema, followed by bleb formation and then desquamation with slow healing
(Wilson et al., 1948). An earlier report by Wilson (1944) reported the onset of nausea, vomiting,
weakness, headache, fatigue, diarrhea, and an "oppressive feeling" in the upper respiratory
passages following exposure to "mild concentrations" of the compound. Several cases of
reversible mild jaundice and one case of slow-to-reverse severe jaundice were identified and
associated with occasional liver tenderness and low-grade anemia, but it was unclear whether
they were related to inhalation or dermal exposures to only AN or to a combination of AN with
other industrial rubber manufacturing chemicals such as butadiene and styrene.
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4.1.2.2. Chronic Exposure
4.1.2.2.1. Cancer epidemiology
4.1.2.2.1.1. Cohort studies
A retrospective cohort study involving workers at a DuPont textile fibers plant in
Camden, South Carolina, was the first epidemiologic study to address the potential
carcinogenicity of AN (O'Berg, 1980). Shortly thereafter, reports from other cohorts of
AN-exposed workers in Germany, the Netherlands, the United Kingdom, and other parts of the
United States, were published. Other smaller-scale occupational cohorts have also been utilized
to assess the association between AN exposure and adverse human health effects.
Early studies did not utilize quantitative measures of AN exposure, nor were they able to
control for smoking, an important consideration when lung cancer is an outcome of interest.
These factors were being taken into account by the late 1980s. This section provides a summary
of some of the major occupational cohort studies conducted both within and outside the United
States. Some studies presented number of observed and expected cases, but did not calculate the
95% confidence interval (CI) corresponding to a standardized mortality or incidence ratio. In
these cases, these values were calculated by EPA.l
The DuPont studies
Researchers from DuPont conducted a series of studies among male workers potentially
exposed to AN (Symons et al., 2008; Wood et al., 1998; Chen et al., 1987; O'Berg et al., 1985;
O'Berg, 1980). The original study was conducted in South Carolina, with a follow-up study a
few years later. A second DuPont plant, located in Virginia, was the site for another cohort
study. Finally, Wood et al. (1998) combined the cohorts from South Carolina and Virginia to
further investigate the relationship between AN exposure and cancer. Most recently, Symons et
al. (2008) updated this combined cohort, adding 11 years of follow-up for a total follow-up
period of over 50 years.
The first epidemiologic study addressing the potential carcinogenicity of AN (O'Berg,
1980) examined cancer incidence and mortality in a cohort of 1,345 male workers from a DuPont
textile plant who were potentially first exposed to AN between 1950 and 1966. The cohort was
chosen from a larger group of more than 10,000 workers, based on examination of work history
(wage roll employees and some salaried employees), recollection of plant supervisors (wage roll
maintenance employees and salaried employees), and survey of salaried employees. No
individual or area exposure monitoring data were available during the period of study. Several
surrogate measures of potential exposure were utilized in the analyses, including job title, initial
1 Calculated using OpenEPI online calculator (http://www.sph.emory.edu/~cdckms/exact-midP-SMR.html'): Boice-
Monson CI are presented.
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period of first exposure, length of exposure, and payroll (assuming that wage roll employees
have higher potential exposure than supervision or salary roll employees).
Workers were followed through the end of 1976, thus allowing a latency period of at least
10 years for all surviving members of the cohort. Cancer incidence was ascertained through a
company (i.e., Dupont) cancer registry for all cohort members who were diagnosed during their
employment, and cause of death (based on death certificate data) was determined for all cohort
members who died either while employed or after termination. Comparison rates for cancer
incidence and mortality were derived from similarly collected data in a company-wide cancer
incidence and mortality registry system. Additional mortality information was supplied by the
Social Security Administration, and death certificates were obtained for all known deaths. Other
standard comparison rates were also utilized and presented; however, the company-wide data
represented the most comparable control group.
After observing only one incident case of cancer in the 217 employees with <6 months of
exposure, examination of the incidence data was concentrated on the 1,128 employees who had
more than 6 months of exposure. Among this group, cancer incidence was stratified by interval
of diagnosis: 1956-1964, 1965-1969, and 1970-1976. The standardized incidence ratio (SIR =
observed + expected), comparing the observed to expected number of cases based on company-
wide incidence rates was 1.26 (95% confidence interval [CI] = 0.82-1.85) for all cancers for the
full study period (1956-1976); when limited to the latest interval (1970-1976), the SIR ) 1.88,
95% CI = 1.17-2.88. Seventeen of the 19 cases were among wage roll employees, for an SIR of
2.05, 95% CI = 1.23-3.21, in this group (Table 4-2).
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Table 4-2. Distribution of select incidence and mortalities among wage
workers and all workers at an AN plant in South Carolina
Cause of
disease/death
Wage workers
Observe
d cases
SIR3
(CI)
Observed
deaths
SMRa
(CI)
All workers
Observed
cases
SIR3
(CI)
Observed
deaths
SMRa
(CI)
Less than 6 months of exposure
All cancers
Respiratory cancer
1
0
0.77
(0.04-3.19)
0.00
o
5
I
1.30
(0.33-3.55)
1.11
(0.06-5.48)
1
0
0.67
(0.03-3.29)
0.00
o
5
I
1.25
(0.32-
3.40)
1.11
(0.06-
5.48)
Greater than 6 months of exposure
All cancers
(1956-1976)
All cancers
(1970-1976)
Respiratory cancer
(1956-1976)
Respiratory cancer
(1970-1976)
20
17
8
6
1.30
(0.82-1.97)
2.05b
(1.23-3.21)
2.35b
(1.09-4.47)
2.86b
(1.16-5.94)
15
11
6
5
1.16
(0.68-1.87)
1.41
(0.74-2.45)
1.30
(0.53-2.71)
1.61
(0.59-3.57)
24
19
8
6
1.26
(0.82-1.85)
1.88b
(1.17-2.88)
1.95
(0.91-3.71)
2.40b
(0.97-4.99)
17
13
7
6
1.13
(0.68-
1.78)
1.44
(0.80-
2.41)
1.35
(0.59-
2.66)
1.71
(0.69-
3.57)
aSIR and SMR standardized against company-wide rates.
bStatistically significant (p < 0.05).
SMR = standardized mortality ratio
Source: Amended from O'Berg et al. (1980).
Eight incident respiratory cancer cases were observed in the cohort of 1,345 workers. All
eight of the respiratory cancers occurred in the group of employees who had >6 months of
exposure, and all were among wage roll employees. Six of the eight respiratory cancers occurred
in the 1970-1976 time interval compared to 2.1 expected based on company-wide rates (SIR =
2.86, 95% CI = 1.16-5.94) (Table 4-2). The numbers were too small to perform any other
cancer-specific analyses.
In order to consider the effects of latency and level of exposure, further analyses were
performed looking at the group of workers who had first been exposed between 1950 and 1952,
when AN was first used at the plant and when exposures were known to have been highest.
Statistically significant differences between the number of observed and expected cases were
found among the wage roll employees in the diagnosis interval from 1970 to 1976, with
17 cancers observed and 5.6 expected (SIR = 3.04, 95% CI = 1.83-4.78), with six of these cancer
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cases being respiratory cancers with an expected value of 1.5 (SIR = 4.0, 95% CI = 1.62-8.32).
This group is presumed to have the longest latency and highest exposures to AN.
Other factors related to significant differences in cancer incidence between observed and
expected cases among the wage roll employees during the time period 1970-1976 were:
(1) being a mechanic (10 observed vs. 3.7 expected), (2) duration of exposure of more than
10 years (9 observed vs. 2.8 expected), and (3) exposure above "low level" (13 cancers observed
vs. 5.5 expected), low level exposure was not defined. Similar relationships were found among
respiratory cancers except for duration of exposure of more than 10 years.
Incidence data among wage roll employees were examined by duration of exposure,
excluding exposures of <6 months. The duration categories were <5, 5-9, and >10 years. The
SIRs for all cancers in these three duration of exposure categories were 0.7 (95% CI = 0.23-
1.79), 1.2 (95% CI = 0.49-2.50), and 2.3 (95% CI = 1.18-4.14), respectively.
A total of 89 deaths was observed in the entire cohort of 1,345 workers by the end of
1976. The mortality analyses using company-wide rates for comparison did not show any large
deviations between observed and expected numbers for all cancers or for respiratory cancers
alone. There were 11 observed and 7.8 expected cancer deaths between 1970 and 1976 for the
wage roll employees with>6 months of exposure (standardized mortality ratio [SMR] = 1.41,
95% CI = 0.74-2.45), and 5 respiratory cancer deaths observed vs. 3.1 expected (SMR = 1.61,
95% CI = 0.59-3.57) (Table 4-2).
The use of company-wide rates as a comparison may be problematic, however, because
the company-wide rates are based on other plants where exposure to carcinogenic agents may be
possible. Additionally, company-wide rates include both exposed and unexposed workers, thus
weakening the ability to observe increased risk among the exposed workers. The use of this
referent, therefore, may result in a downward bias (i.e., weakening or hiding a true association
between AN and cancer). Additionally, the number of incident cancer cases and cancer-specific
mortalities was too small for further analysis. However, the fact that the cancer/respiratory
cancer comparisons appeared to be stronger for workers with longer latency, longer exposure,
jobs with higher exposure, and work during the early years of observation when exposures were
highest add to the weight of the evidence in support of an association.
Another potential limitation of this study was under-ascertainment of incident lung cancer
cases. Two confirmed lung cancer cases that should have been included in the cohort were
omitted because of clerical error. A validity check of a subsample of 465 employees revealed
five more omissions, suggesting that a problem existed with ascertainment of lung cancer cases.
The omission of these cases may have resulted in the underestimation of the risk estimate for the
incidence of lung cancer. Also, subjective exposure assessment could have resulted in
misclassification of exposure. If the exposure misclassification was random, it would be
expected to bias the risk ratio toward the null value. If it was nonrandom (i.e., differential based
on disease), the effect on the risk estimate could result in a upwardly biased association; this type
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of bias is unlikely, however, given the high probability that exposure classification was made
without knowledge of disease status. Other metrics used in this study, including duration of
exposure, time of diagnosis, and job category, support the finding of an association between AN
exposure and cancer. Another limitation is the indication that workers may have been exposed to
other chemicals in addition to AN. Additionally, there was no adjustment for smoking. Finally,
the small sample size and limited follow-up restricted the ability to detect most cancers.
A second DuPont study extended the follow-up of the above cohort through 1983 for
cancer incidence and 1981 for cancer mortality, adding 7 and 5 years of follow-up, respectively
(O'Berg et al., 1985). The exposure assessment was not updated, and case ascertainment and
calculation of expected incidence and mortality were unchanged from the earlier study. To
evaluate dose-response relationships and latency, cumulative exposure was categorized into three
groups (<2, 2-12, and >13 years), and latency was dichotomized into "less than 20 years" or
"20 or more years." Men with <6 months of exposure were included in the analyses. As with
the earlier study, DuPont cancer registry data were used as a comparison for the cancer incidence
rates observed in the exposed cohort. DuPont and U.S. death rates were used as a comparison
for the mortality data in the exposed cohort.
There were 43 incident cancer cases, including 10 lung cancer cases and 6 prostate cancer
cases (Table 4-3). The authors did not comment on the problems with ascertainment of incident
lung cancer cases described in the O'Berg (1980) study. Two additional cases of lung cancer
were identified. The SIRs for all-cancer incidence were similar to those seen in the full study
period of the earlier study (O'Berg et al., 1980) of this cohort. The SIR for lung cancer was
somewhat attenuated from the earlier study: among wage workers, lung cancer SIR = 1.67, 95%
CI 0.85-2.95) (Table 4-3). Analyses were not stratified by year of diagnosis, but the data were
stratified by latency period (<20 and>20 years) and by a cumulative exposure metric
(summation of the product of the number of years and an exposure level measure, with values of
<2, 2-12, and >13). The lung cancer SIR for a latency period of <20 years was based on 3
observed and 2.5 expected cases (SIR = 1.2, 95% CI = 0.38-3.7). For a latency period of >20
years, the SIR was based on 7 observed and 3.5 expected cases (SIR = 2.0, 95% CI = 0.95-4.2).
The cumulative exposure index also showed a pattern of increasing risk with increasing score,
albeit with wide and overlapping confidence intervals: SIR = 0.71 (95% CI = 0.10-5.1) for the 1
observed and 1.4 expected cases with a score of <2; SIR = 1.7 (95% CI = 0.54-5.2) for the 3
observed and 1.8 expected cases with a score of 2-12; and SIR = 2.1 (95% CI = 0.96-4.8) for the
6 observed and 2.8 expected cases with a score of >13.
The number of prostate cancer cases was higher than expected (6 observed, 1.8 expected,
SIR = 3.3, 95% CI = 1.35-6.93). All prostate cancer cases were observed among wage workers
who had at least 20 years of latency (SIR = 5.5, 95% CI = 2.2-11.3).
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Table 4-3. Distribution of select incidence and mortalities among wage
workers and all workers at an AN plant in South Carolina (updated follow-
up)
Cause of
disease/death
All causes
All cancers
Lung cancer
Prostate
Wage workers
Observed
cases
-
37
10
6
SIRab
(CI)
-
1.25
(0.89-1.70)
1.67
(0.85-2.97)
4.00C
(1.62-8.32)
Observed
deaths
139
31
12
1
SMRa
(CI)
1.18
(1.00-1.39)
1.15
(0.79-1.61)
1.17
(0.63-2.00)
1.11
(0.05-5.48)
All workers
Observed
cases
-
43
10
6
SIRab
(CI)
-
1.17
(0.86-1.56)
1.39
(1.41-4.95)
3.33C
(1.35-6.93)
Observed
deaths
155
36
14
1
SMRa
(CI)
1.15
(0.98-1.34)
1.14
0.81-1.56)
1.21
(0.69-1.98)
1.00
(0.05-4.93)
"SIR and SMR standardized against company-wide rates.
bCalculated based on observed and expected values provided.
"Statistically significant (p < 0.05).
Source: Amended from O'Berg et al. (1985).
A total of 155 deaths was reported, 36 of which were attributed to cancer (SMR =1.1,
95% CI = 0.81-1.56); the results for the full sample were similar to those seen in the wage
workers (Table 4-3). Among the wage workers, 12 lung cancer deaths were observed vs. 10.2
expected (SMR =1.17, 95% CI = 0.63-2.00), and 1 prostate cancer death was observed vs. 0.9
expected (SMR = 1.11, 95% CI = 0.05-5.48). The observation that prostate cancer was elevated
in the incidence data but not in the mortality data could reflect the fact that prostate cancer has a
relatively high 5-year survival rate.
In summary, O'Berg et al. (1985) added approximately 7 years of follow-up to the
previous DuPont cohort. As in the earlier study (O'Berg et al., 1980), these data indicate that
period of diagnosis or latency may be important considerations for the interpretation of the
observed associations. The small number of site-specific cancer cases, however, limits the
ability to draw firm conclusions based on these stratified analyses. Additional aspects of the
analysis should also be noted. The population was exposed to chemicals other than AN, and
these additional exposures were not accounted for in the analyses. Information regarding
smoking habits of the workers in this plant compared with the referent groups was not available.
Chen et al. (1987) assembled a cohort of 1,083 mostly white male workers from a
DuPont plant in Waynesboro, Virginia, that produced the acrylic fiber Orion®. The
manufacturing process was similar to that at the Camden plant except there was greater distance
between the process areas at the Waynesboro facility (Wood et al., 1998). Workers employed at
the plant between 1944 and 1970 who had a potential for AN exposure were included. The
cohort consisted of 805 wage roll employees and 278 salary roll employees. Potential exposure
was based on the review of work histories. No quantitative information was available for
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exposure levels to AN; however, each job was classified as having either low, moderate, or high
levels of AN exposure. Because data prior to 1957 were not available, the investigators analyzed
only deaths between 1957 and 1981.
As with previous DuPont cohort studies, morbidity in the exposed cohort was compared
with company-wide cancer incidence rates, while mortality was compared with company-wide
rates and national rates. Death information was gathered for active and pensioned employees,
and the names of terminated employees were submitted to the Social Security Administration for
identification of vital status. Vital status for 20 people (1.8%) was unknown. A total of 92
deaths was observed, with 21 attributed to cancer. For both wage workers and salary workers,
the number of observed deaths was lower than expected (Table 4-4).
No excess of observed cancer deaths was observed in either worker category. Among the
wage workers, the lung cancer SMR was 0.59 (95% CI = 0.22-1.32) (Table 4-4). No lung
cancer deaths were observed among the salary workers. As for other cause-specific deaths, no
significant trends were detected among either wage or salary workers by time period or duration
of exposure. Analyses by the level of AN exposure and cumulative exposure did not show
significant differences between observed deaths and expected values.
Table 4-4. Distribution of select incidence and mortalities among wage and
salary workers at an AN plant in Virginia
Cause of
disease/death
All causes
All cancers
Lung cancer
Prostate
cancer
Wage workers
Observed
deaths
68
18
5
1
SMRusa
(CI)
0.57b
(0.44-0.78)
0.75
(0.44-1.16)
0.59
(0.22-1.32)
1.11
(0.06-5.48)
SMRDuPont
(CI)
0.77b
(0.61-0.98)
0.88
(0.54-1.37)
0.66
(0.24-1.46)
1.11
(0.06-5.48)
Salary workers
Observed
deaths
24
3
0
1
SMRusa
(CI)
0.41b
(0.27-0.60)
0.24b
(0.06-0.66)
Not
determined
2.0
(0.03-2.47)
SMRDuPont '
(CI)
0.66b
(0.43-0.97)
0.31b
(0.08-0.85)
Not
determined
2.0
(0.03-2.47)
Total (incidence)
Observed
cases
-
37
5
5
CTR a,b
fc3±lvDuPont
(CI)
-
1.01
0.72
(0.26-1.61)
2.63
(0.96-5.83)
""Calculated based on observed and expected values provided.
bStatistically significant (p < 0.05).
SMRDuPont = SMR based on DuPont Registry; SMRu s = SMR based on U.S. statistics
Source: Amended from Chen et al. (1987).
Analyses of cancer incidence or morbidity reported 37 incident cancers in the cohort
(Table 4-4). Twenty-seven were among the wage roll employees (vs. 26.0 expected) and
10 were among the salary roll employees (vs. 10.5 expected). No increase was noted in lung
cancer incidence, with 5 cases observed in the total group vs. 6.9 expected (SIR = observed +
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expected = 0.72, 95% CI = 0.26-1.61). There were 5 observed prostate cancer cases compared
to 1.9 expected based on the company-wide database (SIR = 2.63, 95% CI = 0.96-5.83).
In summary, no evidence of an excess in overall cancer incidence or mortality was found
among AN workers. An increase in prostate cancer incidence was noted. As there is a high rate
of survival for prostate cancer cases, cancer incidence rather than mortality is generally preferred
for the evaluation of AN exposure and disease. The small sample size and number of cancer
cases, subjective exposure assessment, lack of smoking information, and lack of stratification by
latency and dose in the data analysis are limitations of this study. There was the possibility of
under-ascertainment of cases, particularly prostate cancer cases, by relying on the DuPont
Cancer Registry (which is limited to active employees) as the source of information on cancer
incidence. The results of this study support the findings of an excess in prostate cancer by
O'Berg et al. (1985), but the aforementioned study limitations may have hindered the
observation of an association between lung cancer and AN.
Wood et al. (1998) combined the Virginia and South Carolina cohorts described above
(Chen et al., 1987; O'Berg et al., 1985) and analyzed 2,428 male workers with an added decade
of follow-up. The exposure assessment was updated, taking into account plant histories,
descriptions of manufacturing processes, and changes that would affect exposures, a matrix of
job titles, and work area names, documentation of personal protective equipment use, personal
and area air sampling data from 1975, and descriptions of working conditions by long-term
employees. Environmental air samples confirmed changes in plant processes, engineering, and
ventilation. Exposure levels for the period before 1975 were estimated from information
provided by a panel of "knowledgeable employees." Wood et al. (1998) estimated peak AN
exposure in parts per million for a 40-hour work week for each job title/work area combination
and averaged this over a year. Peak exposure level categories, reported as the interval averages,
were low (0.11 ppm), moderate (1.1 ppm), high (11 ppm), and very high (30 ppm). Other
exposure variables analyzed included latency (<20 and>20 years of observation), duration of
exposure (<5, 5-9, and >10 years), and cumulative exposure (<10, >10-<50, >50-<100, and
>100 ppm-years). There was no information on exposure to other chemicals or on smoking
status of the subjects.
The cohort included all employees who worked in exposed areas in either plant until the
plants were closed. The DuPont Cancer Registry was used to identify incident cases of cancer
among employees during employment. Thus, incident cancer cases occurring after tenure at
DuPont would not have been identified. Vital status of past employees was determined through
a review of the National Death Index (1979-1991) and the Social Security Administration (all
years). Vital status of living employees was confirmed through pension records, motor vehicle
records, and credit bureau reports. Expected numbers of deaths were derived from the DuPont
mortality files and the U.S. population. The period of follow-up was extended through 1991 in
the South Carolina plant and 1990 in the Virginia plant. At the end of the follow-up period,
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approximately 18% of the cohort was deceased (n = 454). The entire cohort had a total of
72,083 person-years of follow-up for the mortality analyses and 29,461 person-years for the
morbidity analyses. More than half of the study population was born before 1930, and 82% were
first exposed before 1971, thus allowing adequate follow-up to examine latency and adequate
attained age to examine cancers occurring at older ages. A total of 37% of the population had a
cumulative exposure level estimated at 50 or more ppm-years.
The SMR analysis revealed that the 454 deaths observed in the cohort were below the
expected number of deaths based on U.S. rates (SMR = 0.69, 95% CI = 0.62-0.75). A
comparison of the cohort death rates with those derived from the DuPont mortality registry
showed SMRs that were similar to those derived from the U.S. data. There were 126 deaths from
cancer, which include 46 lung cancer deaths, 27 digestive cancer deaths, and 11 prostate cancer
deaths (Table 4-5). The SMR for all cancers was also lower than expected (SMR = 0.78, 95% CI
= 0.64-0.93), indicating a possible healthy worker effect. Among the site-specific cancer
mortalities, the SMR for prostate cancer was 1.29 (95% CI = 0.64-2.30). For respiratory (lung)
cancer, the SMR for the full cohort was 0.76 (95% CI = 0.56-1.02). In the analysis stratified by
latency, the SMR was 0.50 (95% CI 0.21-1.07) for a latency period of <20 years and 0.79 (9%
CI 0.56-1.08) for a period of >20 years. The SMRs stratified by highest exposure level were 0.0
(95% CI not calculated), 0.70 (95% CI 0.14-2.03), 0.71 (95% CI 0.44-1.09) and 1.23 (95% CI
0.80-1.85) for the low (mean 0.11 ppm), moderate (mean 1.10 ppm), high (mean 11.0 ppm) and
very high (mean 30.0 ppm) groups, respectively.
Cancer incidence in the cohort was compared with company-wide incidence rates derived
from the DuPont Cancer Registry. There were no evidence of an elevated incidence for all
cancers, lung cancer, or digestive cancers, although an excess in prostate cancer cases was
reported (SIR = 1.58, 95% CI = 0.82-2.76) (Table 4-5). There were sufficient deaths in several
categories to allow examination of the patterns of incidence by the four measures of exposure
utilized in the mortality analyses. Specifically, the exposure analyses. There was no evidence of
a higher risk with any measure of exposure for respiratory cancer incidence. For prostate cancer,
the SIR was 1.34 (95% CI = 0.58-2.63) for a latency period of > 20 years and 0.83 95% CI =
0.22-2.12) for < 20 years. A pattern of higher risks for prostate cancer incidence was also seen
with increasing duration of exposure (SIR 1.06, 0.88 and 1.97 in the < 5, 5-9 and > 10 years
groups, respectively) and highest exposure level (SIR 0.0, 1.33, 1.52 and 1.92 across 4 exposure
groups, respectively).
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Table 4-5. Distribution of select incidences and mortalities among exposed
workers in two AN plants in South Carolina and Virginia
Cause of
disease/death
All causes
All cancers
Lung cancer
Digestive
cancer
Prostate
cancer
South Carolina cohort
Observed
deaths
271
77
35
12
5
SMRa
(CI)
0.76b
(0.67-0.85)
0.88
(0.69-1.10)
1.06
(0.74-1.47)
0.57b
(0.29-0.99)
1.23
(0.40-2.86)
Virginia cohort
Observed
deaths
185
50
11
15
6
SMRa
(CI)
0.60
(0.51-0.69)
0.66
(0.49-0.87)
0.40b
(0.20-0.71)
0.81
(0.45-1.33)
1.32
(0.48-2.88)
Combined cohort
Observed
deaths
454
126
46
27
11
SMRa
(CI)
0.69b
(0.62-0.75)
0.78b
(0.64-0.93)
0.76
(0.56-1.02)
0.69
(0.45-1.00)
1.29
(0.64-2.30)
Observed
cases
-
101
17
22
12
SIR3
(CI)
-
0.97
(0.79-1.18)
0.81
(0.48-1.28)
0.89
(0.56-1.34)
1.58
(0.82-2.76)
"SIR and SMR standardized against U.S. mortality rates.
bStatistically significant (p < 0.05).
Source: Amended from Wood et al. (1998).
In summary, this study provided better exposure assessment than previous studies of this
group of workers. Additional follow-up and the combination of two small cohorts enhanced the
information available for inclusion; however, the apparent healthy worker effect may have
masked a relationship between AN exposure and death from or incidence of lung cancer or other
cancers. An additional limitation is that the cancer incidence data only includes cases diagnosed
during the active employment period of the workers.
Symons et al. (2008) provided an update based on the combined Virginia and South
Carolina cohorts from Wood et al. (1998), adding 11 years of follow-up. The exposure
assessment was the same as described in Wood et al. (1998) with exposure being based on a job-
exposure matrix, documentation of personal protective equipment use, personal and area air
sampling data, and descriptions of working conditions by long-term employees. Vital statistics
were obtained through the DuPont Epidemiology Registry (covering active and pensioned
employees) and verified by the National Death Index. New in this update was the assignment of
mean intensity values based on estimated intensity categories ranging from <0.2 to >20 ppm.
These intensity categories were coupled with duration of employment to determine cumulative
exposure. For SMR calculations, the expected number of deaths were derived using both the
U.S. population and regional DuPont employees by means of the Occupational Mortality
Analysis Program (OC-MAP) developed by Marsh et al. (1998). For the estimation of relative
mortality risk via a hazard ratio, a log-linear model for AN exposure was assumed for all-cause
and cause-specific outcomes.
Of the 2,559 workers from Wood et al. (1998), 11 workers were found to be exposed to
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AN for <6 months, and thus were excluded from the cohort in Symons et al. (2008). A total of
839 deaths (32%) were observed in this updated cohort of 2,548 male workers. Of these,
240 deaths were due to cancer, with 88 deaths being specific to lung cancer (Table 4-6). One
quarter of the updated cohort was found to be in the mean intensity exposure group of <2.0 ppm,
while, two-thirds were in the 2.0-19.9 ppm mean intensity exposure category. The number of
deaths, and cause-specific deaths in particular, in these exposure groups was not reported. The
SMR for lung cancer mortality was 0.74 (95% CI = 0.60-0.91) in the full cohort; this estimate
increased to 0.93 (95% CI = 0.74-1.16) in the group with cumulative exposure >10 ppm-years.
Table 4-6. Distribution of select mortalities among exposed workers in two
AN plants in South Carolina and Virginia (updated follow-up)
Cause of death
All causes
All cancers
Lung cancer
Prostate cancer
South Carolina cohort
Observed
deaths
481
144
61
12
SMRa
(95% CI)
0.98
(0.89-1.07)
1.00
(0.84-1.18)
1.14
(0.88-1.49)
0.93
(0.48-1.63)
Virginia cohort
Observed
deaths
358
96
27
13
SMRa
(95% CI)
0.85C
(0.76-0.95)
0.81C
(0.67-0.99)
0.64C
(0.42-0.93)
1.12
(0.60-1.92)
Combined cohort
Observed
deaths
839
240
88
25
SMRa
(95% CI)
0.92C
(0.86-0.98)
0.92
(0.81-1.04)
0.92
(0.75-1.14)
1.02
(0.66-1.51)
SMRb
(95% CI)
0.69C
(0.64-0.74)
0.73C
(0.64-0.82)
0.74C
(0.60-0.91)
0.91
(0.59-1.35)
aBased on company regional rates.
''Based on U.S. population rates.
"Statistically significant (p < 0.05).
Source: Amended from Symons et al. (2008).
Symons et al. (2008) also reported hazard ratio estimates for 100-ppm-year increases in
cumulative exposure. In addition to the cumulative exposure variables used in the crude model,
the adjusted model included the variables 'birth period' (6-decade ordinal variable) and
'employment in South Carolina start-up group' (binary indicator) in generating hazard ratio
estimates. The authors noted an increase in all-cause mortality associated with increasing
cumulative exposure to AN in the crude model, but this effect was attenuated in the adjusted
model and was said to reflect in large part an increased risk of cardiovascular mortality among
older workers. Hazard ratios for the cumulative exposure and for the intensity measure for site-
specific cancers are shown in Table 4-7. Colorectal and brain cancers had elevated hazard ratios,
but were not statistically significant in either the crude or adjusted models.
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Table 4-7. Crude and adjusted hazard ratio estimates for cumulative
exposure and adjusted hazard ratio estimates for exposure intensity
among workers in two AN plants in South Carolina and Virginia
(updated follow-up)
Cause of death
All causes (n=839)
All cancers (n=240)
Lung cancer (n=88)
Prostate cancer (n=25)
Colorectal cancer (n=28)
Brain and CNS (n=6)
Lymphatic and
hematopoietic (n=20)
Cumulative Exposure"
Crude
Hazard
ratio
1.12d
1.07
1.04
0.81
1.13
1.37
1.01
95% CI
1.04-1.21
0.92-1.24
0.81-1.33
0.48-1.35
0.74-1.71
0.59-3.16
0.60-1.72
Adjusted0
Hazard
ratio
1.05
1.00
0.95
0.78
1.16
1.03
0.90
95% CI
0.97-1.14
0.86-1.17
0.73-1.23
0.46-1.32
0.75-1.81
0.38-2.78
0.51-1.60
Intensity1"
Adjusted0
Hazard
ratio0
95% CI
Not reported
1.00
1.09
1.45
0.96
0.75-1.3
0.67-1.77
0.56-3.81
0.41-2.28
Not reported
1.09
0.40-2.96
a per 100-ppm-year increases in cumulative exposure
b mean intensity > 10.0 ppm compared with < 10.0 ppm
TVIodel includes exposure term, birth period, and employment in South Carolina start-up group.
dStatistically significant (p < 0.05).
Source: Amended from Symons et al. (2008).
Hazard ratio estimates were also derived based on lagged cumulative exposure (Table 4-
8). The inclusion of an exposure lag or varying length had little effect on the hazard ratios for all
cancers or lung cancer. Increasing exposure lag resulted in a decreasing hazard ratio for
colorectal cancer, but the opposite pattern (i.e., increasing hazard ratio estimates with increasing
lag period) was observed for brain and CNS cancer and lymphatic and hematopoietic cancer. It
should be noted that the number of events for these cancers was very small (Table 4-8).
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Table 4-8. Adjusted hazard ratio estimated for select cancer mortality by
lagged cumulative exposure for 100-ppm-year increases in cumulative
exposure among workers in two AN plants in South Carolina and Virginia
(updated follow-up)
Cause of death
All cancers
Lung
Prostate
Colorectal
Brain and CNS
Lymphatic and
hematopoietic
5-Yr lag
(n = 2,224)
Events
210
75
22
26
4
16
Hazard ratio3
(95% CI)
1.03
(0.87-1.23)
0.95
(0.70-1.28)
0.86
(0.48-1.53)
1.06
(0.64-1.76)
1.38
(0.43-4.47)
1.09
(0.59-2.01)
10-Yr lag
(n = 2,066)
Events
199
72
19
24
4
14
Hazard ratio3
(95% CI)
0.98
(0.80-1.20)
0.84
(0.59-1.19)
0.98
(0.50-1.92)
0.90
(0.49-1.66)
1.55
(0.44-5.47)
1.29
(0.66-2.54)
15-Yr lag
(n = 1,907)
Events
179
63
15
23
4
13
Hazard ratio3
(95% CI)
0.94
(0.74-1.19)
0.80
(0.53-1.22)
0.95
(0.41-2.22)
0.81
(0.40-1.66)
1.96
(0.49-7.84)
1.27
(0.57-2.86)
3Model includes exposure term, birth period, and employment in South Carolina start-up group.
Source: Amended from Symons et al. (2008).
In summary, Symons et al. (2008) followed workers for over 50 years and, among the
33% mortality in the cohort, found no statistically significant observed excess in overall
mortality in these AN workers. Unlike the previous studies based on the DuPont cohort, Symons
et al. (2008) did not provide cancer incidence data or address previous issues with this cohort
such as smoking status and concomitant exposure to other chemicals. Differences between U.S.
population-based SMRs and regional worker-based SMRs indicated that there may still be a
strong healthy worker effect. Although Symons et al. (2008) categorized workers in different
mean intensity exposure and cumulative exposure groups, the authors did not report the observed
number of cause-specific deaths in these categories or provide cause-specific SMR estimates. In
deriving hazard ratio estimates, Symons et al. (2008) adjusted for workers 'employment in South
Carolina start-up group.' Based on O'Berg (1980), it appears that workers in the South Carolina
start-up group may have had higher levels of exposure to AN, and thus inclusion of this variable
may have resulted in an underestimation of the effect of the other exposure metrics..
Overall summary
The DuPont cohort studies were the first to hypothesize an association between AN
exposure and cancer, specifically lung and prostate cancer. The most recent of these studies also
suggest brain and hematopoietic cancers may be relevant sites for further examination, but these
data are limited by the small number of these specific types of cancers. It should be noted that
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the DuPont cohort studies are limited to male workers and lack appropriate unexposed worker
comparison groups to better assess the risk of developing cause-specific diseases, namely cancer.
Even in the most recent cohort study by Symons et al. (2008), smoking and concomitant
exposures were not adequately addressed.
The DuPont cohort studies (other than Symons et al. , 2008) are the only cohort studies
that evaluate the association between AN exposure and cancer incidence rather than relying on
cause-specific death. For cancer incidence, the cohort studies used data from the DuPont cancer
registry which is limited to cancers diagnosed among current (rather than retired or former
workers). In addition, the quality assurance and quality control procedures for diagnoses in this
Dupont system were not described.
Symons et al. (2008) assessed risk based on mortality. As with the other DuPont studies,
the reliance on mortality statistics from the U.S. population as a comparison group is less
desirable given the potential for bias from the healthy worker effect. Symons et al. (2008)
attempted to address this issue by employing DuPont regional workers as a comparison group
instead of using company-wide statistics as in other DuPont studies. Information regarding other
chemical exposures among regional workers would aid in reducing the uncertainty of exposure
to other potential carcinogens in this comparison group.
As with its predecessor, Wood et al. (1998), the most recent DuPont cohort study by
Symons et al. (2008) may also suffer from exposure misclassification, as no exposure monitoring
existed prior to 1975, raising some uncertainty regarding the interpretation of the exposure-
response analyses. However, even given the shortcomings of each of the individual DuPont
studies, they do provide evidence that there may be an association between AN exposure and the
risk of specific cancers, particularly with consideration of latency.
American Cyanamid Company
Collins et al. (1989) conducted a retrospective cohort study of 2,671 men who worked at
two American Cyanamid Company plants in Louisiana and Florida from start-up (1951 for the
Louisiana plant, 1957 for the Florida plant) through December 1973. One facility manufactured
AN and other materials, and the other facility utilized large quantities of AN in the
manufacturing of acrylic fiber. All 2,671 study subjects were followed through the end of
1983 for mortality, thus allowing at least 10 years of follow-up for the cohort. According to
Collins et al. (1989), for exposure estimation, industrial hygiene monitoring, which began in
1977, was considered representative of previous exposure levels, with adjustments made for
changes in practices and engineering controls. Any actual measurements that were available
were also used to tailor job-specific exposure. The investigators created four exposure
categories: 0-<0.01, 0.01-0.7, 0.7-7.0, and >7 ppm/year. Exposed workers were defined as
having a cumulative exposure of >0.01 ppm/year. Smoking information was available from
medical records for 58% of the workers. Each person in the cohort was coded as smoker
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(smoked for >3 months), nonsmoker (smoked for <3 months), or unknown (no information in
medical record on smoking status). The authors used an internally standardized method of
adjustment for age (<45, 45-54, 55-64, 75+ years), race (white, nonwhite), smoking status
(smoker, nonsmoker, unknown), latency (<10, 10-19, 20+ years), and time period (<1965,
>1965). In the calculation of the SMR, the expected number of cause-specific deaths was
derived from U.S. general population data on men.
More than 50% of the exposed workers in this study were followed for at least 20 years.
By the end of the study, a total of 237 deaths (92 from the unexposed group and 145 from the
exposed group) were observed, and death certificates were located for 224 of these workers.
From both groups combined, there were 65 cancers, including 23 respiratory cancers, of which
22 were lung cancers. For all-cancer deaths, the SMR in unexposed group was 1.08 (95% CI =
0.69-1.61); SMR in exposed group was 1.01 (95% CI = 0.74-1.35). A similar pattern was
observed among the lung cancer deaths (SMR in unexposed group = 1.01, 95% CI = 0.44-2.01
[7 cancer deaths]; SMR in exposed group = 1.00, 95% CI = 0.58-1.61 [15 cancer deaths]). The
SMRs for the AN exposure categories for lung cancer were 1.09 (95% CI = 0.51-2.08), 0.63
(95% CI = 0.10-2.06), 0.64 (95% CI = 0.20-1.53), and 1.41 (95% CI = 0.68-2.58) for categories
0-<0.01, 0.01-0.7, 0.7-7.0, and >7 ppm/year, respectively. Using an internally standardized
method that adjusted for smoking, race, latency, age, and time period, the SMRs were 1.11 (95%
CI = 0.52-2.11), 0.72 (95% CI = 0.12-2.36), 0.71 (95% CI = 0.23-1.72), and 1.22 (95% CI =
0.59-2.23), respectively, for the above exposure categories. None of the trend tests was
statistically significant (information regarding details of the trend test not provided).
The inclusion of an analysis for the unexposed group allowed an evaluation of whether
elevations in cancer rates in the exposed group were also observed in the unexposed group, and
thus were not likely to be related to exposure. It should be noted that indirect standardization
(the use of age-specific mortality rates from the standard U.S. population to derive regional
expected deaths) was used, thus hindering the ability to compare SMRs across groups. This
study attempted to quantify exposure levels and control for smoking history. Approximately
50% of the exposed cohort had at least 20 years of follow-up, thus strengthening the possibility
that the study period included sufficient time to assess effects of a long latency period for
specific types of cancers. However, this study may have had insufficient power as portrayed by
the observation of only 15 lung cancer deaths in the exposed group. It should also be noted that
this study, as with many other cohorts assembled to assess the relationship between AN and
cancer, consists of only men.
Unlike other studies that assumed early exposure levels were higher, Collins et al. (1989)
assumed that AN exposure levels in 1977 were representative of the time frame prior to that date.
This assumption may have led to exposure misclassification, resulting in a flattening of the dose-
response gradient. Other limitations included low statistical power to evaluate lung or rarer
cancers, particularly in subgroup analyses, incomplete smoking information, and use of SMRs
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rather than comparable unexposed controls. In summary, this study provided only limited data
that can be used to assess the relationship between AN exposure to cancer deaths.
Synthetic chemical plant in the U.S.
Waxweiler et al. (1981) examined the mortality rates in a cohort of 4,806 chemical plant
workers who were exposed to many potential carcinogens, including AN. The cohort was
identified as all workers at the synthetic chemical plants who were first employed between
1942 and 1973. These workers were followed through the end of 1973. Since the majority of
the cohort (63%) were actually hired before 1954, this allowed for at least 20 years of latency
and follow-up for the majority of the workers. However, it should be noted that some workers
had less than 1 year of follow-up, latency, or exposure. The cohort was described as young, with
only 30% of live workers being over 55 years of age.
Mortality rates in the plant cohort were compared to those derived from the U.S. white
male population. Under the assumption that all workers were considered at risk from the first
day of employment, no difference was noted between the observed deaths and the expected
deaths (556 observed deaths vs. 550.2 expected deaths, SMR = 1.01, 95% CI = 0.93-1.10). The
SMR for all cancers was 1.18 (95% CI = 0.97-1.42), based on 109 observed and 92.5 expected
deaths. Examination by type of cancer revealed higher numbers of respiratory system and CNS
cancers among the workers (respiratory cancer SMR = 1.49, 95% CI = 1.09-1.99; CNS cancer
SMR = 2.09, 95% CI= 1.02-3.84).
A secondary analysis was performed among workers who had at least 10 years of
exposure. The authors reported 39 observed deaths from cancers of the respiratory system
compared to 25 expected (SMR =1.56, 95% CI = 1.12-2.11). A similar elevation had been
observed by the authors in a previous publication (Waxweiler et al., 1976) that focused on the
workers with high vinyl chloride monomer exposure; thus, the authors concluded that the excess
lung cancer risk may not solely be due to exposure to a single chemical.
Serially additive expected dose modeling was performed to determine whether exposure
to chemicals, including AN, was associated with excess lung cancer risk. Job histories were used
to assess the potential for exposure to 19 chemicals routinely used at the plant. Each job was
assigned a qualitative exposure rating (from 0 for no exposure to 5 for intimate contact on the
skin or high inhalation potential) for each year of the study. Thirty-five of the 80 (56%) job
categories ranked had no exposure to AN, with another 30% reporting minimal to low levels of
exposure. An exposure score was derived for each chemical for all workers by summing each
exposure rating for each year. The exposure score for each cohort member dying from
respiratory system cancers was compared to the score for a group of workers with a similar birth
year and year of first hire (the "subcohort" from which the case arose). Scores in the
nonrespiratory death sub cohorts (expected scores) were subtracted from the observed scores in
the respiratory cancer deaths to obtain an observed-minus-expected cumulative dose difference
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per lung-cancer case. For AN, this calculation resulted in a negative unit, reflecting that
expected cumulative dose greatly exceeded the observed cumulative dose. No significant
differences with regard to AN exposure were observed in this analysis.
While this study showed significant increased risks of death from lung and CNS cancer in
a cohort of workers exposed to multiple carcinogens, limitations with the study design hindered
the ability to assess the relationship between the observed deaths and exposure to AN. The
population of workers was potentially exposed to multiple carcinogens, so any effect of exposure
to AN may have been impossible to measure.
Rubber manufacturing plant
Delzell and Monson (1982) analyzed the mortality among workers from a rubber
manufacturing plant in order to determine if potential exposure to AN might be associated with
excess deaths. The study included 327 white males who had been employed at the plant for at
least 2 years between January 1940 and July 1971. These employees were selected from over
15,000 workers because they worked in departments where AN exposure was most likely.
Mortality information was gathered on both active and pensioned workers through July 1978,
allowing at least 7 years of follow-up for most workers in the cohort. Workers without a record
of death were assumed to be alive. Mortality rates were compared with the white male subset of
the U.S. general population rates, stratifying by cause of death, age, and calendar time.
By mid-1978, a total of 74 deaths (~ 22%) were observed, 22 of which were from various
types of cancers. The SMRs for all causes and for total cancer mortality were 0.8 (95% CI =
0.7-1.0) and 1.2 (95% CI = 0.8-1.9), respectively. Nine lung cancer deaths were observed as
compared to the 5.9 expected (SMR = 1.52, 95% CI = 0.74-2.79). The numbers of deaths from
bladder, lymphatic, and hematopoietic cancers exceeded the expected values, but the number of
deaths within each of these cause was small (<5). Lung cancer deaths were examined by
duration of employment and latency. There were no deaths in the group employed for >15 years.
However, 7 lung cancer deaths were observed as compared to the 4.1 expected in the group with
15 or more years of latency since first exposure (SMR 1.71, 95% CI = 0.81-3.6).
This study was based on a small number of study participants (n=327), with less than a
quarter reaching the study endpoint of death. The study provides no quantitative assessment as
to the level of AN exposure. Mixed exposure is also an issue, since it is noted that butadiene,
styrene, and vinyl pyridine were utilized at the plant during the exposure time frame. However,
there is no information in the study as to whether the study participants were potentially exposed
to these other agents. Finally, though lung cancer was specifically analyzed, no mention of data
collection or adjustment for smoking history was reported. These shortcomings limit the weight
that this study carries with regard to assessing the relationship between AN exposure and cancer
mortality.
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The Netherlands cohort
Three successive papers (Swaen et al., 2004, 1998, 1992) reported mortality analyses of a
Dutch cohort consisting of 2,842 male workers employed >6 months from 1956 to 1979 in eight
companies that produced AN, latex polymer, acrylic fiber, AN polymers, resins, or acrylamide.
This cohort was compared to 3,961 workers at a neighboring plant during the same time frame
who were not exposed to any known carcinogens in the normal work setting. Use of this
reference population aided in minimizing the impact of a potential healthy worker effect.
In the original study (Swaen et al., 1992), exposure assessment for most companies was
conducted using a job matrix model and air AN samples collected in 1978-1979. The exposure
assessment took into account changes in the production process, industrial hygiene control
measures, and work procedures over time. The potential for exposure misclassification was
enhanced by the extrapolation of exposure monitoring data from one plant to the other seven
plants, for which there were no exposure monitoring data. For all of the companies involved,
most jobs were classified in the following ranges: 0-0.5, 0.5-1, 1-2, and 2-5 ppm. However,
for some jobs, only two categories could be distinguished: 0-2 and 2-5 ppm. For all exposed
workers in the study, a cumulative measure of exposure was derived by multiplying the average
concentration for a particular exposure class by the number of years in that class. Although
workers were characterized by duration of exposure and duration of follow-up, these variables
were not used in the evaluation of the relationship between AN and cancer deaths.
Both the exposed and the unexposed groups were compared with national Dutch death
rates to generate SMRs. No direct comparison of rates in the exposed and unexposed groups was
undertaken, with the authors citing differences in age and calendar time between the groups as
the rationale. Nearly half of the exposed group worked with AN for at least 5 years, and 26% of
the unexposed group was followed for >20 years after entry into the cohort. Almost 24% of the
exposed cohort was categorized in the highest category of cumulative exposure of>10 ppm-
years, a cutoff higher than the>8 ppm used in the Blair et al. (1998) study.
Mortality information was collected for both groups through the end of 1987, allowing at
least 8 years of follow-up for all surviving study members. A total of 134 deaths was observed
in the exposed group (SMR = 0.78, 95% CI = 0.65-0.92) and 572 deaths in the unexposed group
(SMR = 0.77, 95% CI = 0.71-0.84). As both exposed and unexposed workers were compared to
a national rate, the lower observed deaths may be attributable to the healthy worker effect.
Among the workers exposed to AN, the number of observed and expected cancer deaths (42
observed deaths vs. 50.8 expected) resulted in an SMR = 0.83 (95% CI = 0.60-1.10). Among
the unexposed workers, the number of observed deaths from cancer of the trachea and lung was
lower than expected (i.e., 67 observed deaths vs. 93.3 expected deaths, SMR = 0.72, 95% CI =
0.56-0.90). Among exposed workers in the highest exposure category, the lung cancer SMR
was 1.11 (95% CI = 0.48-2.19). Additional evidence of the healthy worker effect is provided by
the fact that the observed number of deaths among the unexposed group, regardless of disease
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category, is significantly lower than the expected estimates derived from the national population
statistics.
In an updated analysis by Swaen et al. (1998), an additional 8 years of follow-up yielded
156 more deaths in the exposed cohort and 411 more deaths in the unexposed group, bringing the
total number of observed deaths to 290 and 983, respectively. The exposure assessment was not
updated because more recent exposures were considered to be negligible. For both groups, the
total number of cancers observed was lower than expected. The exposed group had 97 deaths
from neoplasms observed vs. 110.8 expected (SMR = 0.88, 95% CI = 0.71-1.06) and the
unexposed group had 332 deaths observed vs. 400.4 expected (SMR = 0.83, 95% CI = 0.74-
0.92). With the increased follow-up time, slight increases in cancer of the brain (SMR = 1.74,
95% CI = 0.63-3.78), large intestine (SMR = 1.26, 95% CI = 0.57-2.39), and all leukemias
(SMR = 1.67, 95% CI = 0.54-3.90) for the exposed group were observed that were not
previously evident.
The association between AN exposure and mortality, all-cancer mortality, and lung
cancer mortality was examined by utilizing the following exposure variables: latency, peak
exposure, cumulative exposure, respirator use, and exposure to other carcinogens. Overall,
among workers exposed to AN, the number of observed deaths did not differ greatly from
expected.
Swaen et al. (2004) revisited this cohort of AN workers and added 5 years of follow-up to
the analysis, for a minimum of 22 years of follow-up. This updated study added 142 new deaths
for the exposed cohort and 360 deaths to the unexposed cohort, bringing the total number of
observed deaths for 432 and 1,343, respectively. The number of deaths was 2.5-fold higher than
in the original cohort study and accounted for over a quarter of the original study population.
The exposure assessment was not updated from the first study.
As in the previous analyses, SMRs were calculated for both the exposed and unexposed
cohorts, using the Dutch general population rates as a comparison (Table 4-9). In contrast to the
reduced risk of lung cancer deaths in the unexposed group (SMR = 0.78, 95% CI = 0.67-0.92),
the SMR for lung cancer in the exposed group, based on 67 observed vs. 62.5 expected deaths
was 1.07 (95% CI = 0.83-1.36). Analyses by peak exposure, respirator use, and possible
exposure to cocarcinogens were also performed, yielding no indication of elevated site-specific
cancer risks in any of the subgroups. Additional analyses were performed by examining the
SMRs for lung cancer as a function of various measures of dose and latency (Table 4-10). A
slightly increasing SMR was seen with increasing levels of exposure (i.e., 0.92, 1.06, and 1.15
for low, medium, and high exposure, respectively).
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Table 4-9. Distribution of select mortalities among AN-exposed and
unexposed workers in the Netherlands
Cause of death
All causes
All cancers
Lung and
trachea cancer
Exposed workers
Observed
432
146
67
Expected
467.8
164.5
62.5
SMR
0.92
0.89
1.07
95% CI
0.84-1.01
0.75-1.04
0.83-1.36
Unexposed workers
Observed
1,343
447
160
Expected
1,545.2
519.8
203.8
SMR
0.87
0.86
0.78
95% CI
0.82-0.923
0.78-0.943
0.67-0.923
"Statistically significant (p < 0.05).
Source: Amended from Swaen et al. (2004).
Table 4-10. Lung cancer mortality among AN-exposed workers in the
Netherlands, stratified by cumulative dose and latency
Dose
All cancer mortality
Observed
Expected
SMR
95% CI
Lung cancer mortality
Observed
Expected
SMR
95% CI
Low (<1 ppm/yr)
<10 Yrs latency
10-20 Yrs latency
>20 Yrs latency
Total
0
7
10
17
1.9
7.5
11.3
20.7
-
0.93
0.88
0.82
-
0.37-1.91
0.42-1.62
0.48-1.31
-
3
4
7
0.7
2.9
4.0
7.6
-
1.03
1.00
0.92
-
0.21-2.97
0.27-2.53
0.37-1.89
Moderate (1-10 ppm/yr)
<10 Yrs latency
10-20 Yrs latency
>20 Yrs latency
Total
8
31
39
78
8.4
32.1
49.4
89.9
0.95
0.96
0.79
0.87
0.41-1.87
0.66-1.37
0.56-1.08
0.69-1.08
1
16
19
36
3.2
12.4
18.2
33.8
0.31
1.29
1.04
1.06
0.40- 1.58
0.74-2.09
0.63-1.63
0.75-1.47
High (> 10 ppm/yr)
<10 Yrs latency
10-20 Yrs latency
>20 Yrs latency
Total
8
25
18
51
6.0
21.1
26.3
53.4
1.33
1.18
0.68
0.95
0.57-2.62
0.77-1.75
0.40-1.08
0.71-1.26
3
12
9
24
2.5
8.4
10.0
20.9
1.20
1.43
0.90
1.15a
0.24-3.44
0.74-2.49
0.41-1.70
0.75-1.683
""Calculated based on reported data.
Source: Amended from Swaen et al. (2004).
In summary, as with previous analyses of this cohort, the interpretation of the results
from this study is limited by the following: potential misclassification of AN exposure because
of the use of current measures to derive past exposures and the use of subjective information
about exposure, use of a population-based control group, pooling of data from factories with
different kinds of AN production and exposures without adjusting for these differences, and lack
of information on smoking.
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Epidemiology studies based on this Dutch cohort provided better exposure assessment
than studies using the "serially additive expected dose" method described by Waxweiler et al.,
(1981). Here, exposure categories represented estimates based, in part, on actual measurements
rather than ordinal ranking. The number of lung cancer deaths observed among the AN-exposed
workers was higher than expected. The low SMRs observed could indicate the presence of a
potential healthy worker effect, making it more difficult to detect an association with AN.
The Dutch studies utilized an unexposed worker cohort for a "comparison" group. The
use of this unexposed group was not fully explored by developing rates by age and calendar time
in the unexposed group to be used as a basis for developing expected rates in the exposed group.
Instead, the unexposed group was compared to national rates and then used as a "standard" to see
if the patterns of SMRs generated in the exposed cohort looked similar to those generated in the
unexposed cohort. Therefore, the true comparison group used in these studies was the national
population. It should be noted that the lung cancer SMR among workers exposed to AN was
found to be higher than that among unexposed workers. Direct comparison between the exposed
and unexposed workers in terms of an RR was not derived, as the demographics between the
exposed and unexposed workers may have differed. Though no explicit details were provided, it
is noted that the number of person-years and crude mortality rate are higher in the unexposed
workers (recruited from a different plant) than in the exposed workers. Thus, the expected
number of deaths for each group, on which the SMRs are based, would be dependent on different
distributions or standards, a situation in which comparing the SMRs between the exposed and
unexposed groups would not be recommended. Additional support is seen in the increasing
SMR with level of exposure; a progressive increase of 0.92 in the low exposure group to 1.15 in
the high exposure group was reported.
Finally, the statistical measure used in all of the Dutch cohort studies was the SMR. This
statistic allows for the comparison of cause-specific deaths and not the incidence of disease.
Using deaths as a surrogate for disease may underestimate the true relationship between AN
exposure and cancer, particularly for cancer sites with a high survival rate or cancer sites that are
not accurately ascertained using death certificate data.
BASF plants in Germany
A mortality study was conducted among workers from 12 BASF plants in Germany
(Thiess et al., 1980). Though none of these plants manufactured AN, this substance, along with
styrene and butadiene, was used in many of their processes. The number of employees in each
plant varied, ranging from 30 to 334. The first uses of AN at these plants did not occur
simultaneously, but rather over a course of 14 years, from 1954 to 1968. A total of 1,469 active
and former employees who had worked for >6 months processing AN were identified for
mortality follow-up. The cohort was followed through May 15, 1978, and death certificates were
obtained and coded for cause of death. There were 1,081 German workers in the cohort, and the
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vital status was traced for 98% of these workers. Tracing was less successful for the 388 foreign
workers in the cohort, with only 56% follow-up.
No measurements of levels of AN exposure were available for use in this study. It was
noted that prior to 1976, AN was handled manually, and as a result, there may have been higher
exposures during that time (Thiess et al., 1980). In later years, closed systems were utilized,
likely leading to reduced exposures. Due to the lack of exposure measures, this study assumed
that workers were occasionally exposed to levels of >20 ppm for short periods, based on the
manual handling of AN at specific work sites prior to 1976. Several other known carcinogens
were used at some of the facilities, and could have presented a confounding effect in terms of
cancer outcomes. One plant (Plant 5) was subsequently excluded from the analysis because of
the potential for concurrent exposure to p-naphthylamine.
The comparison rates used for the majority of the analyses were derived from mortality
rates for the Federal Republic of Germany. The 1,469 workers accounted for a total of
15,350 person-years of follow-up. A total of 89 deaths were observed, compared with the
99 predicted deaths based on the general population rates (SMR = 0.90, 95% CI = 0.72-1.10). A
total of 27 cancer deaths was observed for the entire cohort, compared to the 20.5 expected
deaths (SMR = 1.32, 95% CI = 0.89-1.89).
In the analysis that excludes Plant 5, the number of observed deaths dropped to 74, with
the expected deaths at 78.8 (SMR = 0.94, 95% CI = 0.74-1.17). Among these deaths, 20 were
attributed to cancer, with a calculated expected value of 16.1 deaths (SMR = 1.24, 95% CI =
0.78-1.88). When cause-specific cancer deaths were examined, a difference was noted between
the observed and expected bronchial carcinoma (lung cancer) deaths, regardless of the inclusion
of Plant 5 in the analysis (i.e., SMR =1.9 based on 11 observed deaths vs. 5.7 expected in the
full cohort and SMR = 2.0 based on 9 observed vs. 4.4 expected in the cohort without Plant 5).
The 95% confidence intervals were not provided but these estimates were noted to be statistically
significant.
Neoplasms of the lymphatic and hematopoietic organs were also observed to be elevated,
albeit based on very limited number of cases (i.e., 4 observed vs. 1.7 expected in the full cohort
and 4 observed vs. 1.4 expected in the cohort without Plant 5; these observations correspond to
SMR = 2.4, 95% CI = 0.88-6.3, and SMR = 2.9, 95% CI = 1.1-7.6, respectively). Because of
the small number of deaths in this category, further stratification or examination was precluded.
It should be noted that two of the four deaths in this category were from Hodgkin's disease,
compared with 0.3 expected deaths (SMR = 6.7, 95% CI 1.7-26.6).
Analyses were conducted, excluding Plant 5, for the group of workers who were followed
for at least 5 years before death or loss to follow-up. This comprised 944 workers, with only
seven bronchial carcinomas reported. These carcinoma deaths were stratified by duration of
exposure into three categories: 0-4, 5-9, and >10 years. There were no bronchial carcinomas in
the 0-4 year exposure category, four in the 5-9-year exposure category and three in the highest
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exposure category. For the latter two categories, the SMR5_9years = 3.86 (95% CI= 1.23-9.31)
and SMR>io years = 2.23 (95% CI = 0.57-6.07).
The study of BASF workers showed an approximate two-fold excess in lung-cancer
deaths in workers potentially exposed to AN. The follow-up time was not specifically
quantified; however, examination of the exposure (1954-1968) and follow-up (5/15/1978) dates
indicates that most workers in the cohort were followed for at least 9 years. Only 6% of the
study cohort was recorded as deceased and thus the number of site-specific cancer deaths is
limited. An additional limitation of this study is the fact that the cohort was assembled from
12 different plants, and some of the plants were acknowledged to have concurrent exposures to
known carcinogens such as p-naphthylamine, vinyl chloride, or solvents. The study did not
indicate when these overlaps in exposures could have occurred or how many workers at the plant
could have been exposed to these other agents, and no information on the level of exposure was
provided.
Six factories in the United Kingdom
Two successive studies have evaluated the mortality rates in a cohort of male workers
potentially exposed to AN at six United Kingdom polymerization and spinning factories (Benn
and Osborne, 1998; Werner and Carter, 1981). For both studies, workers were included if they
were employed for at least 1 year between 1950 and 1968. Werner and Carter (1981) examined
the mortality rates in a cohort of 1,111 men drawn from six factories in England, Wales,
Scotland, and Northern Ireland. This cohort was followed through the end of 1978, allowing for
a minimum 10-year follow-up for all surviving workers. The workers included in the study were
deemed to have the potential for the highest level of AN exposure, since they were involved in
either the AN polymerization process or spinning of acrylic fiber. However, no exposure
monitoring data were available for the period of the study. Additionally, there was potential for
concomitant exposure to styrene and butadiene in the work environment.
Each worker in the cohort was classified according to the length of time spent in a high-
exposure job. This categorization resulted in 934 workers with >1 year and 177 workers with <1
year of potential for high exposure to AN. The remainder of the analyses focused on the
934 workers with >1 year in a job with potential for high exposure. Examination of the person-
years distribution by age group in these 934 workers revealed that <2% of the person-years were
observed in persons over age 65, while 65% of the person-years were observed in the 15 to 44-
year age range.
Among the 934 workers, 68 deaths were observed compared to the 72.4 expected based
on mortality rates for the total male population of England and Wales (SMR = 0.93, 95% CI =
0.73-1.18). Of these observed deaths, 21 were attributed to cancer, with 9 specifically attributed
to cancer of the trachea, bronchus, and lung. The SMRs for all malignant neoplasms and cancer
of the trachea, bronchus, and lung are 1.10 (95% CI = 0.72-1.70) and 1.20 (95% CI = 0.58-
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2.17), respectively. Incidentally, the number of observed stomach cancer deaths was higher than
expected (5 observed vs. 1.9 expected, SMR = 2.63, 95% CI = 0.96-5.83).
The distribution of cancer deaths was also examined by age at death (age groups: 15-44,
45-54, 55-64, and >65 years). The all-cancer SMRs for these groups were 1.93 (95% CI =
0.78- 4.02), 1.17 (95% CI = 0.51-2.31), 0.92 (95% CI = 0.37-1.92) and 0.67 (95% CI = 0.11-
2.20), respectively. Of the 14 deaths observed in the 15 to 44-year age group, 3 deaths were
attributed to cancer of the trachea, bronchus, and lung compared to an expected value of 0.7
(SMR = 4.28, 95% CI = 1.09-11.66).
All-cause deaths were also stratified based on the year of first exposure to determine if
cancer risks were higher for those who were exposed in the early years of the factories'
operations (usually a surrogate for longer, and often higher, exposures, as well as longer latency)
compared with more recent hires. Three time periods were examined: 1950-1958, 1959-1963,
and 1964-1968, with the observed number of deaths being 35, 21, and 12, respectively.
Observed and expected deaths from all cancer types, including cancer of the trachea, bronchus,
and lung, were similar in all three time periods, with slightly higher risks seen in the most recent
time period (i.e., 1964-1968). For cancer of the trachea, bronchus, and lung, the SMR for the
time periods 1950-1958, 1959-1963, and 1964-1968 were 1.11 (95% CI = 0.35-2.68), 0.87
(95% CI = 0.15-2.87), and 1.88 (95% CI = 0.48-5.10), respectively.
In order to further examine latency, an analysis of cancer deaths by length of time since
first exposure was performed. No increases in cancers of the stomach or cancers of the trachea,
bronchus, and lung were noted with increasing time since first exposure. Similarly, the observed
increases in cancers of the stomach and of the trachea, bronchus, and lung could not be related to
duration of exposure, year of first exposure, or latency. It should be noted, however, that the
power of the study to assess the significance of these parameters was limited. Furthermore, 25%
of the cohort (foreign workers, who may have had the highest exposure) were lost to follow-up,
potentially introducing a bias resulting from an under-ascertainment of deaths. Another
shortcoming of this study was that the comparison group selected to derive the expected values
may not have been representative of the specific study regions, and the comparisons may have
resulted in a bias downward of the effect estimate because of the healthy worker effect. This
study was also limited by the short length of follow-up and small number of deaths observed.
Aside from the shortcomings of this study, if an effect of AN was to be observed, one would
expect cancer rates to be elevated in the group with the earliest years of first exposure, which
represents the workers with the highest potential for greater exposure, longest follow-up, and
longest latency. The fact that increases were observed in the most recent group (i.e., those first
exposed between 1964 and 1968) weakens the argument that there was an exposure effect
demonstrated in this study, because the latter group probably had lower exposure levels and
fewer than 15 years latency and follow-up.
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In the update of the above study, Benn and Osborne (1998) expanded the sample to
include workers employed in the same six AN polymerization and acrylic fiber spinning
factories from 1969 to 1978 and extended follow-up through 1991, allowing for at least 13 years
of follow-up. Added to the study were craftsmen with possible AN exposure, control laboratory
workers, other possibly exposed workers, and unexposed workers. The sample size was
2,763 men employed for >1 year. The exposure assessment in this study was based on a
threshold limit enacted by the British government in 1981, a few measurements from the late
1970s, and an exposure estimated for the period 1958-1977 by a chemist at one of the factories.
The available exposure measurements were found to be lower than the calculated 8-hour time-
weighted average (TWA) (0.4-2.7 vs. 20 ppm, respectively). Job titles were collapsed into three
AN exposure categories: polymer workers and spinners—"high"; craftsmen, control laboratory
workers, and others with possible AN exposure—"other or medium"; and all others into the
"little or no" AN exposure group. No details were provided on how these categories were based
on the previously described exposure estimates. Unlike the earlier study by Werner and Carter
(1981), death rates for England and Wales were used to calculate expected deaths for the
factories that were located in England and Wales. Scottish rates were used for the factories
located in Scotland and Northern Ireland (rates for Northern Ireland were not available).
The 13 years of follow-up resulted in a total 409 observed deaths compared to 485.5
expected (SMR = 0.84, 95% CI = 0.76-0.93). When stratified by cause of death, no significant
differences were noted between observed and expected deaths (Table 4-11). In the case of all-
circulatory diseases, the number of observed deaths was lower than expected (SMR = 0.86, 95%
CI = 0.75-0.99). Upon stratification by level of AN exposure, a higher SMR for lung, trachea,
and bronchial cancers was seen in the highest exposure group (SMR = 1.41, 0.52 and 0.99 in the
high, possible, and little or no categories, respectively). The increasing trend between stomach
cancer mortality and exposure level was even stronger (Table 4-11), and was reported to be a
statistically significant trend. (This trend was driven by the markedly reduced risk, based on one
observed death, in the lowest exposure group.) The distribution of cause-specific deaths was
also examined by age of death, with age categorized in the same manner as by Werner and Carter
(1981). As was seen in the earlier study, the number of observed deaths from respiratory cancers
was higher than expected (5 observed vs. 0.8 expected, SMR = 6.10, 95% CI = 2.23-13.51)
among the youngest (ages 15-44) cohort of AN-exposed workers,
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Table 4-11. Distribution of select mortalities among AN-exposed and
unexposed workers in the United Kingdom
Cause of
death
All causes
All cancers
Lung,
trachea!,
bronchial
cancer
Stomach
cancer0
Circulatory
disease
Exposure level
High
Oa
170
58
27
7
81
Ea
181.
2
50.1
19.1
4.2
86.9
SMR
0.94
(0.80-1.09)
1.16
(0.89-1.49)
1.41
(0.95-2.03)
1.66
(0.73-3.30)
0.93
(0.74-1.15)
Possible
Oa
97
22
7
3
49
Ea
124.7
35.9
13.3
2.9
59.1
SMR
0.78b
(0.63-0.94)
0.61b
(0.39-0.91)
0.52
(0.23-1.04)
1.03
(0.26-2.81)
0.83
(0.62-1.09)
Little or none
Oa
142
41
19
1
70
Ea
179.6
51.1
19.1
4.3
86.2
SMR
0.79b
(0.67-0.93)
0.80
(0.58-1.08)
0.99
(0.62-1.52)
0.23
(0.01-1.15)
0.81
(0.64-1.02)
Total
Oa
409
121
53
11
200
Ea
485.5
137.1
51.5
11.4
232.2
SMR
0.84b
(0.76-0.93)
0.88
(0.74-1.05)
1.03
(0.78-1.34)
0.96
(0.51-1.68)
0.86b
(0.75-0.99)
aO = observed deaths; E = expected deaths.
bStatistically significant (p < 0.05).
"Statistically significant trend (p < 0.05).
Source: Amended from Benn and Osborne (1998).
Cause-specific mortality data were also analyzed by stratifying by year of first exposure,
time since first exposure, and length of exposure. The year of first exposure differed from the
previous study by Werner and Carter (1981) in that workers were divided into the following
three groups: pre-1960, 1960-1968, and post-1968. No differences in all-cause or cancer deaths
were noted between the observed and expected values. For respiratory cancer deaths in the
subgroup of workers that were exposed post-1968, the SMR was 2.70 (95% CI = 1.18-5.32),
based on 7 observed deaths vs. 2.6 expected deaths. The time since first exposure and length of
exposure variables were divided into the following groups: <5, 5-10, 10-15, and >15 years.
When cause-specific deaths, including all cancers and respiratory cancers, were examined within
these categories, no significant differences were noted between observed and expected values.
However, a significant increasing trend was noted for all deaths and circulatory diseases based
on time since first exposure, but not with length of exposure.
The analysis that focused on cause-specific mortality and the year of first exposure
differed from the previous study by Werner and Carter (1981) in that workers were divided into
the following three groups: pre-1960, 1960-1968, and post-1968. No differences in all-cause or
cancer deaths were noted between the observed and expected values, but a significant increase in
respiratory cancer deaths was noted in the subgroup of workers that were exposed post-1968
(7 observed deaths vs. 2.6 expected deaths, SMR = 2.70, 95% CI = 1.18-5.32).
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This study increased the number of observed deaths sixfold as compared with the number
of deaths reported in the previous study by Werner and Carter (1981). The number of cancer
deaths, including deaths from lung cancer, also rose proportionally. The number of observed
deaths is still limited, however, resulting in imprecise estimates in the stratified analyses for site-
specific cancer mortality risk. As seen in the initial study, cancers of the trachea, bronchus, and
lung (respiratory cancers) were elevated in the youngest age group. While significant, because
of the rarity of lung cancer at these ages, this statistic was based on five subjects in the younger
age group. Also noted was the fact that there was an increased risk of deaths from cancers of the
trachea, bronchus, and lung in the workers who were first exposed in 1969 and later, rather than
those first exposed prior to that year. Smoking history was not addressed, which is of particular
concern when examining lung cancer risk. The analyses of the high-exposure group included
fewer than 800 workers from the total cohort; therefore, power to detect trends and excess deaths
was limited; the SMR for lung, tracheal, and bronchial cancers in the high exposure group was
1.41 (95% CI = 0.95-2.03, n = 27). The finding of an increase in stomach cancer in the high
exposure group is interesting, but the small total number of stomach cancer deaths (n = 11) and
the influence of the reduced risk seen in the lowest exposure group makes a definitive
interpretation of these data difficult. In summary, this study had no quantitative measure of AN
exposure levels, the use of a population reference group could have masked associations, and the
statistical power needed to detect rare cancers, such as stomach cancer, was low.
Acrylic fiber factory in Italy
A retrospective cohort study was conducted with 671 male workers who had at least
12 months of exposure to AN (or mixed exposure to AN and dimethylacetamide) at an acrylic
fiber factory in Venezia, Italy (Mastrangelo et al., 1993). Mortality patterns were examined to
determine whether excess cancer cases were related to these exposures. Occupational exposure
to AN occurred between 1959 and the end of 1988, with mortality tracked until the end of 1990,
thus allowing at least 2 years of follow-up for every person in the cohort. Workers with past
exposure to vinyl chloride or benzidine were excluded from the study cohort. Study participants
were categorized based on their level of AN exposure as follows: (1) high exposure to AN only,
(2) low exposure to AN plus exposure to dimethylacetamide, and (3) episodic exposure to AN
plus exposure to dimethylacetamide. SMRs were calculated based on the observed deaths and
the expected number of deaths in the general population. The expected death rate took into
account age, gender, year, cause, and person-time.
During 1959-1990, 32 deaths (4.7% of total cohort) were reported. A total of 12 deaths
from cancer was observed in the cohort. This observation was slightly higher than the 8.73
expected cancer deaths (SMR 1.37, 95% CI 0.78-2.4). None of these cancer deaths was
observed among the 100 workers who were only exposed to high levels of AN. Of the 12 cancer
cases, there were 2 lung cancer cases and 4 intestinal cancer cases. The number of lung cancer
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cases observed among the 272 workers who had discontinuous, episodic exposure to AN and
dimethylacetamide was higher than expected (2 observed versus 1.2 expected),. The intestinal
and colon cancer cases were equally distributed between the groups that had concomitant
exposure to dimethylacetamide and were significantly higher than expected in both groups.
When analyses were stratified by duration of exposure and time since first exposure, no
relationship was found with regard to all-cause mortality or all cancers. Among individuals
exposed for 1-4 years, significant differences were noted between the number of observed
testicular, rectal, intestinal, and colon cancer deaths and the expected values. However, the
number of cancer-specific deaths was small, with only one or two deaths being attributed to the
above cancer types.
In addition to questions about the comparison group, the ability of this study to inform an
assessment of the relationship between AN and cancer is quite limited due to its small sample
size. The follow-up time is also short, as reflected by the fact that less than 5% of the cohort was
deceased by the end of the follow-up period. These factors combined may contribute to the
study's lack of sufficient power to determine if AN exposure is associated with cancer.
National Cancer Institute (NCI) cohort study
Many studies have investigated the relationship between AN exposure and cause-specific
death using a cohort assembled by the NCI (Starr et al., 2004; Marsh et al., 2001, 1999; Blair et
al., 1998). The studies differed by the types of analyses used, comparison groups, cohort
subsets, or years of follow-up.
Blair et al. (1998) assembled a cohort of 25,460 workers (18,079 white males,
4,293 white females, 2,191 nonwhite males, and 897 nonwhite females) who were employed in
AN production or use beginning in the 1950s through 1983. The cohort included workers who
were employed prior to 1984 and after the start-up of AN operations (between 1952 and 1965) at
one of eight plants located in Alabama, Florida, Louisiana, Ohio, Texas, and Virginia. This
method allowed for the examination of both AN-exposed workers and unexposed workers.
Workers were followed through the end of 1989, allowing a minimum of 6 years of follow-up
from time of first exposure.
Exposure was assessed for each plant by developing a quantitative estimate for each job,
department, and time period. Sources used to develop the estimates were walk-through surveys,
personal and area monitoring data, and interviews with longtime workers. The exposure
assessment (Stewart et al., 1998) for this study was more detailed than for any previously
published study. In Stewart et al. (1998), more than 10,000 estimates were developed for 3,662
job, department, and plant combinations for a 30-year period of time. Individual worker
exposure estimates were developed, including estimates for workers whose exposures were
difficult to estimate because of their movement through all areas of a plant (i.e., maintenance
workers). The estimation procedures were reviewed by the Acrylonitrile Advisory Committee,
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the company, union personnel, government, university experts and other interested parties. The
estimation methods were compared with actual data for validation. Other than AN, Stewart et al.
(1998) qualitatively assessed exposure to 340 substances and reported that 25 of these substances
(7%) had exposure levels with more than 20,000 person-years.
SMR analyses were performed to compare observed mortality in both the exposed and
unexposed groups to expected numbers of deaths based on U.S. race- and gender-specific death
rates. Subsequent analyses used the unexposed worker rates to develop internal comparison rate
ratios adjusted for birth year, plant, calendar time, race, gender, wage, and salary status. These
comparisons alleviate the healthy worker effect that is expected when using external general
population rates. Smoking history was obtained on a sample of workers by using a case-cohort
design to allow statistical adjustment for risk estimates calculated for known smoking-related
cancers, such as lung cancer. A 10% sample of living workers was chosen at random to be
interviewed regarding their smoking history. A 10% sample of persons deceased prior to
1983 was also chosen, and next-of-kin interviews were attempted. Also, all brain and lung
cancer deaths that were not chosen in the 10% sample of deceased persons were also selected for
next-of-kin interview.
At the end of the study, the total person-years for the exposed workers was
348,642, while the person-years for the unexposed workers tallied 196,727. More than 66% of
the members of the cohort had at least 20 years of follow-up. A total of 1,217 exposed workers
and 702 unexposed workers were known to be deceased. For all-cause mortality, lower numbers
of observed deaths than expected were reported for both AN-exposed and unexposed groups
(Table 4-12). The patterns for all-cancer and for lung and tracheal cancer deaths were similar.
The SMR for lung and tracheal cancer for AN-exposed and unexposed workers was 0.9 (95% CI
= 0.8-1.1) and 0.8 (95% CI = 0.6-1.1), respectively (Table 4-12). Among the exposed workers,
the numbers of observed deaths from pancreatic cancer, lymphosarcoma, and reticulosarcoma, as
well as noncancer deaths due to diabetes, cerebrovascular disease, and liver cirrhosis, were
significantly lower than expected values. Lung and tracheal cancer was one of the few causes of
death (out of the 29 specific causes) for which the SMR in the exposed workers was higher than
in the unexposed workers (see RR in Table 4-12). The other causes for which this pattern was
seen (not shown in Table 4-12) were esophageal cancer, stomach cancer, rectal cancer, and
Hodgkin's disease.
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Table 4-12. Distribution of select mortalities among AN-exposed and
unexposed workers, United States (NCI study)
Cause of death
All causes
All cancers
Lung and tracheal cancer
Pancreatic cancer
Cerebrovascular disease
Exposed workers
Observed
1,217
326
134
10
37
SMR
0.7
0.8
0.9
0.5
0.5
95% CI
0.6-0.7"
0.7-0.93
0.8-1.1
0.3-0.93
0.4-0.7"
Unexposed workers
Observed
702
216
59
13
23
SMR
0.7
0.9
0.8
1.2
0.5
95% CI
0.7-0.83
0.8-1.0
0.6-1.1
0.7-2.1
0.4-0.83
RR
(95% CI)
0.9 (0.8-1.0)
0.8 (0.7-1.0)
1.2 (0.9-1.6)
0.4 (0.2-1.0)
0.9 (0.5-1.6)
Statistically significant (p < 0.05).
Source: Amended from Blair et al. (1998).
Data were stratified by time since first exposure (<10, > 10-20, and >20 years) to examine
any evidence of latency effects. Using the unexposed group for comparison, no evidence of
increasing risk with increasing latency was seen for lung cancer mortality (RR 0.4, 1.6, and 0.31
in the < 10, >10-20, and >20 years groups, respectively). The other cancers with a RR >1.0 in
the highest latency period were esophageal cancer (RR 3.6, 95% CI 0.4-29.0; 10 exposed cases),
rectal cancer (RR 3.3, 95% CI 0.4-28.2; 7 exposed cases), breast cancer (RR 1.4, 95% CI 0.3-
7.2; 2 exposed cases), and multiple myeloma (RR 1.4, 95% CI 0.3-6.8; 7 exposed cases).
Cumulative exposure was examined by stratifying the data into five exposure groups:
<0.13, >0.13-0.57, >0.57-1.5, >1.5-8.0, and >8.0 ppm-years. There was no evidence of a dose-
response effect across the quintiles of exposure when death rates for all cancers combined were
examined (RR 0.8, 0.9, 0.8, 0.8, 0.8, respectively). Similarly, none of the individual site-
specific cancers exhibited a clear pattern of increasing risk with increasing exposure. For lung
cancer, the RRs were 1.1, 1.3, 1.2, 1.0, and 1.5, respectively, from the lowest to highest exposure
group.
A sufficient number of lung cancer deaths were available to analyze the relationship
between latency and cumulative exposure. Among the lung cancer deaths that occurred within
10 years since first exposure to AN, there was no difference in the mortality risk seen with
increasing exposure (RRs 0.4 for each of the five exposure groups). For lung cancer deaths
occurring 11-19 years after first exposure to AN, the highest risks were seen in the middle
exposure groups (RR 0.5, 2.6, 2.0, 1.2 and 0.9 for the lowest to highest exposures, respectively).
For the workers with >20 years since first exposure, the RR in the highest cumulative exposure
quintile was 2.1 (95% CI = 1.2-3.8), with a trend p-value = 0.11 for the RRs across the five
groups (1.1, 1.0. 1.2, 1.2 and 2.1).
The risk of lung cancer mortality was also analyzed by other exposure variables,
including duration, intensity, frequency of peak exposures, and cumulative exposures,
considering different lag periods and subgroups of workers. Most of these analyses showed a
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pattern that was similar to that seen with the cumulative, unlagged exposure analyses (i.e., RR
1.1, 1.3, 1.2, 1.0 and 1.5 across quintiles of exposure, adjusting for race, gender, age, calendar
time, and salary-wage classification). None of these analyses yielded a strong exposure-response
gradient or statistically significant estimate in the highest category.
Additional analyses focused on the workers with 20 or more years since first exposure in
order to examine the factors that might contribute to increased risk of lung cancer at the highest
cumulative exposure level in this group. The increased risk was not observed when analyses
were restricted to workers first employed between 1960 and 1969. This could be because
workers who were first employed between 1960 and 1969 would have at least 10 fewer years of
follow-up than those hired before 1960 and would tend to be younger. Also, having a later hire
date would allow workers in this group less time to work and accumulate exposure, so they
would be less likely to be included in the highest quintiles of exposure. This is confirmed by the
fact that 42 lung cancer cases were observed in the two highest quintiles for those first hired
before 1960, and 9 were observed in the two highest quintiles for those first hired between
1960 and 1969. The increase in risk in the highest exposure quintile was evident in both wage
and salary employees and for both fiber and nonfiber plants for this group that was followed for
20 or more years after first exposure. None of the trend tests for cumulative exposure in workers
with 20 or more years since first exposure was significant.
Smoking history was sought on a 10% sample of study subjects. A total of 2,655 workers
was identified for interview, and 1,890 (71%) of this group were interviewed. For lung cancer
deaths, 64 next-of-kin interviews were conducted. Additional analyses were performed by
controlling for smoking status. These analyses did not change previous results, although they did
result in a slight reduction in the risk ratios in the highest quintile of exposure. Blair et al. (1998)
stated that they assumed that if smoking data were available for the full cohort, then the smoking
adjustment would yield the same proportion of effects in the full cohort as in the subcohort.
In summary, the sample size and follow-up time in this study were likely large enough to
detect any substantial elevation of cause-specific cancer deaths. For lung and tracheal cancer
deaths, a RR of 1.2 (95% CI = 0.9-1.6) was observed in AN-exposed workers compared to the
unexposed worker population. The study authors state that increased risk with latency or
cumulative exposure in separate analyses was not demonstrated. However, a statistically
significant RR of 2.1 (95% CI = 1.2-3.81) was observed in the highest cumulative exposure
quintile among workers with >20 years since first exposure.
Marsh et al. (1999) added 7 years of follow-up and focused on analyzing a subset of the
original cohort of Blair et al. (1998). The subcohort was comprised of 992 white male workers
who had worked for >3 months between 1960 and 1996 at a chemical plant in Ohio. Analyses
included the calculation of SMRs and RRs for categories of exposure and latency for lung cancer
and the calculation of SMRs (based on regional mortality rates) and RRs for stomach, prostate,
large intestine, and lymphohematopoietic cancers by cumulative exposure level.
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The exposure assessment for this subcohort is similar to but not as rigorous as the
exposure assessment conducted by Stewart et al. (1998) for the original cohort study by Blair et
al. (1998). Marsh et al. (1999) reported that a panel of industrial hygienists used all of the
exposure data collected to assign calendar time-specific categories to each job title. The
following categories were designated: <0.2, >0.2-2.0, 2.1-20.0, and >20.0 ppm. Job titles were
also assessed for the potential for exposure to nitrogen products in a qualitative manner (i.e.,
"potential" versus "no potential"). For AN exposure, the quantitative evaluations for each job
title were used to compute three time-dependent measures of exposure for each worker who held
any job where AN exposure was possible. These measures included duration of exposure,
cumulative exposure, and average intensity of exposure. Cumulative exposure was recorded in
ppm-years, with workers being assigned to one of the following categories: >0-0.139, 0.14-
0.579, 0.58-1.509, 1.51-7.999 ppm-years, and>8.000 ppm-years. Unlike the study by Blair et
al. (1998), which measured person-year accumulation from the first day of employment at the
plant to the end of 1989 or date of death or last date known to be alive, Marsh et al. (1999)
extended their 1960-1988 job exposure matrix to include new job categories and monitoring data
through 1996. Thus, the exposure assessment by Marsh et al. (1999) covers all jobs held from
the beginning of the plant operation (1960) through the end of 1996.
Marsh et al. (1999) mentioned that other known potential occupational hazards at the
plant included asbestos, 1,3-butadiene, and depleted uranium; however, no further information
was provided on the duration, level of exposure, or the actual opportunity for exposure to these
chemicals. In the exposure assessment by Stewart et al. (1998) on the larger cohort, in which
this subcohort is included, the aforementioned chemicals were not singled out as impactful
potential occupational hazards. In Stewart et al. (1998), only asbestos was among the 25
substances that had exposures with more than 20,000 person-years of exposure; however, the
Blair et al. (1998) study on this larger cohort did not observe any deaths from asbestosis or
mesothelioma that would have been indicative of asbestos exposure.
In Marsh et al. (1999), the total cohort of 992 workers included 474 workers who never
worked in a job where AN exposure was possible and 518 workers who were potentially exposed
to AN in at least one of their jobs. A total of 110 deaths was observed in the cohort through the
end of 1996. Smoking history was collected for 90.3% of the total cohort and 93.2% of the AN-
exposed group using a mail survey and review of the medical records. The prevalence of
smokers was similar between the two groups with 58% of the unexposed workers being smokers
and 62.5% of the exposed workers being smokers. However, the prevalence of smoking in the
exposed group was associated with the level of cumulative AN exposure and increased with
increasing levels of exposure.
Of the 110 deaths observed, 43 deaths were attributed to cancer. The observed number of
cancer deaths did not differ from expected values based on the U.S. mortality rates or the
regional death rates, with SMRs of 0.98 (95% CI = 0.71-1.32) and 0.97 (95% CI = 0.70-1.31),
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respectively. Fifteen of the observed cancer deaths were attributed to respiratory cancers (SMR
= 0.88 and 0.92 for the two comparison groups) (Table 4-13). With the exception of bladder
cancer mortality, none of the other site-specific cancers evaluated showed an increased risk using
either the U.S. mortality rates or the regional mortality rates; however, all the deaths occurred
among workers not exposed to AN.
Table 4-13. Distribution of mortality among AN-exposed workers, United
States (one site from NCI study)
Cause of death
All causes
All cancers
Respiratory cancer
Bladder and other
urinary cancer
Exposed workers
Ob
41
17
9
0
SMR
(95% CI)
0.58a
0.42-0.79
0.88
0.52-1.42
1.28
0.58-2.42
0.00-8.23
Unexposed workers
Ob
69
26
6
4
SMR
(95% CI)
0.78b
0.60-0.98
1.04
0.68-1.53
0.64
0.24-1.40
7.01a
1.91-17.96
Overall cohort
Ob
110
43
15
4
SMRUS
(95% CI)
0.64a
0.53-0.77
0.98
0.71-1.32
0.88
0.49-1.43
4.50a
1.23-11.53
sMRregjona]
(95% CI)
0.69a
0.57-0.83
0.97
0.70-1.31
0.92
0.51-1.51
3.93a
1.07-10.06
RR
(95% CI)
0.74
0.5-1.1
0.87
0.4-1.7
1.98
0.6-6.9
Not
determined
aStatistically significant (p < 0.05).
Ob = observed deaths
Source: Amended from Marsh et al. (1999).
For respiratory cancers, there were nine observed deaths in the exposed group and six in
the unexposed group, leading to a risk ratio of 1.98 (95% CI = 0.6-6.9) (Table 4-13).
Respiratory cancer deaths were further examined using RR regression models that were able to
adjust for age and calendar time as well as one other potential confounding factor, such as year
of hire or smoking. Two variables were marginally associated with lung cancer risk in the total
cohort (exposed and unexposed workers): duration of employment and time since first exposure.
Smoking was not related to lung cancer risk; therefore, it is probable that there was
misclassification of this risk variable, and use of smoking information in models of time-related
exposure variables to control for a smoking effect was not possible. Regression analyses
performed using categories of duration of exposure, cumulative exposure, and average exposure
compared to no exposure are shown in Table 4-14. For each analysis, the exposed worker RR
was approximately 2 times that for unexposed workers across all categories of exposure. Marsh
et al. (1998) did not conduct a trend test to determine if there was any deviation from a montonic
response, and so the only trend value reported is for the cumulative exposure analysis (trend/? =
0.20).
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Table 4-14. Summary of relative regression analyses for cancer of the
bronchus, trachea, and lung, United States (one site from NCI study)
Exposure measure
Duration (yrs)
Cumulative exposure
(ppm-yrs)
Cumulative exposure
(ppm-yrs)
Average exposure (ppm)
Category
Unexposed
>0-4.9
5.0-13.9
14.0+
Unexposed
>0-7.9
8.0+
Unexposed
>0-7.9
8.0-109.9
110.0+
Unexposed
>0-4.9
5.0-11.9
12.0+
Observed deaths
6
3
3
3
6
2
7
6
2
4
o
J
6
o
J
o
J
o
J
RR
1.00
1.71
2.28
2.15
1.00
1.96
2.07
1.00
1.97
2.15
1.97
1.00
1.97
1.70
2.64
95% CI
0.25-8.94
0.35-11.38
0.34-10.70
0.81-12.04
0.58-7.58
0.18-12.10
0.43-9.33
0.31-9.42
0.31-9.54
0.26-8.26
0.42-12.67
Source: Amended from Marsh et al. (1999).
The risk ratio for all-cancer mortality and AN exposure derived by Marsh et al. (1999)
was similar to the risk ratio derived in the larger-scale study by Blair et al. (1998). Despite the
limitations of the small number of observed deaths and the focus on only white males, this study
indicates that there may be an association between AN exposure and increased risk in respiratory
cancer deaths.
Marsh et al. (2001) performed a sensitivity analysis on data from the original cohort of
Blair et al. (1998), examining dependency of lung cancer RR estimates on selection of referent
populations. Exposure categorization from the Blair et al. (1998) study was retained for this
analysis, but the comparison groups differed. Mortality analyses were performed using U.S.
mortality rates and local county rates. SMRs were calculated for both AN-exposed and
unexposed workers.
Upon comparison with U.S. and regional mortality rates, both exposed and unexposed
workers were found to have significantly lower rates of all-cause mortality. Using the U.S.
morality rates, the SMRs for the unexposed group and the exposed group were 0.75 (based on
702 deaths, 95% CI = 0.7-0.8) and 0.66 (based on 1,217 deaths, 95% CI = 0.6-0.7), respectively.
The SMRs for all-cause mortality utilizing regional rates were nearly identical to those using
U.S. rates. Additionally, the number of lung cancer deaths observed among workers, both
exposed and unexposed, was lower than regional estimates, with SMRs for each group of
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workers being 0.74 (95% CI = 0.6-0.9) and 0.68 (95% CI = 0.5-0.9), respectively, compared
with U.S. population rates. These findings indicate a potential healthy worker effect.
Observed lung-cancer deaths were stratified by the cumulative exposure to AN and time
since first exposure to derive RRs and regional-mortality-rate-based SMRs (Table 4-15). Results
were similar to those reported by Blair et al. (1998). Workers with the highest cumulative
exposures and at least 20 years of exposure were twice as likely to have lung cancer as
unexposed workers (RR = 2.1, 95% C.I = 1.2-3.8). The numbers of observed and expected lung
cancer deaths in this highest exposure, longest duration category did not yield a significantly
elevated risk using external comparisons (SMR = 1.07, 95% CI = 0.7-1.6). This illustrates how
the use of an external comparison population can mask an apparent association.
Table 4-15. Distribution of observed lung cancer deaths among AN-exposed
workers, using regional rates for comparison, United States (NCI study)
Cumulative
exposure
(ppm-yrs)
>0-0.13
0.13-0.57
0.57-1.50
1.50-8.00
>8.00
Time since first exposure
Less than 10 yrs
Ob
7
o
J
2
2
1
RR(CI)
0.4
(0.2-1.2)
0.4
(0.1-1.4)
0.4
(0.1-1.6)
0.4
(0.1-2.0)
0.4
(0.1-3.1)
SMR (CI)
0.72
(0.3-1.5)
0.63
(0.1-1.8)
0.70
(0.1-2.5)
0.87
(0.1-3.1)
0.81
(0.02-4.5)
10-19 Yrs
Ob
9
12
10
7
4
RR(CI)
0.5
(0.5-3.2)
2.6a
(1.2-5.7)
2.0
(0.9-4.8)
1.2
(0.5-3.1)
0.9
(0.3-1.2)
SMR (CI)
0.71
(0.3-1.4)
1.18
(0.6-2.1)
1.01
(0.5-1.8)
0.66
(0.3-1.4)
0.54
(0.2-1.4)
At least 20 yrs
Ob
11
11
16
18
21
RR(CI)
1.1
(0.6-2.2)
1.0
(0.6-2.2)
1.2
(0.5-2.1)
1.2
(0.6-2.2)
2.T
(1.2-3.8)
SMR (CI)
0.56
(0.3-1.0)
0.57
(0.3-1.0)
0.71
(0.4-1.2)
0.61
(0.4-1.0)
1.07
(0.7-1.6)
Statistically significant (p < 0.05).
Ob = observed deaths
Source: Amended from Marsh et al. (2001).
To look for plant-specific risks, the lung-cancer deaths in each of the eight study plants
were analyzed separately by cumulative level of exposure. Only one plant showed an increased
risk of lung-cancer deaths among the exposed workers, with the highest level of exposure (i.e.,
>8 ppm-years) having an SMR of 2.68 (based on 10 deaths, 95% CI = 1.3-4.9). There was no
comparison of process differences in the plants, thus raising uncertainty as to whether workers at
this plant were exposed to other potential carcinogens or if these findings were due to chance. In
addition, the small number of observed lung-cancer deaths in the plant-specific exposure
categories lowers the power of this study to discern trends in SMRs that might provide more
information on the association between AN exposure and cancer.
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In summary, Marsh et al. (2001) focused on external comparison groups rather than the
available unexposed worker cohort. The use of an external comparison group, rather than a
cohort more comparable to the exposed population (i.e., internal comparison group), may subject
the relationship between exposure and lung cancer to biases such as the healthy worker effect.
Internal comparison groups are generally preferred since they tend to be less influenced by the
healthy worker effect, other sources of selection bias, and confounding (Checkoway et al., 1989).
The reanalysis by Marsh et al. (2001), particularly the RR results, supports the association, found
by Blair et al. (1998), between AN exposure and increased lung cancer mortality risk among
workers with the highest exposure level, but the healthy worker effect, small sample sizes within
subcategories, use of external controls, and short follow-up may limit the detection of stronger
support for such an association.
Data from this cohort were reanalyzed employing semiparametric Cox regression models
with time-dependent covariates to estimate additional risk of death from lung cancer for several
different AN occupational exposure scenarios (Starr et al., 2004). The "cumulative exposure
estimate" was a time-dependent covariate, and "plant worked" was a time-independent covariate.
The analysis focused on the largest race-sex group (18,079 white males) from the original Blair
et al. (1998) study. The Cox models allowed for the calculation of the cumulative risk of dying
from a disease by a certain age. The outcome measurement used was the risk of dying from lung
cancer by age 70 years. Baseline rates were developed for the unexposed worker population and
for three different AN exposure scenarios: (1) early intense exposure, (2) long moderate
exposure, and (3) late intense exposure. All scenarios provided 50 ppm-years of cumulative
exposure by age 55 years. The increased number of lung cancers per 1,000 workers that would
develop by age 70 was calculated as 0.77-1.56, 0.74-1.50, and 0.68-1.56, respectively (the
ranges reflect the use of different plant-specific baseline rates). The upper bound of additional
risk was in the range of 7.5-15.1 per 1,000 workers with the upper bound on the exposure
parameter being 0.0048 per ppm-working year. It is important to note that this study analysis did
not control for smoking. Also, based on the extent of exposure misclassification, any exposure-
response association may have been underestimated.
In summary, the study by Blair et al. (1998) reported an elevated risk of lung cancer
deaths among AN-exposed workers as compared to the unexposed workers, particularly among
workers in the highest cumulative exposure quintile with >20 years since first exposure (RR =
2.1, 95% CI = 1.2-3.8). However, the short follow-up may have contributed to the study's
inability, overall, to demonstrate an increased risk associated with latency or cumulative
exposure. Similar to Blair et al. (1998), Marsh et al. (1999) found an increased risk of lung
cancer deaths among AN-exposed workers as compared to the unexposed workers, though the
point estimate of 1.98 was not statistically significant. However, Marsh et al. (1999) only
focused on a small subset of the NCI cohort. Though the small sample size and the study
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inclusion of only white males reduces the statistical power and generalizability of the study,
excess lung cancer risk was observed in Marsh et al. (1999).
Marsh et al. (2001) utilized primarily an external comparison group. While they provide
better stability of comparison rates compared to internal controls as used by Blair et al. (1998), a
limitation of external controls is potential reduction in the comparability of the populations (i.e.,
a trade-off between precision and validity). The external comparison group increases the risk of
having potential associations masked by a healthy worker effect or other factors that may be
more similar within an occupational cohort than between the cohort and the general population.
The results from Marsh et al. (2001) indicate the presence of such an effect dampening the
observation of an association between AN and cancer.
Though the studies by Marsh et al. (2001) and Blair et al. (1998) have a relatively large
cohort size, the number of observed deaths is small. These studies, though they rely on cancer-
specific mortality ratios rather than cancer incidence as in the Wood et al. (1998) study, have
several advantages to the Wood et al. (1998) study. The low SMR for overall mortality reported
by Wood et al. (1998) suggests the potential presence of a number of biases, including the
healthy worker effect, incomplete cohort identification, and incomplete ascertainment of the
outcome measure. Furthermore, Blair et al. (1998) provide a better job exposure matrix than
Wood et al. (1998). Although cancer incidence is typically preferred, mortality rates serve as a
good surrogate for incidence for some cancers like lung cancer with relatively low survival rates.
In contrast, these mortality studies may miss associations with treatable cancers such as prostate
cancer.
The small percentage of deaths in the NCI cohort suggest a young mean age, which can
impact statistical power. Thus, the observation of a statistically significant elevation in SMR
among those with the longest latency and high cumulative exposure is noteworthy. Additional
follow-up of this cohort may be useful to further assess the association between AN exposure
and cancer suggested by the excess in lung cancer deaths that has been observed among workers
exposed to high levels of AN.
4.1.2.2.1.2. Case-control studies
A large case-control study conducted in seven European countries investigated the
association between occupational exposure to vinyl chloride, AN, and styrene and the risk of
lung cancer (Scelo et al., 2004). The study included new cases of lung cancer occurring between
1998 and 2002 in 15 centers in six Central and Eastern European countries and in Liverpool in
the United Kingdom. Controls, consisting of subjects hospitalized in general public hospitals in
the same areas as cases, were frequency matched to cases based on age and gender. Controls had
to have been hospitalized within 3 months of diagnosis of the case and could not have cancer or
any tobacco-related diseases. A total of 3,403 cases and 3,670 controls met the study inclusion
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criteria; after exclusions and refusals to participate, the final study group included 2,861 cases
and 3,118 controls.
In order to ascertain exposure and lifestyle information, such as tobacco consumption,
each study participant was interviewed using a standard questionnaire. Exposure assessment was
obtained from the work history portion of the questionnaire, where information was collected for
each job held for >1 year. Experts evaluated the frequency and intensity of exposure to AN,
among other agents, for each job held by each study subject. The following exposure models
were constructed for analysis: (1) duration of exposure, (2) weighted duration of exposure
(which considered both duration and frequency), and (3) cumulative exposure in ppm-years.
Models included age, gender, center, tobacco consumption, and other occupational factors.
A total of 10,555 jobs were held by the 2,861 study participants with lung cancer, and a
total of 11,174 jobs were held by the 3,118 controls. AN exposure was associated with 48 jobs
held by the cases and 26 jobs held by the controls. Thirty-nine of the 2,861 cases and 20 of the
3,118 controls were characterized as being exposed to AN, resulting in an odds ratio (OR) of
2.20 (95% CI= 1.11-4.36) (i.e., cases were 2 times more likely to be exposed to AN than
controls). However, it was found that more than half of the study participants that were exposed
to AN were also exposed to styrene. In the determination of the odds ratios of lung cancer for
exposure to acrylonitrile, the authors noted that adjustments were made for the following in their
analysis: center, gender, age, tobacco consumption, vinyl chloride, styrene, carbon black, and
plastics pyrolysis products. In estimated the OR, unconditional logistic regression models were
fit to the data. Variables, as indicated above, were selected for inclusion into the model when
they appreciably modified the OR. This process thereby adjusts for the effect of that variable on
the association between AN exposure and lung cancer. Further analysis was conducted on
individuals exposed to AN and not exposed to styrene (17 cases and 10 controls). This resulted
in a similar OR estimate but with a wider CI due to the smaller sample size (OR = 2.08, 95% CI
= 0.82-5.27). Increasing linear trends for lung cancer were noted for both weighted duration of
exposure (p = 0.05) and cumulative exposure (p = 0.06) (Table 4-16). Additional analyses
employed a 20-year lag for exposures, which did not change the results appreciably. The authors
reported an increased risk of exposure among lung cancer cases diagnosed before the age of 60
(43% of the cases), where the OR for ever being exposed was 2.79 (95% CI = 1.01-7.70), while
the OR for ever being exposed among those over 60 years was 1.02 (95% CI = 0.35-2.92). An
age-exposure interaction test yielded nonsignificant results.
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Table 4-16. ORs of lung cancer for AN exposure, multisite case-control
study in Europe"
Exposure measure
Cases
Controls
OR
95% CI
Weighted duration of exposure (yrs)
Not exposed
0.01-1.00
1.01-2.25
>2.25
2,822
13
9
17
3,098
9
5
6
1.00
2.03
2.73
2.91
0.72-5.73
0.73-10.20
0.87-9.79
Linear trend: p = 0.05
Cumulative exposure (ppm-yrs)
Not exposed
0.01-0.46
0.47-1.61
>1.61
2,822
13
10
16
3,098
9
4
7
1.00
2.03
2.76
2.87
0.72-5.73
0.68-11.22
0.85-9.66
Linear trend: p = 0.06
""Adjusted for center, gender, age, tobacco consumption, vinyl chloride, styrene, carbon black, plastics pyrolysis
products.
Source: Amended from Scelo et al. (2004).
Because of the personal interviews conducted with each case and control, this study was
able to provide an in-depth assessment of potential confounding factors, such as smoking history
and lifestyle factors. The exposure assessment, though different than for the recent generation of
cohort studies where actual plant measurements were possible, utilized reasonable and
standardized methods to assign various levels of AN exposure to different jobs. As a multi-
industry study, the possibility of exposure misclassification in this study was probably greater
than in single-industry studies, though exposure misclassification is probably nondifferential
with respect to disease and therefore would serve to lessen the outcome measures (ORs)
calculated. A key observation in this study is the fact that, even after adjustment for confounding
factors such as smoking history and exposure to other potential carcinogens, the data suggest a
potential association between AN exposure and lung cancer incidence.
4.1.2.2.1.3. Cross-sectional studies
A detailed cross-sectional study was conducted among workers at a Hungarian AN
factory in June 2000 (Czeizel et al., 2004). Of the 888 employees, 72 employees did not work
during the study time frame and 33 refused to participate. The remaining 783 employees were
interviewed, with information gathered on demographics, lifestyle and habits, occupational
exposures, and history of general and occupational diseases, among other factors. Medical
records aided in the validation of the workers' responses. Workers were categorized into three
groups based on level of contact with AN (i.e., direct exposure, indirect or sporadic exposure,
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and no exposure). Since the interviews were done with living current workers, cancer
information for any worker who had left the plant or died before the study was conducted was
anecdotal.
It is known that 12 former workers from the factory had died from cancer between
1990 and 1999, and none of these deaths was due to lung cancer. Of the 783 workers
interviewed, 12 workers were found in the interview sample that had cancer. Of these, only one
lung-cancer case was identified. This worker, categorized as having direct exposure to AN,
started working at the plant in 1973 and was diagnosed 15 years later. Five of the 12 cancer
cases were among the group (n = 452) thought to have the highest level of exposure to AN, while
4 cancer cases were noted among workers with no exposure to AN. No significant association
between AN exposure and the development of cancer was observed. The study mentioned three
of its shortcomings: persons with serious disorders who died or had premature pensions were
not included, updated information on occupational exposures was unavailable, and there was
difficulty in identifying appropriate controls. It should be noted that workers were not exposed
exclusively to AN, but to a mixture of other chemicals as well. The small number of cancer
incidences and the identified shortcomings hinder the reliability of this study in evaluating the
association between AN exposure and the development of cancer.
Other supporting studies
In 2006, the Ohio Department of Health assessed the burden of cancer among residents of
Addyston, Hamilton County, Ohio, who lived near a thermoplastics manufacturing plant that
emitted AN and 1,3-butadiene into the environment (Ohio Department of Health, 2006). The
study population consisted of invasive cancer cases identified through the Ohio Cancer Incidence
Surveillance System between 1996 and 2003. The incidence of site-specific cancer was
compared to the expected number of cases, the latter being derived from national background
cancer incidence rates from NCI's Surveillance, Epidemiology and End Results (SEER) program
in 1998-2002 and region-specific cancer incidence rates from 1993 to 2003. A total of
55 invasive cancer cases were identified among the 1,010 residents in the area, with an SIR =1.8
(95% C.I. = 1.3-2.3) based on the 1998-2002 SEER age-specific incidence rates. Cancer of the
lung and bronchus was the most common cancer identified (13 cases, 23.6%), followed by
colorectal cancer (10 cases, 18.2%). The number of observed cancer cases in both instances was
higher than expected (lung and bronchus SIR = 3.2, 95% C.I. = 1.7-5.4; colon and rectum SIR =
3.0, 95% C.I. = 1.5-5.6). SIRs based on the region-specific cancer incidence yielded similar
results. Although cancer incidence in this study area was higher than expected, the association
between AN exposure and the incidence of lung cancer may be confounded by other risk factors,
such as smoking, that were acknowledged, but not controlled for in the analyses.
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Overall summary of epidemiology data
The early DuPont studies found potential increases in incidence and/or death from lung
cancer and prostate cancer among persons possibly exposed to AN. These studies were the
impetus for generating several additional studies spanning more than 20 years after the first
publication. Exposure assessment was not fully developed until recently, although older studies
did examine qualitative distributions of exposure. The cohorts ranged from a few hundred
workers to over 25,000 workers. The follow-up period for many of the studies was short,
limiting the ability to detect outcomes such as site-specific cancer mortality with a long latency
period. Over time, the cohort studies increased in power by increasing the number of workers,
length of follow-up, and sophistication of the exposure assessment.
A composite of the major cohort studies reviewed, along with all-cancer SMRs, is
provided in Table 4-17. Table 4-18 summarizes SMRs for lung cancer, a cancer type that has
been assessed in most of the epidemiology studies reviewed. In both tables, the SMRs are based
on the cohort population most likely to be exposed to AN, as the actual cohort sample size in
many cases included an unexposed worker group. As in most studies, the number of deaths on
which the SMRs are based is a fraction (<33%) of the actual cohort studied; thus, the statistical
power of the study is better reflected by the number of observed events (i.e., cause-specific
deaths). These SMRs were evaluated by the size of the exposed cohort, number of observed
deaths, percentage of observed deaths within each study, and year of publication to determine if
any of these factors was associated with increased SMR values. For both all-cancer mortality
and lung-cancer mortality, no discernable association was observed between increased SMRs
and the size of the exposed cohort, number of observed deaths, percentage of observed deaths
within each study, or year of publication (data not shown).
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Table 4-17. Derived SMRs for all-cancer mortality and AN exposure in major cohort studies
Reference
Study population
Comparison group
Potentially exposed
cohort
Observed
deaths
All-cancer
SMRa
DuPont
O'Berg (1980)
O'Bergetal. (1985)
Chen etal. (1987)
Chen etal. (1987)
Wood etal. (1998)
Wood etal. (1998)
Symons et al.
(2008)
Symons et al.
(2008)
Male workers exposed to AN between 1950 and 1966 at a
DuPont Plant in South Carolina and followed through 1976.
As above but updated extended to 1983
Male workers exposed to AN between 1944 and 1970 at
DuPont Plant in Virginia and followed through 1983.
As above.
Combined O'Berg et al. (1985) and Chen et al. (1987)
As above.
Update from Wood et al. (1998) with 1 1 yrs of follow-up
As above.
DuPont Registry
DuPont Registry
White male subset of
U.S. population
DuPont Registry
U.S. general population
DuPont Registry
Regional Dupont
workers
U.S. general population
l,128b
1,345
1,083
1,083
2,559
2,559
2,548
2,548
17
36
18C
18C
126
126
839
839
1.13
(0.68-1.78)
1.14
(0.81-1.56)
0.75
(0.44-1.16)
0.88
(0.54-1.37)
0.78d
(0.64-0.93)
0.86
(0.72-1.02)
0.92
(0.81-1.04)
0.73d
(0.64-0.82)
NCI
Blair etal. (1998)
Blair etal. (1998)
Marsh etal. (1999)
Marsh etal. (1999)
Marsh etal. (2001)
Workers employed in eight AN-producing facilities from
1950s to 1983, followed through 1989.
As above.
Subset of Blair et al. (1998).
As above.
Same as Blair etal. (1998)
U.S. general population
Unexposed workers
County mortality rates
Unexposed workers
Regional mortality rates
25,460
25,460
518
518
25,460
326
326
17
17
-
0.80d
(0.70 -0.90)
RR = 0.80
(0.7-1.0)
0.88
(0.52-1.42)
RR = 0.87
(0.4-1.7)
American Cyanamid Company
Collins etal. (1989)
Male workers at two plants employed between 1951 to 1973,
followed throughl 983.
White male subset of
U.S. population
1,774
43
1.01
(0.74-1.35)
Synthetic chemical plant
Waxweiler et al.
(1981)
Chemical plant workers employed between 1942 and 1973,
followed through 1973.
White male subset of
U.S. population
4,806
101
1.18
(0.97-1.43)
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Table 4-17. Derived SMRs for all-cancer mortality and AN exposure in major cohort studies
Reference
Study population
Comparison group
Potentially exposed
cohort
Observed
deaths
All-cancer
SMRa
Rubber industry
Delzell and Monson
(1982)
White male rubber chemical plant workers employed for at
least 2 yrs between 1940 and mid- 1971, followed through
mid-1978.
White male subset of
U.S. population
327
22
1.20
(0.77-1.79)
Netherlands cohort
Swaenetal. (1992)
Swaenetal. (1998)
Swaen et al. (2004)
Male workers exposed to AN in eight factories for>6 mos
before mid-1979, followed through 1987.
As above, but followed through 1995.
As above, but followed through 2000.
Dutch general population
Dutch general population
Dutch general population
2,842
2,842
2,842
42
97
146
0.83
(0.61-1.11)
0.88
(0.72-1.07)
0.89
(0.75-1.04)
BASF (Germany)
Thiessetal. (1980)
Male workers from 12 plants followed through mid-1978.
German mortality rates
1,469
27
1.32
(0.89-1.89)
Six factories (U.K.)
Werner and Carter
(1981)
Benn and Osborne
(1998)
Male workers employed for at least 1 yr in one of six
factories from 1950 to 1968, followed through 1978.
As above but employed from 1969 to 1978, followed through
1991.
Male mortality rates in
England and Wales
Mortality rates from
different European
countries
934
2,963
21
121
1.10
(0.70-1.65)
0.88
(0.73-1.05)
Acrylic fiber factory (Italy)
Mastrangelo et al.
(1993)
General population
671
12
1.37
(0.74-2.33)
aSMRmay be calculated from article based on available data.
bOnly workers employed for>6 mos.
°Based only on wage workers.
dStatistically significant (p < 0.05).
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Table 4-18. Derived associations between lung cancer and AN exposure in major cohort and case-control
studies
Reference
Study population
Comparison group
Potentially
exposed cohort
Observed
lung cancer
deaths or
cases
Lung cancer
associations3
(full cohort)
Lung cancer associations3
(subsets of cohort)b
DuPont
O'Berg (1980)
O'Bergetal. (1985)
Chen etal. (1987)
Chen etal. (1987)
Wood etal. (1998)
Wood etal. (1998)
Symons et al. (2008)
Male workers exposed to
AN between 1950 and
1966 at a DuPont Plant
in South Carolina and
followed through 1976.
As above, but updated
and extended to 1983.
Male workers exposed to
AN between 1944 and
1970 at DuPont Plant in
Virginia and followed
through 1983.
As above.
Combined O'Berg et al.
(1985) and Chen etal.
(1987).
As above.
Update from Wood et al.
(1998) with llyrs of
follow-up
DuPont Registry
DuPont Registry
White male subset of
U.S. population
DuPont Registry
U.S. general
population
DuPont Registry
Regional Dupont
workers
1,128C
1,345
1,083
1,083
2,559
2,559
2,548
7
14
5d
5d
47
47
88
1.35
(0.59-2.66)
1.21
(0.69-1.98)
0.59
(0.22-1.32)
0.66
(0.24-1.46)
0.74e
(0.55-0.98)
0.89
(0.66-1.17)
0.92
(0.75-1.14)
Most recent period (1970-
1976)with exposure > 6 months:
SMR 1.71 (95% CI 0.69-3.57);
SIR =3. 13 (95% CI 1.14-
6.92)
Latency 20 + yrs : SIR = 2.0
(95% CI 0.95-4.2);
Analysis by latency or time
period not provided.
Latency 20 + years: SIR = 0.92
(95% CI 0.51 -1.53)
Most recent period (1975 -
1983) : SIR 1.11 (95% CI 0.70 -
1.68)
Analysis by latency or exposure
of results using US general
population rates not provided.
Exposure level >30 ppm: SMR
1.23 (95% CI 0.80-1. 85)
Duration > 10 yrs: SMR 0.91
(95% CI 0.59 -1.35)
Exposure > 10 ppm-yrs: SMR =
0.93 (95% CI 0.74 -1.16)
HR per 100 ppm-year increase in
cumulative exposure: 0.95 (95%
CI 0.73-1.23)
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Table 4-18. Derived associations between lung cancer and AN exposure in major cohort and case-control
studies
Reference
Symons et al. (2008)
Study population
As above.
Comparison group
U.S. general
population
Potentially
exposed cohort
2,548
Observed
lung cancer
deaths or
cases
88
Lung cancer
associations3
(full cohort)
0.74e
(0.60-0.91)
Lung cancer associations3
(subsets of cohort)b
Analysis by latency or exposure
of results using US general
population rates not provided
NCI
Blair etal. (1998)
Blair etal. (1998)
Marsh etal. (1999)
Marsh etal. (1999)
Workers employed in
eight AN-producing
facilities from 1950s
to!983, followed
through 1989.
As above.
Subset of Blair etal.
(1998).
As above.
U.S. general
population
Unexposed workers
County mortality rates
Unexposed workers
25,460
25,460
518
518
134
134
9
9
0.90
(0.8-1.1)
RR= 1.2
(0.9-1.6)
1.32
(0.64-2.42)
RR= 1.98
(0.6-6.9)
Analysis by latency or exposure
of results using US general
population rates not provided
Exposure > 8 ppm-yr: RR 1.5
(95% CI 0.9 - 2.4)
Latency 20+ years and exposure
> 8 ppm-yr: RR 2.1 (95% CI 1.2-
3.8)
SMR for cumulative exposure
(ppm-yrs):
Unexposed 0.66 (0.24-1.44)
>0-7.9 1.53 (0.19-5.54)
>8.0 1.27 (0.51-2.63)
Adjusted for time since first
employment, RR for cumulative
exposure (ppm-yrs):
Unexposed 1.00 (referent)
>0-7.9 1.27 (0.10-8.94)
8.0-109.9 1.60 (0.29 -7.57)
110.0 2.19 (0.34 -10.70)
Similar patterns seen with
intensity measures
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Table 4-18. Derived associations between lung cancer and AN exposure in major cohort and case-control
studies
Reference
Marsh etal. (2001)
Study population
Same as Blair et al.
(1998).
Comparison group
Regional mortality
rates
Potentially
exposed cohort
25,460
Observed
lung cancer
deaths or
cases
134
Lung cancer
associations3
(full cohort)
0.74e
(0.62-0.87)
Lung cancer associations3
(subsets of cohort)b
Exposure > 8 ppm-yr: SMR 0.92
(95% CI 0.6 -1.4)
Latency 20+ years and exposure
> 8 ppm-yr: SMR 2. 1 (95% CI
1.2-3.8) (using internal
controls)
American Cyanamid Company
Collins etal. (1989)
Male workers at two
plants employed
between 195 land 1973,
followed through 1983.
White male subset of
U.S. population
1,774
15
1.00
(0.58-1.61)
High exposure (> 7 ppm-yr):
SMR 1.22 (95% CI 0.59-2.23)
Synthetic chemical plant
Waxweiler et al.
(1981)
Chemical plant workers
employed between 1942
and 1973, followed
through 1973; exposure
to other potential
carcinogens likely
White male subset of
U.S. population
4,806
42
1.49e
(1.09-1.99)
Duration of exposure > 10 yrs:
SMR 1.56 (95% CI 1.12 - 2.11)
Rubber industry
Delzell and Monson
(1982)
White male rubber
chemical plant workers
employed for at least 2
yrs between 1940 and
mid- 1971, followed
through mid-1978.
White male subset of
U.S. population
327
9
1.52
(0.74-2.79)
Latency 15 + years: SMR 1.71
(95% CI 0.81-3.6)
Netherlands cohort
Swaen etal. (1992)
Male workers exposed to
AN in eight factories for
>6 mos before mid-
1979, followed through
1987.
Dutch general
population
2,842
16
0.82
(0.48-1.30)
High exposure ( >10 ppm-yr):
SMR 1.11 (95% CI 0.48 - 2.19)
99
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Table 4-18. Derived associations between lung cancer and AN exposure in major cohort and case-control
studies
Reference
Swaenetal. (1998)
Swaen et al. (2004)
Study population
As above, but followed
through 1995.
As above, but followed
through 2000.
Comparison group
Dutch general
population
Dutch general
population
Potentially
exposed cohort
2,842
2,842
Observed
lung cancer
deaths or
cases
47
67
Lung cancer
associations3
(full cohort)
1.10
(0.82-1.45)
1.07
(0.84-1.35)
Lung cancer associations3
(subsets of cohort)b
High exposure ( >10 ppm-yr):
SMR 1.18 (95% CI 0.70 - 1.87)
High exposure ( >10 ppm-yr):
SMR 1.2 (95% CI 0.79 - 1.76)
BASF (Germany)
Thiessetal. (1980)
Male workers from 12
plants followed through
mid-1978.
German mortality
rates
1,469
11
1.85
(0.97-3.22)
SMR for duration of exposure
(yrs):
< 4 0.0 (no observed cases)
5-9 3.86 (95% CI 1.22 -9.31
>10 2.04 (95% CI 0.57 -6.07)
Six factories (U.K.)
Werner and Carter
(1981)
Benn and Osborne
(1998)
Male workers employed
for >1 yr in one of six
factories from 1950 to
1968, followed through
1978.
As above but employed
from 1969 to 1978,
followed through 1991.
Male mortality rates in
England and Wales
Mortality rates from
different European
countries
934
2,963
9
53
1.20
(0.58-2.20)
1.03
(0.78-1.34)
Age at death (15- 44yrs): SMR =
4.28 (95% CI= 1.09-1 1.66)
High exposure group: SMR 1.41
(95% CI 0.95-2.03)
Exposures occurring after 1969:
SMR 2.69 (95% CI 1.18 - 5.33)
Age at death (15- 44yrs): SMR =
6. 10 (95% CI = 2.23-13.5)
Acrylic fiber factory (Italy)
Mastrangelo et al.
(1993)
General population
671
2
0.77
(0.13-2.54)
Stratification identified that both
cases occurred within 9 years of
exposure.
100
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Table 4-18. Derived associations between lung cancer and AN exposure in major cohort and case-control
studies
Reference
Study population
Comparison group
Potentially
exposed cohort
Observed
lung cancer
deaths or
cases
Lung cancer
associations3
(full cohort)
Lung cancer associations3
(subsets of cohort)b
Casae-control study (Europe)
Scelo et al. (2004)
Incident lung cancer canses (n=3 118) and
hospitalized controls (n=3118) from 15 centers
in Europe; Detailed work lifetime work history
and expert exposure assessment
Not applicable
3 118 (cases)
OR 2.20
(95% CI 1.11-
4.36)
Younger cases (under age 60):
OR 2.79, 95% CI 1.01-7.70
Exposure > 2.25 yrs: OR 2.91
(95% CI 0.87 - 9.79)
aSMR (may be calculated from article based on available data) or OR.
bSelected results were primarily for higher exposure groups (based on cumulative exposure, duration, or intensity measures) or longer latency periods, if
available, or for other subgroups with elevated risk estimates, if any.
°Only workers employed for >6 mos.
dBased only on wage workers.
"Statistically significant (p < 0.05).
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To date, the largest cohort assessed in determining the relationship between AN and
cancer is the NCI cohort. The NCI/National Institute for Occupational Safety and Health
(NIOSH) cohort study of Blair et al. (1998) provides the strongest evidence for carcinogenicity,
although this evidence is not conclusive. This study of 25,460 subjects in eight plants was
designed to evaluate a relationship between AN exposure and site-specific cancer, including lung
cancer, an a priori hypothesis. Both AN-exposed and unexposed subjects showed a favorable
all-cause mortality rate compared to the all-cause mortality rate of the U.S. population.
Although only 5% of the cohort had died, this study has a large number of observed deaths that
could be analyzed; the 134 lung cancer deaths observed in this study is two- to three times larger
than any of the other cohort studies. The study addressed known problems occurring with earlier
studies by quantifying exposures, estimating the effect of smoking, and using an internal control
group of unexposed workers.
The study by Blair et al. (1998) is of particular interest because a possible association
between AN exposure and death from lung cancer was observed among workers in the highest
level of AN exposure. Specifically, workers with at least 20 years since first exposure and
exposed to a high level of AN were 2 times more likely to die of lung cancer than unexposed
workers (RR 2.1, 95% CI 1.2-3.8). However, the analyses of this and other exposure metrics in
the Blair et al. (1998) study did not provide strong or consistent evidence for an exposure-
response trend as it was only the highest quintile of exposure in which the association was seen.
The observations of Blair et al. (1998) are supported by the patterns seen in analyses stratified by
exposure level, latency period or age group in numerous other studies (Table 4-18). These
results indicate that the association seen in the full study population may not correctly represent
the association seen in specific higher risk groups. This pattern of results is frequently seen in
epidemiological studies of cohorts experiencing a wide range of exposure scenarios and levels; a
true risk in a small group can be attenuated when subsumed within a much larger group that does
not experience this risk. Furthermore, the two-fold increased risk (OR 2.20, 95% CI 1.11-4.36) in
the large lung-cancer case control study of Scelo et al. (2004) provides additional support for this
association. This study adjusted for effects related to individual smoking history and to a
number of potential coexposures found in a subject's occupational setting. The finding of lung-
cancer risk increasing with increasing exposure duration or with increasing cumulative exposure
in this study provides further evidence of an association with AN. It should be noted, however,
that the most recent follow-up of the Dupont cohort (Symons et al., 2008) and the studies of the
Netherlands cohort (Swaen et al., 2004, 1998, 1992) do not provide evidence of an increasing
risk with increasing exposure, or with other characterstics that could be used to define a higher
risk group.
Within the body of epidemiologic literature examining a potential relationship between
exposure to AN and cancer in occupational cohorts, the following shortcomings are noted: low
power from small numbers of exposed subjects and, in many studies, the small number of cases
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of lung cancer or other specific diseases, lack of quantitative exposure information on individual
study subjects leading to a greater potential for exposure misclassification bias, assessment of an
insensitive outcome (i.e., mortality rather than incidence, particular for treatable diseases such as
prostate cancer), and insufficient follow-up period for cancer latency. In addition, many of the
studies used external comparison groups, and their results were subject to a downward bias from
the healthy worker effect.
Blair et al. (1998) and many other studies observed a lower rate of lung-cancer deaths
among AN-exposed workers compared with the general population. Although this could be
interpreted as evidence of no carcinogenic effect of AN, this type of pattern is likely a
manifestation of the healthy worker effect. Blair et al. (1998) addressed this methodological
issue by conducting an additional set of analyses using internal controls (i.e., workers from the
cohort who were not exposed to AN) to draw inferences about AN risks.
The earlier studies based on the DuPont cohort utilized cancer incidence as well as cancer
mortality data. In these studies, although a potentially better outcome measurement (i.e., cancer
incidence rather than mortality) is used, case ascertainment was limited to diagnoses occurring
during employment, and so would have missed diseases among retirees and others who were no
longer working at the company. These studies thus essentially focused on incidence of disease
among relatively young (i.e., < age 65) workers, and are most directly comparable to the age-
stratified data presented by the studies from the United Kingdom (Benn and Osborne, 1998;
Werner and Carter, 1991) and the European case-control study (Scelo et al., 2004).
4.1.2.2.2. Epidemiological studies of AN in humans (noncancer effects).
Several cross-sectional studies of occupational exposure to AN evaluated subjective
symptoms, physical signs, and clinical chemistry parameters among acrylic fiber workers in
Japan and China. A few studies examined neurological effects. These studies are summarized in
Table 4-19, and briefly described below.
Table 4-19. Epidemiology studies of general symptoms, clinical chemistry, and
neurological outcomes among cohorts of workers exposed to AN
Reference
Study population
Exposure assessment
Toxic effects/outcome
Sakurai et al.
(1978)
Cross-sectional
study; December
1975-March
1976
102 male workers in six
Japanese acrylic fiber
plants with averages of 10-
12 yrs of employment; 62
age- matched controls
Mean 8-hr TWAs in
personal samples:
Group A factories =
0.1 ppm; Group B factories
= 0.5 ppm; Group C
factories = 4.2 ppm
Statistically insignificant increase in the
incidence of palpable liver, reddening of
the conjunctiva and pharynx, and skin
rashes compared with controls; prevalence
of subjective symptoms was not evaluated
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Table 4-19. Epidemiology studies of general symptoms, clinical chemistry, and
neurological outcomes among cohorts of workers exposed to AN
Reference
Muto et al.
(1992)
Cross-sectional
study; April -
December 1988
Follow-up of
plants studied by
Sakurai et al.
(1978)
Kaneko and
Omae (1992)
Cross-sectional
study
Chen et al.
(2000)
Cross-sectional
study
Dong et al.
(2000a)
Occupational
survey
Xiao (2000a)
Occupational
survey
Xiao (2000b)
Occupational
survey
Lu et al. (2005a)
Prevalence study
Study population
157 male shift workers
employed in seven
Japanese acrylic fiber
plants with average of
17 yrs employment;
537 unexposed male shift
workers in polyester fiber
plants, power supply and
finishing branches of the
acrylic fiber plants
504 male workers in 7
Japanese acrylic fiber
plants, averages of 5.6, 7.0,
and 8.6 yrs of employment
at low-, medium-, and
high-level workplace
exposure; 249 unexposed
matched controls
224 workers (180 males
and 44 females) in an
acrylic fiber plant with
average of 13 yrs of
employment;
224 unexposed controls at
a different plant
93 workers at a Chinese
chemical fiber plant;
96 unexposed controls
372 workers exposed to
AN in a chemical factory;
186 unexposed controls
237 workers exposed to
AN in a chemical factory;
184 unexposed controls
Chinese acrylic fiber
workers: 81 monomer
workers (68 male and
13 female); 94 fiber
workers (67 male and
27 female); 174 unexposed
workers
Exposure assessment
Personal sampling
(142/157 exposed; 90.4%):
Group A factories = 0.19
ppm; Group B factories =
1.13 ppm
Group L factories =
1.8 ppm; Group M
factories = 7.4 ppm; Group
H factories =14.1 ppm; air
concentrations reported as
"means" without other
information
Average of multiple
samples in four work areas
= 0.48 ppm
Unclear presentation of
exposure data; air samples
in multiple locations over 3
years: 2-2.8 mg/m3
No data
Workshop A = 7 ppm;
workshop B = 3.3 ppm;
workshop C = 3 ppm
Mean area AN
concentrations:
Monomer work areas =
0.11 ppm (0-1.7 ppm);
fiber manufacture work
areas = 0.91 ppm (0-8.34
ppm)
Toxic effects/outcome
Statistically significant increased
prevalence of subjective symptoms in
Group B factories (e.g., heaviness of the
stomach, decreased libido, poor memory,
irritability, conjunctival reddening, eye
irritation); increased prevalence of
heaviness of the stomach in Group A; no
significantly increased prevalence of
physical signs or abnormal values in urine,
hematological, liver function, or blood
pressure variables in Group A or B
Statistically significant increased
prevalence of subjective symptoms, such as
headaches, tongue trouble, choking lump in
chest, fatigue, general malaise, heavy arms,
and heavy sweating in workers in Groups
L, M, and H factories, compared with
controls
Statistically significant increased
prevalence of subjective symptoms (e.g.,
headache, dizziness, poor memory, choking
feeling in the chest, loss of appetite)
compared with controls; increase in serum
y-GTP and USCN but not in other clinical
chemistry or hematological variables.
Increased micronucleus rate in peripheral
lymphocytes
Statistically significant increased
prevalence of subjective symptoms
(headache, dizziness, sleeping disorders,
and a feeling of choking in the chest)
compared with controls
Statistically significant increase in
prevalence of individuals with serum ALT
activity above threshold
Reduction in whole blood cholinesterase
activity; subjective symptoms such as
neurological disorder, sweating, trembling,
and discomfort in the chest
Small but statistically significant deficits in
tests of neurobehavior in monomer and
fiber workers compared with controls;
significant deficits in tests of mood
(increased scores for anger, confusion,
depression, fatigue, and tension), attention
and response speed, auditory memory, and
motor steadiness; not in tests of manual
dexterity or perceptual motor speed
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Table 4-19. Epidemiology studies of general symptoms, clinical chemistry, and
neurological outcomes among cohorts of workers exposed to AN
Reference
Ding et al. (2003)
Cross-sectional
study (in
Chinese, abstract
in English)
Study population
47 AN exposed workers in
chemistry department of
petrochemical company;
47 unexposed teachers and
staff at a college
Exposure assessment
Geometric mean area
concentration: 0.25 mg/m3
(0.11 ppm); 0-3.7 mg/m3
Toxic effects/outcome
Statistically significant increase in
prevalence of deletion rate in mitochondrial
DNA in exposed workers compared to
controls
Sakurai et al. (1978) performed a cross-sectional health examination in 1976 of 102 male
workers exposed to AN in six acrylic fiber manufacturing factories in Japan. Also examined
were 62 nonexposed age-matched control workers from polyester fiber manufacturing plants,
power supply plants, or finishing branches of the acrylic fiber plants. Exposed and referent
workers were selected randomly from the eligible population at each factory and 99.2% and
96.7% of this group were given a health examination. By design, all exposed subjects were shift
workers who had been exposed to AN in the workplace for at least 5 years, but had no history of
exposure to other chemicals. All subjects underwent medical examinations, and blood and urine
samples were collected for chemical analysis. Clinical chemistry analyses included urinary
protein and other parameters, including Hb, total cholesterol, AST, ALT, alkaline phosphatase,
cholinesterase, y-glutamyl transpeptidase (y-GTP), and lactate dehydrogenase (LDH).
Parameters that measured liver injury were a focus of the clinical examinations because a
previous epidemiological study of 576 Japanese acrylic fiber manufacturing workers from the
same factories exposed to <5 or <20 ppm AN between 1960 and 1970 had reported an increase
in subjective symptoms and mild injury to the liver in exposed workers (Sakurai and Kusumoto,
1972). AN and thiocyanate concentrations in urine were also measured to evaluate individual
exposure levels.
Levels of AN in the air (Sakurai et al., 1978) were measured from "spot" samples (55-
159 stationary air samples per factory over two days) and from exposed subjects wearing
personal samplers. A daily time-weighted average concentration was estimated for each worker
from four 100-minute personal samples. As shown in Table 4-20, the factories and exposed
workers were classified into three groups (A, B, and C) according to their level of AN exposure.
SDs for the means and ranges of the exposure concentrations were not reported. Mean exposure
durations for workers in the three groups of factories were 10.3 years (SD = 4.5) for group A,
10.8 years (SD = 4.4) for group B, and 12.6 years (SD = 2.1) for group C.
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Table 4-20. Industrial AN exposure, levels of AN and thiocyanate in urine,
and prevalence of physical signs of adverse effects in workers exposed to AN
at six acrylic fiber factories in Japan
Factory
group
(n = number
of factories)
A(n=2)
B(n=3)
C(n=l)
Control
Mean 8-hr TWA AN
concentration (ppm)
Spot
samples
2.1
(n=116)
7.4
(n=394)
14.1
(n=98)
—
Personal
samples
0.1
(n=ll)
0.5
(n=37)
4.2
(n=14)
—
AN in urine
(Mg/L)
3.9
(n=35)
19.7
(n=51)
359.6
(n=22)
0
(n=22)
Thiocyanate
in urine
(Mg/L)
4.50
(n=19)
5.78
(n=58)
11.41
(n=14)
4.00
(n=52)
Prevalance of adverse effects in exposed
vs control workers (%)
Reddening of
pharynx or
conjunctiva
19.4 (n= 31)
vs.
18.2 (n= 22)
11.3 (n= 53)
vs.
10.0 (n= 30)
50.0 (n= 18)
vs.
30.0 (n= 10)
-
Palpable
liver
16.1 (n= 31)
vs.
9.1(n=22)
15.1 (n= 53)
vs.
10.0 (n= 30)
38.9 (n= 18)
vs.
30.0 (n= 10)
-
Rashes or
pigmentation
of skin
9.7(n=31)
vs.
9.1(n=22)
3.8(n=53)
vs.
0 (n = 30)
11.0(n=18)
vs.
0 (n = 10)
-
Source: Sakuraietal. (1978).
Although there were some differences in mean age between the groups (38.1 years for
group C, 33.9 years for group B, and 30.5 years for group A), the age distributions of exposed
and control subjects were similar within each group (Sakurai et al., 1978). No meaningful
differences in mean clinical chemistry parameters were found between exposed workers and
controls. Medical histories of exposed workers and controls suggested a transient AN-related
increase in such symptoms as irritation of the conjunctiva and upper respiratory tract, runny
nose, and skin irritation (for example, at the scrotum). Physical examination of subjects
suggested a slight increase in the incidence of palpable liver, reddening of the conjunctiva and
pharynx, and occurrence of skin rashes (see Table 4-20). However, the difference between
groups was not statistically significant. There were no AN-related changes in blood pressure or
neurological findings.
EPA determined 4.2 ppm (average 8-hour TWA from the personal samples of the high
exposure group) as an equivocal lowest-observed-adverse-effect level (LOAEL) for physical
signs of eye or throat irritation, liver enlargement, or skin irritation in male workers with average
durations of 10-12 years of occupational exposure to airborne AN. Limitations to this LOAEL
are that workplace air concentrations across the 10-12 years of exposure were not available,
workers who underwent medical examinations were not necessarily the same as those whose air
and urine were sampled, and self-reported symptoms were not evaluated. Confidence in the
LOAEL is strengthened by the correspondence between the urinary concentrations of biomarkers
of exposure and the AN concentrations measured in air (see Table 4-20).
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Muto et al. (1992) performed another cross-sectional health examination of male workers
in seven Japanese acrylic fiber manufacturing plants in 1988. The seven factories included the
six factories studied in the 1976 cross-sectional health examination of Japanese acrylic fiber
workers (Sakurai et al., 1978). Exposed workers selected for the study were 157 male shift
workers with at least 5 years of experience on production lines. The mean years of exposure for
these workers was 17 ± 6.6 years. A nonexposed control group consisted of 537 male shift
workers in polyester fiber plants, power supply plants, or finishing branches in the acrylic fiber
plants. Controls were similar to exposed workers in average age (42.2 years, controls;
41.9 years, exposed), percentage who drank alcohol (76.5%, controls; 77.1%, exposed), and
percentage who smoked (58.1%, controls; 58.6%, exposed). All subjects underwent a medical
examination that documented past illnesses, work history, exposure to AN, smoking and
drinking habits, subjective symptoms, physical condition, urinalysis, hematology, liver function
blood variables (total bilirubin, AST, ALT, and y-GTP), and chest X-rays. Time-weighted-
average exposure levels were calculated for each worker using area sampling data and time
studies. Personal air samples were collected over a 2-day period for 142 of the 157 exposed
workers. Overall, the TWA AN concentrations for the exposed workers were 0.53 ± 0.52 ppm
(range, 0.01-2.80 ppm) and personal air concentrations were 0.62 ± 0.90 ppm (range, 0.01-5.70
ppm). The mean AN exposure measurements were lower than the exposure measurements made
in 1976 (see Table 4-19), suggesting that workplace air concentrations in the plants had declined
over the 12-year period between the studies.
More exposed workers reported a "heaviness in stomach" (p < 0.01) and decreased libido
(p < 0.05) compared to the referent group (Muto et al., 1992). Reported prevalence for heaviness
of stomach was 29.3% in the exposed group vs. 18.7% in the control group. No significant
differences were found in prevalence for other GI effects such as anorexia, nausea, vomiting,
heartburn, or stomachache.
When workers were grouped into those from factories with mean TWA AN
concentrations below 0.3 ppm (four factories; mean TWA and personal air concentrations of
0.27 and 0.19 ppm, respectively; Group A) or above 0.3 ppm (three factories; mean TWA and
personal air concentrations of 0.84 and 1.13 ppm, respectively; Group B), statistically significant
increased prevalence for the following subjective symptoms (compared with controls) were
found for the workers in the factories with higher exposure levels (Group B): decreased libido
(54.9 vs. 40.2% for controls), poor memory (76.1 vs. 64.0% for controls), irritability (35.2 vs.
24.1% for controls), reddening of conjunctiva (21.1 vs. 11.7% for controls), and eye pain or
lacrimation (32.4 vs. 19.2% for controls). There were no statistically significant differences
between the exposed and control groups in the prevalence of clinically observed physical signs
(skin rashes or reddening of conjunctiva) or abnormal findings in urinalytic, hematological, liver
function, or blood pressure variables. Prevalence of chest X-ray abnormalities were likewise not
different between exposed and control groups.
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The mean personal air concentration of workers in the Group B factories, 1.13 ppm, was
considered by EPA to be a LOAEL for small but statistically significant increased prevalence of
several subjectively reported symptoms (e.g., poor memory and irritability) in the absence of
statistically significant increases in the prevalences of physical signs or abnormal values for a
number of urinalytic, hematological, liver function, or blood pressure variables (Muto et al.,
1992). The average personal air concentration in the Group A factories, 0.19 ppm, was judged to
be a no-observed-adverse-effect level (NOAEL). Like the LOAEL identified in the earlier cross-
sectional health examination of Japanese acrylic fiber workers (Sakurai et al., 1978), a limitation
to the Muto et al. (1992) LOAEL is that historical measurements of air concentrations were not
available.
Kaneko and Omae (1992) performed a cross-sectional health questionnaire study of
neurological and subjective symptoms among exposed and nonexposed workers from seven
acrylic fiber manufacturing plants in Japan. The questionnaire for surveying subjective
symptoms was administered to 1,220 exposed male workers and 757 nonexposed male workers
who were either from the same factory or a close-by factory of the same company. The selected
study population included 504 exposed individuals and 249 unexposed controls. Subjects who
were excluded from the study were workers with administrative or nonshift jobs, a history of
exposure to other chemicals, ages not able to be matched, or incomplete information on the
questionnaire. Workplace air concentrations of AN were measured on 2 consecutive days in
each factory by using a portable gas chromatograph. Factories were grouped into three exposure
groups with the following mean workplace air concentrations: Group L = 1.8 ppm, Group M =
7.4 ppm, and Group H = 14.1 ppm. Further information on the exposure measurements was not
reported (e.g., SD of means, ranges of values, or whether or not the reported concentrations
represented 8-hour TWA concentrations). Mean durations of exposure in the three groups were
5.6, 7.0, and 8.6 years for the L, M, and H groups, respectively.
Neurological status, was assessed among all the workers using the Japanese version of
the Cornell Medical Index with additional questions. Prevalence of neurosis (defined using
Fukamachi's criteria) was slightly higher among the AN-exposed workers compared to control
workers, although the differences were not statistically significant. Subjective symptoms with
significantly higher prevalence in the AN-exposed groups, compared with the referent groups,
included headaches, tongue trouble, choking lump in the throat, fatigue, general malaise, heavy
arms, and heavy sweating. The numbers of subjective symptoms that were significantly more
prevalent in exposed workers were as follows: 8 in group L, 19 in group M, and 14 in group H.
Only the prevalence of one subjective symptom, "often feel a choking lump in the throat," had a
tendency to increase with increasing length of exposure to AN in all factories and in group L.
EPA identified 1.8 ppm as an equivocal LOAEL (Group L mean air concentration) for
statistically significantly increased prevalences of subjective symptoms.
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In a translated study from China, Chen et al. (2000) examined the health effects of
occupational exposure to AN in 224 workers at an acrylic fiber plant. The exposed group
consisted of 180 males and 44 females, with an average age of 38.6 years (range 19-57 years)
and an average of 13 years of service. The average AN concentration in the work areas at the
plant was reported to be 1.04 mg/m3 (0.48 ppm). All subjects were given a physical
examination. The results of these investigations and of subjects' hematological and clinical
chemistry parameters were compared with those of a referent group of 224 workers from a
different, unidentified plant. The authors stated that this group was of similar age, duration of
employment, and smoking habit. Reported symptoms that had significantly higher prevalence in
exposed subjects than the referent group included headache and dizziness (41 vs. 21%), poor
memory (30 vs. 13%), feelings of choking in the chest (13 vs. 8%), and loss of appetite (13 vs.
8%). However, of the other parameters evaluated in this study, all hematological data and all but
one of the clinical chemistry parameters gave similar values to those of controls. The serum
activity of y-GTP was significantly higher (p < 0.05) in exposed subjects than in controls (44.32
± 32.21 vs. 40.22 ±31.06 IU/L). Urine SCN, a marker of AN exposure, also was higher in the
exposed group (47.18 ± 20.66 vs. 43.38 ± 11.88 mmol/L). Finally, the micronucleus rate in
peripheral lymphocytes was higher in the AN exposed workers (2.6% versus 0.62%, p < 0.05).
The study authors identified 0.48 ppm, the average AN air concentration, as a LOAEL for
statistically significantly increased prevalences of subjective symptoms (including headache,
dizziness, poor memory, and loss of appetite) in workers employed in an acrylic fiber
manufacturing plant for an average of 13 years, without significant changes in most
hematological and clinical chemistry variables except for y-GTP activity.
The toxicity of AN in an occupational setting has been the subject of a number of other
reports from China (Dong et al., 2000a; Xiao, 2000a, 2000b), which were collectively submitted
to the U.S. EPA as a Toxic Substances Control Act Test Submissions report (Acrylonitrile
Group, 2000). While these studies are discussed below for hazard identification purposes, some
reports lack sufficient detail to support reported findings of AN-related toxicity. One of the
reports described a statistically significant increased prevalence of symptoms similar to those
found in other studies (e.g., headache, dizziness, sleeping disorders, and choking in the chest)
among 93 workers at a plant with AN concentrations between 2 and 22.79 mg/m3 (Dong et al.,
2000a).
Xiao (2000a) reported the results of an unpublished occupational survey that measured
fasting serum activities of serum glutamate pyruvate transaminase (SGPT, also known as ALT)
in 372 workers exposed to AN for 1-31 years in a chemical factory in China and compared the
levels with those in 186 unexposed administrators and researchers from a research institute. A
significantly higher percentage of individuals with SGPT level >19 umol/L-minute were found
in the exposed group compared with controls (41.13 vs. 4.8%), and exposed males were more
severely affected than exposed females (50.23 vs. 27.82%). Prevalence was higher among males
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with more than five years employment compared to males employed 1-5 years. No data were
provided on the level of AN exposure.
Xiao (2000b) also reported on whole blood cholinesterase activity in 237 workers
exposed to AN in a chemical factory (average age, 37 years) in comparison with those in 184
unexposed workers from a local research institute (average age, 39 years). AN measurements
were provided for three separate workshops, although it is not evident from the report whether
the measurements were taken as spot samples or from personal samplers. The average AN
concentrations in air for the three workshops were 7, 3.3, and 3 ppm, respectively. A
colorimetric method was used to measure cholinesterase activity. The authors reported that
whole blood cholinesterase activity in AN-exposed workers from the three workshops was 47-
65% that of controls (p < 0.05). Health examination results showed there was also an apparent
increase in the incidence of symptoms related to lowered cholinesterase activity in exposed
subjects compared with controls. These symptoms included neurological disorder, excessive
sweating, trembling, and discomfort in the chest.
Lu et al. (2005a) employed the World Health Organization (WHO)-recommended
Neurobehavioral Core Test Battery (NCTB), which includes seven components, to evaluate
neurobehavioral effects of workers exposed to AN in a Chinese plant. The subjects included
81 workers (68 males and 13 females) in the AN-monomer department, 94 workers (67 males
and 27 females) in the acrylic fibers department, and 174 workers (130 males and 44 females) in
the administrative or embroidery departments with no AN exposure. The monomer and fiber
workers represented 96% of the eligible exposed workers in the two departments. Periodic
short-term area sampling between 1997 and 1999 indicated that the geometric means of AN
exposure were 0.11 ppm (range 0.00-1.70 ppm for 390 samples) in the monomer department and
0.91 ppm (range 0.00-8.34 ppm for 570 samples) in the fiber department; no personal sampling
data were collected. As categorized by duration of employment, 23% of monomer workers were
exposed for 1-10 years, 42% for 11-20 years, and 35% for more than 20 years; 47% of fiber
workers were exposed for 1-10 years, 23% for 11-20 years, and 30% for more than 20 years.
Mean durations of employment were not reported for the exposed groups. Monomer workers
were also potentially exposed to cyanide and fiber workers to methyl methacrylate and heat, but
levels of exposure to these possible confounders were not monitored. The exposed workers were
frequency matched on age (within 5 years) and years of education (within 1 year) with the
unexposed workers. Exposed workers (mean age 40.8 years; range 25-53) were slightly older
than unexposed workers (mean age 36.4 years; range 21-53); the percentage of females and
years of education were similar across groups. All subjects were interviewed for demographic
data, general health status, and lifestyle. All tests were conducted by three specially trained
physicians using a Chinese operational guide of the NCTB.
Results of the analysis revealed that exposure to AN had adverse effects for some
components of the NCTB, indicating neuropsychological impairment; scores from the following
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tests were statistically significantly different (p < 0.05) from controls in analyses of covariance
that took into account age, sex, and education level. In the Profile of Mood States test, all scores
for negative moods (anger, confusion, depression, fatigue, and tension) were significantly higher
in the exposed groups than in the unexposed group and higher for monomer workers (41-68%
higher than controls) than for fiber workers (20-44% higher than controls). Simple reaction
time, a test of attention and visual response speed, was longer in the two exposed groups than in
the unexposed group: 16% longer for monomer workers and 10% longer for fiber workers.
Exposed workers performed more poorly (by 21% for monomer workers and 24% for fiber
workers) in the backward sequence of the digit span test, a measure of auditory memory, but
fiber workers had better performance in the forward sequence than monomer and unexposed
workers. Both groups of exposed workers also had a 4% poorer performance in the Benton
Visual Retention test, a measure of visual perception and memory; scores in the Pursuit Aiming
II test, which assesses fine motor skills and perceptual speed, were 14% lower for monomer
workers and 10% lower for fiber workers compared with controls. Exposure to AN had no
significant effect on scores for manual dexterity in the Santa Ana test or for perceptual speed in
the digital symbol test. In examining effects by duration of AN exposure, there was no statistical
relationship for mood scores and duration of exposure. However, there was an insignificant
decrease in the simple reaction test with duration of exposure for both monomer and fiber
workers. Inverse relationships were found between performance and duration in both the digital
symbol test and total scores of the digit span test in the two exposed groups. Decreased
performance with duration of exposure was also found in the Pursuit Aiming II test for exposed
monomer workers.
Several limitations of the study were noted by Lu et al. (2005a) or by EPA. The primary
limitation of the study was the extent of exposure data, with exposure measures based on area
sampling during 1997 to 1999; no contemporaneous personal monitoring data were available.
Although the monomer department exposure levels were somewhat lower (mean 0.11 ppm)
compared with the fiber department (mean 0.91 ppm), it is unclear if these differences were large
enough (or estimated with enough precision) to allow for valid estimation of differing levels of
effects between these groups. Coexposure to cyanide and methyl maethacrylate occurred among
different sets of the exposed workers, but not among the controls; however, information provided
9
by Dr. Lu to EPA indicates these compounds were present in only trace amounts. The NCTB
was developed for populations in Europe and North America, and it is not known to what extent
cultural differences may have affected results of the Profile of Mood States test, shown to be
sensitive to cultural differences. This limitation may affect the sensitivity of the scale in
assessing effects, but would not be expected to produce spurious associations since the controls
were selected from a similar cultural group. EPA determined the results from this study were
2 Email from Dr. Rongzhu Lu, Department of Preventive Medicine, College of Medicine, Jiangsu University, China,
to Dr. Diana Wong, U.S. EPA, dated 5/15/2008.
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consistent with the designation of the average exposure levels for the monomer workers, 0.11
ppm, and the fiber workers, 0.91 ppm, as LOAELs for small deficits in neurobehavioral tests of
mood, attention and response speed, auditory memory, and motor steadiness, but not in tests of
manual dexterity or perceptual motor speed.
In an article published in a Chinese journal (abstract in English), Ding et al. (2003)
evaluated mitochondrial DNA damage in a group of 47 Chinese workers randomly selected from
1,020 active workers in the chemistry department of a petrochemical company (aged 40.0 ±7.1
years). These workers were exposed to AN at a geometric mean concentration of 0.25 mg/m3
(0.11 ppm) (median 0.36 mg/m3 [0.17 ppm], range 0-3.70 mg/m3) for an average of 17.3 ± 3.8
years. An unexposed control group of 47 persons was selected from the teachers and staff of a
college (aged 40.4 ±8.1 years), with an average length of employment of 18.7 ±4.1 years. DNA
was extracted from peripheral blood samples from each subject and evaluated using the
polymerase chain reaction (PCR) with specific primer pairs to detect deletions in mitochondrial
DNA. Deletions in mitochondrial DNA were detected in 8/47 exposed workers compared with
0/47 nonexposed workers. A deletion rate of 0.00225 ± 0.00171 was calculated based on optical
densities from gel scans of the deletion fragment compared with a fragment synthesized by using
primer pairs to a conserved region of mitochondrial DNA; this deletion rate was statistically
significantly different (p < 0.05) from the rate of 0 for the controls. For studying the effects of
aging, a group of 12 healthy nonexposed retirees from governmental organizations (average age
79.15 ± 3.80 years) and a group of 12 healthy nonexposed high school students (average age
14.23 ± 1.52 years) were also examined. Deletion fragments were detected in 3/12 elderly
subjects and 0/12 young subjects. The deletion rate for the elderly was calculated as 0.00193 ±
0.00086, which was not significantly different from that calculated for AN-exposed workers.
The study authors suggested that exposure to AN might have an effect on the molecular
process(es) of aging. The abstract did not report the protocol for selection of controls and the
gender composition in each group. The results identified the mean workplace AN air
concentration, 0.11 ppm, as a LOAEL for increased prevalence of workers with deletions in
peripheral blood mitochondrial DNA compared with controls.
Borba et al. (1996) measured cyanoethylvaline-hemoglobin (CEVal-Hb) adducts as a
marker of AN exposure in three groups of occupationally exposed workers in an acrylic fiber
factory in Portugal. In addition, the induction of CYP450 species was determined by the
excretion of D-glucaric acid (an end product of the glucuronic pathway) in the urine, and
formation of malondialdehyde (MDA) (a final product of lipid peroxidation) in RBCs was
determined as a surrogate measure of oxidative stress. The groups comprised 20 administrative
workers who were not exposed to AN in the same plant, 14 individuals employed in the
continuous polymerization department, and 10 equipment maintenance workers. Considered a
measure of the biologically effective dose, CEVal values in non-smokers were 8.5-70.5 pmol/g
Hb in controls, 635.2-4,603.5 pmol/g Hb for continuous polymerization workers, and 93.9-
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4,746 pmol/g Hb for maintenance workers. The study authors found no indication of AN-
induced CYP450 induction, but a significant increase was seen for the oxidative stress marker in
the group of maintenance workers. Smoking had no influence on these metrics. Borba et al.
(1996) also investigated markers for genotoxicity (viz., gene reversion activity of urine extracts,
chromosomal aberrations (CAs), and sister chromatid exchanges [SCEs]). These are discussed
in the genotoxicity section (Section 4.5.2).
4.1.3. Dermal Exposure
4.1.3.1. Acute Exposure
A case report by Vogel and Kirkendall (1984) described a 24-year-old ship's officer who
was accidentally sprayed with AN when a valve burst while he was unloading the chemical.
Because the man's face, eyes, and body were covered with AN, it is likely that he was exposed
via the oral and inhalation routes as well as to the skin and eyes. Immediate responses to
exposure included dizziness, flushing, and nausea with vomiting. During hospitalization, acute
toxicological impacts included a rapid pulse rate (100 beats/minute) and a respiratory rate of
16/minute. The subject displayed erythema and mild conjunctivitis, tachycardia, and striking
hematological changes (WBC count of 26,400 cells/cm3, of which 76% was polymorphonuclear
leukocytes, 10% lymphocytes, and 7% each basophiles and monocytes). Methemoglobin
(MetHb) concentration was 10.3% on admission. The patient received nitrite/thiosulfate
treatment, underwent dialysis, and, overall, showed steady recovery over his 5-day
hospitalization.
There are a number of additional case studies of the toxic effects of AN resulting from
acute exposure after accidental spillage in the workplace. These support the designation of AN
as a skin irritant. In several cases, erythemas were shown to result from direct dermal contact
with solutions of AN (Davis et al., 1973 [as cited in IPCS, 1983]; Zeller et al., 1969; Wilson et
al., 1948; Dudley and Neal, 1942). The lesions were followed by delayed blistering and burns,
typically 1 or 2 days following exposure (Davis et al., 1973 [as cited in IPCS, 1983]; Zeller et al.,
1969; Babanov et al., 1959; Dudley and Neal, 1942). In some cases, clinically diagnosed
dermatitis was associated with irritation (Bakker et al., 1991; Davis et al., 1973 [as cited in IPCS,
1983]). For example, Davis et al. (1973) (as cited in IPCS, 1983) reported a wide range of
dermal effects from AN contact, including skin dermatitis, local irritation, erythema, swelling,
blistering, and burns. However, dermatitis has not always resulted from AN-induced skin
irritation (Zeller et al., 1969; Babanov et al., 1959; Wilson et al., 1948; Dudley and Neal, 1942).
When Dudley and Neal (1942) investigated the effects of AN exposures to laboratory
animals, an accident in their laboratory resulted in a case of occupational exposure. Symptoms
similar to those described later by Wilson et al. (1948) were reported for a male laboratory
worker who spilled small quantities of liquid AN on his hands. Diffuse erythema on hands and
wrists was evident after 24 hours, with subsequent blistering on the fingertips on day 3. Both
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hands became slightly swollen, erythematous, itching, and painful. By day 10 after exposure, the
skin of the fingers had cracked and peeled and the skin was dry and scaly with large areas of
tender new skin.
Zeller et al. (1969) identified 137 cases of accidental exposure to nitriles when reviewing
industrial hygiene data from Germany on accidental exposures to chemical agents in an
occupational setting over a 15-year period. Of the 137 cases, 66 cases related to AN exposure, of
which 50 cases resulted from direct skin contact and the remaining 16 to inhalation or exposure
to AN vapors. All 66 cases were judged to be minor and did not require hospitalization.
However, up to 3 weeks of recuperation was required as a result of direct skin contact with the
compound. Typically, the first symptoms appeared between 5 minutes to 24 hours after initial
dermal contact with AN. For the most part, the workers complained of burning sensations of the
skin, followed by reddening of the exposed area and the formation of blisters at any point during
the first 24 hours. In one case, AN was thought to have diffused through the leather of a shoe on
which it was spilled, resulting in a delay before blisters appeared and delayed healing. There
was no indication of resorptive damages from the dermal exposure in any of the 50 cases.
In another study (Babanov et al., 1959), blistering was also observed on workers' legs
within 6-8 hours of contact with spilled AN, while a diluted (5%), heated (50°C) solution of AN
caused serious skin burns.
A fatal case of dermal contact exposure with AN was described by Lorz (1950), when
application of a delousing agent containing AN resulted in the death of a 10-year-old girl.
Following dermal application of the delousing agent to the scalp, the girl's head was wrapped in
a cloth and she went to bed. Symptoms of nausea, headaches, and dizziness were followed by
repeated vomiting and coma. Cramps and increasing cyanosis were followed by death 4 hours
after application. A similar case of fatal poisoning was reported by Grunske (1949) in which a
3-year-old girl died after reentry into a home that had been treated with an AN-containing
fumigant. In both of these cases, exposure was likely to have occurred via inhalation as well as
the dermal route.
4.1.3.2. Chronic Exposure
In addition to the known skin irritation effects of acute exposures to high-concentration
liquids and vapors, various chronic dermal exposure effects were reported. There was limited
evidence that skin sensitization resulted in dermal allergies to AN. Therefore, an intrinsic
capacity of AN to act as a skin sensitizing agent was suggested.
The abstract of a Japanese language report by Hashimoto and Kobayasi (1961) discussed
the case of a chemical laboratory worker who developed skin lesions through contact with AN.
Although much of the detail remained uncertain, including period and duration of exposure and
the influence of other chemicals on the subject's condition, the lesions apparently spread across
the subject's body from the contact site, consistent with a direct contact allergic reaction.
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Bakker et al. (1991) reported that 10 employees at an acrylic fiber factory complained of
skin irritation. While five of the subjects developed irritant dermatitis, the other five subjects
gave positive patch tests with AN, suggesting an allergic reaction. This finding also supported
the designation of AN as a sensitizing agent.
Prolonged dermal contact for 6 weeks to methyl methacrylate, a copolymer of AN,
resulted in an allergic skin sensitization reaction in the case of a 27-year-old man who had his
finger splinted for a torn ligament (Balda, 1975). The unpolymerized AN of the Plexidur
copolymer resulted in strong skin irritation response, with erythema and formation of scaly skin
and scattered blisters. Patch testing confirmed that the allergic sensitization was to AN and not
to a copolymer or the polymerization catalyst, benzoyl peroxide. A contributing factor may have
been the man's prior hyperhydrosis episode of the hands that led to blister formation.
Very few studies hinted at a possible desensitization or adaptive effect for prolonged
chronic exposures. Based on investigations carried out over a period from 1965 to 1971, Zotova
(1975) reported complaints of poor health, which included skin irritation, in workers at an AN
manufacturing facility. Gincheva et al. (1977) (as cited in IPCS, 1983) did not find changes in
the health status of a group of 23 men exposed to 4.2-7.2 mg/m3 (2-3.3 ppm) of AN for
exposure durations of 3-5 years. Details of the study were not provided.
4.1.4. Ocular Exposure
Secondary routes of exposure to AN include ocular exposure to either AN liquid or
vapor. For example, in the Wilson et al. (1948) study, subjects exposed to AN at concentrations
varying from 16 to 100 ppm (35-217 mg/m3) for 20-45 minutes demonstrated irritation of all
mucous membranes, including the eyes, nose, and throat. In other studies, blepharoconjunctivitis
was reported in workers exposed to the compound (Delivanova et al., 1978). Of 302 workers
examined over 2 years (138 in 1976 and 164 in 1978), 42 had severe cases of
blepharoconjunctivitis related to AN exposure.
In the case report by Vogel and Kirkendall (1984), the mucous membranes of the eyes of
the 24-year-old man had been sprayed with AN. Among other symptoms, mild conjunctivitis but
no apparent corneal clouding was observed. Eye irritations and nasal discharge also were
reported in workers exposed to relatively high levels of AN at an acrylic fiber plant (Sakurai,
2000; Sakurai et al., 1978).
4.2. SUBCHRONIC AND CHRONIC STUDIES AND CANCER BIOASSAYS IN
ANIMALS—ORAL AND INHALATION
4.2.1. Oral Exposure
4.2.1.1. Subchronic Studies
Single studies in dogs and mice are the only well-documented standard bioassays for the
subchronic toxicity of AN in experimental animals by the oral route (NTP, 2001; Quast et al.,
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1975). A full report of a study in rats is not available (Humiston et al., 1975 [as cited in Quast,
2002]). Additional subchronic studies have evaluated only functional effects of AN exposure in
specific organs, such as the adrenals (Szabo et al., 1984) and the brain (Gagnaire et al., 1998).
These endpoint-specific studies are discussed in Section 4.4.
Humiston et al. (1975) (as cited in Quast, 2002) exposed groups of Sprague-Dawley rats
(10/sex/group) to AN in drinking water at concentrations of 0, 35, 85, 210, or 500 ppm for
90 days. Reported intakes of AN for males/females, respectively, were 0/0, 4/5, 8/10, 17/22, or
38/42 mg/kg-day. Water consumption was decreased in a dose-related manner in both sexes,
with females more affected than males. However, the lowest dose affecting males was not
identified in the summary; in females exposed at 35 ppm, 9 of 26 measured intervals showed
statistically significant decreases compared with controls. In the 210 and 500 ppm groups, there
were significant decreases in water consumption, food consumption, and BW, but no information
was provided as to the magnitude of these changes. Exposure to AN had no effect on
hematology, clinical chemistry, or histopathologic findings; urinalysis results were also
unaffected by treatment, except for an increase in specific gravity that was correlated with
increasing dose. Apparently, a number of treatment-related effects was observed in the 85 ppm
group, but neither their identity nor the magnitude of change were specified. The summary
provided by Quast (2002) did not provide sufficient detail to accurately identify NOAEL or
LOAEL values from the study by Humiston et al. (1975) (as cited in Quast, 2002).
Szabo et al. (1984) evaluated the effect of subchronic oral exposure to AN on the
structure and function of the adrenal gland and intestinal tract in rats. Female Sprague-Dawley
rats (three to four per group) were exposed for 7, 21, or 60 days to AN in drinking water at
concentrations of 0, 1, 20, 100, and 500 ppm, representing approximate daily doses of 0, 0.2, 4,
20, and 100 mg/kg. Other groups of rats received equivalent doses of AN by gavage for the
same duration. Water and food intakes were monitored continuously and BWs were recorded
every fourth day. At the end of the studies, blood samples were taken to measure plasma
corticosterone and aldosterone, and a complete necropsy was carried out on all survivors. The
adrenals, thyroid, liver, and one kidney were weighed. Samples of liver, kidney, lung, brain, and
the entire adrenal and thyroid glands were measured for levels of nonprotein sulfhydryls in tissue
homogenates. Adrenals, "other" endocrine organs, stomach, duodenum, liver, kidney, lung, and
heart were evaluated for histopathology.
BWs were reduced by about 25% in rats exposed to 500 ppm in drinking water up to
60 days, but no reduction in BW was observed in rats receiving the equivalent dose of
100 mg/kg-day by gavage (Szabo et al., 1984). Water intake was reduced in rats exposed to
500 ppm in drinking water but increased at the equivalent gavage dose. Adrenal weights were
slightly lower than in controls in the 7- and 21-day exposure groups, but significant increases in
relative adrenal weights were observed after 60 days of exposure to 500 ppm in drinking water or
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to >0.2 mg/kg-day by gavage. The kidneys were enlarged in rats exposed to 100 or 500 ppm AN
in drinking water for 60 days or 500 ppm for 21 days.
Histopathological lesions included adrenocortical hyperplasia in all rats exposed by
gavage at >0.2 mg/kg-day for 60 days. Enlarged kidneys and hyperplasia of the gastric mucosa
at the junction of the glandular stomach and forestomach was observed in rats exposed to 100 or
500 ppm in drinking water for 21 or 60 days. No incidence data were provided for these lesions.
Plasma corticosterone levels were significantly decreased by both 500 ppm AN in drinking water
or by gavage in the 21-day study. Even the lowest concentration of 1 ppm AN by gavage
resulted in a 50% reduction in corticosterone. Plasma aldosterone concentration was reduced,
starting from 20 ppm AN. In the 60-day study, significant suppression of plasma corticosterone
was observed after exposure to 100 or 500 ppm AN in drinking water and in all exposure levels
given by gavage.
Levels of nonprotein sulfhydryls (glutathione) were increased by 20-50% in the 7- and
21-day studies but were significantly reduced by 20-30% in the adrenals after 60 days of gavage
exposure to 4-60 mg/kg-day. Increases in nonprotein sulfhydryl concentrations in duodenal
mucosa were observed in the 100 or 500 ppm AN exposure groups in the 7- and 21-day studies
and in all doses after 60 days, but not in the control group of rats that received water by gavage.
For the drinking water study, a NOAEL of 20 ppm (4 mg/kg-day) and a LOAEL of 100 ppm
(20 mg/kg-day) were identified for enlarged kidneys and increases in regional hyperplasia of the
gastric mucosa in female rats exposed for 60 days. For the gavage study, no NOAEL was
identified, but a LOAEL of 0.2 mg/kg-day was identified for adrenocortical hyperplasia in rats
exposed for 60 days.
In a separate experiment, Szabo et al. (1984) evaluated age-dependency in the sensitivity
of Sprague-Dawley rats to the action of AN on the adrenals. Groups of weanling rats (40 g) and
adult rats (190 g) were treated with 0, 0.002% (20 ppm), 0.01% (100 ppm), or 0.05% (500 ppm)
of AN in drinking water for 21 days, or were given by the corresponding amount of AN by daily
gavage for 21 days. The animals were sacrificed, and plasma and adrenals were collected as in
the previous experiments. Young, immature rats treated with AN were found to have lower (up
to 66%) levels of plasma corticosterone and aldosterone than adult rats. These differences were
statistically significant with 0.01% AN given by gavage or 0.05% AN in drinking water.
Quast et al. (1975) exposed beagles (four/sex/group) to 0, 100, 200, and 300 ppm AN
(purity >99%) in drinking water for 6 months; the reported calculated doses were approximately
0, 10, 16, and 17 mg/kg-day in males, respectively, and 0, 8, 17, and 18 mg/kg-day in females,
respectively. Dogs were evaluated daily for clinical signs and weighed weekly; food and water
consumption were calculated from the consumption by groups of dogs penned together. Routine
hematology examinations were conducted on all samples taken from all dogs 20 days before
exposure and on days 83, 130, and 179. Blood and urine samples for standard biochemical
analyses were collected from all dogs 8 days before exposure, on days 84, 135, and 176, and at
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termination (day 182 for males and day 183 for females). Serum samples were collected from all
dogs on day 155 to determine total protein concentrations and evaluate the percentages of
specific Igs by electrophoresis. Ophthalmic examinations were conducted pretest and at 85, 112,
and 175 days. All dogs were subjected to gross necropsy at which time organ weights were
recorded for brain, heart, liver, kidneys, and testes; a full set of tissues from all dogs was
examined for histopathology. Samples of liver and kidney were analyzed for nonprotein free
sulfhydryl content.
The following treatment-related effects were observed in dogs exposed to AN in drinking
water for 6 months (Quast et al., 1975). No mortality was observed in the control or 100 ppm
groups, but dogs exposed at 200 ppm (2/4 males and 3/4 females) and 300 ppm (3/4 males and
2/4 females) either died prematurely or were euthanized in a moribund condition. Signs of
toxicity manifested in these dogs included reduced consumption of food and water, decreased
BW, roughened hair coat, and a nonproductive cough. These dogs subsequently exhibited
lethargy, weakness, emaciation, respiratory distress, and terminal depression. Decreased water
consumption was observed throughout the study in males at 300 ppm and sporadically in females
at >200 ppm. Food consumption was reduced in males at 300 ppm and females at >100 ppm. A
supplemental study was conducted on eight female dogs treated with 100 ppm AN for 5 weeks,
and no reduction in food or water consumption was observed.
Substantial reductions in BW were observed among dogs that died (200 and 300 ppm);
group mean weights among surviving male and female dogs were not statistically significantly
different from controls. RBC and Hb counts were reduced by 19-21% in males exposed to
200 ppm for 83 days but not at later time points; females exposed at 300 ppm exhibited 22-26%
decreases in RBC counts after 83 and 130 days of exposure. No treatment-related changes in
hematology parameters were evident after 179 days of exposure. Exposure to AN had no effect
on urinalysis, clinical chemistry, ophthalmoscopic examinations, nonprotein free sulfhydryl
content in liver or kidney, or the electrophoretic behavior of serum proteins. Statistically
significant alterations in organ weight, such as increases in relative kidney weights in males
treated at 100 and 200 ppm, were not biologically significant (less than 10% change).
Histopathologic changes were observed in the esophagus (focal erosions and/or ulcerations in the
middle one-third, dilations, and thinning of the walls) and tongue (increased thickness of
epithelium lining of the dorsal surface) of male and female dogs treated at 200 and 300 ppm.
Lesions of the lung were attributed to parasitic nematode infection that was present in all dogs.
In this study, a NOAEL of 100 ppm (8-10 mg/kg-day) and a LOAEL of 200 ppm (16-17 mg/kg-
day) were identified for early mortality and histopathological lesions of the esophagus and
tongue in male and female dogs.
In a comparative study of the neurotoxicity of nitriles, Gagnaire et al. (1998)
administered AN (>99% purity) at doses of 0, 12.5, 25, or 50 mg/kg-day to male Sprague-
Dawley rats (12/group plus 10 controls) by gavage in olive oil, 5 days/week for 12 weeks. BWs
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were measured weekly. After 3, 6, 9, and 12 weeks of treatment and after an 8-week recovery
period (week 20), rats were evaluated for electrophysiological parameters, with testing occurring
16 hours after dosing (or 48 hours for testing after weekends). Electrical stimulation of the tail
nerve was used to assess four neurological properties, the motor conduction velocity (MCV),
sensory conduction velocity (SCV), and amplitudes of the sensory and motor action potentials
(ASAPs and AMAPs).
One high-dose rat died during the first week of treatment, but no other mortality was
observed. High-dose rats showed reduced BW gain that became significant after the fourth week
of treatment, leading to a terminal BW 17% lower than in controls. BWs of mid-dose rats
became significantly lower than controls after the fifth week, resulting in a terminal weight
approximately 7% lower than controls (as estimated by visual inspection of the data graph); this
weight change was not considered biologically significant. Five of 11 high-dose rats exhibited
significant weakness of the hind limbs that somewhat improved during the recovery period. Rats
exposed to AN showed acute signs of behavioral abnormalities, including salivation, locomotor
hyperactivity, and fur wetting within 1 hour of dosing. Gagnaire et al. (1998) attributed these
findings to cholinomimetic effects possibly caused by AN-induced changes in muscarinic
acetylcholine receptors or by alterations in hepatic metabolism. Exposure to AN had no effect
on MCV or AMAP except for a 40% increase in AMAP observed in high-dose rats at the 9-week
time point. The most consistent effect of AN was a significant reduction in SCV compared with
that in controls (Table 4-21), ranging from 7.5 to 15% between weeks 6 and 12 (p < 0.05 to
p < 0.001); this effect abated following the cessation of exposure, but a 10.6% reduction
compared with controls (p < 0.001) was observed after 8 weeks of recovery (week 20). A 25%
reduction in ASAP was also observed in high-dose rats at week 20, but no effect on this
parameter was observed during the treatment period. A NOAEL of 25 mg/kg-day and a LOAEL
of 50 mg/kg-day were identified in the study by Gagnaire et al. (1998), based on reduced BW
and neurotoxic effects (weakness of the hind limbs and reduced SCV) in male rats exposed to
AN by gavage.
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Table 4-21. Effect on SCV in male Sprague-Dawley rats exposed to AN via
gavage for 12 weeks
Group
(mg/kg)
0
12.5
25
50
SCV (m/s)a
Exposure (weeks)
0
34.2 ±0.7
34.9 ±0.5
34.6 ±0.8
35.8±1.1
3
43.2 ±0.6
42.6 ±0.9
40.0 ±1.0
42.2 ±1.1
6
45.2 ± 1.3
45.7 ±0.7
45.5 ±0.6
41.8 ± 1.3b
9
48.4 ±1.1
47.8 ±0.7
46.0 ±0.9
44.0±1.2C
12
46.6 ±0.8
46.7 ±0.9
44.4 ±0.9
39.8±0.6C
Recovery
20
53.8 ±1.5
51.8 ±0.9
51.3 ±0.7
48.1±0.7C
aValues are means ± SDs, n = 12 for treated rats and n = 10 for controls.
bStatistically significant compared with controls (p < 0.05), as calculated by the authors.
Statistically significant compared with controls (p < 0.001), as calculated by the authors.
Source: Gagnaire et al. (1998).
In a comparative study on the effects of nine compounds related to acrylamide (Barnes,
1970), AN was administered by gavage over a period of 7 weeks to six young adult albino rats of
the Porton strain as 15 daily doses of 30 mg/kg, followed by 7 doses of 50 mg/kg, and finally
13 doses of 75 mg/kg. The time-adjusted average dose administered over the course of the study
was 36 mg/kg-day. Rats were weighed weekly, and their gait and stance when walking on a
sloping nonslippery surface including an ascent up a sloping wooden board were evaluated.
Treated rats were also held by the tail in front of a sloping bar, and tested for the ability to grasp
the bar with the front paws, and then grasp it with the hind feet, a reflex typically lost early in
rats with peripheral neuropathies. No data were provided for this experiment, but the study
author reported that there was no evidence of adverse effects.
In a recent study that explored the neurobehavioral effects of AN in rats (Rongzhu et al.,
2007), male Sprague-Dawley rats (10/group) were exposed to 0, 50, or 200 ppm AN in drinking
water. The study authors estimated AN doses to be 0, 4.03, and 13.46 mg/kg-day. Three
neurobehavioral tests, including the open field test, rotarod test, and spatial water maze, were
conducted to evaluate locomotor activities, motor coordination, and learning and memory,
respectively, prior to initiation of exposure and at 4, 8, and 12 weeks of exposure. Thiocyanate
levels in urine were measured in a minimum of five rats from each group at week 12 and were
reported to be 2.79, 6.10, and 25.03 mg/g creatinine, respectively.
Beginning from the sixth week of AN administration, three rats in the 50 ppm AN group
and five rats in the 200 ppm AN group showed behavioral changes. The coat appearance in all
treated rats was soiled, and the main changes were head twitching, trembling, circling,
backwards pedaling, and decreased home-cage activities. The two treatment groups also showed
less BW gain than the control group. In the open field test, there were no significant differences
in start-up latency among the exposed and unexposed groups. The 200 ppm group consistently
had higher locomotor activity than the control group, from pretreatment to 12 weeks of exposure.
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There were also no changes in the number of rearing and grooming episodes. Therefore, there
were no uniform changes in exploration and locomotion. In the rotarod test, the maximal and the
total and falling latency in the 50 and 200 ppm groups were significantly decreased in a dose-
and time-dependent manner. In the spatial water maze test, rats in the 200 ppm group had
significantly increased training time and training duration, compared with the control and
50 ppm groups. However, these two parameters in the exposed groups returned to close to
control level at the end of the experiment. Rongzhu et al. (2007) suggested that this reversible
phenomena may be caused by tolerance. The study authors concluded that oral exposure to AN
induced neurobehavioral alterations. The neurochemical mechanisms need to be further
investigated. The LOAEL for neurobehavioral alterations is 4 mg/kg-day.
In a subchronic study, the National Toxicology Program (NTP) (2001) treated B6C3Fi
mice (10/sex/group) with 0, 5, 10, 20, 40, or 60 mg/kg-day AN (purity >99%) by gavage in
water, 5 days/week for 14 weeks; adjusted for intermittent exposure (5 days/7 days), the intakes
were 0, 3.6, 7.1, 14.3, 28.6, and 42.9 mg/kg-day. Clinical findings were recorded on day 8 and
once weekly thereafter. BWs were recorded before treatment, weekly, and at study termination.
Necropsies were performed on all animals, at which time organ weights were recorded for heart,
right kidney, liver, lung, spleen, right testis, and thymus. Hematology analysis was conducted on
all mice surviving at the end of the study. Complete histopathologic analyses were conducted on
all control mice and those treated with 40 and 60 mg/kg-day; males receiving 20 mg/kg-day were
also examined. At the end of the study, 10 males/group in the groups receiving 0, 5, 10, and 20
mg/kg-day were selected for reproductive evaluations; the left cauda, left epididymis, and left
testis were weighed and sperm samples were evaluated for sperm counts and motility. Also, 10
females/group in the groups receiving 0, 10, 20, and 40 mg/kg-day were evaluated for vaginal
cytology (estrous cycle and stage length) in the last 12 days of the study before termination.
The following effects were observed in mice exposed by gavage to AN. Aside from one
control male at week 9, there were no deaths at exposures up to and including 20 mg/kg-day.
Mortality at the two highest doses comprised 9/10 males and 3/10 females at 40 mg/kg-day and
all mice treated at 60 mg/kg-day. All deaths occurred on the first day, except for one male in the
40 mg/kg-day group. Slight (2-8%) decreases in BW in treated mice compared with controls
were not biologically significant. Survivors (seven females and one male) receiving 40 mg/kg-
day exhibited lethargy and abnormal breathing immediately after dosing "for several days" but
then appeared to develop tolerance to AN.
Sporadic, statistically significant alterations in hematological parameters included
reductions in platelet counts by 20% in males at 20 mg/kg-day, in leukocyte and lymphocyte
counts (-30% in males at 20 mg/kg-day and -37% in females at 40 mg/kg-day), and in Hb and
RBC counts by 10% in females at 40 mg/kg-day; other statistically significant changes of
doubtful biological significance in females included reductions in RBC counts by 4% in groups
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treated with 5-20 mg/kg-day and in hematocrit by 6% in the 40 mg/kg-day group. Some of these
hematological effects may have been secondary to stomach ulceration observed in some mice.
Absolute and relative heart weights were increased by 20 and 30%, respectively, in males
at 20 mg/kg-day. The absolute weights of the left cauda epididymides were increased by 15% in
groups exposed at 10 and 20 mg/kg-day. However, no histopathological findings were reported
in these organs. The only treatment-related lesions were in the forestomach of females at
40 mg/kg-day: 4/7 with chronic active inflammation (associated with hyperplasia) and 5/7 with
focal epithelial hyperplasia. Two females in this group exhibited focal ulceration of the
forestomach associated with the hyperplasia. No other treatment-related histopathology was
observed. Exposure to AN produced no effects on sperm motility in males at <20 mg/kg-day.
There were no differences in vaginal cytology parameters in females at<40 mg/kg-day. A
NOAEL of 20 mg/kg-day and a LOAEL of 40 mg/kg-day were identified in the NTP (2001)
bioassay, based on hyperplastic lesions in the forestomach of female mice.
4.2.1.2. Chronic Studies
4.2.1.2.1. Quasi (2002) and Quasi et al (1980a). Quast (2002) and Quast et al. (1980a)
conducted a 2-year toxicity and carcinogenicity study in Sprague-Dawley rats (48/sex/group)
exposed to AN (purity >99%) in drinking water at concentrations of 35, 100, or 300 ppm; groups
of 80/sex receiving untreated drinking water served as controls. Additional interim-sacrifice
groups of 10/sex/dose were exposed for 1 year and analyzed under the same protocol as the main
study. The reported intakes of AN were 0, 3.4, 8.5, and 21.3 mg/kg-day for male rats and 0, 4.4,
10.8, and 25.0 mg/kg-day for female rats. Rats were observed daily for clinical signs of toxicity
and were weighed and examined monthly for palpable masses. Food and water consumptions
were determined for 30 rats/sex/group weekly for the first 3 months of the study and for 1 week
during each of the following months: 4, 5, 6, 7, 9, 11, 12, 15, 18, 21, and 24. Hematology
examinations and urinalysis were conducted on 10 rats/sex in the control and the highest dose
group on days 45, 87, 180, and 355. Additional hematology examinations were conducted on
10 rats/sex from all groups on days 544 (males) and 545 (females) and at study termination on
day 724 (males and females); an additional urinalysis was conducted on 10 rats/sex from all
groups on day 181 to evaluate a dose response for increased urine specific gravity observed
previously at the highest dose group. Clinical chemistry examinations were conducted on
10 rats/sex from the control and highest dose groups on days 46 and 356, on 10 rats/sex from all
groups on days 88, 180, and 550, and on all survivors on day 746. All rats, whether dying
prematurely, sacrificed in a moribund condition, or sacrificed on schedule, were subjected to a
gross necropsy, which included an ophthalmologic examination. Necropsies of rats on scheduled
sacrifice included organ weight determinations for brain, heart, liver, kidneys, and testes.
Complete histopathologic examinations were conducted for rats in the control and 300 ppm
groups in the 1-year interim and 2-year studies, and, based on those results, a set of 22 tissues
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was examined microscopically in nearly all rats exposed at 35 and 100 ppm; tumors and other
lesions identified at gross necropsy were also examined microscopically.
Noncancer results
Exposure to AN in drinking water reduced survival, BW, and consumption of food and
water by Sprague-Dawley rats in a dose-related manner (Quast, 2002; Quast et al., 1980a). In
male rats exposed at 300 ppm, survival was significantly reduced compared with controls
(p < 0.05), beginning at 16 months, and none survived after 22 months; survival in other treated
male groups was not significantly different from controls. In female rats, exposure at 300, 100,
and 35 ppm resulted in significantly reduced survival beginning at 10, 12, and 18 months,
respectively, with no high-dose females surviving past month 22; at 24 months, survival was 25,
8.3, 2.1, and 0% for the control and low- to high-dose groups, respectively. The magnitude, time
of onset, and duration of significantly lower BWs in exposed rats compared with controls were
dose related. Only the 17-22 and 17-18% decreases observed in males and females,
respectively, at 300 ppm were biologically significant. Reductions only reached -8-9% in the
100 ppm groups and -6-9% in the 35 ppm groups.
The study authors mentioned that BW comparisons near the end of the study tended to be
confounded by geriatric changes, few surviving rats, and excessive tumor growth in exposed rats.
Food intake (g/rat/day) was significantly lower in 100 and 300 ppm groups compared with
controls beginning during the first week of the study and only in females at 35 ppm.
Significantly lower feed intake values for males and females, respectively, compared with
controls were measured on 12/24 and 16/24 occasions at 300 ppm, 8/24 and 11/24 times at
100 ppm, and 9/24 times in females only at 35 ppm. The largest difference in high-dose rats was
a 25% reduction of feed intake in males after 6 months and a 27% reduction in females after
9 months. Feed intakes during the last weeks of the study were not significantly different among
the different groups.
Water intake was reduced during the first 10 days by 36% in males and 38% in females
exposed to 300 ppm AN and remained significantly lower for all 26 measurements throughout
the study. Significantly lower water consumption was also measured 24/26 times for males and
23/26 times for females at 100 ppm and 13/26 times for males and 21/26 times for females at
35 ppm. Male and female rats in the 100 and 300 ppm dose groups showed a lack of normal
grooming later in the study. In addition, signs of nervous system dysfunction (erratic
movements, trembling, circling, and limb weakness) were observed in the higher dose groups
rats in the absence of end-stage kidney disease or pituitary tumors. These signs of nervous
system dysfunction when they occurred in controls were correlated with end-stage kidney
disease or pituitary tumors and occurred more frequently in males than in females. The study
authors were unable to detect microscopic tumors or lesions that would correlate with the
observed clinical signs.
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AN had no primary effect on hematological parameters, except for lowered RBC counts
associated with blood loss from ulcerated tumors or nutritional anemia during the later stages of
the study. The only effect of AN on urinalysis was a slight, but statistically significant increase
(1.3-3.0%) in urine specific gravity noted for 300 ppm group males at five time points (45-
355 days) and females at all six time points (45-545 days). Significant increase in urine specific
gravity was also observed in 100 ppm male and female rats on day 181 and day 544 (females
only). The lack of significant effect in treated males on days 544 and 724 was attributed by
Quast (2002) to higher incidence of advanced chronic renal disease in controls that resulted in an
inability to concentrate urine normally. All 300 ppm male and female rats were dead by
day 724. No toxicological significant alterations were observed in clinical chemistry parameters
or in absolute or relative organ weights in the few treated rats surviving at termination.
Lesions of the forestomach were the most prominent nonneoplastic histopathologic
effects observed in Sprague-Dawley rats (Quast, 2002; Quast et al., 1980a). In the 1-year interim
sacrifice, the only treatment-related noncancer lesion was squamous cell hyperplasia of the
forestomach, which was observed in 10/10 males and 9/10 females in the 300 ppm group and
4/10 males and 7/10 females in the 100 ppm group. At 2 years, the incidence of forestomach
lesions (hyperplasia and/or hyperkeratosis of the squamous epithelium) was dose related and
significantly elevated in males at > 100 ppm and in females at >35 ppm; incidences were 15/80,
15/47, 44/48, and 45/48 for males and 20/80, 23/48, 41/48, and 47/48 for females in the 0, 35,
100, and 300 ppm dose groups, respectively. Minimal progressive chronic nephropathy was also
elevated in females treated at>100 ppm; incidences were 37/80, 24/48, 37/48, and 38/48 for
control to high-dose groups. In males, minimal progressive nephropathy was elevated only in
the 300 ppm group. However, significant increase in severe progressive nephropathy was
observed in males in the 35 and 100 ppm groups; incidences were 11/80, 13/47, 13/48, and 10/48
for control to high-dose groups. Gliosis of the brain, with or without perivascular cuffing, was
significantly increased in the low- and middle-dose females (8/48 and 13/48, respectively, vs.
2/80 in controls), although the incidence (5/48) did not reach statistical significance in the high-
dose females. In males, it was present in one rat exposed at 100 ppm and three rats exposed at
200 ppm. A NOAEL was not identified in this drinking water study. A LOAEL of 4.4 mg/kg-
day was identified for increases in forestomach lesions, decreased survival, and gliosis in brain in
female rats exposed to 35 ppm AN in drinking water. Table 4-22 summarizes incidence of
nonneoplastic lesions in Sprague-Dawley rats exposed to AN in drinking water for 2 years.
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Table 4-22. Incidence of nonneoplastic lesions in Sprague-Dawley rats
exposed to AN in drinking water for 2 years
AN in drinking water (ppm)
Male rats Dose (mg/kg-d)
Forestomach hyperplasia and/or hyperkeratosis
0
0
15/80 (19%)
35
3.4
15/47 (32%)
100
8.5
44/48 (92%)a
300
21.3
45/48 (94%)a
Kidneys — chronic progressive nephropathy
Severity: minimal
moderate
severe
Female rats Dose (mg/kg-d)
Forestomach hyperplasia and/or hyperkeratosis
10/80 (12%)
10/80 (12%)
11/80(14%)
0
20/80 (25%)
4/47 (9%)
5/47(11%)
13/47 (28%)a
4.4
23/48 (48%)a
7/48 (15%)
10/48 (21%)
13/48 (27%)a
10.8
41/48 (85%)a
16/48 (33%)a
16/48 (33%)a
10/48 (21%)
25.0
47/48 (98%)a
Kidneys — chronic progressive nephropathy
Severity: minimal
moderate
severe
Brain — gliosis and perivascular cuffing
37/80 (46%)
17/80 (21%)
13/80 (16%)
2/80 (3%)
24/48 (50%)
13/48 (27%)
9/48 (19%)
8/48 (17%)a
37/48 (77%)a
8/48 (17)
1/48 (2%)
13/48 (27%)a
38/48 (79%)a
5/48 (10%)
4/48 (8%)
5/48 (10%)
aStatistically significant at/? < 0.05.
Source: Quast(2002).
Cancer results
The drinking water study by Quast (2002) and Quast et al. (1980a) provided evidence of
the carcinogenicity of AN in male and female Sprague-Dawley rats. Tumor findings in the
1-year sacrifices included forestomach papillomas in males: 1/10 at 100 ppm and 7/10 at
300 ppm. In females, the occurrence was 5/10 at 300 ppm. Also observed after 1 year of
exposure were microscopic tumors of the CNS (brain) in 2/10 males and 4/10 females at
100 ppm and 1/10 males and 2/10 females at 300 ppm. Carcinoma of Zymbal gland and benign
and malignant tumors of the mammary gland, also observed after 1 year of exposure, were
considered by the study authors to be related to treatment.
Histopathologic examinations after 2 years of exposure revealed statistically significant
and dose-dependent increase in incidences of several types of tumors (Table 4-23). Squamous
cell papillomas or carcinomas of the forestomach were significantly elevated in males and
females exposed at >100 ppm. Carcinomas of Zymbal gland were significantly increased in
males at 300 ppm and in females at >35 ppm.
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Table 4-23. Selected tumor incidences in response to AN administered to
Sprague-Dawley rats in drinking water for up to 2 years
Tissue type
Astrocytomas only
Combined astrocytomas/glial
cell proliferation
Tongue
(papilloma or carcinoma)
Forestomach
(papilloma or carcinoma)
Zymbal gland
(carcinoma or adenoma)
Small intestine (mucous
cystadenocarcinoma)
Mammary gland (malignant)
Mammary gland
(malignant or benign)
Incidence of tumor formation
Male rats
Female rats
Exposure concentration (ppm)
0
35
100
300
0
35
100
300
Dose (mg/kg-d)
0
1/80
1/80
1/80
0/80
3/80
NDb
ND
ND
3.4
8/47a
12/47a
2/47
2/47
4/47
ND
ND
ND
8.5
19/48
22/48a
4/48
23/48a
3/48
ND
ND
ND
21.3
23/48a
30/48a
5/48a
39/48a
16/48a
ND
ND
ND
0
1/80
1/80
0/80
1/80
1/80
0/80
1/80
58/80
4.4
17/48a
20/48a
1/48
1/48
5/48a
1/48
1/48
42/48a
10.8
22/48a
25/48a
2/48
12/48a
9/48a
4/48a
3/48
42/48a
25.0
24/48a
31/48a
12/48a
30/48a
18/48a
4/48a
10/48a
35/48
""Significantly different from controls (p < 0.05), as calculated by the study authors.
bND = not determined.
Sources: Quast (2002); Quast et al. (1980a).
An increase in papillomas or carcinomas of the tongue in males at 300 ppm was also
considered to be related to exposure. Increases were also observed in the incidences of
malignant tumors of the mammary gland. The incidence of malignant or benign mammary
tumors was significantly increased in the low- and high-dose females but decreased in the high-
dose group. Quast (2002) noted that the lower mammary tumor incidence in the 300 ppm group
was likely due to the marked early mortality in this group, despite the earlier occurrence of these
tumors. The incidence of mammary tumors increased considerably in controls during the latter
portion of the study, when few high-dose females survived.
Two diagnostic categories—focal or multifocal glial cell proliferation (suggestive of
early tumors), and focal or multifocal glial cell tumor (astrocytomas)—were used by Quast
(2002) for tumors present in the brain or spinal cord. These diagnoses were mutually exclusive
and primarily based on the size of the lesion, with glial cell proliferation a smaller-sized lesion
than astrocytoma. For enumerating the astrocytomas of the CNS, the incidence of glial
proliferation was combined with the incidence of astrocytomas and was elevated in both males
and females in all exposed groups (Table 4-23).
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Distribution of astrocytomas in the various regions of nervous tissue in male and female
Sprague-Dawley rats is shown in Table 4-24. No tumors were found from 0 to 12 months in any
location in either male or female rats. As noted by Quast (2002), the sections of cerebral cortex
contained most of the tumors because of the cortex's larger size. However, astrocytomas were
also found in the cerebellum, brain stem, and spinal cord in male and female rats. Smaller-sized
lesions (glial cell proliferation) had a distribution similar to astrocytomas.
Table 4-24. Histopathologic location of astrocytomas in the CNS of male
Sprague-Dawley rats administered AN in drinking water for 2 years
Age at sacrifice
Location of astrocytomas11
Male rats
Cerebral
cortex
Cerebellum
Brain
stem
Spinal
cord
Female rats
Cerebral
cortex
Cerebellum
Brain
stem
Spinal
cord
Oppm
13-18 mos
19-24 mos
Terminal kill
0/23
1/43
0/7
0/23
0/43
0/7
0/23
0/43
0/7
0/22
0/43
0/7
0/11
0/48
1/20
0/11
0/48
0/20
0/10
0/47
0/20
0/11
0/46
0/20
35 ppm
13-18 mos
19-24 mos
Terminal kill
2/14
5/26
1/5
0/14
1/26
0/5
1/14
1/25
1/5
0/14
0/26
0/5
1/13
5/30
3/4
0/13
2/29
0/4
1/13
6/30
2/4
0/13
0/30
0/4
WOppm
13-18 mos
19-24 mos
Terminal kill
3/16
13/26
2/5
1/16
0/26
0/5
2/16
6/26
0/5
1/16
1/26
0/5
5/19
12/24
0/1
2/20
1/24
0/1
1/19
9/24
1/1
1/19
1/24
0/1
300 ppm
13-18 mos
19-24 mos
Terminal kill
10/26
11/18
0/0
0/26
2/18
0/0
2/26
4/18
0/0
1/26
2/18
0/0
11/23
8/11
0/0
2/23
2/11
0/0
4/23
3/11
0/0
3/22
2/11
0/0
aNo astrocytomas were found in any locations from 0 to 12 mos.
Sources: Quast (2002); Quast et al. (1980a).
4.2.1.2.2. Johannsen and Levimkas (2002a) and Biodynamics (1980a): drinking water study.
Johannsen and Levinskas (2002a) and Biodynamics (1980a) carried out another 2-year drinking
water study of AN in Sprague-Dawley rats. Groups of 100 rats/sex/group were exposed to 0, 1,
or 100 ppm AN (100% purity) in drinking water; as calculated by the study authors, average
daily doses were 0, 0.09, and 8.0 mg/kg-day , respectively, for males and 0, 0.15, and
10.7 mg/kg-day, respectively, for females. Ten rats/sex/group were selected from these groups
for interim evaluations at 6, 12, and 18 months and at study term, leaving a maximum of
70/sex/group for lifetime exposure. Although the study was designed to last 24 months, because
of high mortality among high-dose rats, surviving males were necropsied after 22 months and
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females at 19 months to ensure that a sufficient number of animals would be available for
clinical and pathological analyses. Rats were observed daily for morbidity and mortality and
examined and palpated weekly to detect growths. All rats received ophthalmoscopic
examinations in the period before treatment started and again at the time of necropsy.
BWs were recorded weekly for the first 14 weeks, biweekly between weeks 16 and 26,
and monthly thereafter. Food intake and water consumption were recorded over 3-day periods
for about 25 rats/sex/group at the same intervals at which BWs were recorded. Ten rats/sex from
each group were selected for hematological, clinical chemical, and urinalysis examinations at 6,
12, and 18 months and at study termination (22 months for males and 19 months for females).
All rats, whether dying prematurely, sacrificed in a moribund condition, or sacrificed on
schedule, were subjected to a gross necropsy that included preservation of 40 tissues and organs,
gross lesions, and tissue masses for possible histopathological examination. At interim and
terminal necropsies, weights were recorded for selected organs (brain, pituitary, adrenal, gonads,
heart, kidney, and liver) in 10 rats/sex/group. Interim gross necropsies were performed after 6,
12, and 18 months on 10 rats/sex/group and on the remaining rats at termination. For the
scheduled interim and terminal sacrifices, complete histopathologic examinations were
conducted on 10 rats/sex from the control and 100 ppm groups; on all other rats, a limited set of
tissues was examined microscopically that included potential target organs (brain, ear canal,
spinal cord, and stomach) and any tissue masses observed at necropsy.
Noncancer results
Treatment-related noncancer effects were observed in Sprague-Dawley rats exposed to
AN for 19-22 months (Johannsen and Levinskas, 2002a; Biodynamics, 1980a). Statistically
significant increases in early deaths were observed in high-dose male and female rats after
10 months and became particularly severe during the last 5 months of the study. Reduction in
survival compared with controls was not significant in high-dose males at termination
(22 months) but was significant in high-dose female rats terminated at 19 months. Food and
water consumption in high-dose rats was lower than in controls throughout the study. Reduction
in mean BW compared with controls was observed in high-dose males and females throughout
the study (a significant 10% reduction in males and an insignificant 8% reduction in females at
term). Slight reductions in hematocrit, RBC count, and Fib values compared with controls were
observed during the study in high-dose groups but were generally not statistically significant; 3-
6% reductions in Hb values in high-dose males were statistically significant at the 6- and
18-month interim time points. Exposure to AN had no effect on clinical chemistry or urinalysis
parameters or ophthalmoscopic examinations; there were no treatment-related clinical signs
except for some neurological symptoms at the time of terminal necropsy in a few rats with
astrocytomas of the brain or spinal cord. The only significant treatment-related effects on organ
weights were lower absolute pituitary weights in high-dose males at 12 months and in females at
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termination. Significant increase in mean relative kidney and liver weights observed in high-
dose males and females were considered by the study authors to reflect the reduced BWs rather
than target organ effects, since this effect did not occur at other study intervals.
Statistically significant increases in nonneoplastic lesions related to exposure to AN
included renal transitional cell hyperplasia in high-dose females at termination (Table 4-25).
Significant increase in squamous metaplasia of the uterus was observed in high-dose females at
12 months compared with controls (5/10 vs. 0/10), and in low-dose females at 18 months (3/3 vs.
2/10). This effect was not significant in high-dose females at 18 months (4/10), and was not
observed at terminal sacrifice.
Table 4-25. Incidence of nonneoplastic lesions in Sprague-Dawley rats
exposed to AN in drinking water for 2 years
AN in drinking water (ppm)
Male rats Dose (mg/kg-d)
Kidney transitional cell hyperplasia
Oa
0
1/15
1
0.09
2/18
100
8.0
0/12
Forestomach squamous cell hyperplasia
Incidence at termination13
Severity: mild
moderate
severe
14/19
5
7
2
24/26
4
14C
6C
14/15
1
8C
5C
Forestomach squamous cell hyperplasia
Incidence in early deaths or unscheduled sacrifices
Severity: mild
moderate
severe
Female rats Dose (mg/kg-d)
Kidney transitional cell hyperplasia
36/46
9
17
10
0
2/27
31/37
10
15
6
0.15
1/18
37/42
4
18C
15C
10.7
8/14d
Forestomach squamous cell hyperplasia
Incidence at termination13
Severity: mild
moderate
severe
Incidence in early deaths or unscheduled sacrifices
Severity: mild
moderate
severe
39/48
12
18
9
14/20
4
7
3
40/57
8
25
7
6/12
1
2
3
17/17
4
9
4
33/43
1
21d
lld
"Drinking water control.
bMales exposed for 22 mos, females for 19 mos.
Significantly different from controls as calculated by authors (p < 0.05).
dSignificantly different from controls as calculated by authors (p < 0.01).
Sources: Johannsen and Levinskas (2002a); Biodynamics (1980a).
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In addition, the incidence of squamous cell hyperplasia was increased in high-dose males
and females at terminal sacrifice and in spontaneous deaths in high-dose groups (Table 4-25).
The severity of squamous cell hyperplasia of the forestomach was significantly greater
(characterized as moderate or severe) than in controls in male rats at termination after exposure
to 1 or 100 ppm AN; a significant increase in severity in high-dose males was also observed at
12 months (not shown) but not at 18 months. Increased incidences in forestomach hyperplasia
characterized as moderate or severe were also observed in high-dose rats of both sexes that died
prematurely or were sacrificed in a moribund condition. A NOAEL for nonneoplastic effects
was not identified in this study. A LOAEL of 1 ppm (0.09 mg/kg-day) was identified for
increased severity of squamous cell hyperplasia of the forestomach in male Sprague-Dawley rats
exposed to AN in drinking water for 2 years (22 months).
Cancer results
Significant increases in the cumulative incidences of neoplasms of the CNS, ear canal
(Zymbal gland), and forestomach were observed in Sprague-Dawley rats exposed to 100 ppm
AN in drinking water for 2 years (Johannsen and Levinskas, 2002a; Biodynamics, 1980a).
Statistically significant, dose-dependent increases were observed for astrocytomas of the brain in
both sexes and astrocytomas of the spinal cord in females. The astrocytomas varied from
discrete solid masses to localized areas of diffuse infiltrations of neoplastic cells. In addition,
adenomas of Zymbal gland (ear canal) in both sexes, carcinomas of Zymbal gland in both sexes,
and squamous cell papillomas of the forestomach in females were increased. Intestinal
adenocarcinomas were also found in high-dose male and female rats (see Table 4-26).
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Table 4-26. Selected tumor incidences in Sprague-Dawley rats exposed to AN
in drinking water for up to 2 years
Tissue type
Tumor incidence"
Male rats
Female rats
Exposure concentration (ppm)
0
1
100
0
1
100
Dose (mg/kg-d)
0
0.09
7.98
0
0.15
10.7
CNS
Brain (astrocytomas)
Spinal cord (astrocytomas)
Total
2/78
NDC
2/78
3/75
ND
3/75
23/77b
ND
23/77b
0/79
0/76
0/79
1/80
0/79
1/80
32/77b
7/78b
39/78b
Zymbal gland
Carcinomas
Adenomas
Total
1/80
0/80
1/80
0/71
0/71
0/71
14/83b
5/83
19/83b
0/79
1/79
1/79
0/75
0/75
0/75
7/78b
5/78
12/78b
Forestomach
Carcinomas
Papillomas
Total
0/78
3/78
3/78
1/78
2/78
3/78
3/77
8/77
ll/77b
0/80
1/80
1/80
0/79
4/79
4/79
0/79
7/79b
7/79b
Intestine
Adenocarcinomas
0/40
0/34
2/41
0/78
0/79
2/70
"Denominators are calculated from the total number of animals examined (as reported in Table 4 of Johannsen and
Levinskas, 2002a) minus the animals scheduled for sacrifice at 6 and 12 mos. Thus, incidences are for animals
scheduled for the 18-mo, spontaneous deaths and terminal sacrifices (22 mos for males and 19 mos for females). For
Zymbal gland tumors in the 100 ppm male, one carcinoma and adenoma occurred in the 12-mo sacrifice; therefore,
adjustment to the denominator was only made for the 6-mo sacrifice.
bSignificantly different from controls (p < 0.05), Fisher's exact test.
°ND = no data collected.
Sources: Johannsen and Levinskas (2002a); Biodynamics (1980a).
Johannsen and Levinskas (2002a) reported that most of the tumors occurred in rats after
12 months of exposure. CNS astrocytomas were not detected at the 6- or 12-month interim
sacrifices (at which about 10 rats of each gender were sacrificed from each group). However,
brain astrocytomas and Zymbal gland carcinomas were detected in high-dose females as early as
after 6 months of exposure. One each of Zymbal gland carcinoma, adenoma, and squamous cell
carcinoma of the forestomach was detected at the 12-month interim sacrifice in high-dose males.
This finding was consistent with findings in the other drinking water bioassay with Sprague-
Dawley rats, indicating that some male and female rats exposed to 300 ppm AN developed
tumors after only 7-12 months of exposure (Quast, 2002). Mammary gland carcinomas were
detected in female rats scheduled for the terminal sacrifice but not in male or female rats
scheduled for sacrifice at 6, 12, or 18 months. Incidences of female rats (scheduled for the
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terminal sacrifice) with mammary gland carcinomas, however, did not show exposure-related
increases: 13/80, 4/79, and 14/79 for the control, 1 ppm, and 100 ppm female groups,
respectively.
4.2.1.2.3. Johannsen and Levinskas (2002a) and Biodynamics (1980b): gavage study. In
parallel to the chronic drinking water study in Sprague-Dawley rats, Johannsen and Levinskas
(2002a) and Biodynamics (1980b) conducted a lifetime study in which groups of Sprague-
Dawley rats (100 rats/sex/group) received 0, 0.1, or 10 mg/kg-day AN by gavage in water for
7 days/week. The doses were selected to match the daily intake levels in the drinking water
study. The schedule and protocols used in the gavage study for observations of clinical signs,
measurements of food and water consumption and BW, clinical biochemistry analyses, and gross
necropsy and histopathologic analyses were identical to those in the drinking water study.
Ingestion of 10 mg/kg-day AN resulted in increased mortality after 12 months and was
significantly higher in males at termination and in females beginning at month 14 compared with
controls. The study was terminated after 20 months of exposure because of high treatment-
related deaths, with a decreased survival of about 30 and 50% in males and females compared
with their respective controls. BWs showed a significant reduction of about 6% in high-dose
males compared with controls, beginning on week 12 and reaching 13% by termination; there
were no BW effects in exposed females.
Exposure to AN had no effect on food consumption in either sex or on water
consumption in males; high-dose females had slightly increased water consumption during the
first 12 months only. Small reductions in hematocrit, Hb, and RBC counts were observed in
high-dose males at 12 and 18 months and reached statistical significance at terminal sacrifice
(hematocrit, 15%; Hb, 19%; and RBC counts, 18%). Significant increases in absolute mean liver
weights were observed in the high-dose males and females and low-dose males at the 18-month
interval. Relative liver weights for high-dose males and females were significantly increased in
most intervals and might be reflective of lower BWs. Absolute and relative kidney weights of
high-dose males were increased significantly at 18 months and increased insignificantly in both
males and females at study term. The adrenal gland of high-dose males was the organ that
showed the most significant weight increase of 43% at termination.
Noncancer results
Nonneoplastic histopathological effects were observed in male and female Sprague-
Dawley rats exposed to 10 mg/kg-day AN by gavage for 20 months (Table 4-27). A significant
increase in epidermal inclusion cysts was observed in male and female rats, most notably in
animals dying spontaneously. Significant increases in renal transitional cell hyperplasia were
observed in high-dose females at 12-month sacrifice and in high-dose males after 18 months. In
addition, there was a significant increase in the severity of squamous cell hyperplasia of the
132 DRAFT - DO NOT CITE OR QUOTE
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forestomach in high-dose males and females. This effect was most noticeable among rats dying
early or at scheduled sacrifices after at least 12 months. High-dose rats showed a significant
elevation in the incidence of moderate or severe hyperplasia. In this gavage study, a NOAEL of
0.1 mg/kg-day and a LOAEL of 10 mg/kg-day were identified for increased severity of
forestomach lesions (squamous cell hyperplasia) in male and female rats.
Table 4-27. Incidence of nonneoplastic lesions in Sprague-Dawley rats
exposed to AN by gavage for 20 months
Dose (mg/kg-d)
Oa
0.1
10
Male rats
Skin: Epidermal inclusion cysts (in spontaneous deaths)
Kidney: Transitional cell hyperplasia (18 mos)
Heart: Cardiomyopathy (18 mos)
Forestomach: Squamous cell hyperplasia
Incidence at terminal sacrifice
Severity: mild
moderate
severe
Incidence in early deaths or unscheduled sacrifices
Severity: mild
moderate
severe
2/59
3/11
3/11
10/10
2
5
3
46/58
18
19
9
4/68
0/8
No data
No data
46/67
11
26
10
8/56b
10/10C
8/10b
8/10
0
2
6
49/56
2
13C
34C
Female rats
Skin: Epidermal inclusion cysts (in spontaneous deaths)
Kidney: Transitional cell hyperplasia (12 mos)
Heart: Cardiomyopathy (18 mos)
Forestomach squamous cell hyperplasia
Incidence at terminal sacrifice
Severity: mild
moderate
severe
Incidence in early deaths or unscheduled sacrifices
Severity: mild
moderate
severe
0/10
0/10
4/10
8/10
1
5
2
47/57
12
23
12
No data
0/1
No data
No data
63/69
12
38
13
4/10b
4/10b
3/10
9/10
1
3
5
53/59
0
14C
39C
aWater vehicle control.
bSignificantly different from controls as calculated by the study authors (p < 0.05).
Significantly different from controls as calculated by the study authors (p < 0.01).
Sources: Johannsen and Levinskas (2002a); Biodynamics (1980a).
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Cancer results
The carcinogenic effects of AN on Sprague-Dawley rats exposed via gavage were similar
to those in the drinking water studies (Johannsen and Levinskas, 2002a; Biodynamics, 1980b).
Significant increase in tumor incidences were observed in male and female rats dosed with
10 mg/kg-day and included carcinomas of Zymbal gland, astrocytomas of the brain, and tumors
of the squamous epithelium of the forestomach (papillomas and carcinomas in males and
papillomas in females) (Table 4-28). Adenocarcinomas of the intestine were also observed in
high-dose male rats that died or were killed after at least 12 months of exposure. Increases in
carcinomas of the mammary gland were observed in high-dose females that died prematurely.
Table 4-28. Cumulative incidence of tumors in response to AN administered
to Sprague-Dawley rats by gavage for up to 2 years
Tissue type
Brain astrocytomas
Spinal cord astrocytomas
Zymbal gland papillomas
Zymbal gland carcinomas
Zymbal gland adenomas
Forestomach carcinomas
Forestomach papillomas
Intestine adenocarcinomas
Mammary gland carcinomas
Tumor incidence"
Male rats (mg/kg-d)
0
2/80
0/74
0/76
1/76
0/76
0/79
2/79
0/80
0/80
0.1
0/79
0/73
0/73
0/73
1/73
0/77
6/77
0/80
0/78
10
16/77b
1/77
3/76
10/76b
5/76b
18/79b
19/79b
6/80b
0/80
Female rats (mg/kg-d)
0
1/80
0/80
0/65
0/65
1/65
0/79
2/79
0/40
5/80
0.1
2/78
0/75
0/74
0/74
0/74
0/79
4/79
0/40
6/80
10
17/80b
1/79
0/74
9/74b
5/74
1/79
14/79b
1/41
21/80b
""Denominators are calculated from the total number of animals examined (as reported in Table 4 of Johannsen and
Levinskas, 2002a) minus the animals scheduled for sacrifice at 6 and 12 mos; thus, incidences are for animals
scheduled for the 18-mo, spontaneous death, and terminal sacrifices (20 mos).
bSignificantly different from controls (p < 0.05), as calculated by the study authors.
Sources: Johannsen and Levinskas (2002a); Biodynamics (1980a).
4.2.1.2.4. Johannsen and Levinskas (2002b); Biodynamics (1980c). A 2-year drinking water
assay was conducted in F344 rats (Johannsen and Levinskas, 2002b; Biodynamics, 1980c).
Groups of rats (100/sex/group) were given AN (100% purity) in drinking water at concentrations
of 0, 1, 3, 10, 30, or 100 ppm; two groups of 100/sex served as untreated controls. The study
authors reported the equivalent average daily doses of AN as 0, 0.1, 0.3, 0.8, 2.5, and 8.4 mg/kg-
day for males and 0, 0.1, 0.4, 1.3, 3.7, and 10.9 mg/kg-day for females. Rats were observed
twice daily for overt signs of toxicity and examined and palpated weekly to detect growths.
Ophthalmoscopic examinations were carried out before testing and again at the time of necropsy.
BWs were recorded weekly for the first 14 weeks, biweekly between weeks 16 and 26, and
monthly thereafter. Food intake and water consumption were recorded over a 3-day period at the
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same intervals that BWs were recorded for about 25 rats/sex/group for all groups. At intervals of
6, 12, and 18 months and at study termination (26 months for males and 23 months for females),
10 rats/sex/group from each exposed and control group were sacrificed for clinical pathology and
microscopic evaluation.
At 6, 12, and 18 months and at termination, 10 rats/sex from the 100 ppm group and
5 rats/sex from each of the control groups were selected for hematology, clinical chemistry, and
urinalysis examinations. Lower dose group rats were evaluated as needed to determine possible
dose-response relationships. All rats, whether dying prematurely, sacrificed in a moribund
condition, or sacrificed on schedule, were subjected to a gross necropsy that included
preservation of 40 tissues and organs, all gross lesions, and tissue masses for possible
histopathologic examination. At interim and terminal sacrifice, weights were recorded for
selected organs (brain, pituitary, adrenal, gonads, heart, kidney, and liver) for 10 rats/sex from
treated groups and 5 rats/sex from each control group. For the interim sacrifices, complete
histopathologic examinations were conducted on 10 rats/sex from the 100 ppm group and
5 rats/sex from the two control groups. Potential target organs (e.g., brain, ear canal, spinal cord,
and stomach), gross lesions, and tissue masses were examined microscopically in all study
animals. All surviving females were terminated after 23 months and all males after 26 months.
Terminal necropsies included complete microscopic examinations for 10 rats/sex exposed to
100 ppm AN and 5/sex/group for the two control groups. Potential target organs and suspicious
lesions were examined in all animals in other dose groups.
Noncancer results
Treatment-related noncancer effects were observed in F344 rats exposed to AN in
drinking water for 2 years (Johannsen and Levinskas, 2002b; Biodynamics, 1980c). Statistically
significantly early deaths (p < 0.05), compared with controls, were observed in male and female
rats exposed to 100 ppm, beginning after 14 months of exposure. A statistically significant
decrease in survival was also observed in female rats exposed to 30 ppm but not to 10 ppm,
beginning after 18 months of exposure. Exposure to AN did not result in any overt neurological
impairments or ophthalmoscopic findings.
Statistically significant decrease in BWs (-12% lower than controls,/? < 0.01) were
observed in male and female rats exposed to 100 ppm AN. A statistically significant decrease in
BWs of less than 5% was found in male rats exposed to 30 ppm AN and was not biologically
significant. Total food consumption (g/week) was reduced compared with controls in rats
exposed at 100 ppm (more prominently in female rats after week 13), but consumption on a BW
basis (g/kg) for both 100 ppm male and female rats was not significantly different from that for
controls. Total water consumption (mL/week) was significantly lower in the 100 ppm male and
female rats compared with controls. No differences in food and water consumption from
controls were found for groups exposed to <30 ppm AN. Hematological analyses revealed slight
13 5 DRAFT - DO NOT CITE OR QUOTE
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reductions in Hb, hematocrit, and RBC counts in the 100 ppm group, beginning at 12 months,
with Hb significantly lower in female rats at 12 and 18 months (13% at 18 months). Hematocrit
in 100 ppm female rats was 4% lower at 12 months. However, no statistically significant
differences from controls were noted at termination.
Serum alkaline phosphatase levels were significantly elevated by 80 and 65% (p < 0.05),
respectively, in 100 ppm females at 18 and 23 months. Results of clinical chemistry for the
30 and 10 ppm groups were not reported. However, the study authors noted that female rats in
these dose groups also showed significant elevation of this enzyme. A significant 33% increase
in SGPT was also observed in 100 ppm female rats at 18 months but not at other intervals. The
only treatment-related urinalysis finding was a slight increase in urine specific gravity in
100 ppm males at 18 and 26 months. No significant dose-related changes were observed for
absolute organ weights; elevations in relative organ weights (kidney, brain, liver, adrenal, testis,
heart, and pituitary in females) sporadically observed at 100 ppm at intervals throughout the
study were attributed by the study authors to lower BWs rather than target-organ toxicity.
Unlike Sprague-Dawley rats exposed to AN in drinking water (see above, Quast [2002]),
F344 rats did not show forestomach lesions at the 6- or 12-month sacrifices. In rats exposed for
periods >1 year, treatment-related nonneoplastic effects were observed in the forestomach
(Table 4-29) and skin. The incidence of squamous cell hyperplasia or hyperkeratosis of the
forestomach was elevated in male and female rats exposed to 3, 10, and 30 ppm but not to
100 ppm. Johannsen and Levinskas (2002b) suggested that, because the lesions were observed
more frequently in late surviving animals, the reduced survival in 100 ppm group may have been
the reason that no significant increase was observed for that dose group. Because these
forestomach lesions were not observed in rats examined during the 6- or 12-month sacrifices, no
such incidence data are listed in Table 4-29. An increase in epidermal inclusion cysts of the
skin, observed in 4/50 male rats treated at 100 ppm but in no other group, was also considered by
the study authors to be treatment related. A NOAEL of 1 ppm (0.1 mg/kg-day) and a LOAEL of
3 ppm (0.3 and 0.4 mg/kg-day for males and females, respectively) were identified for increases
in squamous cell lesions (hyperplasia and/or hyperkeratosis) of the forestomach in male and
female F344 rats exposed to AN in drinking water for 2 years.
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Table 4-29. Incidences of nonneoplastic lesions in F344 rats exposed to AN
in drinking water for 2 years
AN in drinking water (ppm)
Male rats Dose (mg/kg-d)
Forestomach squamous cell hyperplasia or
hyperkeratosisb
Foci of cellular alteration in liver
Epidermal inclusion cysts
Chronic nephropathy, bilateral
Atrophy of seminiferous tubules, bilateral
Female rats Dose (mg/kg-d)
Forestomach squamous cell hyperplasia or
hyperkeratosisb
Foci of cellular alteration in liver
Chronic nephropathy, bilateral
Oa
0
11/159
7/94
0/61
118/149
150/172
0
4/156
1/84
52/68
1
0.1
3/80
5/31
0/12
57/58d
72/75d
0.1
2/80
2/20
10/19
3
0.3
18/75C
11/51
0/12
62/64c
68/84
0.4
16/80C
4/3 9d
21/26
10
0.8
13/80d
8/45
0/11
52/54d
64/74
1.3
23/74c
3/49
20/29
30
2.5
17/80C
14/50C
0/17
57/60d
78/78c
3.7
13/80C
0/40
23/26
100
8.4
9/77
6/68
4/50d
42/62
64/87
10.9
5/74
1/70
18/50
aDrinking water control.
bMales exposed for 18-26 mos, females for 18-23+ mos; excludes rats sacrificed at 6 or 12 mos. These incidences
were further adjusted to exclude (from the denominators) rats that died between 0 and 12 mos in the study. Rats
dying during this time period were determined from page 6 of Appendix H and Table 1 in Biodynamics (1980c)
and Table 8 in Johannsen and Levinskas (2002b). Unscheduled deaths between 0 and 12 mos in the study occurred
in two female controls, two males at 3 ppm, three females at 10 ppm, and three males and three females at 100 ppm.
Significantly different from vehicle control as calculated by study authors (p < 0.01).
dSignificantly different from vehicle control as calculated by study authors (p < 0.05).
Sources: Johannsen and Levinskas (2002b); Biodynamics (1980c).
Other microscopic findings included increase in chronic nephropathy and atrophy of
seminiferous tubules, which were significant at 1 ppm. However, the study authors did not
consider these effects to be related to treatment, but rather to be due to the sample selection
procedure that only tissues showing questionable or suspicious lesions were selected for
microscopic examination for the intermediate dose groups (1-30 ppm), resulting in the selection
of tissues from more aged animals.
Cancer results
Increases in tumor incidences (Table 4-30) were observed in multiple organs, following
chronic exposure of F344 rats to AN in drinking water (Johannsen and Levinskas, 2002b;
Biodynamics, 1980c). No tumors were detected in rats sacrificed after only 6 months of
exposure. One female rat exposed to 100 ppm for 12 months had an adenocarcinoma of the
mammary gland that the study authors considered to be possibly related to treatment. Treatment-
related tumors, appearing at the 18-month interim sacrifice, included brain astrocytomas
(one/sex/group) in male and female rats exposed to 30 and 100 ppm, squamous cell papillomas
of the ear canal (Zymbal gland) in 2/9 males at 100 ppm and 1/9 females at 30 ppm, and
squamous cell carcinomas of the latter organ in one female at 100 ppm. AN-related neoplastic
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lesions were observed at terminal sacrifice in the brain, spinal cord, forestomach, ear canal
(Zymbal gland), and mammary gland (females only). Significant dose-related elevations in
astrocytomas were observed in the brains of male and female rats exposed to 30 and 100 ppm
and in the spinal cord of male rats exposed to 100 ppm (see Table 4-30). Dose-related increases
in squamous cell adenomas/carcinomas of the ear canal (Zymbal gland) were observed in males
exposed to 30 and 100 ppm and in females at 10-100 ppm. Increased incidences of squamous
cell papillomas or carcinomas of the forestomach, observed in males exposed to 3 and 10 ppm
and in both sexes exposed to 30 ppm, were considered by the study authors to be likely treatment
related despite the lack of a dose response. Fibroadenomas of the mammary gland were slightly
increased in females treated with 10 and 30 ppm. Other lesions considered by the study authors
to be possibly treatment related included a single case of squamous cell carcinoma of the salivary
gland in one male at 100 ppm and squamous cell papillomas of the tongue in female rats
(one/group) exposed to 10 and 30 ppm.
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Table 4-30. Selected tumor incidences in F344 rats exposed to AN in
drinking water for 2 years
AN in drinking water (ppm)
Male ratsb Dose (mg/kg-d)
Brain astrocytoma Total0
Adjustedd
Spinal cord astrocytoma
Ear canal (Zymbal gland) squamous cell
papilloma/adenoma and carcinoma
Forestomach squamous cell papilloma/carcinoma
Mammary gland fibroadenoma
Mammary gland carcinoma
Female ratsb Dose (mg/kg-d)
Brain astrocytoma
Spinal cord astrocytoma
Ear canal (Zymbal gland) squamous cell
papilloma/adenoma and carcinoma
Forestomach squamous cell papilloma/carcinoma
Mammary gland fibroadenoma
Mammary gland carcinoma
Oa
0
2/200c
2/160d
1/196
1/156
2/189
1/147
0/199
0/159
2/48
1/48
0
1/199
1/157
0/197
0/155
0/193
0/157
1/199
1/157
12/65
12/156
1/65
3/156
1
0.1
2/100
2/80
0/99
0/79
1/97
1/76
1/100
1/80
1/4
0/4
0.1
1/100
1/80
0/97
0/78
0/94
0/73
1/100
1/80
5/14
5/80
2/14
4/80
3
0.3
1/100
1/78
0/92
0/70
0/93
0/73
4/97f
4/78f
2/7
0/7
0.4
2/101
2/80
0/99
0/79
2/92
0/73
2/100
2/79
6/14
6/80
0/14
0/80
10
0.8
2/100
2/80
0/98
0/78
2/88
0/67
4/100f
3/80f
2/12
0/12
1.3
4/95
4/75
1/92
1/72
4/90e
0/70e
2/97
2/77
9/16e
8/79
0/16
1/78
30
2.5
10/99e
10/79e
0/99
0/79
7/94-
2/7 r
4/100f
4/80f
0/7
0/7
3.7
6/100e
6/80e
0/96
0/77
5/94e
2/73 e
4/100f
4/80f
10/22f
9/80
3/22
3/80
100
8.4
21/99e
21/76e
4/93f
4/70f
16/93e
14/68e
1/101
1/77
2/45
0/45
10.9
23/98e
23/76e
1/91
1/69
10/86e
8/62e
2/97
2/75
9/49
9/73
2/49
6/73
aDrinking water control.
bMales exposed for 26 mos, females for 23+ mos.
°Total cumulative number of rats with lesion.
dThe denominators for incidences excluded rats from the 6- and 12-mo sacrifices and rats that died before the
appearance of the first tumor for each of three tumor sites: CNS, Zymbal gland, and forestomach. The termination
history reports in Appendix C and the individual animal histopathology reports in Appendix H of the Biodynamics
(1980b) report were examined to determine time of death and tumor occurrence for each of the F344 rats. Times of
first detection of tumors were 419 d for forestomach tumors, 495 d for CNS tumors, and 475 d for Zymbal gland
tumors. Due to the limited number of mammary glands examined in most groups, the adjusted denominators
represented the number of animals that were exposed for more than 12 mos for each group; mammary gland tumor
incidences are for animals scheduled for the 18-mo and terminal sacrifices.
Significantly different from vehicle control as calculated by study authors (cumulative only) (p < 0.01).
Significantly different from vehicle control as calculated by study authors (cumulative only) (p < 0.05).
Sources: Johannsen and Levinskas (2002b); Biodynamics (1980c).
4.2.1.2.5. NTP (2001). NTP (2001) evaluated the toxicity and carcinogenicity of AN (>99%
purity) given by gavage in water to B6C3Fi mice (50/sex/dose) at doses of 0, 2.5, 10, and
20 mg/kg-day, 5 days/week for 2 years. Adjusted for discontinuous exposure (5 days/7 days),
the intakes were 0, 1.8, 7.1, and 14.3 mg/kg-day. The doses were chosen based on results of the
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14-week assay described in Section 4.2.1.1. Mice were observed twice daily for clinical signs
that were recorded on day 29, every 4 weeks, and at study termination. BWs were recorded
before the start of the exposure period, then every 4 weeks, and at study termination. Five male
and five female mice were evaluated at 2 weeks and 3, 12, and 18 months for urinalysis
parameters. All mice were subjected to gross necropsies and complete histopathologic
examinations.
Noncancer results
Survival was significantly reduced in male and female mice treated with 20 mg/kg-day;
38/50, 42/50, 39/50, and 14/50 males and 39/50, 32/50, 39/50, and 23/50 females survived to
104-105 weeks with increasing dose. Despite a tendency for BWs in high-dose mice to be
slightly lower compared with controls after 30 weeks of treatment, AN had no significant effect
on terminal BWs; no treatment-related clinical signs were observed.
Exposure to AN significantly increased the incidences of nonneoplastic lesions in male
and female mice compared with control mice (Table 4-31). Statistically significant increases in
the incidences of mild focal or multifocal epithelial hyperplasia of the forestomach were
observed in males treated with 10 or 20 mg/kg-day and in focal or multifocal epithelial
hyperplasia of the forestomach in females treated with 20 mg/kg-day. The hyperplastic lesions
were often accompanied by focal hyperkeratosis and occasionally associated with chronic
inflammation. The incidence of hyperkeratosis (diffuse or focal) of the forestomach was
statistically significantly elevated in males treated at 20 mg/kg-day. AN-treated males showed a
higher incidence of hyperplasia of the Harderian gland, but only the increase observed at
10 mg/kg-day was statistically significantly different from the control. No significant increase
was found in treated females (Table 4-31). Statistically significant elevations were observed in
the incidences for ovarian cysts in females treated with 2.5-20 mg/kg-day and for ovarian
atrophy in females treated with 10-20 mg/kg-day; for both ovarian lesions, the highest
incidences were observed at 10 mg/kg-day. A NOAEL was not identified for noncancer effects
in this gavage study. A LOAEL of 1.8 mg/kg-day (adjusted for continuous exposure) was
identified for increased incidence of ovarian cysts in female mice. A NOAEL of 1.8 mg/kg-day
and a LOAEL of 7.1 mg/kg-day were identified for increased incidences of forestomach
hyperplasia in male mice.
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Table 4-31. Incidence and severity of nonneoplastic lesions in B6C3Fi mice
exposed by gavage to AN for 2 years
Adjusted dose (mg/kg-d)a
Ob
1.8
7.1
14.3
Males
Number of male mice examined
Forestomach hyperkeratosis, diffuse/focal
Forestomach epithelial hyperplasia, focal
Harderian gland hyperplasia
50
T (2.5)d
2 (3.0)
1 (2.0)
50
3 (2.0)
4 (2.3)
4 (2.3)
50
7 (1.7)
8e (2.0)
r (3.4)
50
12f(1.8)
9f(1.9)
4 (2.3)
Females
Number of female mice examined
Forestomach hyperkeratosis, diffuse/focal
Forestomach epithelial hyperplasia, focal, or multifocal
Harderian gland hyperplasia
Ovarian atrophy
Ovarian cyst
50
2(1.5)
2(1.5)
5 (3.0)
6 (3.0)
12 (2.3)
50
1 (2.0)
2 (3.0)
4 (3.3)
8 (3.9)
20e(2.3)
50
2 (2.0)
5 (1.8)
6 (2.2)
45f (4.0)
27f(2.1)
50
4 (2.0)
r (i.6)
8 (3.5)
40f (4.0)
19e(2.1)
aDoses administered 5 d/wk.
bVehicle control.
°Number of mice with lesion.
dAverage severity grade of lesions in affected mice: 1 = minimal, 2 = mild, 3 = moderate, 4 = marked.
"Significantly different from vehicle control as calculated by authors (p < 0.05).
Significantly different from vehicle control as calculated by authors (p < 0.01).
Source: NTP(2001).
Cancer results
In mice exposed to AN, treatment-related carcinogenic effects (Table 4-32) were
observed in the forestomach, Harderian gland, and, less consistently, lung and ovary of treated
female mice (NTP, 2001). Histopathologic examination revealed significantly increased
incidences of forestomach squamous cell papillomas and in overall incidence of papillomas or
carcinomas in male and female mice treated with 10 or 20 mg/kg-day. Forestomach squamous
cell carcinomas alone were significantly increased in males at>10 mg/kg-day and in females at
20 mg/kg-day. Incidences of Harderian gland adenoma and adenomas or carcinomas were
significantly elevated in males at >2.5 mg/kg-day and in females at>10 mg/kg-day. The overall
incidence of alveolar/bronchiolar adenomas or carcinomas was significantly elevated in females
treated at 10 mg/kg-day but not at the highest dose. Increase in benign or malignant granulosa
cell tumor of the ovary was observed in females treated with 10 mg/kg-day AN. However, the
increase was not statistically significant. Significant positive trends (p < 0.001) were observed in
both male and female mice for the incidences of forestomach squamous cell papilloma or
carcinoma and Harderian gland adenoma or carcinoma; there was also a significant positive
trend (p = 0.029) for alveolar/bronchiolar adenoma or carcinoma in treated females. An inverse
relationship between tumor latency and dose was observed for forestomach squamous cell
papillomas or carcinomas in females and for Harderian gland adenomas or carcinomas in males.
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Table 4-32. Incidences of selected neoplastic lesions in B6C3Fi mice exposed
by gavage to AN for 2 years
Adjusted Dose (mg/kg-d)a
Control11
1.8
7.1
14.3
Males0
Number of mice examined
Forestomach
Squamous cell papilloma (includes multiple)
Squamous cell carcinoma
Squamous cell papilloma or carcinoma
Harderian gland
Adenoma (includes bilateral)
Carcinoma
Adenoma or carcinoma
50
3
0
3
5
1
6
50
4
0
4
16d
1
16f
50
19d
8d
26e
24d
4
27e
50
25d
9d
32e
27d
3
30e
Females0
Number of mice examined
Forestomach
Squamous cell papilloma (includes multiple)
Squamous cell carcinoma
Squamous cell papilloma or carcinoma
Harderian gland
Adenoma (includes bilateral)
Carcinoma
Adenoma or carcinoma
Lung
Alveolar/bronchiolar adenoma or carcinoma
Ovary
Benign or malignant granulosa cell tumor
50
3
0
3
10
1
11
6
0
50
6
1
7
10
0
10
6
0
50
24d
1
25e
25d
3
26e
14g
4
50
19d
lld
29e
23d
2
25e
9
1
aDoses administered 5 d/wk.
bVehicle control.
°Number of mice with tumor.
dSignificantly different from vehicle control as calculated by authors (p < 0.01).
Significantly different from vehicle control as calculated by authors (p < 0.001).
Significantly different from vehicle control as calculated by authors (p = 0.014).
8Significantly different from vehicle control as calculated by authors (p = 0.039).
Source: NTP(2001).
4.2.1.2.6. Gallagher et al. (1988). In a cancer bioassay, groups of 20 male Sprague-Dawley-
derived CD rats were exposed to 0, 20, 100, or 500 ppm AN in drinking water for 2 years
(Gallagher et al.,1988). As calculated from the reported AN concentration and average daily
drinking water consumption data, the intakes of AN were 0, 1.5, 7.1, and 28 mg/kg-day. Rats
were weighed weekly, and feed intake and water consumption were measured for 1 week each
month. All rats, whether dying prematurely or sacrificed at termination, were subjected to a
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complete gross necropsy. Tissues exhibiting gross lesions and a selection of organs (liver,
stomach, adrenal, kidney, heart, brain, pituitary, and lung) were examined for histopathology.
Significantly reduced survival compared with controls was observed in the high-dose
group after 1 year of exposure but not in the other treatment groups. The last high-dose rat died
shortly before the scheduled termination. BWs were significantly reduced in high-dose rats after
8 months and in mid-dose rats after 16 months; compared with terminal BWs of the control
group, reductions of >50% at 500 ppm and >20% at 100 ppm were estimated from the data
graph. Exposure to AN had no effect on food consumption, but water consumption was reduced
in a dose-related fashion (by 7.5, 11.3, and 30% in the low- to high-dose groups). A NOAEL of
1.5 mg/kg-day and a LOAEL of 7.1 mg/kg-day were identified for reduced BW in rats exposed
to AN for 2 years. This study was limited as to its usefulness for noncancer risk assessment
because few tissues were evaluated for histopathology and noncancer histopathology was not
reported.
Gallagher et al. (1988) reported dose-related increases for the incidences of tumors of the
forestomach and ear canal (Zymbal gland). Four of 20 high-dose rats (20%) had papillomas of
the squamous epithelium of the forestomach, a tumor type not observed in the other groups.
Squamous carcinomas of Zymbal gland were observed in 1/20 (5%) mid-dose rats and
9/20 (45%) high-dose rats but not in low-dose rats or controls. This study provided evidence of
the carcinogenicity of AN to the forestomach and Zymbal gland in male rats. Limitations of the
study in evaluation of AN carcinogenicity included small group sizes, lack of testing in females,
and only limited tissues were evaluated in gross and histopathological examination.
4.2.1.2.7. Signer et al. (1986). In an interim (18-month) report of a 2-year drinking water study
(Bigner et al., 1986), F344 rats were exposed to AN at targeted concentrations of 0 ppm
(51 males and 49 females), 100 ppm (50 rats/sex), and 500 ppm (two subgroups, with a total of
197 males and 203 females). One group of the 500 ppm rats, 147 males and 153 females, was
allocated for studies of morphology, tumor biology, and karyotyping, leaving the remaining rats
for studies of comparative survival and tumor incidences. By using reference average BWs for
F344 rats in a chronic-duration study and an allometric equation deriving drinking water
consumption from BW (U.S. EPA, 1988), the average daily doses of AN were calculated as 0,
12.9, and 64.5 mg/kg-day for males and 0, 14.8, and 74.2 mg/kg-day for females. Rats were
observed twice daily for neurological signs and death and were weighed weekly. Complete
gross necropsies were conducted on all rats. Brains were evaluated for tumors using light and
electron microscopy. The only endpoints related to noncancer effects reported for this study
were mortality, BW, and clinical signs. Incomplete quantitative data were provided for
mortality, which was reported to show dose-response effect in both sexes, occurring earlier in
100 and 500 ppm females than in males. A total of 215 high-dose rats died between month 6 and
18, whereas only "a few" male and female controls were reported to have died by month 18. No
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quantitative results were provided for the magnitude or statistical significance of BW reductions
that affected high-dose male rats by the third week and high-dose females "slightly" later. BW
reductions in the mid-dose group occurred in males after 2 months and females sometime during
the second year. The incidences of neurological clinical signs (paralysis, head tilt, circling, and
seizures) were dose related, affecting a total of 45/400 rats treated at 500 ppm, 4/100 rats at
100 ppm, and 0/100 controls exposed for 12-18 months. The reported data were insufficient to
accurately identify a NOAEL or LOAEL for noncancer effects in this study. A final report of
this study was not located.
In 215 rats in the 500 ppm group that died between months 6 and 18, the types of tumors
frequently found included s.c. papillomas, papillomas of the forestomach, and tumors in Zymbal
gland, but no incidence data were provided for these lesions (Bigner et al., 1986). A statistically
significant increase in tumors of the brain bearing similarity to astrocytomas was observed in rats
exposed to 500 ppm AN, with 49 primary brain tumors observed among the 215 rats dying
between months 6 and 18; incidences in other treatment groups were not reported. The tumors
were observed mostly in the cerebral cortex (about 75%) and also in the brain stem and
cerebellum. When the brain tumors were classified according to size, 10/49 of these tumors
were larger than 5 mm, 28/49 were between 1 and 5 mm in diameter, and 11/49 were detected
only microscopically. Although this study provided support for the carcinogenic effect of AN at
multiple sites in rats, the lack of numerical results rendered it inadequate for the purpose of
quantifying cancer risk.
4.2.1.2.8. Maltoni et al (1988,1977). In a study conducted at the Ramazzini Institute, Maltoni
et al. (1988, 1977) exposed Sprague-Dawley rats (40/sex) to 5 mg/kg AN by gavage in olive oil,
3 days/week for 52 weeks; a control group of 75 rats/sex received olive oil alone on the same
schedule (study designated BT203). After the exposure period, rats were maintained without
further treatment for the rest of their natural lives (the study ending on week 131). Rats were
examined 3 times daily for their general health status and were subjected to a clinical
examination for gross changes every 2 weeks. Rats were weighed every 2 weeks during the first
year and monthly thereafter. All rats were examined by gross necropsy. All tissues with gross
lesions and a limited set of 12 tissues/organs from each rat were examined microscopically.
No statistically significant increases in tumors were observed in treated rats in this study;
however, decreases in tumor latency or increased incidence were observed for some tumors types
identified in other studies of AN. Papillomas and acanthomas of the forestomach were observed
in 1/40 treated males (latency 92 weeks), 4/40 treated females (average latency 97.5 weeks), 0/40
control males, and 1/75 control females (latency 54 weeks). Maltoni et al. (1977) considered the
increase in forestomach tumors to be treatment related. Gliomas (a category that includes
astrocytomas) appeared in 1/40 treated females (latency of 33 weeks), 0/40 treated males, 2/75
control females (average latency 104 weeks), and 1/74 control males (latency 98 weeks).
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Tumors of the mammary gland and Zymbal gland were reported in treated animals, but the
incidences were not elevated over the control. NTP recently released a memorandum (Malarkey
et al., 2010) that discussed differences of opinion between NTP scientists and the Ramazzini
Institute in the diagnoses of certain cancers reported in a methanol study conducted by the
Ramazzini Institute. See Section 5.4.4.3 for additional information on the use of the Maltoni
study in this assessment.
4.2.1.2.9. Friedman andBeliles (2002); Litton Bionetics (1992). In a three-generation
reproductive toxicity study in Sprague-Dawley rats, Friedman and Beliles (2002) and Litton
Bionetics (1992) provided further evidence of the possible carcinogenicity of AN (methods and
reproductive/developmental findings of this study are presented in Section 4.3.2.1.3). In this
study, 15 males and 30 females per dose level (FO parents) were exposed to 0, 100, and 500 ppm
AN in drinking water for 100 days prior to mating for 6 days. As calculated by the study
authors, the concentrations were equivalent to doses of 0, 11, and 37 mg/kg-day for males and 0,
20, and 40 mg/kg-day for females. A subset of FO, Fl, and F2 females underwent two cycles
each of breeding, then were held for a further 20 weeks after weaning of the second litter prior to
termination.
Considering each generation separately, Fib female breeders in the 500 ppm group
showed statistically significant increased incidences of brain and Zymbal gland tumors, while
increased incidences of tumors in FO or F2b female breeders in the 100 or 500 ppm groups
compared with controls were consistent with the Fib responses, but not statistically significant
(Table 4-33). These were relatively small groups, however, with low power to detect responses
as high as 10% statistically significant. Across all of the generations, there was a statistically
significant increasing trend in both tumor types, supporting the conclusion that exposure to AN
for up to 51 weeks at 100 or 500 ppm in drinking water was associated with increased tumor
incidence in female Sprague-Dawley rats.
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Table 4-33. Incidence of tumors in female Sprague-Dawley rats exposed to
AN in drinking water for up to 48 weeks
Generation
FO
Fib
F2b
Total
Tumor incidence in rats
Brain astrocytomas
Zymbal gland
Exposure concentration (ppm)
0
100
500
0
100
500
Dose (mg/kg-d)
0
0/19
0/20
0/20
0/59b
20
1/20
1/19
1/20
3/59
40
2/24
4/17a
1/20
7/6 la
0
0/19
0/20
0/20
0/59b
20
0/20
2/19
0/20
2/59
40
2/24
3/17a
3/20
8/6 la
""Significantly different from controls (p < 0.05), as calculated by the authors.
bStatistically significant by Cochran-Armitage trend test, p < 0.01.
Sources: Friedman and Bellies (2002); Litton Bionetics (1992).
Tables 4-34 and 4-35 summarize the noncancer and cancer findings, respectively, from
chronic oral studies of AN in rats and mice.
Table 4-34. Summary of chronic oral toxicity studies of AN: noncancer
findings in rats and mice
Strain
number/sex
Exposure
route/
duration"
Doses
Effects
NOAEL/
LOAEL3
References
Comments
Rats
F344
100/sex/group
Unexposed
controls =
200/sex
Sprague-
Dawley
100/sex/group
Drinking
water/
2yrs
Drinking
water:
22 months
(M);
19 months
(F)
M: 0,0.1,0.3,
0.8,2.5,
8.4 mg/kg-d;
F: 0,0.1,0.4,
1.3,3.7,
10.9 mg/kg-d
M: 0, 0.09,
8.0 mg/kg-d;
F: 0,0.15,
10.7 mg/kg-d
Squamous cell
hyperplasia/
hyperkeratosis of
the forestomach;
decreased survival
in high-dose
groups, increase in
epidermal inclusion
cysts of the skin
Squamous cell
hyperplasia of the
forestomach,
decreased survival
in high-dose
groups, reduction
in absolute/relative
pituitary weight (F)
NOAEL =
0.1 mg/kg-d;
LOAEL =
0.3 mg/kg-d
(M)
0.4 mg/kg-d
(F)
LOAEL =
0.09 mg/kg-d
(M)
Johannsen and
Levinskas
(2002b);
Biodynamics
(1980c)
Johannsen and
Levinskas
(2002a);
Biodynamics
(1980a)
10 rats/sex/group
were taken from
exposed and
control groups for
interim sacrifice at
6, 12, and 18 mos
10 rats/sex/group
were taken from
exposed and
control groups for
interim sacrifice at
6, 12, and 18 mos
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Table 4-34. Summary of chronic oral toxicity studies of AN: noncancer
findings in rats and mice
Strain
number/sex
Sprague-
Dawley
100/sex/group
Sprague-
Dawley
48/sex/group;
Controls:
80/sex
F344
50/sex low-
dose;
197 males and
203 females
high-dose.
Controls:
51 -males and
49-females:
Exposure
route/
duration"
Gavage/
20 months
Drinking
water/
2yrs
Drinking
water/
lifetime
Doses
0,0.1,
10 mg/kg-d
M: 0, 3.4, 8.5,
21.3 mg/kg-d;
F: 0,4.4,
10.8,
25.0 mg/kg-d
M: 0, 12.9,
64.5 mg/kg-d;
F: 0, 14.8,
74.2 mg/kg-d
Effects
Squamous cell
hyperplasia of the
forestomach, small
reductions in
hematocrit, Hb, and
RBC count in high-
dose males,
increase in
absolute/relative
liver weight in
high-dose groups
Hyperplasia/
hyperkeratosis of
the forestomach,
reduced survival
and BW, minimal
progressive
nephropathy,
gliosis of the brain
Neurological signs
NOAEL/
LOAEL3
NOAEL =
0.1 mg/kg-d;
LOAEL =
10 mg/kg-d
LOAEL =
4.4 mg/kg-d
(F)
No data
References
Johannsen and
Levinskas
(2002a);
Biodynamics
(1980b)
Quast (2002);
Quast et al.
(1980a)
Bigner et al.
(1986)
Comments
10 rats/sex/group
were taken from
exposed and
control groups for
interim sacrifice at
6, 12, and 18 mos
Doses calculated
using default
assumptions (U.S.
EPA, 1988);
noncancer effects
not evaluated in
the study
Mice
B6C3FJ
50/sex/group
Gavage/
2yrs
0, 2.5, 10,
20 mg/kg-d,
5d/wk;
continuous
exposure-
adjusted
doses: 0, 1.8,
7.14,
14.3 mg/kg-d
Reduced survival
in high dose group;
hyperkeratosis/
hyperplasia of the
forestomach;
ovarian cysts and
atrophy (F)
LOAEL:
1.8 mg/kg-d
(F); NOAEL
= 1.8 mg/kg-
d;(M)
LOAEL =
7.1 mg/kg-d
(M)
NTP (2001)
aM = male; F = female.
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Table 4-35. Summary of chronic oral toxicity studies of AN: cancer
findings in rats and mice
Strain
number/sex
Exposure
route/
Duration
Doses"
Effects"
References
Comments
Rats
F344
100/sex/group
Unexposed
controls =
200/sex
Sprague-
Dawley
100/sex/group
Sprague-
Dawley
100/sex/group
Sprague-
Dawley
48/sex/group;
Controls:
80/sex
F344
50/sex low-
dose; 197 M
and 203 F
high-dose.
Controls:
51 M and 49 F
Sprague-
Dawley
40/sex
Control:
75/sex
Sprague-
Dawley
20 M/group
Drinking
water/
2yrs
Drinking
water/
22 months
(M);
19 months (F)
Gavage/
20 months
Drinking
water/
2yrs
Drinking
water/
lifetime
Gavage/
3 d/week for
52 weeks; rats
maintained
without
treatment until
natural death
(study ended
week 131)
Drinking
water/2 yrs
M: 0,0.1,0.3,
0.8, 2.5, and
8.4 mg/kg-d;
F: 0,0.1,0.4,
1.3, 3.7, and
10.9 mg/kg-d
M: 0,0.09, and
8.0 mg/kg-d; F:
0,0.15, and
10.7 mg/kg-d
0,0.1, and
10 mg/kg-d
M: 0,3.4,8.5,
and 2 1.3 mg/kg-
d;F: 0,4.4,
10.8, and
25.0 mg/kg-d
M: 0, 12.9, and
64.5 mg/kg-d; F:
0, 14.8, and
74.2 mg/kg-d
0, 5 mg/kg in
olive oil
0, 1.5, 7.1, and
28 mg/kg-d
Males: increase in brain
astrocytomas and Zymbal
gland tumors, and
forestomach tumors; females:
increase in brain
astrocytomas, Zymbal gland
tumors, forestomach tumors,
and mammary gland tumors
Increases in tumors of the
CNS, Zymbal gland, and
forestomach
Increase in brain
astrocytomas, and tumors of
Zymbal gland, forestomach,
and intestine; mammary
gland carcinomas in females
Increases in CNS tumors,
squamous cell papillomas or
carcinomas of the
forestomach, Zymbal gland
carcinomas, benign and
malignant mammary gland
tumors, tongue tumors, small
intestine tumors
Brain astrocytomas, Zymbal
gland tumors, and papillomas
of the forestomach
Females: increases in
forestomach tumors; not
statistically significant
compared to controls
Increases in tumors of the
forestomach and Zymbal
gland
Johannsen
and
Levinskas
(2002b);
Biodynamics
(1980c)
Johannsen
and
Levinskas
(2002a);
Biodynamics
(1980a)
Johannsen
and
Levinskas
(2002a);
Biodynamics
(1980b)
Quast (2002);
Quast et al.
(1980a)
Bigner et al.
(1986)
Maltoni et al.
(1977)
Gallagher et
al. (1988)
10 rats/sex/group
were taken from
exposed and
control groups for
interim sacrifice at
6, 12, and 18
months
10 rats/sex/group
were taken from
exposed and
control groups for
interim sacrifice at
6, 12, and 18
months
10 rats/sex/group
were taken from
exposed and
control groups for
interim sacrifice at
6, 12, and 18
months
Doses calculated
using default
assumptions (U.S.
EPA, 1988); no
tumor incidence
data were provided
Study limitations:
single dose, short
exposure period
Study limitations:
small group size,
lack of testing in
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Table 4-35. Summary of chronic oral toxicity studies of AN: cancer
findings in rats and mice
Strain
number/sex
Sprague-
Dawley
15 M/group,
30 F/group
Exposure
route/
Duration
Drinking
water/
48 weeks
(3 -generation
reproduction/
developmental
study)
Doses"
0, 11, 37mg/kg-
d (M); 0, 20, and
40 mg/kg-d (F)
Effects"
Increases in brain and
Zymbal gland tumors
References
Friedman and
Beliles
(2002); Litton
Bionetics
(1992)
Comments
females, and only
limited tissues were
examined for
histopathology.
Study suggested
increased
susceptibility to the
carcinogenicity of
AN from early -life
exposure (see
Section 4.8.1).
Mice
B6C3FJ
50/sex/group
Gavage/
2yrs
0, 2.5, 10, and
20 mg/kg-d,
5d/wk,
continuous
exposure-
adjusted doses:
0, 1.8, 7.14, and
14.3 mg/kg-d
Increase in tumors of the
forestomach and of Harderian
gland
NTP (2001)
Overall incidence
of alveolar/
bronchiolar
adenomas or
carcinomas was
significantly
elevated in F at
10 mg/kg-d, but not
at 20 mg/kg-d.
aM = male; F = female.
4.2.2. Inhalation Exposure
4.2.2.1. Subchronic Studies
The only subchronic inhalation study of AN in animals was a comparative study of the
neurotoxicity of nitriles (Gagnaire et al., 1998). Groups of 12 male Sprague-Dawley rats were
exposed (whole body) to AN vapor at concentrations of 25, 50, or 100 ppm, 6 hours/day, 5
days/week for 24 weeks. A control group of 10 male rats was exposed to filtered air. BWs were
measured weekly. Following 4, 8, 12, 16, 20, and 24 weeks of exposure and an 8-week recovery
period (week 32), rats were evaluated for the same electrophysiological parameters that were
tested in parallel experiments on orally exposed rats (see Section 4.2.1.1). As in the companion
study, electrophysiological testing was performed at least 16 hours after daily exposure (waiting
period was 48 hours for weekends). Electrical stimulation of the tail nerve was used to assess
MCVs, SCVs, ASAPs, and AMAPs. No mortality was observed during the treatment period,
but, during the first and second week of the recovery period, 2/12 rats in the 100 ppm group died
and one rat each in the 100 and 25 ppm groups was euthanized in week 31 because of tumors in
the neck. BW gain in the 100 ppm group was significantly lower than in controls in weeks 4, 8,
16, and 21-24, such that the BW was 11% lower at the end of week 24.
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Rats exposed to AN did not develop weakness of the hind limbs or disturbances in gait.
After 1 or 2 weeks of exposure, rats exposed at >50 ppm exhibited clinical signs of gross toxicity
(wet fur and excessive salivation but not hyperactivity). Excessive salivation was attributed by
the study authors to a cholinomimetic effect of AN. Exposure to AN had no effect on
neurophysiological parameters during the first 8 weeks and no effect on the AMAP at any time
during the study. Statistically significant concentration-dependent SCV reductions of-9%
compared with controls were observed in the 100 ppm group from weeks 12-24 (Table 4-36). In
week 12, an -7% reduction was observed in the 50 ppm group, and in week 24 the SCV was
reduced by 5% in the 25 ppm group (not biologically significant) and by >8% in the 50 and
100 ppm groups. The ASAP was significantly reduced by 14.5-20% in the 50 ppm group and by
29-30% in the 100 ppm group from weeks 16-24. After recovery in week 32, a 21% reduction
in ASAP persisted in the 100 ppm group. Sporadic reductions in MCV were observed in
100 ppm rats beginning in week 16 (11% reduction) but were not concentration dependent in
week 24 or after recovery in week 32. A LOAEL of 25 ppm was identified for reductions in
SCV in rats exposed to AN by inhalation for 24 weeks. A NOAEL was not identified.
Table 4-36. Effect on SCV in male Sprague-Dawley rats exposed to AN via
inhalation for 24 weeks
Exposure
(ppm)
0
25
50
100
SCV (m/s)a
Exposure (weeks)
0
35.0 ±0.5
35.2 ±0.4
35.3 ±0.6
35.8 ±0.5
12
49.7 ±0.8
48.2 ±0.7
46.3±0.8C
45.3±1.0C
16
49.9 ± 1.0
47.8 ± 1.0
48.0 ± 1.1
46.2 ± 0.7b
20
50.3 ±0.5
50.2 ±0.7
50.5 ±0.6
48.1±0.7b
24
53.3 ±1.0
50.5±0.8b
49.1±0.5d
48.4±1.0d
Recovery
32
53.4 ±0.6
51.8 ±0.8
51.3 ±1.0
50.4 ±0.8
aValues are means ± SDs (n = 12 for treated, n = 10 for controls).
bStatistically significant compared with controls (p < 0.05) as calculated by the study authors.
Statistically significant compared with controls (p < 0.01) as calculated by the study authors.
dStatistically significant compared with controls (p < 0.001) as calculated by the study authors.
Source: Gagnaire et al. (1998).
4.2.2.2. Chronic Studies
4.2.2.2.1. Dow Chemical (1992a) and Quasi et al. (1980b). Dow Chemical (1992a) and Quast
et al. (1980b) evaluated the effects of AN in Sprague-Dawley rats (100/sex/group) exposed by
inhalation at concentrations of 0, 20, or 80 ppm (0, 43.4, or 173.6 mg/m3) 6 hours/day,
5 days/week for 2 years. Additional groups of 7 and 13 rats/sex/group were exposed and
sacrificed at 6 and 12 months, respectively. BWs were determined 10 times during the first
3 months and monthly thereafter. Rats were observed daily for clinical signs and mortality, and
beginning after 6 months, were examined for palpable masses and dental condition. Moribund
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animals or those with ulcerating tumors were sacrificed and subjected to gross necropsy.
Hematology and urinalysis examinations were conducted on 10 rats/sex/group on days 174/175,
365/366, 616/617, and 727/727 for males/females, respectively. The hematology determinations
of male rats was also conducted on day 183 to verify observations made on day 174. Clinical
chemistry analyses were conducted on 10 rats/sex/group on days 176, 372, and 735. Water
consumption was determined for representative male and female rats in each group for the first
8 months. At terminal sacrifice, all rats received an ophthalmologic examination and necropsy
during which organ weights were recorded for brain, heart, liver, kidneys, and testes. Complete
histologic examinations were carried out on all rats in the control and 80 ppm groups at terminal
sacrifice. More than 80% of rats in the 20 ppm group were examined for gross lesions, and
23 selected organs and all tissues with grossly recognized tumors were collected for
histopathology examination. Because there were signs of upper respiratory tract irritation in the
nasal turbinate, about 10 rats/sex/group from the terminal sacrifice were evaluated by light
microscopy. In addition, because of brain lesions observed during drinking water studies, nine
sections from various regions of the CNS of all rats were examined microscopically.
Noncancer results
Inhalation exposure to AN resulted in significant concentration-related noncancer effects
compared with controls (Quast et al., 1980b). Statistically significant (p < 0.05) decreases in
survival with respect to controls were observed in males after 6 months of exposure at 80 ppm
and in females after 10 months of exposure at 80 ppm or 22 months at 20 ppm. The numbers of
rats surviving at termination (out of 100/sex) were 18, 14, and 4 males and 22, 9, and 1 females
in the control, 20 ppm, and 80 ppm groups, respectively. BWs were decreased by about 10-15%
in male and female rats after 9 months of exposure to 80 ppm AN. A significant decrease of less
than 10% was also observed in 20 ppm female rats. By the end of the study, BWs in 20 and 80
ppm females were not significantly different from those in controls.
Hb and RBC counts were significantly lower (by -9%) in rats exposed to 80 ppm for 4-
8 months but not later. However, the study authors considered these changes to be a secondary
effect of reduced growth, tumor formation, and hemorrhage, resulting from exposure and not due
to bone marrow toxicity. Statistically significant increases in water consumption were observed
in both exposed groups of male and female rats during the first 6 months of the study and were
consistent with slightly decreased urine specific gravity measured in 80 ppm groups during that
period. No significant effects on urinalysis parameters were observed after 6 months. Exposure
to AN had no consistent significant effect on clinical chemistry parameters or results of
ophthalmoscopic examinations. A significant elevation (about 26%) in blood urea nitrogen
(BUN) was observed in the 20 and 80 ppm female rats on day 176. SGPT was also elevated by
57% in the 80 ppm females at that time. However, no significant findings of these parameters
were found upon subsequent evaluation at a later time interval. Hence, Quast et al. (1980b)
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considered these changes as secondary responses and not indications of direct renal or
hepatotoxicity from AN exposure. Significant increases in relative weights of brain, heart, and
testes of male rats exposed at 80 ppm were considered by the study authors to be a consequence
of the reduction in BW.
Gross pathological examinations found statistically significantly findings in the nasal
turbinates, lungs, teeth (malocclusion), and liver of the 80 ppm rats. Gross observation of male
rats indicated significant increase in minimal chronic nephropathy in the 80 ppm group
(40/100 vs. 24/100). Significant increase in pneumonia, atelectasis, or edema was found in the
20 and 80 ppm males (14/100, 27/100, and 30/100 for 0, 20, and 80 ppm groups, respectively).
Gross observations of noncancer changes included enlarged liver in female rats, with incidence
of 2/100, 9/100, and 7/100 in 0, 20, and 80 ppm groups, respectively. (The increased incidence
in the 20 ppm group was statistically significant.)
Statistically significant increases in the incidence of histopathologic lesions of the nasal
turbinates were observed in all rats exposed at 80 ppm and most rats in the 20 ppm group. These
effects were considered by the study authors to be the result of irritant effects of AN
(Table 4-37). These lesions appeared in rats sacrificed after at least 13 months (usually
19 months) of exposure. These inflammatory and degenerative changes included hyperplasia,
flattening, focal erosion, and squamous metaplasia of the respiratory epithelium and hyperplasia
of mucus secreting cells. Flattening of the respiratory epithelium in females and hyperplasia of
mucus-secreting cells in males were both significantly increased at the 20 ppm exposure level.
Table 4-37. Incidence of histopathological lesions of the nasal turbinates in
Sprague-Dawley rats exposed to AN via inhalation for 2 years
Response
Concentration (ppm)
0
20
80
Incidence
Males
Suppurative rhinitis
Hyperplasia of respiratory epithelium
Focal erosion of mucous lining
Squamous metaplasia of the respiratory epithelium
Hyperplasia of mucus-secreting cells
Focal inflammation
Flattening of the respiratory epithelium
0/11
0/11
0/11
0/11
0/11
0/11
0/11
1/12
4/12
0/12
1/12
7/12a
1/12
2/12
5/10a
10/103
4/10a
7/10a
8/10a
1/10
3/10
Females
Suppurative rhinitis
Hyperplasia of respiratory epithelium
Focal erosion of mucous lining
Squamous metaplasia of the respiratory epithelium
Hyperplasia of mucus-secreting cells
1/11
0/11
0/11
0/11
0/11
0/10
2/10
1/10
2/10
2/10
2/10
5/10a
1/10
5/10a
8/10a
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Table 4-37. Incidence of histopathological lesions of the nasal turbinates in
Sprague-Dawley rats exposed to AN via inhalation for 2 years
Response
Focal inflammation
Flattening of the respiratory epithelium
Concentration (ppm)
0
20
80
Incidence
2/11
1/11
6/10
7/10a
7/10a
8/10a
""Statistically significant (p < 0.05), as calculated by the study authors.
Source: Quastetal. (1980b).
Increase in acute suppurative pneumonia was observed in the lungs of the 80 ppm male
rats during the 7-12-month time interval. A nonsignificant increase was also observed in the
20 ppm group.
Other histopathological observations included an increase in the incidence of gliosis and
perivascular cuffing in the brain of high-dose rats (either sex) and, in males only, minimal
chronic focal progressive nephrosis and formation of keratinized cysts in the thyroid gland
(Table 4-38). Incidences of focal necrosis of the liver were increased in 20 and 80 ppm female
rats. The incidence of AN-related hyperplasia and hyperkeratosis of the nonglandular portion of
the stomach did not achieve statistical significance in either sex but was statistically significant
(p < 0.05) by Fisher's exact test when the data were combined. There were concentration-related
increases in the incidences of several lesions that were secondary to other effects of AN exposure
(numerical data not provided here). These included hepatocellular atrophy without fatty changes
and atrophy of mediastinal fat in 80 ppm male rats, attributed by the study authors to the
decreased feed intake of the rats at the end of the study, and extramedullary hematopoiesis of the
spleen in females, a consequence of the reductions in RBC and Hb counts. An increase in
lymphoid hyperplasia in males at 80 ppm was interpreted by the study authors to be secondary to
tumors of the ear canal and inflammatory changes in the nasal turbinates. A NOAEL was not
identified in this study. A LOAEL of 20 ppm was identified for increased lesions of the nasal
turbinates (hyperplasia of mucus-secreting cells in males and flattening of the respiratory
epithelium in females) and focal necrosis in liver of female rats exposed to AN vapor for 2 years.
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Table 4-38. Incidence of dose-related noncancerous histopathological lesions
in Sprague-Dawley rats exposed to AN via inhalation for 2 years
Response
Concentration (ppm)
0
20
80
Incidence
Males
Focal nephrosis (progressive)
Minimal chronic nephropathy
Thyroid cyst
Gliosis and perivascular cuffing (brain)
Acute suppurative pneumonia
Pulmonary changes
Hyperplasia of the nonglandular epithelium (stomach)
22/100
24/100
4/95
1/100
0/100
15/100
8/98
24/100
21/100
9/97
2/99
4/100
26/1003
7/100
48/1003
40/1003
13/963
7/99a
10/1003
20/100
16/99
Females
Gliosis and perivascular cuffing (brain)
Focal necrosis in liver
Vasculization of spinal myelin (minimal)
Hyperplasia of the nonglandular epithelium (stomach)
0/100
3/100
42/100
2/99
2/100
16/1003
67/1003
3/99
8/1003
10/1003
47/100
7/97
""Statistical significance (p < 0.05), as calculated by the study authors.
Source: Quastetal. (1980b).
Cancer results
In rats chronically exposed to AN vapor, there were significant increases in tumors at
multiple sites, several of which also had been affected in oral exposure bioassays (Quast et al.,
1980b). In males and females exposed to 80 ppm AN, there were increased incidences of
astrocytomas of the brain as well as glial cell proliferation that was considered an earlier stage in
the progression to astrocytomas. The incidence data in Table 4-39 combine the incidences of
astrocytomas and glial cell proliferation. The incidence of CNS tumors was also significantly
increased in females exposed to 20 ppm AN and insignificantly increased in males exposed to 20
ppm AN. Carcinomas of Zymbal gland were significantly elevated in the 80 ppm rats of both
sexes. Other tumor increases observed in the 80 ppm groups were squamous cell papillomas or
carcinomas in the tongue and carcinomas of the intestinal tract of males and adenocarcinomas of
the mammary gland in females. An increase was observed in forestomach tumors in 80 ppm
male rats and nasal turbinate tumors in 80 ppm female rats. Both types of tumors were
considered by the study authors to be treatment related, although the increase was not
statistically significant. Table 4-39 presents incidences of AN-induced target organ-specific
tumor formation. All tumors were found in exposed rats that died or were sacrificed after
12 months of exposure, with the exception of one male rat with a CNS tumor and three female
rats with mammary gland adenocarcinomas that died between 7 and 12 months.
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Table 4-39. Cumulative incidence of tumors in Sprague-Dawley rats
exposed to AN via inhalation for up to 2 years
Tissue
All brain/CNSa
Zymbal gland
Intestinal tract
Mammary gland (adenocarcinomas)
Mammary gland (total, benign and malignant)
Forestomach (squamous cell papilloma)
Tongue
Nasal turbinate (carcinoma in respiratory
epithelial region)
Males (ppm)
0
0/96
2/96
4/96
0/100
4/100
1/98
1/95
20
4/93
4/93
3/93
0/100
5/100
1/100
0/14
80
22/82b
ll/82b
17/82b
1/100
7/100
4/99
7/82b
Not increased
Females (ppm)
0
0/93
0/93
20
8/99b
1/98
80
20/89b
ll/89b
Not increased
9/93
88/100
0/99
0/96
0/11
8/98
95/100
0/99
0/9
0/98
20/99b
85/100
1/97
1/91
2/10
"Incidence data include all brain/CNS tumors (astrocytomas and glial cell proliferation). Male tumor incidence data
are from Tables 22, 25, and 26 of the Quast et al. (1980b) report; female tumor incidence data are from Tables 31,
34, and 35 in the same report. For all incidence data, the denominators excluded rats dying earlier than 12 mos in
the study. These data were ascertained from Tables 22, 25, 31, 34, and 35 in the original study report by Quast et
al. (1980b).
bSignificantly different from controls (p < 0.05) as calculated by the study authors.
Sources: Dow Chemical (1992a); Quast et al. (1980b).
An apparent decrease in the incidence of tumors of the pituitary, adrenals, thyroid, and
pancreas of male and female treated rats, and testes of males, was observed when compared with
controls.
4.2.2.2.2. Maltoni et al (1988,1977). In studies conducted at the Ramazzini Institute, Maltoni
et al. (1988, 1977) reported the results of three cancer bioassays in Sprague-Dawley rats exposed
to AN by inhalation. In the first (designated BT201 by the authors), 30 rats/sex/group were
exposed to 0, 5, 10, 20, and 40 ppm AN for 4 hours/day, 5 days/week for 52 weeks; the animals
then were allowed to complete their natural life spans, with the final deaths occurring in week
136 (Maltoni et al., 1977). Rats were examined 3 times weekly for general health status and
subjected to a clinical examination for gross changes every 2 weeks. Rats were weighed every
2 weeks during the exposure period and monthly thereafter. All rats were subjected to gross
necropsy. Histopathologic examinations were conducted on all gross lesions and a selection of
about 12 organs and tissues, including the Zymbal glands, interscapular brown fat, salivary
glands, tongue, lungs, liver, kidneys, spleen, stomach, intestine, bladder, and brain. Only the
incidence of neoplastic lesions was reported.
Exposure to AN had no significant effect on survival or BWs in male or female rats.
Increased incidences of gliomas, forestomach tumors, Zymbal gland carcinomas, and mammary
tumors were reported in the treated group; however, increases in the incidences of these tumors
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in treated rats were not statistically significant. Gliomas were found in 1/30 and 2/30 males at 20
and 40 ppm, respectively, but not in controls and other exposed groups of male and female rats.
The average latency of the gliomas was shorter at the higher concentration (84 weeks in 20-ppm
males vs. 63.5 weeks in 40-ppm males). Forestomach papillomas and acanthomas were found in
0/30, 1/30, 2/30, 0/30 and 3/30 male rats at 0, 5, 10, 20, and 40 ppm, respectively. In females,
the incidence of forestomach tumors was 0/30, 1/30, 2/30, 1/30, and 0/30 at 0, 5, 10, 20, and 40
ppm, respectively. The average latency of forestomach tumors ranged from 103 to 124 weeks.
Zymbal gland carcinomas were found in 1/30 males in the 10-ppm group and 1/30 females in the
20-ppm group. No Zymbal gland carcinomas were found in the control and other dose groups.
The incidence of benign and malignant mammary gland tumors in treated females rats was
increased over controls, but the increase was not dose related (5/30 [controls], 10/30 [5 ppm],
7/30 [10 ppm], 10/30 [20 ppm], and 7/30 [40 ppm]) (Maltoni et al., 1977).3
In two additional cancer bioassays, Maltoni et al. (1988) exposed Sprague-Dawley rats to
AN by inhalation beginning in gestation. In the first bioassay (designated BT4003), 54 adult
pregnant females, beginning on gestation day 12, were exposed to 60 ppm AN for 4 hours/day,
5 days/week for 7 weeks and then 7 hours/day, 5 days/week for 97 weeks. A group of 60
unexposed adult females served as controls. Gestation was permitted to proceed normally and
the offspring were exposed on the same schedule as the dams. The exposed offspring included
67 males and 54 females; the controls included 158 males and 149 females.
Overall, there was a statistically significant treatment-related increase in the percentage
of dams with malignant tumors at all sites (37 vs. 15%). Increased incidences in exposed dams
compared with controls were observed for several sites (Zymbal gland carcinomas, mammary
gland carcinomas, malignant mammary gland tumors, extrahepatic angiosarcomas, and
encephalic gliomas), but none of these was statistically significantly different from controls. No
hepatomas were observed in exposed or unexposed dams. These tumor results are summarized
in Table 4-40.
3 Discrepancies were noted in the incidence of mammary gland tumors for controls and 40-ppm females as reported
in a subsequent report of this study (Maltoni et al., 1988). In the 1988 publication, the incidence of mammary gland
tumors was reported as 20% (6/30) in control females and 26.7% (8/30) in 40-ppm females. Also in the 1988
publication, encephalic gliomas were reported in both male and female rats at 20- and 40-ppm, whereas the 1977
publication reported gliomas in male rats only; glioma incidences reported in Maltoni et al. (1988) were: 20-ppm
females—3.3% (1/30); 40-ppm females—3.3% (1/30); 20-ppm males—3.3% (1/30); 40-ppm males—6.7% (2/30)..
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Table 4-40. Comparison of carcinogenic effects of chronic exposure to AN at
60 ppm starting either in utero or in adulthood, in Sprague-Dawley rats
Stage
during
exposure
Adult only
Starting at
GD 12
Exposure
protocol
Chronic0
Unexposed
controls
Chronic
Subchronice
Unexposed
controls
Sexa
F
F
M
F
M+F
M
F
M+F
M
F
M+F
Number
of rats at
startb
54
60
67
54
121
60
60
120
158
149
307
Percent with tumor
Brain
tumors
(encephalic
gliomas)
5.5
0.0
16.4
18.5
17.3
5.0
3.3
4.2
1.3
1.3
1.3
Zymbal
gland
carcinomas
5.5
1.7
14.9C
1.8
9.1
6.7
1.7
4.2
1.3
0.0
0.7
Hepatomas
0.0
0.0
7.5d
1.8
4.9
1.7
0.0
0.8
0.6
0.0
0.3
Malignant
mammary
tumors
5.5
3.3
0.0
16.7d
7.4
0.0
6.7
3.3
1.9
5.4
3.6
Extra-
hepatic
angio-
sarcomas
1.8
0.0
4.4
5.5d
4.9
5.0
1.7
3.3
0.6
0.0
0.3
aF = female; M = male.
bAnimals were allowed to live until spontaneous death.
°Chronic: 60 ppm AN for 4 hrs/d, 5 d/wk for 7 wks (starting during gestation), followed by 7 hrs/d for 97 wks.
dStatistically significantly higher than corresponding control incidence, p < 0.05.
eSubchronic: 60 ppm AN for 4 hrs/d, 5 d/wk for 7 wks (starting during gestation), followed by 7 hrs/d for 8 wks.
Source: Maltonietal. (1988).
In contrast, chronically exposed male and female offspring showed statistically
significant increases in the incidences of malignant tumors of the mammary gland in females,
extrahepatic angiosarcomas in females, hepatomas in males, Zymbal gland carcinomas in males,
and encephalic gliomas (see Table 4-40).
In the second bioassay (designated BT4006), Sprague-Dawley rats were initially exposed
to AN under the same exposure conditions as bioassay BT4003; however, exposure of the
offspring (127 males and 114 females) ended after 15 weeks (Maltoni et al., 1988). This group
of offspring was exposed for 4 hours/day, 5 days/week for 7 weeks starting on GD 12, followed
by exposure for 7 hours/day, 5 days/week for 8 weeks. All animals were kept under observation
until spontaneous death, at which time they were examined for the presence of tumors. The
control group of offspring was the one used in experiment BT4003 (158 males and 149 females).
There was a statistically significant increase in the total incidence of malignant tumors in
exposed offspring compared with controls for both males (31.7 vs. 17.1%, p < 0.05) and females
(35.0 vs. 17.4%, p < 0.01). Increased incidences of the following tumors were observed when
compared with controls, although the incidences were not statistically significantly different:
Zymbal gland tumors in males and females combined; extrahepatic angiosarcomas in males;
encephalic gliomas in males and females combined; and hepatomas in males (see Table 4-40).
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NTP recently released a memorandum (Malarkey et al., 2010) that discussed differences
of opinion between NTP scientists and the Ramazzini Institute in the diagnoses of certain cancers
reported in a methanol study conducted by the Ramazzini Institute. See Section 5.4.4.3 for
additional information on the use of the Maltoni et al. studies in this assessment.
4.3. REPRODUCTIVE/DEVELOPMENTAL STUDIES—ORAL AND INHALATION
4.3.1. Studies in Humans
Four studies of reproductive history of women occupationally exposed to AN in chemical
factories in China were identified; two of these also included data pertaining to birth outcomes
among partners of male workers in these factories. Two other studies examined effects on
testosterone and sperm parameters among male workers, and one other case-control study
examined congenital abnormalities in relation to proximity to a factory that used AN. These
studies are summarized in Table 4-41, and briefly described below.
Table 4-41. Epidemiology studies of reproductive and developmental
outcomes among cohorts of workers exposed to AN
Reference
Wu et al.
(1995)
Dong et al.
(2000b);
(update of
Dong and Pan
[1995])
Li (2000)
Czeizel et al.
(2000, 1999)
Xu et al. (2003)
Ivanescu et al.
(1990)
Study population
477 female AN workers;
527 workers
548 male and 391 female
workers in a Chinese
chemical fiber plant with
averages of 1 1.0 and
10.4 yrs of employment;
496 male and 427 female
unexposed controls
379 female AN
manufacturing workers
employed an average of 14
yrs; 511 unexposed controls
Case control study of babies
born with congenital
abnormalities in the vicinity
of an AN-using factory
30 AN-exposed workers,
age range = 25-30 yrs, with
exposure time of 2.8 yrs; 30
unexposed controls, age
range = 24-35 yrs
39 subjects (May 1975),
109 subjects (March 1976),
149 subjects (May 1977)
Exposure assessment
No data
Workplace air
concentrations = 0.05-
7.1 ppm; midpoint =
3.6 ppm
Average workplace air
concentration =
7.5 ppm; range = 0-
70 ppm
No data
Average workplace air
concentration =0.37
ppm
No data
Toxic effects/outcome
Increased vomiting, anemia, preterm delivery,
and birth defects in pregnant females
Statistically significantly increased prevalence
of adverse reproductive outcomes (increased
stillbirths, birth defects, premature deliveries,
and sterility) in female workers compared
with controls; similar results with respect to
stillbirth and birth defects seen for wives of
male workers; see Table 4-23
Statistically significant increased prevalence
of adverse reproductive outcomes (sterility
[2.6 vs. 0.8%], pregnancy complications [20.8
vs. 7.1%], premature deliveries [11.6 vs.
4.7%], and congenital defects [25.4 vs. 4.2%])
in female exposed workers compared with
controls
Incidence of undescended testis possibly
related to proximity to the AN-using facility
Statistically significant decrease in sperm
density (75 x 106 ml'1 vs. 140 x 106 ml'1);
statistically significantly higher comet sperm
nuclei (28.7 vs. 15%); increase in sex
chromosome disomy (0.69 vs. 0.35%)
Reduced serum testosterone
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Wu et al. (1995) reported statistically significant increases in the incidence of pernicious
vomiting, anemia, preterm delivery, and birth defects in 477 females exposed to AN in the
workplace compared with 527 controls; increased risk of birth defects and preterm birth was seen
in logistic regression analyses of these data. Only limited information is currently available
concerning details of this study, however, because it was published in a Chinese journal with
only the abstract translated to English. EPA did not obtain a translation of the full article
because more recent papers with similar data were available.
Dong et al. (2000a,b) examined reproductive outcomes in 548 male and 391 female
workers at a chemical fiber plant in China. (This is an expanded follow-up to an earlier study
conducted in this plant, described in Dong and Pan, 19954). The mean concentration of AN in
the air in work areas of the plant was over 2 mg/m3 in most areas of the plant, and ranged from
O.llto 15.5 mg/m3 (0.05-7.1 ppm; midpoint = 7.8 mg/m3 [3.6ppm]). Controls in the study
were 496 male and 427 female workers who had not been exposed to AN. Average employment
durations for the exposed males and females were 11.0 ± 4.5 and 10.4 ±3.8 years, respectively.
The average ages of the male and female workers were 33.8 ± 4.6 and 32.4 ±3.9 years,
respectively. The age, length of service, and lifestyle of the controls were similar to the exposed
workers. Data were collected using a structured interview with questions pertaining to medical
history, smoking and alcohol use, occupational history, and menstrual and reproductive history.
This information was also obtained for the female spouses of the exposed workers and controls.
There were 614 pregnancies and 574 live births among the 548 wives of exposed men,
and 510 pregnancies and 494 live births among the 496 wives of the control men (Dong et al.,
2000b). The rate of stillbirths among all pregnancies (n = 13, 2.1% among exposed wives and n
= 3, 0.59% in control wives) and the rate of birth defects among live births (n = 9, 15.7 per 1000
among exposed wives and n = 3, 6.1 per 1000 among control wives) were more than doubled in
the exposed group. Similar or even stronger associations were seen in the analysis of
reproductive outcomes among female workers and their controls. There were 413 pregnancies
and 375 live births among the 391 female exposed workers, and 439 pregnancies and 416 live
births among the 427 female controls. A more than twofold increase was seen in the rate of
several outcomes. The rate of premature births was 8.2% (n = 34) in exposed and 3.9% (n=17)
in controls and the rate of stillbirth was 2.7% (n = 11) in exposed and 1.1% (n = 5) in controls.
Among live births, the rate of birth defects was 21.3 per 1000 (n = 8) in exposed and 4.8 per
1000 (n = 2) among controls, and the newborn mortality rate was 10.7 per 1000 (n = 4) among
exposed and 4.8 per 1000 (n = 2) among controls. All of these differences, except the mortality
rates, were statistically significant (p < 0.05). Based on these results, EPA identified the
midpoint of the range of workplace air concentrations, 3.6 ppm, as a LOAEL for statistically
significantly increased prevalence of adverse reproductive outcomes (increased stillbirths, birth
1 The publication date is variously reported in the available translations as 1993 or 1995.
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defects, and premature deliveries) in female exposed workers employed for an average of 11
years.
Li (2000) used a similar study design to examine reproductive outcomes in 379 female
workers in a Chinese AN manufacturing company and in 511 unexposed control workers from a
bed sheet factory and a biological research institute in 1991. The average age and duration of
employment of the exposed workers were 33.95 years (22.75-54.83 years) and 14.10 years
(3.25-34.45 years), respectively. Average age and employment duration of the control workers
were reported to be similar to the exposed group. Monthly workplace air concentration data for
1989 and 1990 were provided by the factory. The average AN air concentration was reported as
16.35 mg/m3 (7.5 ppm, range = 0-152.88 mg/m3 [0-70 ppm]). Statistically significant increased
prevalences of the following reproductive outcomes were found, compared with controls:
sterility (2.64 vs. 0.78%), pregnancy complications (20.8% vs. 7.14%), premature deliveries
(11.62 vs. 4.72%), and congenital defects (25.4 vs. 4.2%). Li (2000) also divided the exposed
group into female workers with and without exposed male partners. Prevalence rates for several
adverse reproductive outcomes (pregnancy complications, premature delivery, late delivery,
stillbirths, and congenital deficits) were statistically significantly elevated in females with
exposed partners, compared with those with nonexposed partners. EPA considered the average
workplace AN air concentration, 7.5 ppm, as a LOAEL for increased prevalence of adverse
reproductive outcomes in female AN manufacturing workers employed for an average of 14
years.
The epidemiological report and follow-up by Czeizel et al. (2000, 1999) examined the
environmental distribution of congenital abnormalities in 30 settlements within a 25 km radius of
an AN factory, drawing on data from the Hungarian Congenital Abnormality Registry covering
46,326 infants born between 1980 and 1996. A number of time-space-specific clusters of
abnormalities were identified among the subjects, the most striking of which was the incidence
of pectus excavatum in the community of Tata between 1990 and 1992. This effect was
associated with an OR of 78.5 (95% CI 8.4-729.6). Other clusters of congenital abnormalities in
the vicinity of the factory were undescended testis in the community of Nyergesujfalu between
1980 and 1983 (OR = 8.6, CI= 1.4-54.3) and at Esztergom between 1981 and 1982 (OR = 4.2,
CI = 1.3-13.5) and clubfoot in the Tata community between 1980 and 1981 (OR = 5.5, CI = 1.5-
20.3). Exposure data, and specifically exposure variation by area and time, were not included in
the analysis. The study authors stated that there was a technological change at the AN factory in
1984 that resulted in greater environmental protection, implying that releases of AN to the
environment were less after 1984 than before the change. This suggested that the cluster of
pectus excavatum obtained at Tata between 1990 and 1992 was unlikely to have been due to AN
exposure. However, the high incidence of undescended testis in Nyergesujfalu between 1980
and 1983 may have been an environmental phenomenon, because the region-wide incidence of
this congenital abnormality appeared to decrease with increasing distance from the factory. In
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general, however, it was difficult to draw conclusions about a link between maternal exposure to
AN and the incidence of congenital abnormalities from the data in this study because of a lack of
exposure data.
Xu et al. (2003) performed conventional sperm analysis according to WHO guidelines
and investigated DNA strand breakage and sex chromosome aneuploidy in spermatozoa of
30 AN-exposed workers compared with 30 unexposed controls (recruited from the general
population). The age of exposed workers ranged from 25 to 30 years and the age of the controls
ranged from 24 to 35 years. All of the subjects were non-smokers and non-regular drinkers, with
no chronic disease or exposure to chemotherapy or radiotherapy. The mean concentration of AN
at exposure sites was reported to be 0.8 ± 0.25 mg/m3 (0.37 ppm); the mean duration of exposure
was 2.8 years. Sperm density was significantly lower in the exposed group (75 x 106/mL) than
in the control group (140 x 106/mL). Sperm number per ejaculum was 205 x 106 in the exposed
group, significantly lower than the 280 x 106 spermatozoa in the control. There were no
significant differences between the groups in semen volume, sperm motility, viability, or
morphology. Xu et al. (2003) used single cell gel electrophoresis (comet assay) to monitor the
incidence of DNA strand breakage of sperm cells. The rate of comet sperm nuclei was 28.7% in
the exposed group, significantly higher than in the control group (15.0%). Mean comet tail
length was 9.8 um in exposed workers but 4.3 um in control workers. The frequency of sex
chromosome aneuploidy in sperm cells was analyzed using fluorescence in situ hybridization
(FISH). Sex chromosome disomy was found to be 0.69% in the exposed group, significantly
higher than 0.35% in controls. XY-bearing sperm was the most common sex chromosome
disomy, with an average rate of 0.37% in exposed vs. 0.20% in controls. XX- and YY-bearing
sperm accounted for an additional 0.09 and 0.23% of sperm in exposed vs. 0.05 and 0.10% in
controls, respectively. Xu et al. (2003) concluded that AN exposure affected semen quality
among occupationally exposed persons by the induction of DNA strand breakage and sex
chromosome nondisjunction.
Ivanescu et al. (1990) used a radioimmunoassay to measure serum levels of testosterone
in three groups of male workers exposed to AN in a chemical factory. Blood samples were taken
from 39 subjects in May 1975, from 109 subjects in March 1976, and from 149 subjects in May
1977. Subjects were between 19 and 40 years old and had been employed at the facility from
6 months to 10 years. Controls in the study consisted of 145 unexposed men ages 17 to 49 years.
There were five groups of controls (37 blood donors, 23 new workers, 84 workers exposed to
other chemicals at 3 other plants in the region, including 23 workers using natrium cyanid, 22
workers using cyan derivatives, and 39 workers using Pyrolisis technology). The three groups of
exposed subjects had average serum testosterone concentrations ranging from 3.5 to 4.1 ng/mL.
This compared with average values ranging from 5.4 to 7.3 ng/mL in different subsets of the
145 control subjects. Although the time of blood sampling during the day was variable in this
study and the circadian rhythm of testosterone was unknown, testosterone concentrations in sera
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of exposed groups were lower than in control groups of the same month. However, no data were
presented in the report on the level of exposure to AN or other chemicals.
4.3.2. Studies in Animals
4.3.2.1. Oral Studies
Assessments of reproductive/developmental effects of AN in orally exposed animals are
derived from standard toxicity assays in rats and mice, standard developmental toxicity assays in
female rats, and a three-generation reproductive toxicity assay in male and female rats.
4.3.2.1.1. Standard toxicity assays: reproductive organ pathology. As described in
Section 4.2.1, standard 2-year oral toxicity assays in rats revealed no evidence for increased
reproductive histopathology in males or females exposed to AN at doses as high as 8-25 mg/kg-
day (Johannsen and Levinskas, 2002a, b; Quast, 2002; Biodynamics 1980a, b, c; Quast et al.,
1980a). In addition, there was no evidence for adverse effects of AN on functional reproductive
parameters (sperm morphology, estrous cycle) or the histology of reproductive organs in male
B6C3Fi mice treated with AN by gavage 5 days/week at 20 mg/kg-day or in females at
40 mg/kg-day for 14 weeks (NTP, 2001). In this subchronic study, the weights of the left cauda
epididymides were significantly elevated compared with controls in male mice exposed to
10 and 20 mg/kg-day, but this effect was not considered biologically significant in the absence of
histopathology. In the companion 2-year gavage assay (NTP, 2001), no reproductive
histopathology was observed in male mice treated 5 days/week with AN at doses as high as
20 mg/kg-day, but effects were observed in females (Table 4-33). The incidence of ovarian cysts
was significantly elevated at 2.5, 10, and 20 mg/kg-day, and the incidence of atrophy of the
ovary increased at 10 and 20 mg/kg-day. Atrophy was severe and was characterized by lack of
histologically evident follicle and corpus luteum development with a predominance in interstitial
tissue. The lowest dose of 2.5 mg/kg-day in this study was identified as the LOAEL for ovarian
cysts and atrophy.
4.3.2.1.2. Developmental toxicity assays. Dow Chemical (1976b) evaluated developmental
effects in pregnant female Sprague-Dawley rats (29-39/group) that received 10, 25, or 65 mg/kg
AN by gavage in water on GDs 6-15; an additional group of 43 controls received water alone.
Dams were observed daily for clinical signs and weighed on GDs 6, 10, 16, and 21. Food and
water consumption were monitored at 3-day intervals on GDs 6-21, at which point all animals
were sacrificed and maternal liver weights were recorded. Among the reproductive parameters
evaluated were the numbers and positions of live, dead, and resorbed fetuses and the number of
implantation sites. Developmental toxicity was evaluated by the weight, sex ratio, and crown-
rump length of the fetuses. One-third of the fetuses in each litter were evaluated for visceral
malformations and soft tissue abnormalities; the rest were examined for skeletal alterations.
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Maternal toxicity of AN was most evident in the high-dose group. Systemic effects
observed only at this dose included a single maternal death on GD 6, an increase in clinical signs
(hyperexcitability and excessive salivation) during the dosing period, BW gain reduced 46%
compared with controls by GD 15 (22% by GD 21), water consumption significantly increased
by an unspecified amount on GDs 6-20, and a statistically significant 11% increase in absolute
(but not relative) liver weight compared with controls. Food consumption was significantly
reduced by an unspecified amount compared with controls in high- and mid-dose dams on
GDs 6-8. At necropsy, most (number not specified) dams dosed with 65 mg/kg-day and
3/33 dams dosed with 25 mg/kg-day displayed a thickening of the nonglandular portion of the
stomach. Pregnancy rate was significantly decreased among dams given 65 mg/kg-day AN, with
only 20 of 29 dams producing litters. Uterine staining revealed implantation sites in four
additional dams. A NOAEL for maternal toxicity was identified as 10 mg/kg-day AN, and
25 mg/kg-day was the LOAEL for hyperplasia of forestomach.
Exposure to AN had no statistically significant effect on the average numbers of
implantations/dam, live fetuses/litter, or resorptions/litter; the average sex ratio of litters was not
reported. At 65 mg/kg-day, there were statistically significant reductions in fetal BW (by 7.4%)
and crown-rump length (by 1.8%) compared with controls. The high-dose group also showed a
significant increase in external malformations (short tail in 6/17 litters and short trunk in
3/17 litters); there was also a significant increase in skeletal malformations (missing vertebrae)
(Table 4-42). The defect ranged in severity from the absence of a single lumbar vertebra to the
absence of all sacral and lumbar vertebrae and most thoracic vertebrae. Although the incidence
of malformations at 25 mg/kg-day was not statistically significant (Table 4-42), a LOAEL of 25
mg/kg-day was identified for fetal malformation in Sprague-Dawley rats. The NOAEL for fetal
toxicity was 10 mg/kg-day.
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Table 4-42. Incidence of fetal abnormalities among litters of Sprague-
Dawley rats following maternal exposure to AN on GDs 6-15
Type of malformation
External and skeletal malformations
Visceral malformations
External malformations
Short tail
Short trunk
Imperforate anus
Visceral abnormalities
Right side aortic arch
Missing kidney, unilateral
Anteriorly displaced ovaries
Skeletal malformations
Missing vertebrae
Missing two vertebrae and two ribs
Hemivertebrae
Total malformed
AN (mg/kg-d)
0
10
25
65
Number of fetuses affected/number of litters examined
443/38
154/38
388/35
135/35
312/29
111/29
212/17
71/17
Number of fetuses (litters) affected
1(1)
0(0)
0(0)
0(0)
1(1)
0(0)
1(1)
7(1)
0(0)
8(2)
0(0)
0(0)
0(0)
0(0)
0(0)
0(0)
0(0)
0(0)
0(0)
0(0)
2(2)
0(0)
0(0)
1(1)
0(0)
1(1)
2(2)
7(2)
0(0)
10(4)
8(6)a
3(3)a
2(2)
1(1)
1(1)
1(1)
8(6)a
0(0)
0(0)
8(6)a
""Significantly different from controls (p < 0.05), as determined by the study authors.
Source: Murray et al. (1978).
Behavioral teratogenicity of AN was examined in the progeny of pregnant Wistar rats
(15/group) that received 0 or 5 mg/kg-day AN by gavage on GDs 5-21 (Mehrotra et al., 1988).
Dams were weighed at intervals, and food and water intakes and the length of gestation were
recorded. At parturition (postnatal day [PND] 0), litters were culled to four males and four
females. On PND 1, pups were sexed and examined for gross external anomalies; litters with
fewer than two/sex were rejected. Eight pups from four litters were examined for behavioral
abnormalities, including tests for spontaneous locomotion and passive avoidance. Excised brains
of 21-day-old pups were assayed for biogenic amines (noradrenaline, dopamine, and
5-hydroxytryptamine) and for the activities of Na+,K+-adenosine triphosphatase (ATPase),
monoamine oxidase, and acetylcholinesterase.
Exposure to AN at 5 mg/kg-day had no significant effect on BW or food and water
intakes in dams or on reproductive parameters such as gestational length, number of viable
offspring, or pup sex ratio. No AN-induced effects were observed on postnatal morphological
development, pup BWs, developmental indices (eye opening or incisor eruption), or tests of
neurological impairment (righting reflex, cliff avoidance, and grip strength). The levels of
biogenic amines were significantly altered in specific brain regions as a result of exposure.
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Levels of 5-hydroxytryptamine were increased by 32% in the pons medulla and decreased by
44% in the corpus striatum and 30% in the hippocampus. Levels of noradrenaline were
decreased by 40% in the pons medulla and increased by 81% in the hippocampus. Gestational
exposure to AN resulted in a statistically significant 49% reduction in brain levels of monoamine
oxidase in 21-day-old pups; levels of acetylcholinesterase and Na+,K+-ATPase were not
significantly affected. The biological significance of the enzyme and neurotransmitter changes
was unclear, given that no treatment-related behavioral effects were observed. However,
Mehrotra et al. (1988) noted that alterations in the levels of biogenic amines may become more
prominent after prolonged exposure or exposure to higher doses of AN.
In a study comparing developmental toxicities of aliphatic nitriles in vitro and in vivo,
Saillenfait and Sabate (2000) administered a single dose of 0 or 100 mg/kg AN by gavage in
olive oil to groups of four pregnant Sprague-Dawley rats on GD 10. Dams were sacrificed on
GD 12, and the numbers of uterine implantation sites and fetuses with heartbeats were recorded.
Viable fetuses were examined for defects of the allantois, trunk, and misdirected caudal
extremity (left-sided), which the investigators had found to be typical in embryos exposed to
sodium cyanide. Maternal effects of all of the nitriles (including AN) included increases in
clinical signs (piloerection, prostration, and/or tremors) and unspecified maternal BW loss
between GDs 10 and 12. AN exposure had no effect on the numbers of implants/litter or live
embryos/litter, but significantly increased the incidence of overall poor and abnormal
development and the incidence of misdirected allantois or allantois, trunk, and caudal extremity
misdirected (Table 4-43). The study authors suggested that maternal production of the
metabolite cyanide may have contributed to the developmental toxi city of AN since the
characteristic defects that occurred in embryos exposed to cyanide were found in AN-exposed
embryos. The single applied dose of 100 mg/kg is a LOAEL for maternal and fetal effects.
Table 4-43. Morphological alterations in GD 12 fetuses of Sprague-Dawley
rats exposed to 100 mg/kg AN on GD 10
Response
Implants/litter
Live embryos/litter
Total embryos examined
Group (number of litters examined)
Control (n = 4)
13.25 ±1.71
13.0 ± 1.83
52
100 mg/kg (n = 4)
13.75 ±0.96
12.5 ±1.0
50
Allantois, trunk, and/or caudal extremity misdirected
Embryos affected/embryos examined
Litters affected/litters examined
0/52
0/4
13/463
3/4
""Significantly different from controls (p < 0.01) as calculated by the reviewers (Fisher's exact test).
Source: Saillenfait and Sabate (2000).
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4.3.2.1.3. Reproductive toxicity assay. Friedman and Bellies (2002) and Litton Bionetics (1992)
conducted a three-generation reproductive study in Sprague-Dawley rats (groups of 15 males and
30 females) exposed to 100 or 500 ppm AN in drinking water for 100 days before mating.
Groups of 10 male and 20 female controls received untreated drinking water. As calculated by
the study authors, the concentrations were equivalent to average doses of 0, 11, and 37 mg/kg-
day for males and 0, 20, and 40 mg/kg-day for females. The calculated average doses, averaged
across sexes, were 0, 16, and 39 mg/kg-day. Rats were observed daily, especially for signs of
neurotoxicity (e.g., abnormal gait). Water consumption in the FO generation was measured twice
a week, food intake was measured weekly, and BWs were recorded every 2 weeks. After
100 days of exposure, rats were paired for mating for 6 days; females not bred after 6 days were
mated to another proven breeding male from the same exposure group. The offspring (Fla) of
the first mating were examined on PNDs 0, 4, and 21. Litters were culled to 10 pups on PND 4
to achieve an equal sex ratio. Litter BWs were recorded on PND 4 and individual pup weights
were recorded on PND 21. Two weeks after removal of the Fla offspring, FO females were
remated to produce the Fib litter, and those not used in breeding Fla were also mated to produce
Fib pups to ensure a sufficient number of offspring for the F2 generation, although these animals
should have been discarded according to the original study design. The original study design
stipulated discarding Fla pups (those not scheduled for breeding) at weaning, but, because of
high mortality in the 500 ppm group, they were retained to ensure sufficient time-mated females
to use as foster mothers (one-half of each high-dose litter was fostered onto untreated females).
All FO males used for breeding were discarded after the second mating, and those not used in
breeding were discarded when the Fla litter were weaned.
At weaning (21 days), one male and one female from the unfostered Fib litter (or from
the Fla litter, if needed) were selected as potential breeders for the F2 generation. All
Fl offspring exposed to AN were checked daily for mortality, and necropsies were performed on
all animals either found dead or killed in a moribund condition. The FO dams and females not
producing a litter were exposed for an additional 20 weeks after weaning the Fib pups. After
that period, they were sacrificed and the sciatic nerve, gastrocnemius muscle, brain, and gross
lesions were examined for histopathology. Fla and Fib rats were sacrificed in week 95 of the
study and necropsied for the examination of papillomas in the stomach and intestines. The
protocol for the subsequent generations (Fib parents and F2a and F2b offspring; F2b parents and
F3a and F3b offspring) was as described for the first generation. F2 females, after an additional
20 weeks of exposure following the weaning of the F3b litters, were necropsied, and sciatic
nerve, gastrocnemius muscle, brain, stomach, and gross lesions were evaluated for
histopathology. F3b rats (10/sex) in the control and 500 ppm groups were randomly selected for
histologic examination.
The following discussion of results does not include cancer data from this study, which
were presented separately in Section 4.2.1.2.9. In FO parents, BWs were lower than controls
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after 4 weeks of exposure to 500 ppm, the reduction reaching 22% in males and 17% in females
by week 10. This BW reduction was accompanied reduction in food intake by -18% in 500 ppm
males and 8% in 500 ppm females for the first 10 weeks. Water consumption was reduced in
males and females by about 50% at the high dose and 20-25% at the low dose for the first
10 weeks. Reduced water consumption and possibly BWs may have been a result of reduced
palatability of treated water. Exposure to AN had no effect on the incidence of neurological or
other clinical signs, male or female fertility indices, or the duration of mating or gestation of
FO parents.
Compared with controls, no significant changes in fertility index or gestational index
were observed in any of the exposed generations (Table 4-44) (Friedman and Beliles, 2002;
Litton Bionetics, 1992). However, poor fertility among controls for the Fib (50-60%) and
F2b (60-70%) parents might have limited the capability of this study to detect differences in
fertility between control and treated rats (Table 4-44). The durations of mating and gestation
were also unaffected, except the mating duration of F2b rats for F3a generation. Significant
decreased viability and lactation indices were observed in the 500 ppm FO parents for the
Fla generation, due to the deaths of pups between 1-4 and 5-21 days, respectively. Significant
decreases in viability index were also observed in 100 and 500 ppm FO parents for the
Fib generation and in 500 ppm F2b parents for the F3a generation (Table 4-44). In addition,
decreases of about 10-40% in pup weights were observed in the Fla, Fib, F2a, F2b, F3a, and
F3b generations in the 500 ppm groups (Table 4-45). Overall, the results identified a drinking
water concentration of 100 ppm (16 mg/kg-day) as a reproductive toxicity LOAEL for decrease
in lactation index in F2a generations. Although no effects on fertility were observed in this
three-generation study, poor fertility among controls in Fib and F2b parents might have limited
the sensitivity of this assessment. The study also identified a LOAEL of 100 ppm for reduced
viability in FO parents.
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Table 4-44. Group-specific reproductive indices in three generations of
Sprague-Dawley rats receiving AN in drinking water
Generation/group
Malea'b
Femalec'd
Fertility index
Gestation index"
Viability index'
Lactation index8
FO parents for F la
Control
100 ppm
500 ppm
10/10
8/10
10/10
18/20
16/20
16/20
18/18
16/16
16/16
185/186
197/201
166/17711
138/150
139/150
95/143h
FO parents for Fib
Control
100 ppm
500 ppm
10/10
10/10
13/15
16/20
17/20
22/28
16/16
17/17
22/22
186/186
182/202h
99/109h
137/150
132/139
87/99
Fib parents for F2a
Control
100 ppm
500 ppm
5/10
7/10
8/10
10/20
11/20
14/20
10/10
11/11
14/14
107/109
116/124
133/140
91/91
95/104h
107/1 14h
Fib parents for F2b
Control
100 ppm
500 ppm
6/10
5/10
8/10
10/20
8/20
14/20
10/10
8/8
14/14
101/101
93/97
138/138
82/82
70/73
123/123
F2b parents for F3a
Control
100 ppm
500 ppm
6/10
9/10
10/10
14/20
13/20
15/20
14/14
13/13
15/15
161/161
157/158
157/16611
128/131
124/124
134/135
F2b parents for F3b
Control
100 ppm
500 ppm
9/10
10/10
10/10
14/20
15/20
17/20
14/14
15/15
17/17
170/176
198/198
170/178
106/108
117/119
115/125
"Doses in males: 0, 11, or 37 mg/kg-d for 0, 100, or 500 ppm as calculated by the study authors.
fertility index in males = number of males producing a litter/number mated.
Doses in females: 0, 20, or 40 mg/kg-d for 0, 100, or 500 ppm, as calculated by the study authors.
fertility index in females = number of pregnant females/number mated.
eGestation index = number of litters born/number of pregnant females.
Viability index = number of pups that survived to PND 4/number of pups born alive.
8Lactation index = number of pups surviving to weaning/number of pups alive on PND 4.
hSignificantly lower than controls (p < 0.05), as calculated by the study authors.
Sources: Friedman and Beliles (2002); Litton Bionetics (1992).
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Table 4-45. Group-specific pup weights in three generations of Sprague-
Dawley rats receiving AN in drinking water
Generation-concentration
F la-control
Fla-100 ppm
Fla-500 ppm
Fib-control
Fib- 100 ppm
Flb-500 ppm
F2a-control
F2a-100 ppm
F2a-500 ppm
F2b-control
F2b-100 ppm
F2b-500 ppm
F3a-control
F3a-100 ppm
F3a-500 ppm
F3b-control
F3b-100 ppm
F3b-500 ppm
Weight (g)
D4
11
10
9a
10
9
10
11
10
9a
11
10
9
10
9
8a
10
10
8a
D 21 (males)
42
40
28a
39
36
34a
39
39
30
53
46
30a
43
43
30a
50
47
32a
"Significantly different from controls (p < 0.05), as calculated by the study authors.
Sources: Friedman and Beliles (2002); Litton Bionetics (1992).
4.3.2.1.4. Male exposure reproductive toxicity studies. Tandon et al. (1988) evaluated
reproductive toxicity in male CD-I mice that received 0, 1, or 10 mg/kg-day AN by gavage in
saline for 60 days. The testes of six mice/group were examined histopathologically, and
homogenates of pooled testes (four testes) in each group were assayed for the activities of
sorbitol dehydrogenase (SDH), acid phosphatase, LDH, glucose-6-phosphatase dehydrogenase,
and p-glucuronidase.
Exposure to AN decreased the epididymal sperm count by 21 and 45% at the low- and
high-dose group, respectively. However, the decrease was statistically significant (p < 0.05)
only with the 10 mg/kg-day group. Histopathological examination of the testes did not reveal
changes in the low-dose group. In high-dose mice, degenerative changes were seen in 40% of
seminiferous tubules. In addition, the testes of mice dosed with 10 mg/kg-day showed a 12%
increase in the activity of LDH, a 22% decrease in the activity of SDH, a 37% increase in the
activity of P-glucuronidase, and a 16% decrease in the activity of acid phosphatase compared
with controls. AN exposure had no effect on the activity of testicular glucose-6-phosphatase
dehydrogenase. The study authors suggested that the changes in the activities of LDH and SDH
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were related to the AN-induced degeneration of germinal epithelium. In this study, a NOAEL of
1 mg/kg-day and a LOAEL of 10 mg/kg-day were identified for the toxicological effects of AN
on the testes of male CD-I mice.
In a range-finding acute toxicity study for a dominant lethal assay, groups of 10 male
F344 rats received 45, 60, 68, 75, or 90 mg/kg-day AN by gavage in 0.9% saline daily for 5 days
(Working et al., 1987). Rats were observed for a total of 42 days from the first administration.
No deaths were observed in groups receiving 45 or 60 mg/kg-day. In the higher dose groups, the
study was terminated early on account of mortality: 40% at 68 mg/kg-day (terminated on day 6),
30% at 75 mg/kg-day (terminated on day 4), and 30% at 90 mg/kg-day (terminated on day 2).
As a result of this range-finding study, 60 mg/kg-day was selected as the maximum tolerated
dose for the dominant lethal assay. In this assay, groups of 50 male F344 rats received AN by
gavage at 0 or 60 mg/kg-day in 0.9% saline for 5 days. BWs of males were recorded three times
during the week of treatment and then weekly during the mating period of 10 weeks; males were
caged with a different female each week for 6 days. A transient reduction in mean BW (-4%)
compared with controls was observed in rats on treatment days 3 and 5 and on the fourth post-
treatment day; BW gain in treated rats was equivalent to controls beginning the second week of
observation. AN exposure in males had no effect on the incidence of pre- or postimplantation
losses, indicating a negative result in the dominant lethal assay. AN also had no effect on the
fertility of exposed males in any postexposure week.
4.3.2.2. Inhalation Exposure
Information about reproductive/developmental toxicity in animals exposed to AN by
inhalation comes from a standard chronic toxicity assay in rats, developmental toxicity assays in
rats, a two-generation reproductive toxicity study in rats, and a dominant lethal assay in male
mice.
4.3.2.2.1. Standard toxicity assays. As described in Section 4.2.2.2.1, no reproductive
histopathology was observed in male or female Sprague-Dawley rats that were exposed to AN at
concentrations as high as 80 ppm 6 hours/day, 5 days/week for up to 2 years (Dow Chemical
Co., 1992a; Quastetal., 1980b).
4.3.2.2.2. Developmental toxicity assays. Haskell Laboratory (1992a) evaluated developmental
toxicity in groups of 30 pregnant Sprague-Dawley rats exposed (whole body) to 0, 40, or 80 ppm
AN vapor for 6 hours/day on GDs 6-15. The parameters examined were the same as those for
the oral exposure study by these authors described in Section 4.3.2.2.2. Inhalation exposure to
40 or 80 ppm AN did not result in deaths, changes in appearance, gastric thickening, or increase
in terminal liver weight in dams. Maternal BW gain was significantly reduced during GDs 6-15
by 47% in the 40 ppm group and by 58% in the 80 ppm group (a reduction of about 20% in both
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groups for GDs 6-21). Both exposure groups exhibited significant (unspecified) decreases in
food consumption on GDs 6-9 (but not later intervals) and increases in water consumption on
GDs 9-20. In this study, a NOAEL for maternal effects was not identified, but a LOAEL of 40
ppm was identified for reduced BW gain in dams.
Gestational exposure to AN had no significant effect on any of the reproductive
parameters (pregnancy rates, numbers of implantations, live fetuses, or resorptions) and no effect
on fetal BW or crown-rump length measurements. Furthermore, no single major malformation
occurred at significantly higher incidence in AN-exposed rats vs. controls. However, as shown
in Table 4-46, there was an increase in the incidence of total major malformations when they
were considered collectively (p < 0.06) for the high-dose group (present in 6/35 litters).
Malformations observed in litters of 80 ppm group included short tail, missing vertebrae, short
trunk, omphalocele, and hemivertebra. In this study, a NOAEL of 40 ppm and a LOAEL of
80 ppm were identified for increases in total malformations in rats.
Table 4-46. Incidence of fetal malformations among litters of Sprague-
Dawley rats exposed to AN by inhalation
Type of malformation
External and skeletal malformations
Visceral malformations
AN concentration (ppm)
0
40
80
Number of fetuses/number of litters examined
421/33
140/33
441/36
148/36
406/35
136/35
Number of fetuses (litters) affected
External malformations
Short tail
Short trunk
Imperforate anus
Omphalocele
0(0)
0(0)
0(0)
0(0)
0(0)
0(0)
0(0)
1(1)
2(2)
1(1)
0(0)
1(1)
Visceral abnormalities
Right-sided aortic arch
Missing kidney, unilateral
Anteriorly displaced ovaries
0(0)
0(0)
0(0)
0(0)
0(0)
0(0)
0(0)
0(0)
1(1)
Skeletal malformations
Missing vertebrae
Missing two vertebrae and two ribs
Hemivertebrae
Total malformed
0(0)
8(1)
0(0)
8(1)
0(0)
2(1)
0(0)
3(2)
2(2)
7(2)
1(1)
ll(6)a
"Significantly different from control, as calculated by the study authors (p = 0.06).
Source: Murray et al. (1978).
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Saillenfait et al. (1993) included AN in a survey of the relative developmental toxicities
of inhaled aliphatic mononitriles in rats. Pregnant Sprague-Dawley rats (20-2 I/group) were
exposed (whole body) for 6 hours/day to 0, 12, 25, 50, and 100 ppm AN vapor on GDs 6-20.
Dams were observed daily throughout pregnancy and BWs were recorded on GDs 0, 6, and 21.
All subjects were sacrificed on GD 21, and the uteri were weighed and opened to assess the
numbers of implantations, resorption sites, and live and dead fetuses. The fetuses were
examined for external abnormalities and then split into two equal groups for examination of
skeletal or visceral anomalies.
AN exposure did not cause premature deaths in the dams, but exposure to >25 ppm
caused concentration-related, statistically significant reductions (by 16-45%) in overall BW gain
and concentration-related losses in absolute BW (exclusive of gravid uterus weight) in dams. In
this study, 12 ppm is a NOAEL and 25 ppm is a LOAEL for reduced BW in dams.
AN exposure had no effect on reproductive parameters (pregnancy rate, average number
of implantations, numbers of live fetuses, incidences of nonsurviving implants, or resorptions per
litter). Concentration-dependent, statistically significant reductions in average fetal weight (by
5-15% compared with controls) were observed at >25 ppm. There were no significant increases
in the incidences of external, visceral, or skeletal anomalies in the exposed groups and one
control fetus. In this study, a NOAEL of 12 ppm and a LOAEL of 25 ppm were identified for
significantly reduced fetal BW.
In a dominant lethal assay for AN, Zhurkov et al. (1983) continuously exposed male ICR
mice (20/group, whole body) to AN at concentrations of 0, 20, or 100 mg/m3 (0, 9.1, or 46 ppm)
for 5 days. After exposure, each male was mated to two unexposed females for 8 weeks.
Females were sacrificed between GDs 13 and 15, and a number of reproductive and
developmental parameters were monitored, including the percentage of pregnant females,
number of corpora lutea, implantations, and live and dead fetuses per female as well as total and
pre- and postimplantation mortality. Exposure to AN did not result in any dominant lethal effect
on male germ cells nor did it cause adverse pre- or postimplantation outcomes. The highest
exposure level in this study, 46 ppm, was a NOAEL for reproductive toxicity. This study was
limited by sparse descriptions of methods and a lack of quantitative reporting of results.
4.3.2.2.3. Two Generation Reproductive Toxicity Study. In a two-generation reproductive
toxicity study of AN via the inhalation route (Nemec et al., 2008), Sprague-Dawley rats (FO
generation, 25/sex/group) were exposed to AN vapor via whole-body inhalation at 0, 5, 15, 45,
and 90 ppm for 6 hours/day, 7 days/week for 10 weeks. These animals were randomly bred to
produce an Fl generation. Following weaning on postnatal day (PND) 28, animals selected to be
parents from the Fl generation were similarly exposed. Exposure of the Fl parents at 90 ppm
was terminated after 16 to 29 days due to excessive systemic toxicity in the males. The
remaining four groups of the Fl generation followed the same breeding procedure as the FO
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generation (25 animals/sex/group). The FO and Fl generations were about 8 weeks old and 4
weeks old at initiation of their respective exposures. The FO and Fl males were exposed for 10
weeks prior to mating and throughout mating until one day prior to euthanasia, and the FO and Fl
females were exposed for 10 weeks prior to mating and throughout mating, gestation, and
lactation until one day prior to euthanasia. Exposure of the FO and Fl dams was suspended for
five days following parturition (lactation days (LDs) 0 to 4), to avoid confounding nesting and
nursing behavior and neonatal survival during early postnatal development. Exposure of the
dams resumed on LD 5.
To reduce variability among the litters, large litters were randomly reduced to 10
pups/litter (5/sex when possible) on PND 4. Each male pup selected as a parent for the Fl
generation was examined for balanopreputial separation beginning on PND 35, and each selected
Fl female pup was examined for vaginal perforation beginning on PND 25. Plasma and red
blood cell (RBC) cholinesterase levels were measured on 10 rats/sex of the FO parental
generation from the control and 90-ppm groups and from 10 rats/sex of the Fl parental
generation from the control, 5-, 15-, and 45-ppm groups.
Sperm samples from the right epididymis were collected from each adult FO and Fl male
and evaluated for the percentage of progressively motile sperm. Sperm morphology was
evaluated by light microscopy.
Surviving FO and Fl adults were euthanized and necropsied following completion of
weaning of their offspring (Fl and F2 pups, respectively). Microscopic evaluations (with
emphasis on developmental and reproductive organs) were conducted on the following tissues
for 10 randomly selected FO and Fl parental animals per sex from the control and high-exposure
groups: adrenal glands, prostate, brain, pituitary, seminal vesicles, right epididymis, right testis,
vagina, cervix, coagulating gland, uterus, oviducts, ovaries, nasal cavities, lungs. Additionally,
gross lesions from all FO and Fl animals in the control, 5-, 15-, and 45-ppm groups were
microscopically evaluated. On PND 28, a complete necropsy similar to that performed on
parental animals was conducted on Fl pups not selected for AN exposure and on F2 pups.
Clinical sign of irritation (clear/red material around the nose, eyes, and mouth and on the
forelimbs) were observed for the FO males and females exposed to 90 ppm throughout the
exposure period within 1 hour following completion of daily exposure, but generally did not
persist to the following day. Wet, cool tails were also noted for these animals within 1 hour
following exposure; and to a greater extent in the males.
Body weight gains for the 45- and 90-ppm FO males were statistically reduced relative to
controls during the first three weeks of exposure, resulting in a statistically significant decrease
in body weight gain (up to 11.8%) throughout the FO generation. Food consumption for these
males was also decreased. Decreased food consumption and body weight gains were also noted
for FO females exposed to 45 and 90 ppm during the first 2 weeks of treatment and throughout
gestation, resulting in a statistically significant decrease (about 4.5%) in body weight for 45-ppm
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females at study week 2, and 90-ppm females throughout the 10 week premating period and
gestation (7.5% to 9.1%).
Body weight gains in the 45-ppm Fl males were slightly reduced during the first 3 weeks
of AN exposure, but the effects were less pronounced than in the FO males. Body weights for the
45-ppm Fl males were decreased by up to 9.4% during study weeks 18 to 26.
No adverse exposure-related effects were observed on the number of days between
pairing and coitus, gestation length, or reproductive performance (fertility, mating, copulation,
and conception indices) in FO and Fl generations. A slight (6%) and statistically significant
decrease in sperm motility (including progressive motility) was observed for the FO males
exposed to 90 ppm AN when compared to controls. A decrease (6 to 9.5%) in sperm motility
(including progressive motility) was also observed for the FO and Fl males exposed to 45 ppm
AN when compared to controls, although the decrease was not statistically significant.
Statistically significant and exposure-related increases (up to 8%) in absolute and relative
anogenital distances were observed for the Fl males exposed to 45 and 90 ppm AN (see Table
4-47). Slight delays (up to 8%) in the acquisition of sexual developmental landmarks
(balanopreputial separation in males and vaginal patency in females) and lower body weights on
the day of acquisition (relative to control group) were observed for Fl males in the 45- and 90-
ppm groups and Fl females in the 90-ppm group (see Table 4-47).
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Table 4-47. Summary of developmental landmark evaluations in Fl and F2 offspring
in two generation inhalation reproductive study
Acrylonitrile exposure level
Endpoint
0 ppm
5 ppm
15 ppm
45 ppm
90 ppm
Absolute anogenital distance on PND 1 (mm)
Fl males
F2 males
Fl females
F2 females
3.45 ±0.378
4.66 ±0.305
1.82 ±0.163
2.56 ±0.213
3.54 ±0.317
4.59 ±0.411
1.79 ±0.128
2.56 ±0.277
3.49 ±0.343
4.66 ± 0.422
1.77 ±0.170
2.55 ±0.308
3.67±0.205a
4.49 ±0.325
1.82 ±0.175
2.48 ±0.195
3.66 ±0.202
NA
1.78 ±0.126
NA
Normalized anogenital distance on PND 1 (relative to cube root of pup body weight)
Fl males
F2 males
Fl females
F2 females
1.76 ±0.168
2.40 ±0.148
0.95 ±0.091
1.34 ±0.124
1.81 ±0.158
2.37 ±0.197
0.94 ± 0.077
1.36 ±0.138
1.80 ±0.170
2.38 ±0.198
0.94 ±0.092
1.34 ±0.164
1.88±0.119a
2.32 ±0.156
0.95 ±0.108
1.30 ±0.093
1.90±0.094b
NA
0.95 ±0.064
NA
Balanopreputial separation (PND)
Fl males
44.6 ±3. 11
44.2 ±2.87
45.6 ±3.07
46.1 ±3.77
46.9 ±3.05
Body weight at acquisition of EPS (g)
Fl males
205.4 ± 20.73
196.3 ±18.59
201 ±20.09
191.8±15.23a
169.2 ±15.45a
Vaginal patency (PND)
Fl females
34.3 ±1.54
34.2 ±1.91
34.5 ±1.67
34.8 ±2.60
37.0±2.92b
Body weight at acquisition of VP (g)
Fl females
107.6 ±10.45
103. 3 ±10.89
102.7 ±8.51
102.7 ±13. 3
99.2 ±9.38
Note: The number of pups evaluated ranged from 17-25/group for anogenital distance and 24-25/group for
balanopreputial separation and vaginal patency.
EPS: balanopreputial separation; VP: vaginal patency.
NA: not applicable because the Fl 90-ppm group was terminated prior to breeding.
"Statistically significant atp < 0.05; bstatistically significant atp < 0.01.
Source: Nemec et al. (2008)
Slightly decreased male pup weights in the 5-, 15-, and 45-ppm groups on PND 28 were
statistically significant relative to controls, resulting in lower overall weight gain during PNDs 1
to 28. However, the decrease did not appear to be dose-related.
RBC cholinesterase activity was unaffected in males and females exposed to 90 ppm AN
in the FO generation and 5, 15, and 45 ppm AN in the Fl generation. Plasma cholinesterase
activity in the FO females exposed to 90 ppm AN was statistically significantly lower (by 40%)
than control. Plasma cholinesterase activity in the Fl females was lower than control in the 5-,
15-, and 45-ppm group, but the decrease (40%) was statistically significant only for the 15-ppm
group.
An increase (up to 10%) in liver weights occurred in the 90-ppm FO males (statistically
significant) and females and the 45-ppm Fl males. Nemec et al. (2008) stated that statistically
significant increases in relative liver weight and brain weight were also observed, and could not
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be explained by decreased body weight alone (data not provided). Statistically significant
decreases in thyroid gland weight (about 10%) were found in Fl females at 5 ppm and 15 ppm.
Histopathologic alterations were found in the nasal tissues of FO males and females at 45
ppm, Fl males at 5, 15, and 45 ppm, and Fl females at 15 and 45 ppm (see Table 4-48). The
lesions showed exposure-related response in incidence and severity, and included
respiratory/transitional epithelial hyperplasia, subacute inflammation, squamous metaplasia,
and/or degeneration of the olfactory epithelium.
Table 4-48. Histologic changes in nasal tissues of adult CrlrCD (SD) rats after
continuous exposure to test atmospheres of acrylonitrile
FO males3
Exposure level (ppm)
0
5
15
45
Fl males
Exposure level (ppm)
0
5
15
45
FO females3
Exposure level (ppm)
0
5
15
45
Fl females
Exposure level (ppm)
0
5
15
45
Nasal level I
Total number
examined
Hyperplasia,
respiratory/Transi-
tional epithelium
Minimal
Mild
Moderate
Severe
Metaplasia,
squamous
Minimal
Mild
Moderate
Severe
Inflammation,
subacute
Minimal
Mild
Moderate
10
0
-
-
-
0
-
-
-
-
0
-
-
-
10
0
-
-
-
0
-
-
-
-
0
-
-
-
10
0
-
-
-
0
-
-
-
-
0
-
-
-
10
10b
-
6
4
o
J
o
J
-
-
-
2
1
1
-
10
2
2
-
-
0
-
-
-
-
2
2
-
-
10
6
3
2
1
2
-
1
1
-
4
3
1
-
10
10b
-
5
5
8b
-
2
5
1
9b
-
6
3
10
10b
-
1
5
4.
8b
2
5
1
-
9b
5
3
1
10
2
-
1
1
1
1
-
-
-
4
-
3
1
10
0
-
-
-
0
-
-
-
-
0
-
-
-
10
0
-
-
-
0
-
-
-
-
0
-
-
-
10
7
4
o
J
-
0
-
-
-
-
1
1
-
-
10
0
-
-
-
0
-
-
-
-
0
-
-
-
10
0
-
-
-
0
-
-
-
-
0
-
-
-
10
7b
2
5
-
6b
2
o
J
1
-
6b
5
1
-
10
9b
5
2
1
1
4
1
1
2
-
3
1
2
-
Nasal level II
Total number
examined
Degeneration,
olfactory
epithelium
Minimal
Mild
Moderate
Hyperplasia,
respiratory/Transiti
onal epithelium
10
0
-
-
-
10
0
-
-
-
10
0
-
-
-
10
6b
2
4
-
10
0
-
-
-
0
10
0
-
-
-
0
10
0
-
-
-
0
10
5b
2
2
1
1
10
0
-
-
-
10
0
-
-
-
10
0
-
-
-
10
6b
2
4
-
10
0
-
-
-
10
0
-
-
-
10
0
-
-
-
10
8b
2
6
-
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Table 4-48. Histologic changes in nasal tissues of adult CrlrCD (SD) rats after
continuous exposure to test atmospheres of acrylonitrile
Mild
FO males3
Exposure level (ppm)
0
-
5
-
15
-
45
-
Fl males
Exposure level (ppm)
0
-
5
-
15
-
45
1
FO females3
Exposure level (ppm)
0
-
5
-
15
-
45
-
Fl females
Exposure level (ppm)
0
-
5
-
15
-
45
-
Nasal level III
Total number
examined
Degeneration,
olfactory
epithelium
Minimal
Hyperplasia,
epithelial,
respiratory
Minimal
10
10
10
10
10
10
10
10
10
10
10
10
10
0
0
10
0
0
10
0
0
10
1
1
1
1
Nasal level IV
Total number
examined
Degeneration,
olfactory epithelium
ivniH
10
10
10
10
10
10
10
10
10
10
10
10
10
0
10
0
10
0
10
1
1
Additional organs evaluated histologically in 10 randomly selected animals/sex in the control and 45-ppm groups
included the brain, right epididymis, pituitary, prostate, seminal vesicles, adrenal medulla, adrenal cortex, right testis,
coagulating gland, oviducts, ovaries, uterus, cervix, and vagina (data not presented).
aNasal tissues from 90-ppm FO animals were not examined.
bStatistically significant atp < 0.05.
Source: Nemec et al. (2008)
EPA identified a NOAEL of 15 ppm (8.14 mg/m3, adjusted for continuous exposure)for
reproductive/developmental toxicity, based on increases in anogenital distance in Fl males on
PND 1 and decreases in body weight of Fl males at acquisition of balanopreputial separation.
However, the study LOAEL for the Fl generation was 5 ppm (2.7 mg/m3, adjusted for
continuous exposure) based on histologic changes in nasal tissues; a NOAEL could not be
identified. The NOAEL for FO generation was 15 ppm (8.14 mg/m3, adjusted for continuous
exposure), based on histologic changes in nasal tissues.
4.3.2.3. Intraperitoneal Administration
In a translated study in Chinese, the effect of AN on spermatogenesis was examined in
groups of male Kunming mice (10 3-4-week-old pubertal and five 6-8-week-old adult mice per
group) treated by i.p. injection (Liu et al., 2004). The animals received AN (>99% purity) at
doses of 1.25, 2.5, or 5.0 mg/kg-day (1/24, 1/12, or 1/6 of the LD50) for 5 days. Negative
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controls received physiological saline (10 mL/kg) daily for 5 days, whereas positive controls
received a single injection of 40 mg/kg CP (not identified but presumably cyclophosphamide).
On day 35, mice were sacrificed, the left testicle was selected from five mice per group, and
testicular cell suspensions were evaluated using flow cytometry.
The following dose-related effects observed in the 2.5 and 5.0 mg/kg-day groups were
statistically significantly different from the negative control group (p < 0.05); the same effects
were observed in positive controls (Liu et al., 2004). AN decreased the percentage of haploid
testicular cells (indicative of completed spermatogenic meiosis) by 14.3 and 15% in mid- and
high-dose adult mice, but a slight reduction was not statistically significant in pubertal mice.
The percentage of apoptotic testicular cells was significantly increased by 58.7 and 74.8% in
mid- and high-dose pubertal mice and by 81.5 and 108% in mid- and high-dose adult mice. The
percentages of spermatogenic epithelial cells in Go/Gi phase were significantly reduced by
18 and 20% in mid- and high-dose pubertal mice and by 35.5 and 40.5% in mid- and high-dose
adult mice. The percentage of spermatogenic epithelial cells in G2/M phase was significantly
elevated in adult mice by 31.4 and 32.8% at the mid- and high doses, respectively. AN exposure
had no effect on the percentage of spermatogenic epithelial cells in S phase. The NOAEL and
LOAEL for suppression of spermatogenesis were 1.25 and 2.5 mg/kg-day, respectively.
The teratogenic effects of AN were investigated in groups of pregnant golden hamsters
that were exposed via an i.p. injection on GD 8 (Willhite et al., 1981). The actual numbers of
dams per group were not reported, but data were presented for 3-6 litters in the exposed groups
and 12 litters for the controls injected with sodium chloride. Dose levels were 0, 0.09, 0.19,
0.47, 1.23, and 1.51 mmol/kg (equivalent to 0, 5, 10, 25, 65, and 80 mg/kg). No adverse clinical
symptoms were observed in the dams exposed to up to 1.23 mmol/kg (65 mg/kg) AN by GD 14,
when the dams were sacrificed. No malformations were observed in the offspring. In contrast,
dams exposed to 1.51 mmol/kg (80 mg/kg) showed intense dyspnea, gasping, incoordination,
hypothermia, salivation, and convulsions for 1-5 hours after injection. Additionally, several
fetal abnormalities and malformations were observed at this exposure level, including
encephaloceles and fused or bifurcated ribs. Coadministration of 8.06 mmol/kg STS and
1.51 mmol/kg AN induced neither maternal toxicity nor fetal malformations. Coadministration
of 8.06 mmol/kg STS and a higher dose of AN (1.88 mmol/kg, 100 mg/kg) prevented toxic
symptoms in dams but not the teratological effects in fetuses. The study authors concluded that
the teratogenic action of AN was related to the metabolic release of cyanide.
The effect of AN on rat liver cytochrome P-450 and serum hormone levels were studied
in male Sprague-Dawley rats (4/group) injected with 33 mg/kg AN (i.p.) for 3 consecutive days
(Nilsen et al., 1980). Control animals were treated with 0.9% sodium chloride. Blood were
collected immediately after the animals were sacrificed and serum luteinizing hormone (LH),
follicle stimulating hormone (FSH), and prolactine (PRL) were measured by radioimmunoassay.
Serum corticosterone levels and liver microsomal cytochrome P-450 were also determined.
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A significant increase (3%) in BW was reported in the AN treated group when compared
with controls (Nilsen et al., 1980), but relative liver weight was not increased. A significant
decrease in liver microsomal content of CYP450 was observed. Serum corticosterone measured
24 hours after the last AN injection was decreased by 70%, while FSH levels were increased by
118%. Serum PRL levels were decreased by 63% when compared with the controls, while
serum LH levels were unchanged. Nilsen et al. (1980) suggested that the increase in FSH might
be secondary to impaired spermatogenesis in the testes of treated rats.
4.3.3. In Vitro Studies
The embryotoxicity of AN was evaluated in cultures of whole rat embryos by Saillenfait
and coworkers in a series of studies (Saillenfait et al., 2004, 1992; Saillenfait and Sabate, 2000).
As described initially in Saillenfait et al. (1992), the experimental system involved incubating
day 10 embryos from pregnant Sprague-Dawley rats for 26 hours in whole organ culture in a
medium containing AN at concentrations of 76-760 umol/L. The growth-related parameters
evaluated were functional yolk-sac circulation, yolk-sac diameter, crown-rump length, head
length, number of somites, number of malformed embryos, incidences of abnormal brain,
malformed caudal extremities, delayed yolk-sac circulation, and defective flexion. The effects of
metabolic activation on the embryotoxicity of AN were evaluated by the inclusion of S9 and
cofactors (NADPH, glucose-6-phosphate) for CYP450-dependent biotransformation in the
incubation system.
Exposure to AN induced concentration-related effects on growth and development.
Functional yolk-sac circulation (circulating erythrocytes) was reduced at>304 uM and was
completely absent at 760 uM. Crown-rump length was reduced at 304 uM and could not be
measured at higher concentrations. A concentration-dependent increase in the incidence of
malformations was observed following in vitro exposure of day 10 fetuses to AN at 152 uM and
above. AN at 152 and 304 uM induced malformations in 53 and 100% of the exposed embryos,
respectively. Malformations primarily consisted of a shortened caudal extremity (significant at
152 uM) and a reduction of the brain (achieving significance at 304 uM). Other general
malformations were delayed development of the yolk-sac circulation and defective flexion.
Furthermore, growth retardation and severity of malformations induced by 304 uM AN were
enhanced by the presence of S9 and cofactors, suggesting a role for oxidative biotransformation
in AN embryotoxicity. Addition of 0.1-2.2 mM GSH in the incubation medium reduced the
embryotoxic effects of AN in this system.
Similarly, a later experiment by this research group (Saillenfait and Sabate, 2000)
demonstrated that day 10 rat embryos exposed to AN for 46 hours in vitro showed concentration-
related effects on growth, development, and morphology. Reduced growth (reduced yolk sac
diameter and crown-rump length) occurred at 100 uM. The head length and somite number were
reduced at 125 uM. Abnormal development (reduced prosencephalon and mesencephalon) and
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maxillary process defects occurred at>125 uM AN. Defects of the rhombencephalon and the
auditory system were observed less frequently. These effects were characteristic of those
developed by embryos exposed to sodium cyanide in culture. Addition of microsomes and
NADPH to the culture medium containing 150 or 175 uM AN enhanced the observed growth
retarding and dysmorphogenic effects, especially in increasing the incidence of
rhomb encephal on and auditory system defects.
Saillenfait et al. (2004) evaluated the effects of eight aliphatic nitriles on the viability and
differentiation of cultured limb bud cells from Sprague-Dawley rat embryos on GD 13. Limb
bud micromass cultures were exposed for 5 days to AN at concentrations between 0.01 and
0.45 mM (0.5 and 23.9 ug/mL), with or without microsomal activation, after which they were
evaluated for cytotoxicity (neutral red uptake assay) and differentiation of chondrocytes (number
and total surface of foci with Alcian blue staining were used as indicator of cell differentiation).
The concentrations that inhibited cell viability and cell differentiation by 50% of concurrent
untreated controls were determined.
The ICso values were 0.24 mM (13 ug/mL) for cytotoxicity (viability) and 0.33-0.38 mM
(18-20 ug/mL) for differentiation (total surface of foci and number of foci), respectively. The
ratios of ICso for cytotoxicity and differentiation were 0.6 for number of foci and 0.7 for total
surface of foci. Microsomal activation had no effect on the results with AN. In parallel
experiments, the ICso for cytotoxicity in cultured 3T3 cells (differentiated mouse fibroblast cell
line) exposed to AN was 0.065 mM. The relative potency of the tested nitriles in this limb bud
cell culture system matched previously published results for cultured whole embryos, but not
necessarily for in vivo teratogenicity (false negative results were obtained for two of the eight
nitriles). Saillenfait et al. (2004) characterized the response of AN in the micromass culture
assay as equivocal, since it depended on the criteria used to define a positive result. According
to one criterion, AN might be considered to have teratogenic potency because its ICso was less
than 50 ug/mL. However, under the "twofold rule" that defines a positive result by a value
>2 for the ratio (ICso for cytotoxicity)/(ICso for differentiation), AN would be classified as
having poor potential developmental hazard. As suggested by comparison of the ICso for
cytotoxicity, embryonic rat limb bud cells were not more vulnerable to AN than differentiated
3T3 mouse fibroblasts. These suggestive results are consistent with the observations in vivo (see
Section 4.3.2) that fetal toxicity from AN occurs only at exposure levels that cause maternal
toxicity.
4.4. OTHER DURATION- OR ENDPOINT-SPECIFIC STUDIES
4.4.1. Acute Toxicity Data
Closely similar values have been reported for the oral LD50 in rats: 78 mg/kg (Benesh
and Cherna, 1959), 93 mg/kg (Smyth and Carpenter, 1948), 90 mg/kg (Sprague-Dawley rats)
(Younger Labs, 1992), and 81 mg/kg (male CF Nelson rats) (Vernon et al., 1990). Lower value
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has been published for the oral LD50 in mice: 25 mg/kg (H-strain, sex not stated) (Benesh and
Cherna, 1959) and 36 and 48 mg/kg in male and female mice, respectively (strain not stated)
(Tullar, 1947, as cited in IPCS, 1983). These findings indicate that mice are more susceptible to
the acute toxicity of AN than are rats. Published LD50 values for guinea pigs (56 mg/kg)
(Jedlicka et al., 1958) and rabbits (93 mg/kg) (Paulet and Desnos, 1961) are in the same narrow
range as those for rats and mice. For the inhalation route, median lethal concentration (LCso)
values in rats of 217 ppm (470 mg/m3) (Knobloch et al., 1971) and 333 ppm (777 mg/m3)
(Haskell Laboratory, 1992b) have been reported. However, exposing Sprague-Dawley rats to
1,008 ppm AN for 1 hour failed to produce mortality (Younger Labs, 1992). A 4-hour
LCso value of 946 ppm (2,053 mg/m3) was reported for Sprague-Dawley rats following nose-
only exposure (WIL Research Laboratories, 2005); no mortality was observed in male or female
rats at exposures as high as 775 ppm (1,682 mg/m3) in this study. However, ataxia, labored
respiration, and hypoactivity were observed at this exposure concentration and higher.
LCso values (duration not reported) of 138 ppm (300 mg/m3) in mice and 456 ppm (990 mg/m3)
in guinea pigs have been reported (Knobloch et al., 1971).
As set forth in the English abstract of an article in Polish, Knobloch et al. (1971)
investigated the acute and subacute toxicity of AN in BN mice, Wistar rats, and guinea pigs
(strain not given). For rats, s.c. and i.p. LDso values of 80 and 100 mg/kg, respectively, were
reported. For mice, the LDso was 34 mg/kg.
An earlier subacute study examined the effect of oral AN administration on the liver of
sodium PB-pretreated (400 umol/kg) or Aroclor 1254-pretreated (300 umol/kg) male and female
Sprague-Dawley rats that received 100 or 500 ppm AN in drinking water for 21 days (Silver et
al., 1982). Other pretreated animals were given 0, 50, 75, 100, or 150 mg/kg AN by gavage for
up to 3 days. Some liver-related biochemical changes were noted as a result of AN exposure,
including a dose-dependent reduction in hepatic nonprotein sulfhydryl concentrations (maximal
reduction of 81% at the highest dose of 150 mg/kg). Serum SDH activity was increased by
fourfold 24 hours after administration of 150 mg/kg AN. On the other hand, SGPT activity was
not significantly altered. Pretreatment with PB or Aroclor 1254 resulted in only a slight
enhancement of AN-induced elevation of serum SDH or SGPT activities. In the experiment in
which female rats were pretreated with either vehicle, PB or Aroclor 1254, and treated with
100 ppm or 500 ppm AN for 21 days, there was a 60% increase in serum SDH activity in
animals receiving 500 ppm AN in drinking water.
In the 3-day studies, there was some evidence of focal superficial necrosis of the liver in
rats receiving high gavage doses (100 or 150 mg/kg). This effect was associated with the
presence of hemorrhagic gastritis and distention of the forestomach. However, while the
biochemical and pathology changes implied a limited perturbation of the liver by AN, light
microscopy showed only minor changes in the histopathology of the organ and no ultrastructural
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changes to the liver were evident when evaluated by electron microscopy. Silver et al. (1982)
concluded that liver did not appear to be a target organ in the acute or subacute toxicity of AN.
Dudley andNeal (1942) exposed several species of laboratory animals—rat, guinea pig,
rabbit, cat, dog, rhesus monkey—to AN vapor for 0.5-8 hours at concentrations ranging from
0.063 to 5.3 mg/L (30-2,445 ppm). In rats (Osborne Mendel, sex not stated), 1,260 ppm was
found to be an effective lethal concentration when administered for 4 hours. During or after an
8-hour exposure, 320 ppm AN was fatal. A 4-hour exposure to 260 ppm AN was fatal in rabbits
(strain and sex not stated), while 1,160 ppm appeared to be a fatal concentration in guinea pigs.
Other lethal concentrations were 600 ppm for 1.5 hours in cats and 110-165 ppm for 3 hours in
dogs (breed not given, male and female), while Rhesus monkeys (males and females) tolerated
90 ppm AN with only minor, transitory adverse effects (skin redness, sleepiness). In all species
except guinea pig, AN caused an initial respiratory stimulation, followed by rapid, shallow
breathing, nasal exudate, and watering eyes. In rats, the most striking symptom was reddening
of the skin in the less prominently haired regions (nose, ear, feet). Reddening of the skin was
also observed in the other species, again with the exception of guinea pigs. The study authors
concluded that, in five of the six species investigated, AN-induced symptoms resembled cyanide
poisoning. By contrast, AN acted as a severe pulmonary irritant in guinea pigs. Consistent with
these findings, sodium nitrite (a cyanide antidote) protected five of the species studied, but not
guinea pigs, from AN toxicity. The study authors postulated that guinea pigs may metabolize
AN differently than rats, rabbits, cats, dogs, or monkeys, possibly by transformation to acrolein
rather than cleavage of the cyano group. However, Dudley and Neal (1942) were unable to
extract any cyanide from the tissues of animals that had succumbed to AN exposure.
Knobloch et al. (1971) evaluated the acute toxicity of AN via inhalation exposure
(duration not given) in three species. The LCso in mice was 0.30 mg/L (138 ppm), in rats
0.47 mg/L (217 ppm), and in guinea pigs 0.99 mg/L (456 ppm). These values confirmed the
roughly twofold species difference between rats and guinea pigs, but were about threefold lower
than values reported by Dudley and Neal (1942).
Two studies by Gut et al. (1985, 1984) evaluated the subchronic toxicity of AN in rats
when administered via inhalation. In both studies, male Wistar rats were exposed to 280 mg/m3
AN by inhalation 8 hours/day for 5 days (Gut et al., 1985, 1984). As shown in Table 4-50, BW
was decreased in rats exposed to AN for 5 days. The absolute weight of liver decreased, while
brain weight remained unchanged. Hence, the relative liver weight was significantly decreased,
while relative brain weight increased due to the BW decrease. However, there were no
significant histopathological changes in the lungs, livers, kidneys, or adrenals. The
concentrations of glucose, pyruvate, and lactate were elevated in the brain and blood of exposed
rats (Table 4-49). Treatment-related reductions in the concentrations of serum cholesterol and
triglycerides were found in test animals. The study authors considered the increase in blood
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glucose concentration to be the most sensitive indicator of AN exposure under these
experimental conditions (Gut et al., 1984).
Table 4-49. Effects of AN on organ weight, clinical chemistry, and
biochemical parameters when administered to male Wistar rats via
inhalation
Parameter (units)
BW(g)
Relative liver weight (g/100 g BW)
Relative brain weight (g/100 g BW)
Serum triglyceride (mmol/L)
Serum cholesterol (mg/dL)
Controls (n)
341 ± 18 (8)
3. 39 ±0.29 (8)
0.564 ± 0.06 (8)
2.36 ±0.51 (10)
72.3 ± 8.5 (10)
AN(n)
287 ± 24a (8)
2.79 ± 0.08a (8)
0.647 ± 0.076a (8)
1.47±0.47b(10)
54.6 ± 12.9b (10)
Pyruvate (mmol/L)
Blood
Brain
0.080 ±0.012
0.039 ±0.007
0.128±0.028a(5)
0.068±0.014a(5)
Lactate (mmol/L)
Blood
Brain
Glucose (mmol/L)
2.259 ± 0.577 (5)
6.002 ± 0.258 (5)
4.10 ±0.21
4.032±1.061a(5)
8.034±0.471a(5)
10.22 ± 1.26C(5)
aSignificantly different from controls (p < 0.05).
bSignificantly different from controls (p< 0.001).
'Significantly different from controls (p < 0.01).
Source: Gut etal. (1984).
Gut et al. (1984) further examined the dose-response effects of AN on blood glucose
levels by giving food-deprived male Wistar rats a single 12-hour exposure of 0, 57, 125, or
271 mg/m3 AN. Blood glucose levels were dose-dependently increased at the end of the
exposure period (4.33 ± 0.64 at 57 mg/m3; 10.88 ± 3.74 at 271 mg/m3 vs. 3.46 ± 0.29 mg/m3 in
controls) but declined to normal or below normal levels 24 hours after exposure
(1.79 ± 0.84 at 271 mg/m3 vs. 4.73 ± 0.34 mg/m3 in controls). The simultaneous increase in the
pyruvate and lactate concentrations suggests that AN affected carbohydrate metabolism on the
level of glycolysis as well as on the citric acid cycle level.
In a subsequent study using exposure regimens identical to those in Gut et al. (1984), Gut
et al. (1985) monitored AN-induced changes in sulfhydryl concentrations in the liver and brain
and the appearance of thioethers (AN-mercapturic acid) and thiocyanate in the urine. GSH
concentrations in the liver were significantly reduced compared with controls (3.86 ± 0.41 vs.
7.66 ± 0.58 umol/g) in rats exposed to AN for five consecutive daily 8-hour exposures at
280 mg/m3. No GSH depletion was evident in the brain. The protein sulfhydryl level remained
unchanged in the brain but was increased by 17% in the liver, although not statistically
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significant. One out of 15 treated animals died, and the study authors suggested that GSH
depletion could have been responsible.
Urinary excretion of thioethers and thiocyanate was proportional to the inhaled
concentration. The ratio between urinary thioethers and thiocyanate in exposed rats was about
2:4 and was not influenced by the exposed AN concentrations (Gut et al., 1985).
Bhooma et al. (1992) evaluated the effect of AN on the procoagulant activity in
pulmonary alveolar macrophages of rats. Six male Wistar rats/group were exposed to 100 ppm
AN 5 hours/day for 5 days. Animals were sacrificed at 1-28 days after the last exposure. The
lungs were lavaged, and alveolar macrophages were collected from the broncho alveolar lavage
(BAL) fluid. Procoagulant activity (the ability of a cell or cell products to accelerate the
conversion of fibrinogen to fibrin) in macrophage and BAL fluid was determined. The
procoagulant activity in the isolated macrophages from exposed animals was about 10-fold
higher than in controls 1 day after AN exposure and then slowly declined to control levels by
day 28 after exposure. Procoagulant activity of BAL fluid remained unaltered in rats sacrificed
up to 7 days after exposure but was elevated in those sacrificed 14 and 28 days after exposure.
The study authors noted that acute lung injury often resulted in deposition of fibrin in alveolar
spaces (Bachofen and Weibel, 1977) and that fibrin and its degradation products have been
implicated in contributing to pulmonary inflammation (Malik et al., 1979; Cuterman et al.,
1977). Since this study showed macrophage-associated procoagulant activity in the lung
following inhalation to AN, the study authors suggested the alveolar macrophages participated in
the pulmonary deposition of fibrin.
Other experimental studies have used acute dosing protocols to study toxicological
effects of AN on various target organs. These studies are summarized in the following sections.
4.4.1.1. Effects of AN on the GI Tract
Two studies have reported GI hemorrhage and gastric erosion in rats administered with
single doses of AN. In the first report (Ghanayem and Ahmed, 1983), a single dose of 50 mg/kg
AN was administered to Sprague-Dawley rats (orally or s.c.). GI bleeding was observed 3 hours
after treatment, with no significant difference in the amount of GI blood loss resulting from s.c.
or oral administration. Thus, AN-induced GI bleeding was not a result of direct irritation of AN
on the GI tract. A time-course study of a single 50 mg/kg dose of AN administered by the s.c.
route indicated that the amount of blood recovered from the stomach was significantly higher
than that of controls at 1, 2, and 3 hours after treatment, while the amount recovered from the
intestinal contents was significantly higher than in controls at 2 and 3 hours. Dose-response
study of s.c. administration of AN and GI bleeding indicated that significantly higher GI
bleeding than in controls occurred at 40, 50, and 70 mg/kg AN, with the maximum at 50 mg/kg
AN.
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Pretreatment of rats with the CYP450 enzyme inducer, PB, decreased the AN-induced GI
blood loss by 55%, whereas pretreatment with Aroclor 1254 increased blood loss by 240%
(Ghanayem and Ahmed, 1983). In contrast, pretreatment of rats with CYP450 inhibitors, cobalt
chloride or SKF 525 A, prior to AN administration produced significant decreases in blood loss
of 10 and 40%, respectively. Pretreatment of rats with diethylmaleate (DEM), a known depletor
of GSH, prior to AN administration produced no significant change in GI bleeding. In addition,
s.c. administration of a sublethal dose (6 mg/kg) of KCN did not induce GI bleeding when
compared with controls, whereas 50 mg/kg AN produced significant GI bleeding. The study
authors concluded that metabolic activation of AN to a reactive metabolite other than cyanide
(probably CEO) by CYP450 was a prerequisite for AN to induce gastric hemorrhage.
In the second paper, Ghanayem et al. (1985) studied the mechanism of AN-induced
gastric mucosal necrosis in the glandular stomach. Male Sprague-Dawley rats were treated with
a single s.c. dose of AN (50 or 30 mg/kg), and the glandular stomach was removed and evaluated
for histopathology and GSH concentrations. The liver was also removed for GSH determination.
Gastric erosion severity index (GEI) was obtained for each exposure group by multiplying the
mean severity score by the incidence of gastric necrosis in the group. Calculated GEI was found
to be dose and time dependent: higher at 50 mg/kg than 30 mg/kg AN 3 hours after
administration and in rats killed 3 hours as compared with rats killed 1 hour after the same dose
of AN.
Subcutaneous administration of 30, 40, or 50 mg/kg AN also caused a significant
decrease in hepatic GSH concentration 3 hours after treatment, with greater decrease at lower
dose (30 mg/kg) than in high dose (50 mg/kg). A significant decrease in gastric GSH
concentrations was observed 3 hours after treatment at 40 and 50 mg/kg AN. Pretreatment of
rats with various metabolic modulators (CYP450 monooxygenase and GSH) before
administration showed that there was a significant inverse relationship between gastric GSH
concentration and AN-induced gastric erosions. P450 inducers (Na PB and Aroclor 1254) alone
increased GSH levels in the liver. Pretreatment of rats with these P450 inducers inhibited
AN-induced gastric necrosis, and partially blocked AN-induced gastric GSH depletion. SKF
525 A, a CYP450 inhibitor, caused a slight depletion of gastric and hepatic GSH and potentiated
the AN-induced gastric necrosis and GSH depletion. In contrast, cobaltous chloride, another
inhibitor of CYP450 enzyme, inhibited AN-induced gastric necrosis and increased both the
hepatic and gastric GSH concentrations. In rats treated with DEM, a depletor of GSH and AN,
or DEM + cysteamine + AN, GEIs were increased up to fivefold when compared with rats
treated with AN alone. Mucosal erosion severity in these two groups was also greatly increased,
with more than 50% of rats showing severe or extensive lesions. Pretreatment of rats with
sulfhydryl-containing compounds (cysteine or cysteamine) protected against AN-induced gastric
necrosis and blocked the depletion of gastric GSH.
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In addition, AN-induced gastric erosions could be prevented by pretreatment with
atropine, a muscarinic receptor blocker, suggesting the involvement of muscarinic receptors in
the AN-induced gastric mucosal necrosis (Ghanayem et al., 1985). Activation of acetylcholine
muscarinic receptors is known to increase gastric acid secretion and cause gastric erosions.
Because muscarinic receptors are known to contain sulfhydryl groups in their active site (Ikeda
et al., 1980; Aronstam et al., 1978), Ghanayem et al. (1985) hypothesized that AN inactivated
critical sulfhydryl groups and caused gastric erosions by locally modulating muscarinic
acetylcholine receptors in the stomach.
More recently, Ahmed et al. (1996a) showed accumulation of AN-derived radioactivity
in intestinal contents and intestinal mucosa following i.v. injection of 2-[14C]-AN to rats. A
recent study by Jacob and Ahmed (2003a) also demonstrated that AN and or its metabolites
accumulated and covalently interacted in GI mucosa of male F344 rats treated either i.v. or orally
with 2-[14C]-AN. These studies supported the hypothesis that AN-induced injury of the GI
mucosa is not due to direct irritation by AN but by metabolic incorporation and macromolecular
interaction of AN in these tissues.
4.4.1.2. Effects of AN on the Kidney
The acute nephrotoxic effect of AN was investigated in single exposure studies in rats
and hamsters by inhalation or i.p. administration. Intraperitoneal injection of Chinese hamsters
with 30 mg/kg AN increased kidney weight and renal GSH concentration 24 hours after injection
(Zitting et al., 1981). Kidney deethylation activity was also decreased. Rouisse et al. (1986)
administered i.p. doses (0, 10, 20, 40, 60, or 80 mg/kg) of AN to male F344 rats (six/dose group).
Urinary volume was increased two- to threefold during the 24-hour period following
administration for all dose groups. Urinary glucose was about 6 times higher in the 20 mg/kg
group than in controls, and 40-60 times higher in the higher dose groups. Urinary excretion of
N-acetyl-p-D-glucosaminidase was increased at the highest doses, up to 80% over controls in the
80 mg/kg group. An increased number of lysosomes or dense bodies in renal proximal tubules
was seen under the light or electron microscope. In the inhalation study, a similar array of
toxicological and clinical chemistry effects relating to kidney structure and function was
observed in male rats (seven/group) exposed to 200 ppm AN for 4 hours (Rouisse et al., 1986).
4.4.1.3. Effects of AN on the Adrenal Gland
An experimental model for the toxicity of AN investigated the capacity of single i.v.
doses of AN to induce acute hemorrhagic necrosis of the adrenal gland (Szabo et al., 1976). This
condition has a parallel in man, the Waterhouse-Friderichsen syndrome, which is characterized
by thrombocytopenia, disseminated intravascular coagulation, and the appearance of fibrin
degradation products in the circulation (Szabo et al., 1976). The condition has been induced in
female Sprague-Dawley rats by a single injection of 150 mg/kg AN into the jugular vein (Szabo
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et al., 1980) and typically has been marked by massive "bilateral apoplexy," a usually fatal
hemorrhagic necrosis of the adrenal glands that occurs in 90-100% of the rats within 90-
150 minutes. Thrombocytopenia and a range of associated clinical signs, including tremors,
cyanosis, and ultimately respiratory failure, were observed. Light and electron microscopy
showed that an early event in the onset of this condition was damage to the vascular endothelium
of the adrenal cortex, with parenchymal injury as a late event. However, these AN-induced
lesions could be prevented by pretreatment with the a-adrenergic antagonist phenoxybenzamine,
the a,p-blocker labetalo, or the 11-p-hydoxylase inhibitor metyrapone. Elevation of tissue
sulfhydryl levels by cysteine or GSH reduced the adrenal apoplexy. Dopamine concentrations in
the adrenals increased over time. Because depletion of catecholamines by reserpine, or
medullectomy, could prevent the chemically induced adrenocortical necrosis, the study authors
proposed that cortical damage resulting from AN was associated with vasoactive amines released
from the medulla and/or with metabolites of AN.
4.4.1.4. Effects of AN on Neurological Endpoints
Neurotoxic effects were induced in male Sprague-Dawley rats receiving single gavage or
s.c. doses of 20, 40, or 80 mg/kg AN (Ghanayem et al., 1991). Two distinct phases of
neurological response were evident. The early phase was cholinomimetic in nature, with signs
such as salivation, lacrimation, polyuria, miosis, vasodilatation, gastric secretion, and diarrhea.
The second phase developed 4-5 hours later with toxic signs, including depression, convulsions,
and respiratory failure, followed by death at the higher doses. These CNS effects were observed
in rats treated with 40 and 80 mg/kg and were similar to those caused by cyanide. Pretreatment
of animals with 1 mg/kg atropine, an acetylcholine muscarinic antagonist, abolished the
cholinomimetic toxicity, implicating an involvement of the cholinergic system in some aspects
of acute AN neurotoxicity. Since effects were observed even at the lowest dose, a NOAEL was
not identified, and 20 mg/kg was the LOAEL.
In another study, male Sprague-Dawley rats that were administered s.c. doses of
112 mg/kg AN (LD90) showed a biphasic response consisting of an early phase with tremors and
seizures about 100 minutes after dose administration, followed by severe clonic convulsions that
preceded death at about 3-4 hours (Benz and Nerland, 2005). The effects of preadministered
inhibitors of oxidative metabolism by CYP450 (80 mg/kg SKF 525A, 75 mg/kg
1-benzylimidazole, or 100 or 200 mg/kg metyrapone) and an alternative CYP450 substrate,
ethanol (5,000 mg/kg), on the acute convulsions were examined. Although blood levels of
cyanide, and the development of the first phase of tremors and seizures, were inhibited by
1-benzylimidazole or ethanol, treatment with these two agents did not prevent the terminal
convulsions or the death of rats injected with 112 mg/kg AN. Ethanol, being a CNS depressant,
decreased the incidence of terminal convulsion (5/17 vs. 15/17). These results suggested that the
initial phase of the acute neurotoxically lethal effects may have been due to cyanide, which is
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released via the CYP450 metabolic pathway for AN, and that the second phase was mediated by
the parent compound.
Because ethanol showed some effect on lessening the second phase response to AN
(although it did not prevent lethality), several anticonvulsants were examined for their ability to
counteract the acute neurotoxicity and lethality of 112 mg/kg AN. Administration of PB
(25 mg/kg) or phenytoin (150 mg/kg) (but not 144 mg/kg valproic acid) markedly inhibited the
lethal response to 112 mg/kg AN: 9/10 rats died following administration of AN alone or AN
plus valproic acid, whereas 1/10 and 2/10 rats died following administration of AN plus PB or
AN plus phenytoin, respectively. The protection by phenytoin and PB against convulsion and
lethality was not due to inhibition of metabolism of AN to cyanide, since only phenytoin was
able to lower blood cyanide levels (by about 32%), and was much less effective than
1-benzylimidazole or ethanol (97 and 94%, respectively).
4.4.1.5. Effects of AN on Hearing
AN is one of a number of organic compounds that have been shown to promote noise-
induced hearing loss (NHL) in rats (Fechter, 2004; Fischel-Ghodsian et al., 2004). The
ototoxicity of AN was examined in several experiments in male Long-Evans rats exposed by s.c.
injection (Fechter et al., 2003). The AN used in these experiments was stabilized to minimize
the accumulation of peroxides. Ten rats were anesthetized and surgically prepared for the
assessment of the compound action potential (CAP), which represents the synchronous neural
activity elicited by primary auditory neurons (spiral ganglion cells) and directly measures
auditory threshold sensitivity. Auditory thresholds at each test frequency were measured for
each rat under anesthesia on 20 different occasions before treatment with AN to determine
baseline auditory thresholds. Auditory thresholds for 11 test frequencies between 2 and 40 kHz
were recorded at 5-minute intervals up to 100 minutes postinjection with AN. The acute effect
of 50 mg/kg s.c. AN on auditory sensitivity was studied in five rats; five control rats were
injected only with water. In a second study, the effects of AN on permanent NIHL were
evaluated in six experimental groups (six rats each) that received the following: no treatment;
50 mg/kg AN alone, two injections of 50 mg/kg-day AN on 2 consecutive days, noise alone
(108 dB octave-band noise for 8 hours), single injection of 50 mg/kg AN immediately followed
by noise, and exposure to noise after a second injection of AN.
The acute study showed that exposure to 50 mg/kg AN s.c. alone elevated auditory
threshold temporily and produced a 10-20 dB loss in auditory threshold sensitivity (temporary
threshold shift) in the stimulus range of 8-40 kHz (Fechter et al., 2003). This transient loss
reached a maximum within 10-20 minutes after injection but returned to control levels within
75-100 minutes. In the study on permanent auditory threshold shifts, when rats were tested
3 weeks following AN and AN + noise treatment, exposure to AN alone did not produce a
persistent loss in auditory threshold sensitivity. AN-treated rats had slightly reduced auditory
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threshold compared with controls, indicating slightly increased sensitivity (within 5 dB of
controls), but the difference was not statistically significant. Noise treatment alone elevated
auditory thresholds by less than 20 dB at all tested frequencies. However, rats given two
injections of AN followed by noise in the high frequency range of 12-40 kHz exhibited auditory
impairments averaging 27 dB (maximally 40 dB) compared with controls; this shift was
statistically significant compared with other groups (controls, noise alone, AN alone, or no
treatment). Rats given a single injection before exposure to noise showed an average 11 dB
threshold shift in the high-frequency range (12 and 40 kHz). This study demonstrates that
exposure to AN exacerbated NIHL.
Fechter et al. (2003) also assessed blood cyanide and glutathione levels in brain, liver,
and paired cochleae in additional groups of rats exposed to AN. Groups of five rats given 20, 50,
or 80 mg/kg s.c. AN produced peak levels of cyanide, a metabolite of AN, in the blood at 1 hour
(for 20 and 50 mg/kg) and 2 hours (for 80 mg/kg) following injection. Cyanide levels returned
to baseline values within 2, 3, and 4 hours, respectively. Since AN produced maximal auditory
threshold impairment within 20 minutes of administration, before blood cyanide level peaked at
1 hour, the acute ototoxic effect of AN was not likely associated with elevated cyanide levels.
When rats were given a single injection of 50 mg/kg AN, maximal reductions in
glutathione levels (measured between 15 minutes and 8 hours postinjection) were detected in the
brain by 15 minutes (-50%), in the liver by 1 hour (-80%), and in the cochlea by 2 hours (-
45%). Cochlear glutathione levels remained depressed for about 4 hours. Recovery of
glutathione levels to near control levels was achieved in all tissues by 8 hours. Since AN
induced transient cochlear function loss that peaked within 10-20 minutes after injection and
recovered within 75-100 minutes, the acute ototoxic effect of AN could not be associated with
GSH level in the cochlea. Fechter et al. (2003) concluded that, while AN-induced oxidative
stress in the cochlea may play a role by which AN promotes NIHL, the acute ototoxic effect of
AN might reflect other unidentified toxic action of AN in the cochlea.
Fechter et al. (2004) extended their evaluation of the potentiation effect of AN exposure
on NIHL. Two experiments were conducted: the first experiment studied the effects of a single
AN exposure on permanent NIHL, while the second one evaluated the effects with five daily AN
and noise exposures. In both experiments, six male Long-Evans rats were assigned to each
treatment group: AN alone, noise alone, AN + noise, and untreated controls. AN exposed
groups were given s.c. injections of 50 mg/kg-day AN, with or without 4 hours of exposure to
noise (105 dB), and then assessed for auditory threshold sensitivity 4 weeks later.
In the single-day treatment study, AN alone did not alter the auditory threshold, but a
single AN exposure in combination with noise significantly elevated auditory thresholds by an
average of 10 dB above the effect of noise alone. Noise exposure alone increased auditory
thresholds compared with controls by an average of 10 dB in the range of 12-40 kHz (Fechter et
al., 2004). In the 5 consecutive-day treatment study, AN plus noise treatment exacerbated the
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impairment induced by noise alone by an average of 30-45 dB at frequencies between 20 and
40 kHz and by no more than 10 dB at frequencies <16 kHz. Similar to the single-exposure
study, repeated AN exposure had no effect on auditory thresholds assessed 4 weeks later.
Repeated noise exposure elevated auditory thresholds about 17 dB above control rats.
When the reactive-oxygen scavenger phenyl-N-tertiary-butylnitrone (PEN) (100 mg/kg
i.p.) was injected twice daily for 5 days prior to exposure to noise alone and rats were assessed
4 weeks later, PEN reduced the magnitude of hearing loss. When PEN was injected daily prior
to the injection of AN and again following noise exposure, the magnitude of hearing loss was
equivalent to that exhibited by rats treated by noise alone. (No significant difference was found
between rats receiving noise alone and those receiving PEN plus noise.) These results indicated
that reactive oxygen species (ROS) were responsible for the ototoxic effects of AN. These
findings are consistent with the suggestion that mitochondria injury within cochlear cells,
primarily or secondarily through oxidative stress, may be a common feature of ototoxicity
induced by chemicals and noise (Fischel-Ghodsian et al., 2004).
Results from a recent study indicated that hearing loss from AN and noise exposure
involves histologic damage to hair cells on the surface of the organ of Corti (Pouyatos et al.,
2005). Groups of five male Long-Evans rats were given s.c. injections of 0 or 50 mg/kg-day AN
for 5 consecutive days, with or without exposure 30 minutes later to noise for 4 hours/day (95 or
97 dB octave-band noise at 8 kHz). Hearing dysfunction of these rats were then assessed by:
(1) distortion product otoacoustic emissions (DPOAEs) before exposure, as well as 1 hour and
4 weeks postexposure; (2) CAP for auditory threshold sensitivity 4 weeks after the last treatment;
and (3) number of hair cells on surface preparation of the organ of Corti 4 weeks after treatment.
Permanent effects on these endpoints (i.e., effects observed 4 weeks following the last treatment)
were only observed in rats exposed to both AN and noise and not in rats exposed to AN or noise
alone. Permanent effects from combined exposure to AN and noise included auditory threshold
shifts (13-16 dB between 7 and 40 kHz), a decrease in DPOAE amplitudes (up to 25 dB at
19 kHz), and significant outer hair cell (OHC) loss in the cochleae. With the AN plus 97 dB
treatment, average OHC loss was 20, 16, and 9% in the first, second, and third rows,
respectively, in the areas corresponding to frequencies ranging from 13 to 47 kHz. Similar
effects were found in the AN plus 95 dB treatment group. This study demonstrated AN could
potentiate NIHL at noise levels that are relevant to human exposure.
Pouyatos et al. (2007) further proposed that AN exacerbated NIHL by decreasing
antioxidant defenses of hair cell. This hypothesis was tested in a study in which the capability of
specific antioxidants in the protection of the cochlea of male Long-Evans rats treated with
50 mg/kg AN s.c. 30 minutes prior to the daily noise exposure of 97 dB sound pressure level
4 hours/day for 5 days. Sixty-five Long-Evans rats (2-12/group) were exposed to different
combinations of noise, AN, and antioxidants; AN alone or AN + STS (150 mg/kg i.p.), a CN
inhibitor; AN + 4-methylpyrazole (4MP,100 mg/kg i.p.), a drug that blocks CN generation by
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competing with CYP2E1; AN + L-N-acetylcysteine (L-NAC) (4 x 400 mg/kg, orally), a
pro-GSH drug; noise (97 dB octave band of noise [OBN]/8 kHz) alone; noise and STS, 4MP, or
L-NAC; or noise plus AN and antioxidants. To evaluate auditory impairment, DPOAEs and
CAPs were measured prior to experimental treatment and 3 days and 4 weeks after treatment. At
the end of exposure, cochleae were harvested for histologic examination. Additional rats (n =
64) were used to measure cochlear and liver GSH and blood CN levels at different time points
after treatment.
At 3 days postexposure, similar auditory loss was found in animals exposed to AN +
noise, STS + AN + noise, 4MP +AN + noise, and L-NAC + AN + noise. The maximum shifts
averaged 25-30 dB between 12 and 32 kHz. At 4 weeks postexposure, animals exposed to
L-NAC + AN + noise recovered to baseline levels above 25 kHz. However, at lower
frequencies, L-NAC did not prevent auditory loss caused by AN + noise exposure. Animals
received combined exposure to AN + noise, and AN + noise + STS or AN + noise + 4MP
showed little change in producing auditory loss.
In addition, the cochleae from rats exposed to AN and noise demonstrated substantial
damage in the basal half of the organ of Corti. Mean OHC loss averaged 35% in the three rows
in the region corresponding to frequencies above 12 kHz. Neither STS nor 4MP pretreatment
protected against OHC loss caused by AN + noise. However, pretreatment with L-NAC reduced
the OHC loss caused by AN + noise in the region corresponding to 25 kHz and above.
Liver GSH level was depleted by 63% 1 hour after AN injection. Cotreatment with STS
or 4MP reduced GSH level about 80% at 1 hour and 52% at 3 hours. However, with L-NAC
pretreatment, GSH level was reduced only by 23 and 20% at 1 and 3 hours, respectively.
Similarly, whereas AN treatment depleted cochlear GSH levels to undetectable levels at 1 hour,
STS pretreatment had no effect on GSH depletion. 4MP pretreatment only slightly reduced GSH
depletion. However, L-NAC pretreatment not only protected but induced an increase in GSH
levels above control levels. Pouyatos et al. (2007) concluded that, since L-NAC cotreatment
reduced auditory loss and OHC loss from AN + noise treatment, GSH is involved in the
protection of the cochlea against ROS generated by moderate noise levels. However, CN did not
appear to be involved in this potentiation.
4.4.2. Immunological Effects of AN
A case study by Balda (1975) described a human subject who developed contact
dermatitis following the use of a Plexidur finger splint. Investigators obtained a positive result in
a patch test with AN, which is one of the constituents of the polymer Plexidur. This observation
suggested that AN might induce immunological response in exposed subjects.
The immunotoxicity of AN was evaluated in groups of six male CD-I mice given a
single oral dose in water or repeated oral doses for 5 or 14 days (Ahmed et al., 1993). In each of
these experiments, a positive control group received a single immunosuppressive dose of
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225 mg/kg of cyclophosphamide i.p. In the first experiment, mice given a single gavage dose of
0 or 13.5 mg/kg AN in water were evaluated 3 or 5 days later for BW, relative organ weights
(thymus, spleen, liver, and kidney), number of viable splenocytes, and total and differential
WBC counts in spleen cells suspension (Ahmed et al., 1993). AN slightly reduced BWs in mice
(6-7%) after 3 and 5 days, but the difference was not biologically significant. In treated mice,
relative spleen weights were increased by 50% after 5 days, whereas relative thymus weights
were decreased by 42% after 5 days; relative liver weight was decreased by 11% on day 3 but
was 9% higher than controls on day 5. AN treatment reduced total leucocytes by 42-57%,
lymphocytes by 39 and 65%, monocytes by 38 and 50%, and neutrophils by 67 and 73% on
days 5 and 3, respectively.
In the second experiment, mice received 0 or 6.75 mg/kg-day of AN in water on
5 consecutive days and were evaluated on day 6 (Ahmed et al., 1993). In mice treated for
5 days, BWs were reduced by 27%, relative liver weights were decreased by 10%, and relative
spleen weights were increased by 25%. Total splenocytes were decreased by 51%, with total
blood leucocytes and lymphocytes reduced by 48 and 68%, respectively; circulating monocytes
were increased by 88%. The immunotoxic effects of AN in both studies were comparable to
effects elicited by the known immunosuppressant cyclophosphamide.
In a third experiment, groups of male CD-I mice (six/group) were treated with AN by
gavage in water for 14 days (Ahmed et al., 1993). There were a total of nine groups: three
groups dosed with AN at 1.35, 2.7, or 5.4 mg/kg-day only, three groups dosed the same way but
immunized with sheep red blood cells (SRBCs) on day 9 of exposure, and three concurrent
control groups (normal, SRBC-immunized, and a positive control group treated intraperitoneally
with cyclophosphamide [225 mg/kg] 1 day before immunization). Mice were evaluated for total
BW, relative organ weights (thymus, spleen, liver, and lung), number of splenocytes, total and
differential WBC counts in spleen cell suspensions, and histopathology of lymph nodes, lung,
and intestinal Peyer's patches. Subsets of spleen lymphocytes (T and B cells) were enumerated
by flow cytometry. Spleen cells from mice immunized with SRBCs were evaluated in a plaque-
forming cell assay.
In nonimmunized mice, AN decreased BW by 12-15% and increased relative thymus
weights by 43% at 2.7 mg/kg-day and relative lung weights by 39% at 5.4 mg/kg-day. AN
increased splenocyte viability by 49-264% in a dose-independent manner. AN also caused dose-
independent increases of 126-293% in relative spleen weight at all doses. Total leukocyte
counts/spleen were increased by 171, 119, and 107% at the low to high doses; lymphocyte
counts/spleen were significantly reduced by 25-45% at all doses, while monocyte and neutrophil
counts/spleen were increased concomitantly by as much as 78-fold. Reductions in lymphocyte
subsets were observed at all doses: T-cells by 40-53%, B-cells by 32-36%, T-helper cells by
40-59%, and T-suppressor cells by 49-62%. Histological examination revealed severe
enlargement of mesenteric lymph nodes and intestinal Peyer's patches, abscesses and massive
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necrotic damage in the lung, and swellings in the brachial lymph nodes in mice treated with AN
at all doses but more aggressive in the 5.4 mg/kg group. No incidence data were reported for
these lesions. Some of the mice treated with 5.4 mg/kg-day died rapidly. Microbiological
examination revealed the swelling in the lymph nodes was related to migration of normal
intestinal flora, which the study authors attributed to immunosuppressive effects of AN.
In the SRBC-immunized mice, AN treatment did not affect the BW. The relative weight
of lung, liver, and thymus showed inconsistent and dose-independent increases. The viable
spleen cells showed 77% increases only in the 2.7 mg/kg group. The lymphocytic count showed
dose-independent decreases. However, the neutrophilic and monocytic counts showed large
increases at all doses. Decreases in lymphocyte subsets were found in all three doses of AN:
T-cells by 42-45%, B-cells by 37-50%, T-helper by 43-54%, and T-suppressors by 40-48%.
The IgM antibody plaque forming cell response was decreased by 55, 23, and 65 after treatment
with 1.35, 2.7, and 5.4 mg/kg AN, respectively. The lowest dose used in this study, 1.35 mg/kg-
day, was a LOAEL for immunotoxicity (suppression of humoral and cell-mediated immunity) in
mice treated for 14 days.
Hamada et al. (1998) further investigated the immunotoxicity of AN by administering
2.7 mg/kg-day AN (1/10 the LD50) to male CD-I mice orally for either 5, 10, or 15 days. All
mice were injected with 100 mg/kg bromodeoxyuridine (BrdU) i.p. 1 hour before sacrifice. An
immunohistochemical assessment of the number of cells capable of producing IgA in different
intestinal compartments as a result of AN administration was conducted. Uptake of
3H-thymidine into splenocytes derived from treated animals and stimulated with different
mitogens—phytohemagglutinin (PHA), concanavalin-A (con-A), or lipopolysaccharide (LPS)—
was measured. The rate of proliferation of gut epithelial cells of different intestinal
compartments was determined by the incorporation of BrdU in newly synthesized DNA of
S-phase cells.
The mitogenic response of mouse splenocytes to PHA, con-A, and LPS was affected by
AN exposure as shown by a significant decrease in [3H]-thymidine incorporation (Table 4-50).
The decreases were 68-79% with con-A and 35-57% with LPS, depending on the time intervals.
However, uptake of [3H]-thymidine by splenocytes after stimulation with PHA was markedly
reduced only after 15 days of exposure. These results suggested that AN induced systemic
suppression of humoral immunity (by the decrease in mitogen response to LPS) and cell-
mediated immunity (by the inhibition of mitogen response to con-A and PHA).
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Table 4-50. Time course of the effect of AN administration on
[3H]-thymidine uptake into mouse splenocytes under the influence of
different mitogens in vitro
Mitogens
PHAa
con-Aa
LPSa
Control
1,331 ±163
14,927 ± 972
1,225 ±112
D of AN treatment
5
1,203 ± 265
4,756 ± 532b
803 ± 87b
10
1,434 ±35
4,792 ± 946b
486 ± 2b
15
898 ± 32b
3,198±448b
522±31b
"Values are counts/min, mean ± standard error of the mean; n = 4.
bSignificantly different from controls (p < 0.05) as calculated by the authors.
Source: Hamadaetal. (1998).
Inhibition of [3H]-thymidine uptake by stimulated spleen lymphocytes may indicate a
systemic immunosuppression by AN. Such an effect could also occur locally, as indicated by a
reduction in the number of IgA-producing cells in all intestinal compartments following AN
administration. The counts of IgA-producing cells were reduced by 56-77% in the duodenum,
44-67% in the jejunum, and 60-62% in the ileum. Another local effect of AN was demonstrated
by the increased incorporation of BrdU into epithelial cells of the duodenum (threefold) and
ileum (1.6-fold) of AN-treated animals. This result indicated that the rate of cell proliferation
was markedly increased following oral AN administration, even as a result of short-term
treatment. The study authors considered this to be the result of a regenerative response to
chemically induced intestinal injury and suggested that AN-induced immunosuppressive effect
systemically and locally in the gut, as well as increases in the rate of cell proliferation, may
contribute to carcinogen!city of AN in the gut.
Summary
AN caused contact dermatitis in a human subject exposed via the use of Plexidur finger
splint. In mice, AN suppressed cell-mediated and humoral immunity systemically and locally in
the intestine. Table 4-51 summarizes immunotoxicity studies of AN.
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Table 4-51. Summary of immunotoxicity studies of AN
Test
species
Human
subject
Male CD-I
mice
(n=6)
Endpoint/effect
Contact
dermatitis
Immuno-
suppressive
effect
Exposure
concentration/condition
Use of a Plexidur finger
splint
Study 1:
Single dose of 0 or
13.5 mg/kg AN in water
(oral)
225 mg/kg
cyclophosphamide as
positive control
Study 2:
0 or 6.75 mg/kg-d AN in
water
225 mg/kg
cyclophosphamide as
positive control.
Study 3:
a. Three groups dosed with
1.35, 2.7, or 5.4 mg/kg-d
AN
b. Three groups dosed with
1.35, 2.7, or 5.4 mg/kg-d
AN, but immunized with
SRBCs on d 9 of exposure
c. Three concurrent control
groups (normal, SRBC-
immunized)
d. a positive control group
treated intraperitoneally
with cyclophosphamide
(225 mg/kg) Id before
immunization
Exposure
duration
ND
3or5d
5d
14 d
Results
Positive patch test
Increased relative spleen by
50%, decreased relative
thymus weight by 42%,
reduction in total leucocytes,
lymphocytes, monocytes, and
neutrophils.
Increased relative spleen
weights by 25%; decreased
total spenocytes by 51%;
reduction in total blood
leucocytes and lymphocytes by
48 and 68%, respectively.
In nonimmunized mice:
increased spleen weight at all
doses, increased total
leukocyte counts, reduced
lymphocyte counts/spleen, and
increased monocyte and
neutrophil counts/spleen.
Enlargement of mesenteric
lymph nodes and intestinal
Peyer's patches, swellings in
the brachial lymph nodes for
all dose groups.
In the SPvBC-immunized mice:
viable spleen cells showed
77% increase in the 2.7 mg/kg
group. Decreases in
lymphocyte count, increases in
neutrophilic and monocytic
counts, and decreases in
plaque forming cell response
were observed. The
LOAEL for immunotoxicity
was 1.35 mg/kg-d.
References
Balda, 1975
Ahmed etal.,
1993
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Table 4-51. Summary of immunotoxicity studies of AN
Test
species
Male
CD-mice
Endpoint/effect
Systemic and
local immune -
suppression
Exposure
concentration/condition
2.7 mg/kg-d AN orally
Exposure
duration
5, 10, or
15 d
Results
Decrease 3H-thymidine
incorporation into splenocytes
in: (1) mice treated with
mitogens con-A or LPS,
suggesting systemic
suppression of cell-mediated
immunity, and (2) mice treated
with mitogen PHA, suggesting
systemic suppression of
humoral immunity.
Local immunosuppression in
the gut, as indicated by: (1)
reduction in the number of
IgA-producing cells in all
intestinal compartments, and
(2) increased proliferation of
epithelial cells of the
duodenum.
References
Hamadaetal.,
1998.
4.5. MECHANISTIC DATA AND OTHER STUDIES IN SUPPORT OF THE MODE OF
ACTION
4.5.1. Mode-of-Action Studies
Studies have been conducted with the primary purpose of evaluating the potential
mechanisms by which AN and/or its metabolites produce noncancer and cancer effects in
experimental animals. Potential mechanisms of toxicity have been investigated for the following
noncancer effects: GI effects, effects on Hb and metabolism in RBCs, neurotoxicity, oxidative
stress, and immunotoxicity. Studies that investigated the potential mechanisms for the
carcinogenicity of AN have addressed the following effects: formation of DNA adducts,
oxidative stress, intercellular communication, and cell proliferation. In addition, genotoxicity
studies are described in Section 4.5.2.
4.5.1.1. Noncancer Endpoints
4.5.1.1.1. GI Effects. Ghanayem and Ahmed (1983) administered AN to male Sprague-Dawley
rats by both s.c. or oral routes, in each case observing significant gastric bleeding (see Section
4.4.1.1). The response appeared to peak at 50 mg/kg, the effect being maximal at 2 hours after a
50 mg/kg s.c. dose. Different CYP450 inducers had different effects: Aroclor 1254 more than
doubled the effect, while PB reduced the bleeding by half. CYP450 inhibitors SKF 525A and
cobalt chloride were even more effective, with cobalt chloride reducing gastric bleeding to
almost untreated levels. GSH depletion by DEM did not prevent hemorrhaging. The effect was
not due to cyanide release from AN, since KCN administration (6 mg/kg) did not induce
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bleeding. The results indicated that AN-induced hemorrhaging in rat stomach (forestomach and
glandular) and intestine was caused by metabolic activation of AN to a reactive metabolite other
than cyanide (probably CEO).
As described in Section 4.4.1.1, the follow-up study by Ghanayem et al. (1985)
demonstrated that pretreatment of rats with sulfhydryl-containing compounds or atropine
protected the rats from AN-induced lesions. The findings suggested that modulation of
muscarinic receptors by AN increased gastric acid secretion and caused gastric mucosal erosions.
Ghanayem et al. (1997) evaluated the effect of forestomach cell proliferation and
apoptosis in male F344 rats treated with 0, 11.7, or 22.8 mg/kg AN by gavage for 6 weeks (see
Section 4.5.1.2.4). AN induced a dose-dependent increase in epithelial cell proliferation in the
forestomach, as determined by the incorporation of BrdU into S-phase DNA.
4.5.1.1.2. Effects on Hb and metabolism in RBCs. Farooqui and Ahmed (1983b) studied the
effect of AN on Fib and metabolism in RBCs both in vivo and in vitro. Male Sprague-Dawley
rats (three/group) were administered a single oral dose of 80 mg/kg aqueous AN, and blood was
collected 1 hour after treatment. Other groups of animals were sacrificed at 3, 6, and 24 hours
after dosing. Mean cell Fib concentration, hematocrit, and platelet counts were reduced to
78, 79, and 71% of the controls 1 hour after dosing. GSH levels were lowered significantly
within 1 hour. A reduction in the intracellular concentration of 2,3-diphosphoglyceric acid and
increases in intracellular levels of ATP, pyruvate, and lactate were observed. The increase in
intracellular pyruvate and lactate, end products of glycolysis, suggested an increase in the
metabolic rate in RBCs as a result of exposure to AN (Table 4-52). A significant decrease in the
activity of 2,3-diphosphoglycerate mutase, an erythrocyte enzyme, was found in treated animals
(3.61 ± 0.21 lU/g Hb in treated animals after 1 hour vs. 4.57 ± 0.23 lU/g Hb in controls). In
general, the intermediates studied returned to normal values between 6 and 24 hours.
Table 4-52. Effect of AN on RBC metabolic intermediates following a single
oral dose
Intermediates
(umol/mL blood)
2,3 -Diphosphoglycerate
Adenosine triphosphate
Pyruvate
Lactate
GSH
Control
3.6 ±0.3
0.49 ±0.04
58 ±6
1,478 ±131
1.39±0.11
AN treatment
Ihr
2.8 ±0.2a
0.60±0.05a
227±21a
9,932 ±21T
0.28 ± 0.02a
3hrs
3.0 ±0.4
0.64 ± 0.06a
187±13a
1,889 ±129a
0.34±0.04a
6 hrs
3.2 ±0.3
0.66 ± 0.06a
175±16a
1,955 ± 160a
0.47±0.05a
24 hrs
3.7 ±0.5
0.58 ±0.05
167±lla
2,079 ±151a
1.27 ±0.15
"Significantly different from controls (p < 0.05) as calculated by the study authors.
Source: Farooqui and Ahmed (1983b).
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In another study, GSH concentrations in RBCs from male Sprague Dawley rats treated
with a single oral dose of 46.5 mg/kg [2,3-14C]-AN were 10% lower than control in 1 hour,
followed by a slow recovery of 10% in 5 hours (Farooqui and Ahmed, 1983b). Extensive
covalent binding of AN to Hb (about 1.7 umol equivalents of AN bound/mL RBCs) in 1 hour
was also observed.
For the in vitro study, isolated RBCs from male Sprague-Dawley rats were incubated
with 5 mmol/L AN at 37°C to investigate the ability of AN to interact covalently with Fib and
deplete GSH, and the effect of such depletion on Hb (Farooqui and Ahmed, 1983b). GSH in the
supernatant and in the RBCs was estimated by measuring nonprotein sulfhydryl groups.
Additionally, GSH conjugates with AN and levels of Hb and MetHb were monitored.
Incubation of rat RBCs with AN caused a depletion of more than 85% of intracellular
GSH within 1 hour. No hemolysis occurred during the incubation period. In addition, GSH was
not detected in the incubation media. Most of the GSH in the RBCs was converted to S-
cyanoethyl GSH. This GHS-AN conjugate was present in the RBCs after 24 hours, suggesting
that it was not metabolized further at least for 24 hours.
The conversion of Hb to MetHb was about 6% in 1 hour and 8% in 3 hours (Farooqui and
Ahmed, 1983b). MetHb levels in controls during this time period were 0.61-0.87%. The rate
and extent of MetHb reduction to Hb in AN-treated RBCs was also determined to investigate if
this protective mechanism against such oxidative damage to Hb was affected. Incubation of
nitrite-treated RBCs (for conversion of Hb in RBCs to MetHb) with AN resulted in a significant
decrease in MetHb reduction, with a 70% decrease in RBCs treated with 10 mM AN when
compared with controls. In the same experiment, AN initiated hemolysis of RBCs at a
concentration of <0.1 M.
Farooqui and Ahmed (1983b) suggested that the effects of AN on RBC metabolism were
related to the availability of GSH. Oxidative stress induced by depletion of GSH as a result of
AN exposure may have stimulated the rate of RBC metabolism, based on increase in the end
products of glycolysis. Another possible explanation could be the impaired permeability of the
erythrocyte membrane due to extensive covalent binding of AN, resulting in the retention of
metabolic products. Since the levels of ATP and 2,3-diphosphoglycerate were altered and these
two intermediates regulate the oxygen dissociation curve, it was concluded that chronic exposure
to AN may lead to methemoglobinemia, damage to the RBC membrane, and impaired delivery
of oxygen to tissues.
Farooqui et al. (1990) provided in vitro data on the effect of AN on lipid metabolism in
RBCs. RBCs containing oxyhemoglobin (HbO), MetHb, or carboxyhemoglobin (HbCO) were
obtained from male Sprague-Dawley rats. HbO-containing RBCs were taken directly from the
RBC pellet; MetHb-containing RBCs were produced by incubating packed RBCs with 0.5%
sodium nitrite; while RBCs containing HbCO were prepared by blowing carbon monoxide over a
20% suspension (volume/volume [v/v]) of RBCs until the visible spectrum of red cell lysate
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reached a maximum at 570 nm. All preparations of RBCs were incubated for 1 hour with 10
mM AN and variable additions of glucose. Following incubation, supernatants were removed
and used to estimate lipid peroxidation by measuring the concentration of conjugated dienes; the
red cell pellets were used to determine Hb.
Incubation of HbO-containing RBCs with AN resulted in the formation of MetHb, loss of
intact Hb, and membrane lipid peroxidation. The availability of glucose to HbO-containing
RBCs during incubation reduced the formation of MetHb by 33% and the loss of intact HbO by
33% but increased lipid peroxidation by 35%. As a positive control, HbO-containing RBCs were
incubated with 0.1 mM t-butyl hydroperoxide, a strong GSH depleter. The formation of MetHb
and non-intact Hb in the positive control was 3 and 5 times higher than that in AN-incubated
RBCs. Incubation of MetHb-containing RBCs with AN also resulted in the loss of intact Hb and
membrane lipid peroxidation. However, availability of glucose resulted in only 13% increase in
membrane lipid peroxidation. With or without glucose, lipid peroxidation was about 3 times
higher in incubations of HbCO-containing RBCs with AN than the other two RBC preparations.
The extent of lipid peroxidation in RBCs and isolated RBC membranes was dependent on the
concentrations of AN.
Farooqui et al. (1990) also demonstrated an inverse relationship between GSH
concentrations and lipid peroxidation in RBCs incubated with AN. A 75% reduction in GSH
levels in RBCs was observed as a result of AN incubation for 2 hours, with the half-life of GSH
depletion being less than 22 minutes. The concentration of lipid peroxides increased by 274%
over control levels during the same period.
In another experiment, total and Na+/K+-ATPase activity was measured in isolated rat
RBS membranes incubated with 25 mM AN at different temperatures. Farooqui et al. (1990)
showed that Na+/K+-ATPase activity was reduced in the isolated RBC membranes incubated
with AN (release of inorganic phosphate: 87.9 ±9.1 vs. 145.6 ± 13.1 nmol/mg protein per hour
at 37°C; and 4.3 ± 0.7 vs. 87.9 ± 9.1 nmol/mg protein per hour at 15°C). The degree of AN-
induced inhibition of ATPase was temperature dependent. The Km of Na+/K+-ATPase (3.5 mM
at 37°C) was barely affected by the AN treatment, while the Vmax was significantly lower (-36%)
than that of controls. This noncompetitive inhibition of Na+/K+-ATPase by AN was proposed to
be the result of changes in the physicochemical properties of RBC membrane macromolecules
subsequent to irreversible binding of AN to membrane proteins.
4.5.1.1.3. Neurotoxicity. Campian et al. (2002) evaluated the capacity of AN to inactivate the
important glycolytic enzyme, glyceraldehyde-3-phosphate dehydrogenase (GAPDH) in vitro in
order to elucidate the mechanism for the acute toxicity of AN. GAPDH was incubated with AN
(final concentrations ranged from 50 to 400 uM), and the activity of GAPDH was assayed at
different time points. An irreversible inhibition of GAPDH activity was obtained. Incubation of
GAPDH with 200 uM AN resulted in 90% loss of activity in about 60 minutes. The second-
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order rate constant for inhibition of GAPDH activity was measured as 0.2 M"1 s"1 at pH 7.4 and
37°C.
The site of AN incorporation was identified by using matrix-assisted laser desorption
ionization time-of-flight mass spectrometry to analyze tryptic digests of control and AN-labeled
GAPDH (Campian et al., 2002). Inactivation of GAPDH was due to covalent binding of AN to
cysteine 149 at the active site of the enzyme. This finding demonstrated the specificity of AN
binding to cysteine residues and, more widely, implied the ability of AN to impair glycolytic
ATP production in vivo. Campian et al. (2002) speculated that the combination of glycolytic
ATP production impairment with inhibition of mitochondrial ATP synthesis by the AN
metabolite cyanide could result in metabolic arrest. This might have profound consequences for
the toxicity of AN in sensitive tissues such as the brain.
In a follow-up study (Campian et al., 2008), male Sprague-Dawley rats (3-5/group) were
injected subcutaneously with LDgo (115 mg/kg) AN and the brains of treated rats were frozen via
head immersion (HI) in liquid nitrogen or funnel freezing (FF) technique when respiration
ceased. Only minor decreases in ATP of 5% (FF) and 21% (HI) were found when respiration
ceased, although phosphocreatine was decreased by 74% (FF) and 80% (HI), possibly due to
inhibition of creatine kinase by AN. Campian et al. (2008) concluded that no toxicological
relevant depletion of ATP occurred when respiration ceased in AN treated rats. Hence, the acute
lethality of AN was not due to brain metabolic arrest.
Dorman et al. (1996) investigated the mechanisms by which AN and CEO exert
neurotoxic effects in primary dissociated cerebrocortical cell cultures prepared from GD16-18
CD rats. Mature 7-day old cultures were exposed to 0-10 mM AN or 1-2 mM CEO for 8 h at
37°C. Cytotoxicity was evaluated by measuring leakage of LDH, GSH depletion, inhibition of
acetylcholinesterase (AChE) and histopathology.
Both AN and CEO induced dose-dependent increased in cytoxicity. Significant increase
in LDH-leakage was observed in neural cultures following 8 h exposure to 2.5-10 mM AN or
0.125 - 1 mM CEO. Significant reduction in GSH only occurred following exposure to 5 mM
AN. Thus, Dorman et al. (1996) concluded that GSH depletion probably played a limited role in
the development of AN toxicity. No change in AN-induced cytotoxicity was observed following
cotreatment with the cytochrome P-450 inhibitor 1-phenylimidazole, indicating minimal
metabolism of AN to CEO in the neural cells. AChE inhibition was not observed following 8 h
exposure up to 2.5 mM AN. Widespread necrosis of the small, round cholinergic neurons was
present in the cultures treated with 2.5 mM AN or 0.125 mM CEO for 8 h. Astrocytes,
pyramidal neurons, and bipolar neurons were mostly unaffected. On the other hand, treatment
of cell cultures with >0.6 mM cyanide resulted in extensive loss of cytochrome oxidase activity
in the large pyramidal neurons without a concomitant increase in LDH leakage. Dorman et al.
(1996) concluded that AN may selectively destroy cholinergic neurons and induce neurotoxicity
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in rats, and that AN metabolism to cyanide was not a prerequisite for the development of AN-
induced neurotoxicity.
Satayavivad et al. (1998) provided some evidence for the alterations of central
muscarinic functions from subchronic exposure to AN. Male Wistar rats (10/group) were
injected subcutaneously with 0, 1, or 25 mg/kg AN for 5 days/week for 8 weeks. The authors
studied the impact of AN on motor behavioral activities by using a computerized system to chart
the movements of rats within the cage. The system provided information on a total of 11 motor
behavioral parameters, such as distance traveled/time interval, resting time, time spent moving,
and number of clockwise and counterclockwise rotations. This test model can be used to detect
subtle changes of central muscarinic receptors during exposure to cholinomimetic agents.
Evaluations were carried out on each animal 1 hour after treatement with AN on day 5 of weeks
1, 2, 4, 6, and 8.
Both doses of AN were associated with marked decreases in all motor activities and a
concomitant increase in the resting time of treated rats compared with controls (Satayavivad et
al., 1998). There was a decrease in all motor parameter values at weeks 1 and 2, but most of
these effects were diminished by weeks 4 and 6. However, the effects of the high dose of AN
were more pronounced and longer lasting. The incidence of clockwise and counterclockwise
rotations in high-dose rats was reduced compared with controls throughout the study (0.7 ± 0.3
[high dose], 2.6 ± 0.8 [low dose], and 2.8 ± 0.4 [controls] counterclockwise rotations/10-minute
study period after 8 weeks).
The effects of intramuscular injection of the muscarinic receptor antagonist, atropine, and
the reversible acetylcholinesterase inhibitor, physostigmine, with and without concurrent AN
administration, also were evaluated in this system. Atropine administration (10 mg/kg) was
associated with increases in motor activity that were enhanced by AN treatment at 25 mg/kg.
Physostigmine (0.5 mg/kg) caused reductions in motor activity irrespective of AN
administration. Satayavivad et al. (1998) concluded that AN possesses cholinomimetic effects,
one of which might include the down-regulation of muscarinic receptors. This would explain the
marked increase in the response to atropine. Since AN did not inhibit the activity of
acetylcholinesterase, the cholinomimetic effect of AN might be mediated by the release of
acetylcholine from nerve endings.
Jacob and Ahmed (2003b) studied AN-induced neurotoxicity by exposing proliferating
normal human astrocytes (NHAs) in culture to 25-400 uM AN for 12 hours. Assessment was
then made on cell viability, levels of endogenous antioxidants, GSH, catalase, levels of ROS, and
secretion of tumor necrosis factor (TNF-a), a cellular marker for oxidative stress and oxidative
damage to nuclear DNA. Treatment with 25-50 uM AN had no significant effect on viability of
the astrocytes. At 100-400 uM AN, 15-42% reduction in cell viability was observed (as
indicated by trypan blue exclusion). Reduced viability was further substantiated by 8-40%
increased cytotoxicity (as indicated by leakage of LDH). The morphology of astrocytes was
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normal at concentrations up to 200 uM, but cells exposed to 400 uM AN showed a larger
number of swollen nuclei and enlarged membrane structures.
Intracellular levels of GSH were not affected at 25 and 50 uM AN. However, a
significant dose-dependent decrease in GSH was observed at 100, 200, and 400 uM AN (18, 28,
and 35% lower than controls, respectively). A concomitant increase in levels of oxidized GSH
(glutathione disulfide [GSSG]) was observed (Jacob and Ahmed, 2003b). The ratio of GSH to
GSSG was reduced from the control value of 37 to 18, 7, 3, and 2 at 50, 100, 200, and 400 uM
AN, respectively. Compared to control level, catalase activity increased 21% at 100 uM AN, but
declined to 37% below control levels at 400 uM.
Significant increases in measures of oxidative stress (four- to sevenfold increase in the
generation of ROS) and oxidative DNA damage (greater than twofold increase in
8-oxodeoxyguanosine [8-oxodG]) were observed at 200-400 uM AN. Treatment at 400 uM
significantly increased the release of the inflammatory cytokine, TNF-a, by 30% compared with
controls. The observation that compromised antioxidant defense mechanisms (depletion of
glutathione, increase in GSSG, inhibition of catalase) occurred at the same exposure
concentration as reduced cell viability supported the hypothesis that oxidative stress in astrocytes
was a possible mechanism for neurotoxic effects of AN exposure.
In a translated Chinese study (Lu et al., 2005b), levels of monoamine neurotransmitters
and their metabolites were measured in the striatum and cerebellum of the brains of male
Sprague-Dawley rats (n = 10 per group, 7 selected randomly for measurement) exposed to 0, 50,
or 200 ppm AN in drinking water for 12 weeks. The study authors estimated the administered
doses to be 4.0 and 13.5 mg/kg-day for the 50 and 200 ppm groups, respectively. Monoamine
oxidase activity in the cerebral cortex was also measured.
Compared with control values, average dopamine levels in the striatum were decreased
by 76 and 64% in the 50 and 200 ppm groups, respectively, and by 46 and 18% in the
cerebellum. The decreases in dopamine levels were statistically significant only in the striatum.
No statistically significant exposure-related changes were observed in average levels of the
dopamine metabolite, 3,4-dihydroxyphenylacetic acid, in the striatum or the cerebellum except
for increased levels (by about 32%) for the 50 ppm group in the cerebellum compared with
controls. Average concentrations of serotonin were decreased by about 38 and 49% in the
striatum in the 50 and 200 ppm groups, respectively, and by about 41 and 68% in the cerebellum.
These changes were statistically significant only in the striatum. No statistically significant
exposure-related changes were observed in the average levels of the serotonin metabolite,
5-hydroxyindoleacetic acid, in the average levels of norepinephrine in the striatum or the
cerebellum, or in the average activities of monoamine oxidase. The observed changes in the
endpoints examined in this study are of uncertain biological relevance; as such, the study does
not identify NOAELs or LOAELs suitable for health hazard identification or dose-response
assessment. However, the striatum and cerebellum are major centers for movement control,
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balance, and coordination. Thus, an abnormality in the neurotransmitters will cause malfunction
in the movement coordination of an organism. The study authors suggested that a decrease in
dopamine will reduce the environmental adaptability of rats and provide neurotransmitter
evidence for the effect of AN exposure on neurobehavior. Serotonin plays a role in the
maintenance of sleep and the emotional and psychological states of the body. Thus, reduced
serotonin level in the exposed groups may also have neurobehavioral effect.
4.5.1.1.4. Oxidative stress. Farooqui and Ahmed (1983b) demonstrated that exposure of rats to
AN resulted in depletion of GSH and induction of oxidative stress in RBCs (see Section
4.5.1.1.2). Farooqui et al. (1990) provided in vitro data that AN induced GSH depletion, lipid
peroxidation, and enhanced concentration of MetHb in rat RBCs.
In another study, the cytotoxic effect of AN-potentiated oxidative stress in rat alveolar
macrophages was investigated (Bhooma and Venkataprasad, 1997). When alveolar macrophages
isolated from male Wistar rats were incubated with 200 nM to 20 uM AN at 37°C for up to
4 hours, a dose-dependent loss of viability was observed. Incubation of alveolar macrophages
with 10 uM AN increased the release of H2O2 by 44% when compared with controls. This effect
was abolished by the addition of the antioxidant enzymes superoxide dismutase (SOD) or
catalase to the incubation medium. In addition, while exposure of alveolar macrophages to
10 uM AN resulted in 42% viability, addition of SOD exerted a marked protective effect with a
resultant viability of 79%, suggesting cell injury induced by AN is mediated by the production of
highly toxic OH radicals.
The role of oxidative stress and lipid peroxidation in the induction of the toxic effects of
AN was studied by a number of other research groups. In a study designed to evaluate the
possible role of free-radical-mediated lipid peroxidation in the etiology of AN-induced acute
adrenal necrosis, Silver and Szabo (1982) monitored the formation of MDA and conjugated
diene concentrations in the adrenal glands and other tissues (brain, liver, stomach, duodenum) of
female Sprague-Dawley rats treated with an i.v. dose of 150 mg/kg AN. Controls received i.v.
injections of aqueous 0.1% Tween 80. Rats were killed at 15, 30, 60, or 90 minutes after
injection of AN, and the liver, adrenal glands, brain, stomach, and duodenum were removed.
While no effects on the monitored parameters were observed in the mitochondria or microsomes
from adrenal gland, brain, duodenal mucosa, or glandular stomach mucosa of rats 30 minutes
after injection of AN, conjugated diene concentrations were elevated by 60% in hepatic
microsomes and 30 and 40% in gastric mitochondria and microsomes. Therefore, although lipid
peroxidation is unlikely to be involved in the pathogenesis of AN-induced adrenal necrosis, AN
can cause lipid peroxidation in other target organs.
The cytotoxicity and oxidative stress induced by AN were examined in cultured
colonocytes from male and female Sprague-Dawley rats (Mohamadin et al., 2005). These cells
were exposed to AN in the concentration range of 0.1-2.0 mM for 60 minutes or incubated with
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1.0 mM (the concentration that reduced viability by 50%) for different time intervals up to
180 minutes and then assayed for LDH leakage, cellular GSH levels, and lipid peroxidation.
Cell viability was reduced (assessed by trypan blue exclusion method) after 60-minute exposure
to 0.5 mM AN and higher; incubation with 1.0 mM reduced viability as early as 60 minutes and
by 72% at 180 minutes. Concentration-dependent increases in plasma membrane damage, as
assessed by leakage of LDH, and enhanced lipid peroxidation (production of thiobarbituric acid-
reactive substances [TEARS]) were observed following 60-minute exposures at all levels (0.1-2
mM); exposure to 1.0 mM AN produced a 2.5-fold increase in leakage after 60 minutes.
Significant reductions in glutathione levels resulted from 60-minute exposure to >1.0 mM
AN. The time course study revealed that exposure to 1.0 mM AN caused a significant decrease
in GSH concentration by 30 minutes that kept decreasing until 180 minutes, when the
experiment was terminated. In additional experiments, AN-induced membrane damage and lipid
peroxidation were reduced, but not totally abolished, by cotreatment with thiol-containing
compounds (GSH, N-acetyl-L-cysteine (NAC), and dithiothreitol (DTT)) and antioxidant
enzymes (SOD and catalase) as well as the iron chelator desferrioxamine (DFO) and the
hydroxyl radical scavenger DMSO. Mohamadin et al. (2005) suggested that the observed
protective effects of GSH, NAC, and DTT could be attributed to interaction with ROS, binding
to toxic metabolites, and/or enhancement of cellular GSH synthesis, while depletion of iron by
DFO could indirectly prevent cell damage by inhibiting the generation of hydroxyl radical.
Pretreatment of colonocytes with either SOD or catalase inhibited LDH leakage by about 23 and
54%, respectively, when compared with colonocytes treated with AN alone. These antioxidant
enzymes reduced TEARS production that was induced by AN by 17 and 45%, respectively.
Since these antioxidants could not restore the normal level of LDH leakage or TEARS
production, Mohamadin et al. (2005) concluded that in addition to lipid peroxidation, other
factors contributed to AN-induced cytotoxicity.
Mahalakshmi et al. (2003) evaluated the potential protective effect of taurine (TAU)
against AN-induced oxidative stress in rat brain. Male Wistar rats (six/group) were exposed to
0 or 100 ppm AN (average intake 8-10 mg/kg-day) in drinking water for 14 or 28 days.
Additional groups of rats received TAU (10 g/kg in diet) alone or along with AN treatment for
14 or 28 days. AN had no effect on BW gain or weights of liver or brain. AN treatment for
14 days increased levels of TEARS by about 13% in plasma and 30% in brain and increased
levels of lipid hydroperoxides by 38% in plasma and 31% in brain; values after 28 days were
similar, except for a 45% increase in lipid hydroperoxides in plasma. AN exposure also
increased the percentage of DNA fragmentation detectable in brain by 60 and 81% after 14 and
28 days, respectively. TAU completely prevented the AN-induced increases in levels of TEARS
and lipid hydroperoxides in plasma and brain, but afforded only partial protection from
AN-induced increases in DNA fragmentation, which was significantly increased to 30 and 19%
after 14 and 28 days, respectively. These results demonstrated that, even as TAU completely
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prevented the formation of ROS and oxidative stress, DNA fragmentation still occurred. Thus,
other factors in addition to oxidative stress caused DNA fragmentation as a result of AN
exposure.
AN significantly reduced activity levels of enzymatic antioxidants after 14 days (28-day
values were similar to 14-day values but slightly lower): SOD (50% lower in hemolysate, 30%
lower in brain), catalase (50% in hemolysate, 62% in brain), glutathione peroxidase (39% in
hemolysate, 60% in brain), and GST (20% in hemolysate and brain). TAU also partially
protected against these AN-induced reductions of enzymatic antioxidant activities, resulting in
reductions of less than 11% in rats treated with AN and TAU for 14 days when compared with
controls. There was insignificant reduction of <5% in rats treated with AN and TAU for 28
days. AN exposure for 14 days also lowered levels of nonenzymatic antioxidants, such as
ascorbic acid (by 30% in plasma and brain), a-tocopherol (by 40% in plasma and brain), and
GSH (by 45% in plasma and 28% in brain); reductions were slightly greater for the 28-day
exposure. In animals treated with TAU and AN for 14 days, the AN-related reductions in
nonenzymatic antioxidants were less than 10% compared with controls. These experiments
demonstrated that oral exposure to AN in drinking water increased oxidative stress in the brain
and that TAU partly protected against AN-induced oxidative stress by increasing the activities of
enzymatic antioxidants and replenishing nonenzymatic antioxidants.
Carrera et al. (2007) were not able to reproduce the results obtained by Mahalakshimi et
al. (2003) when oxidative stress parameters were measured in Wistar rats treated with AN in
vivo. Male Wistar rats (12/group) were treated with 0 or 200 ppm AN in drinking water for 14
days. The estimated daily dose was 30 mg/kg-day, using water factor of 0.15 L/kg-day (USEPA,
1988). Brains were excised and homogenized, and lipid peroxidation (as measured by nmol
MDA/mg protein), catalase activity, GSH levels, and proteins in brain tissue were measured. No
differences were found in lipid peroxidation products, catalase activity, and reduced and oxidized
GSH levels in the control and treated rats. Carrera et al. (2007) concluded that there was no
evidence of oxidative damage in the brain of AN-treated rats at the studied dose of AN.
In another study (El-Sayed et al., 2008), the effect of hesperidin (HES), an antioxidant
flavonoid, on AN-induced oxidative stress in rat brain was investigated. Male Swiss rats
(8/group) were treated with 50 mg/kg-day AN via oral gavage for 28 days (Group II). Control
rats received distilled water (Group I). Group III rats received 200 mg/kg-day HES i.p. for 28
days. Group IV rats received 200 mg/Kg-day HES i.p. 24 h before starting AN treatment and
concomitantly with AN treatment. At study termination, brain GSH content, MDA content, and
enzymatic antioxidant parameters (SOD, CAT, glutathione peroxidase (GSH-Px), and GST)
were measured. Histopathological examination was conducted on brain samples from two rats in
each group.
Brain lipid peroxides levels measured as MDA was increased by 107% in the AN-treated
rats, accompanied by a 63% decrease in brain GSH content as compared with controls (El-Sayed
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et al., 2008). On the other hand, pretreatment of rats with HES prior to AN administration
resulted in 55% reduction in brain MDA content and 183% increase in brain GSH content when
compared with the AN-treated group. In addition, significant decreases in enzymatic antioxidant
parameters were found in the brain of AN-treated rats, with SOD, CAT, GSH-Px, and GST
decreased by 43, 64, 52, and 43%, respectively, when compared with controls. Pretreatment
with HES and coadministration with AN attenuated the reduction of enzymatic antioxidants
levels, resulting in elevations of SOD, CAT, GSH-Px, and GST by 73, 169, 197, and 71%,
respectively, when compared with the AN-treated group.
Histopathological examination of brain sections indicated damage to neuronal cells of the
brain in AN-treated rats, as manifested by edema and interstitial neuronal atrophy with
perineuronal vacuolation. Pretreatment of the rats with HES nearly normalized the
histopathological changes induced by AN. This study indicated that treatment of rats with 50
mg/Kg-day AN via oral gavage induced oxidative stress in the brain, and that pretreatment with
the antioxidant HES might have protective role against AN-induced oxidative stress in the brain.
Zhang et al. (2002) studied the mechanisms by which AN induced oxidative stress and
found evidence that CYP450 metabolism is required for AN to affect the activities of antioxidant
enzymes, such as catalase, xanthine oxidase, or SOD. Syrian hamster embryo (SHE) cells were
incubated for 4, 24, and 48 hours with subcytolethal doses of AN (0, 25, 50, or 75 ug/mL), and
the effects of AN on enzymatic and nonenzymatic antioxidants were monitored. All three
concentrations of AN increased ROS (hydroxyl radicals, measured as 2,3-dihydroxybenzoic acid
formation from salicylic acid) levels in SHE cells at all time points, with concurrent depletion of
GSH after 4 hours of treatment, ranging from 80 to 66% reduction. GSH levels in all AN-
treatment groups returned to control values after 24 hours of treatment and increased only in the
SHE cells treated with 75 ug/mL AN after 48 hours of treatment.
Inhibition of the antioxidant enzymes was temporal. Decreased catalase activity was
observed following treatment with 50 and 75 ug/mL AN for 4 hours at 72 and 52%, respectively.
However, catalase activity was increased with 25, 50, and 75 ug/mL AN treatment for 24 hours
by 82, 138, and 182%, respectively. Similar increases were observed with 48 hours of treatment.
SOD activity was decreased only in SHE cells treated with 75 ug/mL AN for 4 hours.
The activity of xanthine oxidase, the enzyme that generates the superoxide radical and
hydrogen peroxide via the oxidation of hypoxanthine or xanthine by oxygen, was also
monitored. Xanthine oxidase activity was increased by 47% in SHE cells treated with 75 ug/mL
AN for 24 hours. After 48 hours of treatment, both 50 and 75 ug/mL AN significantly increased
xanthine oxidase activity. The addition of 0.5 mM of ABT (a suicide inhibitor of CYP450) to
the system prevented the AN-induced decrease in catalase activity after 4 and 24 hours.
Cotreatment of AN with ABT also blocked the increase in xanthine oxidase in SHE cells after
48 hours of treatment. ABT alone had no effect on catalase or xanthine oxidase activity. In
addition, AN had no effect on catalase or SOD activities in the absence of metabolic source
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(SHE cell homogenate). Thus, AN-induced oxidative stress in SHE cells involved decrease of
antioxidants and activation of the oxidant enzyme xanthine oxidase. Taken together, these data
suggest that AN induced oxidative stress via its oxidative metabolism.
Nerland et al. (2003) proposed another mechanism by which AN might induce oxidative
stress in their study on the covalent binding site of AN to rat liver CAIII. Two-dimensional
polyacrylamide gel electrophoresis and autoradiography were used to locate proteins in male rat
liver cytosol that were radiolabeled after s.c. administration of 115 mg/kg of [2,3-14C]-AN to
male Sprague-Dawley rats. The intensely labeled spots in the autoradiogram were identified as
CAIII. Analysis of the trypic fragment established that only cysteine 186 in the CAIII was
labeled. Thus, AN selectively bound to the cysteine 186 residue of CAIII in rat liver. Nerland et
al. (2003) also showed that over 50% of rat CAIII had participated in scavenging the toxicant
AN. CAIII has been proposed to protect cells against oxidative stress by scavenging reactive
xenobiotics, thereby reducing covalent binding to more critical macromolecules. Thus, AN
exposure would impair this protective function by covalently binding to cysteine 186 of CAIII.
In another study, the cytotoxicity of AN was related to disturbances in intracellular ionic
homeostasis and induction of oxidative stress. Mikhutkina et al. (2004) investigated the role of
disturbance in cell Ca2+ homeostasis in AN-induced blebbing of thymocyte plasma membrane
and apotosis. Blebbing of cell membrane develops in the initial stage of cell damage and is a
sign of apoptosis and necrosis. A component of apoptogenic and necrogenic factors on the cell
is Ca2+ imbalance, a result of disturbed activity of ion channels and intracellular Ca2+ stores.
n, 94-
Exhaustion of intracellular Ca stores increases the activity of store-activated (capacitance) Ca
channels that allow influx of Ca2+ into cells. Exhaustion of Ca2+ stores is a stimulus for
apoptosis.
When exposed to 5 mM AN in vitro for 1 hour, thymocytes isolated from male albino
mice exhibited twofold increases in the incidence of blebbing of the plasma membrane, followed
by apoptosis (as detected by the expression of phosphatidylserine on the outer membrane) and
necrosis (as indicated by membrane permeability to propidium iodide) (Mikhutkina et al., 2004).
The initial and terminal blebbing of the plasma membrane peaked by 15 and 45 minutes of
incubation, respectively. The dynamics of terminal blebbing correlated with the accumulation
of MDA in the incubation medium. Cotreatment with compounds (e.g., caffeine and procaine)
9-1-
that regulate activity of intracellular Ca stores modulated the cytotoxic effect of AN, suggesting
9-1-
to Mikhutkina et al. (2004) that AN induced the release of Ca from the endoplasmic reticulum.
Preincubation of thymoctyes with 10 uM isoptin, a blocker of voltage-dependent ion
channels, decreased AN-induced apoptosis and necrosis but increased the number of cells in
AN-induced secondary necrosis and decreased the intensity of oxidative stress (as indicated by
the levels of MDA). This effect was related to the ability of isoptin (an antioxidant) to suppress
AN-induced generation of free radicals. Cotreatment of AN with 10 (jM SKF 96365, a blocker
of calcium release-activated channels that were dependent on the activity of voltage-dependent
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ion channels, increased the apoptogenic, but not the necrogenic, activity of AN. (SKF96365 had
low prooxidant activity.) The cells treated with SKF 96365 and AN developed intensive
blebbing at various periods of incubation. Mikhutkina et al. (2004) interpreted the results to be
that isoptin, via K+ channels, only partially inhibited store-activated Ca2+ entry into cells through
calcium release-activated channels and in turn, protected cells from apoptosis. However,
secondary necrosis (the process associated with cells unable to form apoptotic bodies) developed
in thymocytes with suppressed blebbing. On the other hand, inhibition of these channels with
94-
SKF 96365 completely blocked Ca entry and promoted progression of apoptosis. The results
of these experiments suggested that AN caused disturbances in intracellular calcium balance that
led to plasma membrane blebbing, apoptosis, and necrosis of thymocytes.
A recent study on the effects of AN on primary-cultured astrocytes from Wistar male rats
(Carrera et al., 2007) indicated that AN-induced cellular damage was not due to oxidative stress,
since antioxidants did not prevent AN-induced toxicity. In this study, primary cultured rat
astrocytes were treated for 1 hour with or without trolox (TRX) (100 uM), TAU (5 mM), NAC
(20 mM), estradiol (10 uM), or melatonin (MEL) (10~3 M, 10~5 M, or 10~7M). They were then
exposed to 2.5 mM AN (which induced about 40% cell death) for 24 hours. Cell viability was
determined by measuring the activity of LDH released by damaged cells into the medium. GSH
levels were also measured.
Among the antioxidants included in this study, only NAC, a sulfhydryl donor, prevented
the decrease of the number of viable cells in the culture. None of the other antioxidants
prevented cell death induced by AN treatment. Moreover, 10~5 M MEL and TAU increased the
toxicity of AN in astrocytes. Treatment of astrocytes for 4 hours with AN partially depleted
intracellular GSH, whereas pretreatment with 20 mM NAC recovered GSH content to the control
levels. Carrera et al. (2007) concluded that protective effect of NAC was due to increase in the
intracellular pool of GSH, leading to an increase in GSH conjugation with AN, subsequent
decrease in availability of AN to be metabolized to CEO and cyanide, and resulting reduction in
toxicity. Carrera et al. (2007) also concluded cellular toxicity induced by AN could not be
prevented only by antioxidants.
Oxidative stress summary
The role of oxidative stress in AN-induced cytotoxicity was evaluated in both in vitro and
in vivo studies (Mohamadin et al., 2005). AN was reported to induce oxidative stress and
decreased the viability of NHAs in cell cultures (Jacob and Ahmed, 2003b) and cultured rat
colonocytes (Mohamadin et al., 2005). However, other causes contributed to AN-induced
cytotoxicity as antioxidants could not restore the normal level of LDH leakage or TEARS
production.
Zhang et al. (2002) studied the mechanisms by which AN induced oxidative stress in
SHE cells and suggested that AN induced oxidative stress via its oxidative metabolism. Other
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proposed mechanisms by which AN might induce oxidative stress included covalent binding of
AN to cysteine 186 of rat liver CAIII (Nerland et al., 2003), and disturbances in intracellular
ionic homeostasis (Mikhutkina et al., 2004).
Drinking water studies in Wistar rats yielded contradictory results. Mahalakshmi et al.
(2003) reported exposure to 8-10 mg/kg-day AN significantly reduced enzymatic antioxidant
(SOD, glutathione peroxidase, catalase) levels in rat brain after 14 or 28 days, and that TAU
prevented the AN-induced increases in levels of TEARS and lipid hydroperoxides in plasma and
brain. However, Carrera et al. (2007) were unable to reproduce the results obtained by
Mahalakshimi et al. (2003) in another 14-day drinking water study in Wistar rats. No differences
were found in lipid peroxidation products, catalase activity, and reduced and oxidized GSH
levels in the control and treated rats. Carrera et al. (2007) concluded that there was no evidence
of oxidative damage in the brain of rats treated with 30 mg/kg-day AN for 14 days.
In an oral gavage study of male Swiss albino rats, treatment of rats with 50 mg/Kg-day
for 28 days induced increased levels of MDA in the brain, and decreased brain GSH content and
enzymatic antioxidant levels (El-Sayed et al., 2008). Pretreatment with HES (200 mg/Kg, i.p.)
attenuated the effects of AN treatment.
4.5.1.1.5. Immunotoxicity. Zabrodskii et al. (2000) studied the mechanisms for cell and
humoral immunosuppressive effect of AN. CBA mice were treated by a single s.c. injection at
one-half the LD50 (28 mg/kg) of AN. One day after treatment, mice were tested for delayed-type
hypersensitivity (DTH) reactions: a-naphthylbutyrate esterase activity in splenocytes, number of
antibody-producing cells in spleen, and the paw edema test. AN treatment suppressed primary
cell immune response, as demonstrated by a 61% reduction of the DTH reaction (edematous paw
weight). Secondary DTH response (the number of esterase-positive splenocytes) was reduced by
28%. Humoral immune response was also decreased, as shown by a 56% reduction of the
number of antibody-producing cells. The cholinesterase reactivator, bispyridinium dioxime
(dipyroxime), rehabilitated paw swelling completely but only partially restored the numbers of
antibody-producing cells and esterase-positive splenocytes.
The role of cytochrome c oxidase a3 in the mitochondrial respiration enzyme system of
immunocompetent cells was evaluated using hydrogen cyanide, a metabolite of AN, and its
antidote anticyan. Anticyan, an agent that converts Hb to MetHb, also improved the DTH effects
of AN but less efficiently than dipyroxime. A combination of both agents reversed the
immunotoxic effects of AN completely. Zabrodskii et al. (2000) suggested that AN-induced
immunotoxic effects were the result of its anticholinesterase activity targeted at T lymphocytes
and general toxicity associated with inhibition of cytochrome c oxidase a3 of the immunocytes.
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4.5.1.2. Cancer Effects
4.5.1.2.1. Formation ofDNA adducts. There is evidence that AN and its epoxide metabolite,
CEO, can form a number of different DNA adducts in vitro. Solomon and Segal (1985)
demonstrated the nonenzymatic alkylation of calf thymus DNA by AN. Following a 40-day
incubation of AN at 37°C and pH 7.4, the reaction products were identified as cyanoethyl
adducts of guanine and thymine and carboxyethyl adducts of adenine and cytosine. The major
adducts were 1-carboxyethyl adenosine (26%), 7-cyanoethyl-guanine (26%), imidazole ring-
opened 7,9-bis cyanoethyl guanine (19%), 3-cyanoethyl thymine (16%), and N6-cyanoethyl
adenine (8%).
CEO formed a number of DNA adducts more quickly when incubated with DNA in vitro
(Solomon et al., 1993; Yates et al., 1993; Hogy and Guengerich, 1986). When calf thymus DNA
was incubated with CEO at pH 7.0-7.5 and 37°C for 3 hours (Solomon et al., 1993),
N7-(oxoethyl)guanine (110 nmol/mg DNA), N3-(2-hydroxy-2-carboxyethyl)deoxyuridine (80
nmol/mg DNA), and smaller amounts of adenine and thymine adducts were produced. The
adducts formed from CEO were different than those formed from AN. The order of reactivity
with CEO was guanine > cytosine > adenine > thymine. In addition to reacting with N7 of
guanine to form N7-(oxoethyl)guanine, CEO reacted to a great extent with the cytosine residue,
resulting in the detection of a major adduct, N3-(2-hydroxy-2-carboxyethyl)deoxyuridine.
Solomon et al. (1993) proposed that the uracil adduct was formed from an initial cytosine adduct
through hydrolytic deamination of cytosine to uracil.
Yates et al. (1993) also characterized an adduct formed when calf thymus DNA was
incubated with 150 mM CEO at 37°C for 3 hours. The adduct was identified as N3-(2-cyano-
2-hydroxyethyl)deoxythymidine. This adduct was also formed when 150 mM [2,3-14C]-CEO
reacted with 10 mM deoxythymidine in vitro (Yates et al., 1993). Subsequent degradation of this
adduct yielded N3-(2,2-dihydroxyethyl)deoxythymidine.
Guengerich et al. (1981) showed that when 1 mM [1-14C]-AN and [2,3-14C]-AN were
incubated at 37°C with 1.5 mg/mL calf thymus DNA in the presence of rat liver microsomes or a
reconstituted CYP450 system, DNA adducts were formed as measured by irreversible binding of
radioactive label to DNA. Formation of DNA adducts was enhanced by the presence of
NADPH. Only trace levels of DNA adducts were detected when incubated with rat brain
microsomes, and the reaction was not NADPH dependent, probably due to insignificant
metabolism of AN by brain microsomes. Guengerich et al. (1981) also showed nonenzymatic
irreversible binding of labeled CEO to calf thymus DNA. The extent of binding for
[2,3-14C]-labeled CEO was three- to fivefold greater than that for [l-14C]-labeled CEO.
Moreover, when 100 mM CEO was incubated with 50 mM adenosine for 24 hours at 37°C,
l,N6-ethenoadenosine was formed. In another experiment, an unidentified product was formed
when CEO was incubated with cytidine.
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Data from Peter et al. (1983a) showed the extent of irreversible binding of AN or its
metabolites to DNA to be lower after purification of isolated DNA and RNA by column
chromatography on hydroxyapatite. In their in vitro incubation experiment, [2,3-14C]-AN was
incubated with rat-liver microsomes, NADPH and DNA or RNA. Irreversible AN binding was
found to be 3 nmol/hour per mg DNA, when ethanol precipitation or phenol extraction of DNA
was used. When the isolated DNA was further purified by column chromatography on
hydroxyapatite, irreversible binding was found to be 0.15 nmol/hour per mg DNA.
Peter et al. (1983a) also administered [2,3-14C]-AN intraperitoneally to male Wistar rats
and measured the incorporation of radioactivity into hepatic RNA bases. The amount of
radiolabel from AN associated with hepatic DNA was lower than that from labeled vinyl
chloride. When hepatic DNA from these treated rats was isolated and hydrolyzed, however,
chromatography on PEI-cellulose showed two 14C peaks that did not correspond to known
standards. Thus, Peter et al. (1983a) concluded that AN or its metabolites could alkylate DNA,
although these DNA adducts had not yet been identified.
Yates et al. (1994) also characterized the products formed when 150 mM CEO reacted
with 50 mg/mL nucleotides for 3 hours at 37°C in vitro. The reaction of CEO with
5'-monophosphates of deoxyguanosine, deoxyadenosine, deoxycytidine, or deoxythymidine
resulted in the formation of at least one adduct for each nucleotide. These CEO-nucleotide
adducts were characterized as 2-cyano-2-hydroxyethyl phosphodiesters. The reaction of
deoxyguanosine-5'-monophosphate (dGMP) also produced a second adduct, N7-(2-cyano-
2-hydroxyethyl)-dGMP. Yates et al. (1994) suggested that the cy ano-hydroxy ethyl-
phosphodiester adduct could induce single and double DNA strand breaks (as observed when
CEO was incubated with pBR322 plasmid DNA) via interaction of the adduct's p-hydroxyl-
group with the DNA phosphate backbone.
Irreversible binding of AN or its metabolites to DNA in vivo has also been studied.
Farooqui and Ahmed (1983a) investigated the ability of [2,3-14C]-AN or its metabolites to bind
to macromolecules in rats in vivo. [2,3-14C]-AN was administered in a single dose to male
Sprague-Dawley rats (three to four per group) via gavage at a dose of 46.5 mg/kg in water.
Animals were sacrificed at 1, 6, 24, and 48 hours after dosing. Organs, including liver, kidney,
brain, spleen, and stomach, were dissected out and frozen rapidly. Nucleic acids extracted from
the homogenates were applied to the hydroxyapatite column for separation of RNA and DNA
fractions. Radioactivity was detected in the extracts of RNA and DNA from liver, stomach, and
brain. Bound radioactivity (as pmol equivalent AN/mg DNA) to DNA in the brain was the
highest at 24 hours (119 pmol/mg DNA), followed by stomach (81 pmol/mg DNA), and liver
(25 pmol/mg DNA). Bound radioactivity in all three organs plateaued at 24 hours and remained
unchanged thereafter. RNA from liver showed the highest amount of radioactivity, followed by
brain and stomach. Bound radioactivity in liver RNA peaked at 6 hours after dosing, whereas
RNA from the brain and stomach showed maximal radioactivity at 24 hours. Binding to proteins
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was extensive and time dependent. In the first hour after oral dosing, the highest protein binding
occurred in spleen and stomach, followed by liver, kidney, and brain. After 6 hours, protein
binding plateaued until 48 hours. At 6 hours, the highest protein binding occurred in the spleen,
followed by liver, stomach, and kidney, with brain having the lowest protein binding. The study
authors developed numerical indices for the AN-derived DNA alkylation; covalent binding
indices were 5.9, 51.9, and 65.3 for liver, stomach, and brain, respectively, after 24 hours.
Ahmed et al. (1992a) demonstrated the covalent binding of radiolabel from [2,3-14C]-AN
to testicular DNA of male Sprague-Dawley rats after a single oral dose of 46.5 mg/kg of [2,3-
14C]-AN. In a time course study, bound activity was shown to be greatest after 30 minutes (8.93
± 0.80 umol AN bound/mol nucleotide), with covalent binding index of 10.15. Using an
identical experimental protocol, Ahmed et al. (1992b) demonstrated the capacity of AN to bind
covalently to DNA in the lung. Covalent binding of radioactivity to DNA increased with time
and was maximum at 12 hours after a single oral dose, with a covalent binding index of 3.48.
Binding was associated with a 55-72% decrease in replicative DNA synthesis at time points up
to 24 hours after dosing.
Hogy (1986) and Hogy and Guengerich (1986) studied the in vivo interaction of AN and
CEO with DNA. Male F344 rats (3/group) were administered AN (50 mg/kg, i.p.) or CEO (6
mg/kg, i.p.) and were sacrificed after 2 hours. The brain and livers were removed and frozen,
and DNA and RNA were isolated from these tissues. For detection of N7-(2-oxoethyl)guanine,
DNA samples were reductively tritiated with NaB3H4; the adduct was released by neutral
thermal hydrolysis and purified by thin-layer chromatography for subsequent quantitation by
liquid scintillation counting. N7-(2-oxoethyl)guanine was detected in liver DNA at 3.1 x 107
nucleotides per alkylation in AN-treated rats and 6.9 x 107 nucleotides per alkylation in CEO-
7 8
treated rats (Table 4-53). For brain DNA, N -(2-oxoethyl)guanine was detected at 2.4 x 10
nucleotides per alkylation in AN-treated rats and at 1.1 x 109 nucleotides per alkylation for CEO-
treated rats. The values obtained from the brain DNA samples were near the limit of detection.
Thus, the presence of these adducts in the brain could not be unequivocally verified.
Table 4-53. Detection of N7-(2-oxoethyl)guanine after i.p. administration of
50 mg/kg AN or CEO to male F344 rats
Compound
AN
CEO
Formation of N7-(2-oxoethyl)guanine
Liver
fmol/mg DNA
109 ±71
48 ±15
Nucleotides/alkylation
3.1x 107
6.9 x 107
Brain
fmol/mg DNA
14
3
Nucleotides/alkylation
2.4 x 108
1.1 x 109
Source: Hogy (1986).
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In the same study (Hogy, 1986; Hogy and Guengerich, 1986), rat DNA samples were
also analyzed by HPLC with a fluorescence detector for l,N6-ethenoadenosine and
l,N6-ethenodeoxyadenosine. These adducts were not detected with limits of detection estimated
to be 3 pmol/mg DNA and 1 pmol/mg RNA, respectively.
In another study, Prokopczyk et al. (1988) administered 50 or 100 mg/kg s.c. AN to male
F344 rats (10/group). DNA was isolated from liver and brain after 2 hours (50 mg/kg group) or
6 hours (100 mg/kg AN). An HPLC assay with fluorescence detector was used to detect
7-(2-cyanoethyl)guanine (detection limit: 1 per 5 x 104 guanine) and O6-cyanoethylguanine
(detection limit: 1 per 7 x io4 guanine). Neither adduct was detected. These two adducts were
not formed when CEO was incubated with calf thymus DNA in vitro (Solomon et al., 1993).
4.5.1.2.2. Oxidative stress. Jiang et al. (1998) evaluated the ability of AN to induce oxidative
stress in the brain cortex of rats. Male Sprague-Dawley rats (at least nine/group) were exposed
to 0, 5, 50, 100, or 200 ppm AN in drinking water for either 14, 28, or 90 days. These
concentrations were selected from those used in chronic bioassays (see Section 4.2.1.2). As
calculated by the study authors, average daily doses of 0, 0.6, 5.1, 8.9, or 15.0 mg/kg-day were
ingested over the 90-day treatment period by the control to high-dose groups, respectively. At
study termination, brains and livers were weighed. The brain cortexes and livers from six
rats/group were evaluated for the following oxidative endpoints: oxidative DNA damage
(8-hydroxy-2'-deoxyguanosine levels), lipid peroxidation (MDA levels), levels of nonenzymatic
antioxidants (glutathione and vitamin E), and the activities of enzymatic antioxidants (catalase,
SOD, and glutathione peroxidases). 8-Hydroxy-2'-deoxyguanosine has also been referred to as
8-oxo-7,8-dihyro-2'-guanosine (Murata et al., 2001) and 8-oxodeoxyguanosine (Whysner et al.,
1998a) and will be referred to as 8-oxodG in this document. ROS formation, as measured by the
formation of 2,3-dihydrobenzoic acid from salicylic acid in brain cortex and liver, were
evaluated in three rats/group injected with salicylic acid in saline 12 hours before termination.
In rats exposed to AN at concentrations as high as 200 ppm for up to 90 days, no effects
were noted on viability. A statistically significant reduction of 9% in BW was observed in the
200 ppm group after 90 days. No differences were observed in brain or liver weights nor were
any of the measures of oxidative damage or antioxidant levels in the liver for all groups at any
sampling times. AN exposure resulted in statistically significant increases in oxidative stress
parameters and decreases in antioxidants levels in the brain cortex. Levels of hydroxy free
radicals (ROS) were significantly elevated in a dose-related manner in the 50-200 ppm groups,
beginning at 14 days and persisting until the 90-day time point; at 200 ppm, the increase was
approximately fivefold over controls. Levels of 8-oxodG in the cellular DNA of brain cortex
were significantly elevated (three- to fourfold compared with controls) in a dose-related manner
in the 100 and 200 ppm groups, beginning at 14 days of exposure, and two- to threefold in the
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50 ppm group, beginning after 28 days of exposure. MDA levels were significantly increased
(1.5-fold) in the 200 ppm group after 14 days but at no other time point.
Slight but significant decreases in levels of vitamin E (about 20%) and glutathione (about
20%) in the brain cortex were observed in the 50-200 ppm groups but were statistically
significant only at the 14-day time point. Statistically significant, dose-dependent reductions (by
up to 60%) in catalase activity were observed in brain cortex at all exposure levels after 14 days,
at >100 ppm after 28 days, and at >50 ppm after 90 days; the transient effect at 5 ppm was not
considered to be biologically significant. Reductions in SOD levels in the brain (by up to 30% at
200 ppm) were dose related and statistically significant at>50 ppm after 14 days and only at 200
ppm after 28 or 90 days. In this study, a NOAEL of 5 ppm (0.6 mg/kg-day) and a LOAEL of 50
ppm (5.1 mg/kg-day) were identified for increases in levels of oxidative damage (increased ROS
and DNA damage) and reductions in the antioxidant enzymes catalase and SOD in the brain of
rats exposed to AN in drinking water. Jiang et al. (1998) suggested that observed oxidative
stress in rat brain cortex following AN treatment could be induced by: (1) generation of free
radicals from AN or its metabolites, (2) binding of AN or its metabolites to free radical
scavengers (e.g., GSH, vitamin E), (3) modulation of the activity and/or synthesis of antioxidant
enzymes (e.g., SOD, catalase), and/or (4) interference with electron flow through the respiratory
chain via inhibition of cytochrome C oxidase by the cyanide ion, a metabolite of AN. However,
these potential mechanisms need to be further investigated.
In a follow up study, Pu et al. (2009) examined the potential for AN to induce oxidative
DNA damage in rats. These investigators also examined whether blood could serve as a
surrogate for the biomonitoring of oxidative stress induced by AN in target tissues (in particular
brain) of exposed populations. Male Sprague-Dawley rats (9/group) were treated with 0, 3, 30,
100, or 200 ppm AN in drinking water for 28 days. N-acetyl cysteine (NAC), an acetylated
precursor of glutathione, was coadministered at a dietary concentration of 0.3% to one group of
rats receiving 200 ppm AN in drinking water to evaluate its protective effect against potential
AN-induced oxidative stress. At the end of treatment, animals were sacrificed and blood
samples were collected immediately. The standard alkaline comet assay was used as a measure
of DNA damage in the brain cortex and white blood cells (WBCs) of 3 rats/group. The standard
comet assay detects single and double strand breaks, cross links, oxidative DNA damage,
apurinic/pyrimidinic sites and DNA repair (Smith et al., 2006; Collins, 2007). The
formamidopyrimidine DNA glycosylase (fpg)-modified comet assay was used as a measure of
oxidative DNA damage in the brain cortex and WBCs of different treatment groups. 8-oxodG
levels in brain tissues and WBCs were also measured by HPLC with electrochemical detection.
To detect the presence of ROS, 2,3-dihydroxybenzoic acid (2,3-DHBA) was measured in WBCs
and brain.
Pu et al. (2009) observed no increase in DNA damage in rat WBC and brain tissue
following exposure to AN with or without NAC coadministration using the alkaline comet assay.
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Dose-dependent increases in DNA damage were observed in rat WBC and brain following
exposure to AN using the fpg-modified alkaline comet assay. These increases in DNA damage
were not observed in WBC and brain of rats exposed to 200 ppm AN and NAC. Pu et al. (2009)
interpreted the results as indicative of oxidative DNA damage in AN-exposed rats, and further
that the absence of oxidative DNA damage in rats coadministered NAC reflected the antioxidant
action of NAC. Significant increases in the level of 8-oxodG were found in WBC and brain of
rats treated with 100 and 200 ppm AN; no increase was observed in rats treated with 200 ppm
AN and NAC. 2,3-DHBA was increased in the WBC of rats treated with AN in a dose-
dependent manner, but was not detectable in the brain of treated rats. In addition, the ratios of
reduced to oxidized glutathione (GSH/GSSG) were reported to be significantly lower in rats
treated with 30, 100 and 200 ppm AN, but not in a dose dependent manner. Pu et al. (2009)
concluded that AN induced oxidative stress and DNA damage in male Sprague-Dawley rats and
that the fpg-modified comet assay in WBC correlated with target tissue oxidative DNA damage.
Whysner et al. (1998a) also evaluated oxidative DNA damage (increased levels of
8-oxodG) in the brains of male Sprague-Dawley and F344 rats exposed to AN in drinking water
for durations of either 21 or 94 days. In the first experiment, male Sprague-Dawley rats
(20/group) were exposed to 0, 3, 30, or 300 ppm AN in drinking water for 21 days (Whysner et
al.,1998a). Based on default values for the BW of male Sprague-Dawley rats in a subchronic
study (0.267 kg) and an allometric equation linking water consumption to BW (U.S. EPA, 1988),
the average intakes of AN were calculated as approximately 0, 0.43, 4.3, and 43 mg/kg-day for
the control to high-dose groups, respectively. At termination, brains, livers, and forestomachs
were excised from five rats/group each for DNA isolation. These tissues were analyzed for
GSH, cyst(e)ine, and 8-oxodG levels in nuclear DNA. TEARS levels in brains were determined
for another five rats/group. Brain homogenate from another five rats/group were analyzed for
cytochrome oxidase activity (in the mitochondria fraction), catalase (in supernatant), and
glutathione peroxidase (in homogenate). Tissues from another five rats/group were examined for
histopathology, but results were to be published separately.
In rats exposed to 30 or 300 ppm AN, statistically significant increases in 8-oxodG levels
in nuclear DNA were measured in brains (approximately twofold increase compared with control
values for both dose groups) and livers (approximately 1.4-fold increase for both dose groups
when compared with controls). Also observed in the 300 ppm dose group were increases in the
level of cyst(e)ine in the forestomach (approximately twofold) and brain (approximately 50%
increase). There were no exposure-related effects on levels of glutathione in the brain and liver
or on levels of cytochrome oxidase, catalase, glutathione peroxidase, or TEARS in the brain.
However, glutathione level in the forestomach was increased about 1.8-fold in the 300 ppm dose
group compared with controls. In this study, a NOAEL of 3 ppm (0.43 mg/kg-day) and a
LOAEL of 30 ppm (4.3 mg/kg-day) were identified for increased oxidative DNA damage in the
brains and livers of Sprague-Dawley rats. Whysner et al. (1998a) concluded that the formation
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of 8-oxodG from AN exposure did not involve disruption of antioxidant defense or lipid
peroxidation. In addition, the absence of effect on brain cytochrome oxidase activity in exposed
rats indicated lack of inhibition by cyanide, a metabolite of AN. Thus, Whysner et al. (1998a)
concluded that cyanide-induced metabolic hypoxia did not appear to be involved in the
generation of ROS by AN administered in drinking water.
Whysner et al. (1998a) conducted a parallel 21-day drinking water study in male F344
rats (10/group). Concentrations of AN in drinking water were of 0, 1, 3, 10, 30, or 100 ppm.
Based on default values for the BW of male F344 rats in a subchronic study (0.18 kg) and an
allometric equation linking water consumption to BW (U.S. EPA, 1988), the average doses of
AN were estimated at 0, 0.16, 0.47, 1.6, 4.7, and 15.6 mg/kg-day in the control to high-dose
groups. An additional group of rats received 5 mg methylnitrosourea (MNU) per kg/week via
i.v. injection. At termination, rat brains were evaluated for levels of 8-oxodG in nuclear DNA,
cytochrome oxidase, glutathione, and cyst(e)ine. Levels of all these parameters in the brains of
AN-exposed rats and MNU-exposed rats were not significantly different from controls. The
levels of 8-oxodG in the brains of 3-100 ppm AN dose groups were the same (about 1.3-fold
higher than control values). However, the increases were not statistically significant. The
highest drinking water concentration of 100 ppm AN (15.6 mg/kg-day) was a NOAEL for
oxidative effects in the brains of F344 rats.
In the subchronic experiment by Whysner et al. (1998a), male Sprague-Dawley rats
(10 rats/group) were exposed to 0 or 100 ppm AN in drinking water for 3, 10, 31, or 94 days.
Two additional groups of Sprague-Dawley rats (six/group) were exposed to 5 mg MNU/kg/week
or 5 mg MNU/kg/week +100 ppm AN. Brains were assayed for levels of 8-oxodG, cytochrome
oxidase, and glutathione. Levels of 8-oxodG were also measured in the livers of rats exposed for
3, 10, or 94 days.
No effects on brain levels of cytochrome oxidase or glutathione were observed in any
dose group up to 94 days. The 8-oxodG levels were significantly increased by 77% following
exposure to 100 ppm AN for 3, 10, and 94 days. Administration of 5 mg/kg MNU (a DNA-
reactive carcinogen that produces glial cell tumors in rats) did not increase the level of 8-oxodG
in the brain of treated rats but increased 8-oxodG in the liver after 10 days. However,
coadministration of 100 ppm AN and MNU increased the 8-oxodG level in the brain after 31 and
94 days when compared with controls. Whysner et al. (1998a) proposed that AN-induced
generation of ROS and resultant oxidative DNA damage represented one possible mode of action
for the neoplastic process in the rat brain. However, as discussed in Section 4.7.3.3.1, results
from this study did not support the proposed oxidative stress mode of action.
The ability of AN to induce oxidative stress and oxidative DNA damage was also studied
in a rat glial cell line and cultured rat hepatocytes in vitro (Kamendulis et al., 1999a). In parallel
experiments, DITNC1 rat astrocytes and rat hepatocytes were incubated with sublethal
concentrations (up to 1 mM) AN in vitro for 4 or 24 hours. The following were measured at the
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end of the incubation period: the 8-oxodG levels in total cellular (nuclear and mitochondrial)
DNA; generation of ROS (measured as an increase in 2,3-DHBA); and levels of MDA, GSH,
and antioxidant enzymes activity.
In rat astrocytes, significant increases (up to 3.8-fold) in the production of 8-oxodG over
control was observed after 4 hours and up to 3.9-fold after 24 hours. No increase in 8-oxodG
formation was observed in rat hepatocytes at all AN concentrations and time examined. The
induction of 8-oxodG by AN in rat astrocytes was reversible. Following removal of AN after
24 hours of treatment, 8-oxodG levels returned to control values at all studied concentrations in
24 hours. Intercellular production of ROS was found to be increased 2- to 2.6-fold after 4 and
24 hours at 0.1 and 1.0 mM AN. No increase in ROS generation was found in rat hepatocytes at
any AN concentration or exposure duration. No significant change in MDA formation (as
indicator of lipid peroxidation) was found in either cell type following treatment with AN.
A significant decrease in cellular GSH levels was observed in rat astrocytes treated with
0.1 and 1.0 mM AN for 4 hours (25-36% of control) and 24 hour (43-61% of control); and in
SOD activity in astrocytes treated for 4 hours with 1 mM AN (39% reduction over control) and
for 24 hours with 0.1 and 1 mM AN (38-40% reduction over control). No significant decrease
in catalase and glutathione peroxidase activities was observed in treated astrocytes. Cotreatment
with L-2-oxothiazolidine-4-carboxylic acid (OTC), a precursor to GSH biosynthesis, or with
vitamin E, an antioxidant, reduced 8-oxodG and ROS formation induced by AN treatment.
These effects were not evident in isolated hepatocytes treated with AN. Results from this study
were in agreement with results from the in vivo study by Jiang et al. (1998) on rat brain cortex.
Since the formation of 8-oxodG and ROS observed after AN treatment in this study was
temporal, dose dependent, and reversible following removal of AN in the culture medium,
Kamendulis et al. (1999a) suggested that these are established properties of tumor-promoting
agents. Kamendulis et al. (1999a) proposed that AN-induced astrocytomas in rats were produced
via tumor promotion mechanisms.
Jacob and Ahmed (2003b) also demonstrated the ability of AN to induce oxidative stress
and oxidative DNA damage in NHA culture (see Section 4.5.1.1.3).
Pu et al. (2006) measured direct and oxidative DNA damage in cultured Dl TNC1 rat
astrocytes treated with 0-2.5 mM AN for 24 hours. Direct DNA damage was measured by the
standard alkaline comet assay and oxidative DNA damage was measured with a fpg-modified
comet assay. 7-Ethoxyresorufm-O-deethylase (EROD) and CYP2E1 activities in the astrocytes
were measured.
At 2.5 mM AN, a 40% decrease in cell viability was observed. No increase in direct
DNA damage (measured as increase in tail moment) was observed at any AN concentration.
Hydrogen peroxide (20 uM) was used as positive control, and significant increase in DNA
damage was found. On the other hand, a threefold increase in oxidative DNA damage was
observed in astrocytes treated with 1 mM AN. Supplementation of 1 mM AN with three
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different antioxidants—vitamin E (150 uM), TRX (water-soluble analog of vitamin E, 150 uM),
and epigallocathechin-3 gallate (a polyphenol from green tea, 5 uM)—reduced AN-induced
oxidative DNA damage by 65, 54, and 65%, respectively.
As expected, only low level CYP2E1 activity was measured in the astrocytes and was
about 10% of that measured in mouse liver. EROD activity was measured in rat astrocytes as an
indicator of CYP1A and was about 500-fold higher than that for CYP2E1. In addition, a 90%
reduction in the activities of EROD and CYP2E1 was observed when the astrocytes were
cotreated with AN and 0.5 mM ABT for 24 hours. Cotreatment of astrocytes with 1 mM AN and
0.5 mM ABT also prevented the increase in oxidative DNA damage induced by AN, indicating
the metabolism of AN by P450 was required for the production of oxidative stress. GSH
depletion induced by 4 and 24 hours of treatments with DL-buthionine [S,R]-sulfoximine, a
selective inhibitor of y-glutamylcysteine synthetase, enhanced the oxidative DNA damage
induced by AN by 44-160% over control. On the other hand, cotreatment with 2.5 mM OTC, a
precursor for GSH biosynthesis, reduced AN-induced oxidative DNA damage by 63-85%.
Exposure to 0.1-0.5 mM cyanide also increased oxidative DNA damage (Pu et al., 2006).
Murata et al. (2001) investigated the enhancing effect of AN on the formation of
8-oxodG in calf thymus DNA, induced by hydrogen peroxide (H2O2) and Cu(II). Calf thymus
DNA was incubated with various concentrations of H2O2plus CuCb in the presence or absence
of AN for 30 minutes. The level of Cu(II)-mediated 8-oxodG formation increased with
increasing concentration of H2O2. The addition of AN (0.1-0.5%) enhanced the formation of
8-oxodG by hydrogen peroxide and Cu(II) in a dose-dependent manner, whereas AN itself did
not cause DNA damage. The enhancing effect of AN was more marked in double-stranded than
^9
in single-stranded DNA. Further experiments with [ P]-labeled DNA showed that addition of
AN enhanced the site-specific DNA damage at guanine residues, particularly at the 5'-site of the
GG and GGG sequences while H2O2/Cu(II) induced piperidine-labile sites at thymine, cytosine,
and guanine residues. Electron spin resonance spectroscopy showed that a nitrogen-centered
radical was generated from AN during incubation with hydrogen peroxide and Cu(II). Murata et
al. (2001) proposed that AN enhanced H2O2-mediated DNA damage via nitrogen-centered
radical formation. Thus, AN may enhance endogenous oxidative stress, although AN itself does
not have the ability to induce oxidative DNA damage.
4.5.1.2.3. Intercellular communication. Kamendulis et al. (1999b) investigated the effect of
AN on gap junction intercellular communication (GJIC) in Dl TNC1 astrocytes (a rat astrocyte
transformed cell line) and primary rat hepatocytes in culture. Noncytolethal concentrations of
AN (0.01-1.0 mmol/L) were incubated with Dl TNC1 astrocytes or primary rat hepatocytes.
GJIC was determined by microinjection of lucifer yellow CH into cells. Dye coupling was
quantitated by determining the number of recipient cells in contact with microinjected cells that
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showed communication. The reversibility of AN inhibition of GJIC was also evaluated by
replacement of AN with fresh medium.
Following 2 hours of treatment, AN at 0.1 and 1 mmol/L inhibited GJIC in Dl TNC1
astrocytes, which were putative target cells. After treatments for 4 and 24 hours, all
concentrations of AN (0.01-1.0 mmol/L) inhibited GJIC. The inhibitory effect of AN in this
system was dose dependent at all time points, reversible by removal of AN, and partially
suppressed by the presence of antioxidants such as vitamin E (0.1 mmol/L). After treatment with
both vitamin E and AN for 24 hours, inhibition of GJIC was reduced by 23%. On the other
hand, AN did not inhibit GJIC in primary cultured hepatocytes at all concentrations and
durations of treatment.
AN also caused a concentration-dependent decrease in cellular GSH content in both
Dl TNC1 astrocytes and rat hepatocytes (Kamendulis et al., 1999b). Cotreatment with
5 mmol/L OTC (a precursor of GSH synthesis) reduced inhibition of GJIC by AN in Dl TNC1
astrocytes following 4 and 24 hours of exposure, with the greatest reduction observed at
1.0 mmol/L AN (up to 68%). However, depleting GSH by L-BSO (an inhibitor of intracellular
GSH synthesis) alone without AN did not affect GJIC in rat astrocytes. Thus, depletion of GSH
alone in astrocytes was not sufficient for the observed decrease in GJIC by AN.
Inhibition of intercellular communication by AN was implied in an inhibition of
metabolic cooperation assay. Two studies (Elmore et al., 1985; Umeda et al., 1985) evaluated
AN by using 6-thioguanine (6-TG) sensitive wild type Chinese hamster V79 cells and 6-TG-
resistant cloned cells that are the hypoxanthine guanine phosphoribosyl transferase (hprT)
mutant of the V79 cell line. The hprT mutant cannot phosphorylate several purine analogues,
including 6-TG, and are therefore resistant to cell killing by the purine analogue. When the WT
cells are cultured at a density that permits frequent contact with the mutant cells, metabolic
cooperation (i.e., gap junction formed between these cells) allows the transfer of nutrients and
phosphorylated purine analogue from the WT cells to the mutant cells and decreases probability
of recovery of the mutant cells in purine analog selective medium. The principle of the assay is
based on the fact that recovery of 6-TG-resistant cells cocultivated with 6-TG-sensitive cells in
6-TG-containing medium increased by addition of compounds that inhibit metabolic
cooperation. This assay evaluates if the test agent can modulate gap junctional communication.
AN inhibited gap junction formation slightly and dose dependently (Umeda et al., 1985).
Average recovery of the mutant cells was 22% in the control, 33% at 1 mM AN, and 39% at
2 mM AN. Umeda et al. (1985) considered AN as positive in the metabolic cooperation assay.
In the study by Elmore et al. (1985), AN produced positive responses at noncytotoxic
concentrations of 10-50 ug/mL after incubation for 3 days.
4.5.1.2.4. Cell proliferation. Ghanayem et al. (1997) examined the effects of AN on
forestomach cell proliferation and apoptosis in male F344 rats (12/group) administered either 0,
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0.22, or 0.43 mmol/kg (0, 11.7, or 22.8 mg/kg) by gavage for 6 weeks. Six rats from each dose
group were used to assess BrdU incorporation in the stomach, the remaining six rats from each
group were used to assess BrdU incorporation in hepatocytes. Proliferation of forestomach
squamous epithelial cells was evaluated with light microscopy by determining the number of
cells (nuclei) per unit length muscularis and by quantitating BrdU-stained cells.
AN was shown to induce a dose-dependent increase in epithelial cell proliferation in the
forestomach, as determined by the incorporation of BrdU into S-phase DNA. The increase in
forestomach mucosal cell proliferation was significant at both the low and high doses of AN. No
cellular proliferation was detected in the liver and glandular stomach, which are not target organs
of AN carcinogenicity.
Hyperplasia, a possible indicator of enhanced cell proliferation, was significantly
increased in the forestomach squamous mucosa of the high-dose group (about 60% above
vehicle-treated controls); the increased in the low-dose group (7% above controls) was not
statistically significant. The effects of AN on cellular proliferation in the forestomach of treated
rats were also evaluated by quantitative determination of BrdU incorporation into S-phase DNA,
using immunohistochemical staining. The increase in forestomach mucosal cell proliferation
was significant at both the low- and high-dose groups. In addition, the effect of AN on apoptosis
was determined by in situ end labeling of tissue sections. Apoptotic bodies were observed in the
forestomach of rats treated with high-dose AN. No increase in apoptosis was detected in the
liver or glandular stomach of control or treated rats. Thus, AN induced a significant increase in
forestomach apoptosis at the high dose, coupled with an increase in hyperplasia.
Ghanayem et al. (1997) proposed that disruption of the normal balance between cell
proliferation and apoptosis in favor of enhanced forestomach cell proliferation (as reflected by
hyperplasia of the forestomach epithelium) probably contributed to the pathogenesis of
AN-induced forestomach tumors. This suggestion was supported by the observations that cell
proliferation in forestomach squamous mucosa of treated rats occurred at doses that caused
forestomach tumors in rats and that cell proliferation was selective and only occurred in the
target organ forestomach but not in liver.
In a recent study, Chantara et al. (2006) evaluated whether AN induced extracellular
signal-regulated kinase (ERK) activation in human neuroblastoma SK-N-SH cells. The
activation of ERKs belonging to the mitogen-activated PK family has been implicated to play
crucial roles in cell proliferations and is involved in many steps of tumor progression (Fang and
Richardson, 2005; Platanias, 2003; Seger and Krebs, 1995). Dysfunction of the ERK signaling
pathways was shown to play a pivotal role in the development of many cancers, including
leukemia and colon cancer. Active forms of ERK1/2 were found to be dually phosphorylated at
threonine and tyrosine. To investigate whether AN could activate ERK, the effect of AN on the
activation-association phosphorylation of ERK1/2 was measured in SK-N-SH cells. Treatment
with 400 ug/mL AN for 1 hour increased the activation-associated phosphorylation of ERK1/2.
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Further increase in ERK1/2 phosphorylation was observed with 3 and 24 hours of incubation.
Furthermore, increase in ERK phosphorylation was found to be dependent on AN concentration.
When SK-N-SH cells were treated with specific mitogen-activated/ERK-activating kinase
(MEK) inhibitors, PD98059 (10 uM) and U0126 (10 uM), for 1 hour prior to treatment with
400 ug/mL AN for 24 hours, activation of ERK by AN was significantly abolished. Thus,
Chantara et al. (2006) concluded that AN induced ERK1/2 phosphorylation in SK-N-SH cells via
activation of MEK.
The role of muscarinic receptors in AN-mediated ERK activation was also investigated
by pretreating SK-N-SH cells with or without 10 uM atropine, a muscarinic receptor antagonist,
followed by incubation with AN for 24 hours. Carbachol, a muscarinic receptor agonist, was
used as a positive control. Previous studies suggested that expression of muscarinic receptors
can induce cell proliferation by activating the ERK1/2 pathway (Jimenez and Montiel, 2005).
The results showed that 1 mM carbachol induced ERK1/2 activation, and this effect was reduced
by atropine pretreatment. However, AN-induced ERK activation was not significantly altered by
atropine pretreatment, suggesting that muscarinic receptor stimulation may not be directly
involved in the observed AN-induced ERK activation (Chantara et al., 2006).
Chantara et al. (2006) also investigated whether oxidative stress generated by various
stimuli might result in activation of ERK by studying the effects of antioxidants on AN-induced
ERK activation. Three non-enzymatic antioxidants were used: NAC, ascorbic acid, and water-
soluble vitamin E (TRX) were used. When SK-N-SH cells were pretreated with 20 mM NAC
for 10 minutes or 1 mM ascorbic acid or 1 mM TRX for 1 hour prior to addition of 400 ug/mL
AN for 24 hours, no reduction in AN-induced ERK activation was found. Therefore, Chantara et
al. (2006) concluded that the activation of ERK by AN observed in SK-N-SH cells was not
mediated via an oxidative stress-dependent mechanism.
To determine if AN-induced ERK activation was mediated via PKC, PKC was inhibited
via several methods. In addition to applying the PKC inhibitors, GF109203X
(bisindolymalcimide) and rottlerin, PKC was also depleted by prolonged incubation of the cells
with phorbol 12-myristate 13-acetate (PMA). Inhibition of PKC by GF109203X significantly
reduced the increase in ERK1/2 phosphorylation to 26% of that caused by AN alone. Similarly,
rottlerin and prolonged treatment with PMA reduced the activation of ERK by AN. Therefore,
the study authors concluded that PKC played an important role in AN-induced ERK activation in
SK-N-SH cells. In summary, this study demonstrated that AN activated ERK1/2 in a
PKC-dependent manner and that oxidative stress and muscarinic receptor activation were
probably not involved in ERK1/2 activation by AN.
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4.5.2. Genotoxicity Studies
4.5.2.1. Studies in Humans
Eight studies have examined the genotoxicity of AN in vivo in occupationally-exposed
populations.
Xu et al. (2003) detected DNA strand breakage in AN-exposed workers (using single-cell
gel electrophoresis), with the rate of comet sperm higher in exposed workers than in the controls.
There were also significant differences in the frequencies of XX-, YY-, and XY-bearing sperm
between the exposed and control groups, and an increase in the frequency of sex chromosome
disomy. See Section 4.3.1 for a more complete summary of this study.
Fan et al. (2006), in an article translated from Chinese, evaluated the application of a
micronucleus test to detect genetic damages in buccal mucosal cells of AN-exposed workers.
The low concentration (average concentration 0.522 mg/m3 AN) exposed group consisted of
41 healthy male workers with direct contact with AN in a chemical plant that produced AN (by
the oxidation of propylene, ammonia, and air) in Shanghai. Since the entire propylene-ammonia
oxidation process was carried out in a closed system of pipes and automated technology, the
chance of contact with AN was primarily at the time of on-site sampling and pipe inspection.
The average age of this exposed group was 37.4 years; and the range and average exposure
duration were 1-33 and 15.7 years, respectively.
The intermediate (average concentration 1.998 mg/m3) exposed group consisted of
47 healthy male workers in an acrylic fiber factory in Shanghai. AN was used as a raw material
in the synthesis of polyacrylonitrile by polymerization. The average age of workers was
39.8 years and the range and average exposure durations were 1-33 and 17.2 years, respectively.
The control group consisted of 31 healthy male workers with no exposure to any known mutagen
or AN and living in the same community. Their average age was 37.2 years and the average
working duration was 16.7 years. The rates of alcohol consumption and cigarette smoking were
similar in the exposed and control groups.
Buccal mucosal cells were collected from second scrapings from the mucous membrane
on the inside of both cheeks of study subjects after rinsing their mouths with clean water. Blood
samples were also collected for measurement of micronuclei (MN) in peripheral blood
lymphocytes. The rates of occurrence of MN in buccal mucosal cells in the intermediate
concentration exposed, low concentration exposed, and control groups were 4, 3.68, and 2.03%,
respectively; the rates in both exposed groups were significantly higher than in the control group
(p < 0.05). The rates of occurrence of MN in peripheral blood lymphocytes in the intermediate
and low concentration exposed and control groups were 4.23, 2.44, and 2.48%, respectively.
The rate of occurrence of MN in blood lymphocytes in the intermediate exposed group was
significantly different from that in the control group (p < 0.05).
To investigate the relationship between AN exposure and the rate of occurrence of MN
and to eliminate the possibility of the presence of other confounding factors, a multivariate linear
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regression analysis was conducted. The results indicated that the cumulative exposed amount of
AN, the recent exposed amount of AN, and the extent of cigarette smoking were important
factors in the rate of occurrence of MN in buccal mucosal cells and blood lymphocytes. Fan et
al. (2006) concluded that the micronucleus test of buccal mucosal cells could replace the
micronucleus test of lymphocytes in the peripheral blood as a screening test for genetic damage
in AN-exposed workers.
Borba et al. (1996) evaluated urinary genotoxicity of three groups of workers in an AN
fiber production plant (exposed group 1: 14 workers in the continuous polymerization section;
exposed group 2: 10 equipment maintenance workers; control group: 20 administrative workers
from the same plant). Urine extracts were used in the Ames test (using TA 98, +S9) to assess
gene reversion activity. No differences in urinary genotoxicity were found in the three groups.
Additionally, there were no significant differences in the incidence of SCE in peripheral
lymphocytes among the groups. However, the maintenance workers had a higher incidence of
CAs (consisting of gaps and breaks in both chromatids and chromosomes) in lymphocytes than
the controls (p < 0.003). These effects were also increased in production workers but not to
statistically significant levels, possibly due to the comparatively low number of subjects in the
study. The significantly higher incidence of CAs in maintenance workers were in agreement
with the highly significant levels of CEVal-Hb adduct in the same population (see Section
4.1.2.2). The maintenance workers also had significantly higher levels of erythrocyte MDA, an
indicator of lipid peroxidation, than the other two group of workers.
Ding et al. (2003) compared deletion frequencies of mitochondrial DNA in peripheral
lymphocytes in 47 workers exposed to a mean workplace concentration of 0.11 ppm AN and
47 nonexposed workers using PCR techniques (this study is described more fully in Section
4.1.2.2). No deletions were detected in the nonexposed group, but a deletion frequency of 17%
was detected in the exposed group. In a separate experiment on presumably nonexposed
individuals, no deletions in mitochondrial DNA were detected in samples from 12 high school
students, whereas the deletion frequency was 25% in samples from 12 elderly persons.
Consistent with the hypothesis that damage to mitochondrial DNA contributes to degenerative
diseases related to aging, the study authors suggested that occupational exposure to AN may
induce mitochondria DNA deletion in cells that are related to aging.
Using the FISH technique with probes for chromosomes 1 and 4, Beskid et al. (2006)
examined patterns of CAs in cultured lymphocytes from blood samples of 61 AN-exposed male
workers involved in the polymerization of Indian rubber and 49 nonexposed control subjects.
Stationary monitoring in the workplaces indicated AN air concentrations of 0.05-0.3 mg/m3 for a
group of 39 exposed workers sampled in 2000 and 0.05-0.7 mg/m3 for another group of
22 exposed workers sampled in 2003. A 38% increase in frequency of aberrant cells in
AN-exposed workers was found to be statistically insignificant. However, the number of
reciprocal translocations increased by 53% (p < 0.05) in the AN-exposed group. In addition, a
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significant increase in a relative number of insertions was found in the AN-exposed group.
Furthermore, chromosomal specificity was observed in lymphocytes with aberrations on
chromosome 1 and 4. In the AN-exposed group, the proportion of cells with aberrations on
chromosome 1 decreased significantly (58.8 vs. 73.8% in the control subjects, adjusted to age
and smoking), but aberrations on chromosome 4 increased (47.0 vs. 29.4% in the controls).
In an earlier study by the same research group, the frequency of CAs in peripheral blood
samples was studied in 45 male rubber polymerization workers exposed for the last 3 months to
0.05-0.3 mg AN/m3, 23 matched controls living in the same region (control group I), and
33 unexposed controls from Prague (control group II) (Sram et al., 2004). Subjects were
interviewed and completed questionnaires on demographic data, occupational and environmental
exposures, smoking habits, medications, X-ray examinations, viral infections, and alcohol
consumption within the 3 months before sampling. Cytogenetic analysis was conducted using
two methods. Conventional chromosomal analysis was used to quantify CAs (chromatid plus
chromosome breaks and chromatid plus chromosome exchanges), the number of aberrant cells
(those with breaks and exchanges, gaps not included), and the aberration frequency (number of
breaks per cell). The FISH technique, using probes for chromosomes 1 and 4, was employed to
quantify translocations. Conventional analysis did not detect any differences in the frequency of
CAs in exposed workers compared with either control group. FISH detected no differences in
the frequencies of aberrations or translocations in exposed workers compared with matched
controls (control group I), but the frequencies in both groups were significantly elevated
compared with unexposed controls from Prague (control group II). Sram et al. (2004) concluded
that occupational exposure to 0.05-0.3 mg/m3 AN did not present a significant genotoxic risk
and attributed higher frequencies in the exposed group and control group I to undetermined
factors present in the region in which the petrochemical industries are located but absent in
Prague.
Rossner et al. (2002) evaluated the effect of AN on the levels of p53 and p21WAF1
proteins in the blood plasma of 49 workers (average age 44 years, 88% males, 12% females)
exposed to 0.05 to 0.3 mg AN/m3 for the last 3 months in the petrochemical industry. Forty-nine
subjects matched for age and gender and living in the same area, but not working in the
petrochemical industry, were used as controls. No differences in p53 and p21WAF1 expression
between the exposed group and the control group were found. Rossner et al. (2002) suggested
that a possible explanation for these results was that very low exposure levels of AN did not
result in the induction of p53 and p21WAF1 gene expression. The AN exposure concentrations
were low in petrochemical plants when compared with exposure concentrations in the acrylic
fiber plants (For example, Lu et al. [2005] reported a geometric mean AN exposure
concentration of 1.97 mg/m3 in an acrylic fiber plant).
Thiess and Fleig (1978) surveyed 18 workers at a plant that was used for manufacturing
copolymers of styrene and AN, styrene, AN, and butadiene and also for synthesis of organic
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intermediates. The workers had been exposed to AN, on average, for 15.3 years and could have
been exposed simultaneously to styrene, ethylbenzene, butadiene, and other chemicals. The
control group consisted of 18 workers who had not been exposed to AN. Atmospheric
monitoring conducted between 1963 and 1974 revealed AN concentration of about 5 ppm, with
the possibility of higher peak values occurring in connection with special tasks. Between 1975
and 1977, exposure to AN had been reduced to an average of 1.5 ppm. The incidence of CAs in
the lymphocytes was measured in these 18 workers and in age-matched nonexposed controls.
The numbers of aberrant metaphases (gaps and iso-gaps included) were 5.5 ± 2.5% in exposed
workers and 5.1 ± 2.3% in controls. When gaps were not included, the numbers dropped to
1.8 ± 1.3% in exposed workers and 2.0 ± 1.6% in controls. The differences were statistically not
significant. The potential for confounding exposures, small group size, and uncertainty in
exposure levels are limitations of this study.
4.5.2.2. In Vivo Tests in Mammals
Rats
Oral administration of up to 40 mg/kg AN or i.v. administration of up to 98 mg/kg AN to
male Sprague-Dawley rats did not induce MN in bone marrow and peripheral blood, respectively
(Morita et al., 1997). However, Wakata et al. (1998) demonstrated induction of MN in bone
marrow polychromatic erythrocytes of Sprague-Dawley rats (four/group) treated twice with
124.8 mg/kg AN i.v. and sampled 24 hours after treatment. A negative result was obtained in
peripheral blood. In another study (Rabello-Gay and Ahmed, 1980), male Sprague-Dawley rats
treated orally with 16 daily doses of 40 mg/kg-day AN showed no increase in CAs in the bone
marrow over controls.
Irreversible binding of radioactivity from [2,3-14C]-AN to DNA in brain, stomach, and
liver of male Sprague-Dawley rats was reported 24 hours after a single oral dose of 46.5 mg/kg
(Farooqui and Ahmed, 1983a). DNA alkylation was significantly higher in the target organs,
brain and stomach (119 and 81 pmol/mg DNA at 24 hours, respectively), than in the liver
(25 pmol/mg DNA). The covalent binding indices in the liver, stomach, and brain at 24 hours
after dosing were 5.9, 51.9, and 65.3, respectively. Similarly, covalent binding of [2,3-14C]-AN
or its metabolite to testicular (Ahmed et al., 1992a), lung (Ahmed et al., 1992b), and gastric
tissue DNA (Abdel-Rahman et al., 1994b) has been reported in male Sprague-Dawley rats
treated with a single oral dose of 46.5 mg/kg AN. Maximum covalent binding of radioactivity to
gastric DNA occurred at 15 minutes after dosing and occurred at 0.5 and 12 hours for testicular
and lung DNA, respectively. Alkylation of hepatic DNA was also reported when a single dose
of 0.2 mmol [2,3-14C]-AN was administered to male Wistar rats intraperitoneally (Peter et al.,
1983a). Two 14C peaks that did not cochromatograph with any known standards were observed
when the DNA hydrolysate from rat livers was chromatographed on PEI-cellulose column.
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In another study, 0.6 mg/kg [2,3-14C]-CEO was administered to one F344 rat
intraperitoneally (Hogy and Guengerich, 1986), and the rat was sacrificed after 1 hour. Covalent
binding to both liver and brain protein was found, but no covalent binding to nucleic acids could
be detected at the level of 0.3 alkylations per 106 bases. In the same study, three male F344 rats
were administered 50 mg/kg AN i.p., and three other rats were administered 6 mg/kg CEO i.p.
(Hogy and Guengerich, 1986). The rats were sacrificed after 2 hours. N7-(2-oxoethyl)guanine
was measured in liver DNA at the level of 0.032 and 0.014 alkylations/106 bases for CEO- and
AN-treated rats, respectively. In the brains of treated rats, the levels of N7-(2-oxoethyl)guanine
were not above the limit of detection. Since DNA adduct was detected in liver DNA, covalent
binding to DNA had to occur. The method used in the study was not sensitive enough to detect
low levels of alkylation of nucleic acids, probably due to the small amount of DNA obtained
from one rat and the loss of DNA during isolation. A method for correction of contaminating
protein via quantitative amino acid analysis in the DNA sample may have allowed a stringent
determination of DNA-bound material.
When the alkaline comet assay was used to detect DNA lesions, Sekihashi et al. (2002)
demonstrated DNA damage in the forestomach, colon, kidney, bladder, and lung of Wistar rats
treated with a single dose of 30 mg/kg AN i.p. but not in the brain or bone marrow.
Oral exposure to 100 ppm AN in drinking water for 14 or 28 days significantly increased
the level of cellular DNA fragmentation in the brain of male Wistar rats (Mahalakshmi et al.,
2003). Other aspects of this study are discussed in Section 4.5.1.1.4.
AN-induced unscheduled DNA synthesis (UDS) was demonstrated in several studies in
rats. In a study by Hogy and Guengerich (1986), male F344 rats (12/group) received a single
sublethal dose of 50 mg/kg AN in saline by gavage, followed by hydroxyurea to arrest
replicative DNA synthesis but allowing excision repair DNA synthesis. Two hours after dosing,
the animals received s.c. methyl [3H]-labeled thymidine. This dose was repeated after 2 hours,
and half the dose was given again after 2 more hours for a total dose of 3.0 mCi/kg of BW. The
animals were sacrificed 2 hours after the last methyl [3H]-labeled thymidine. A significant
occurrence of UDS was found in the livers but not in the brains of AN-treated rats.
Hogy and Guengerich (1986) also studied the effect of treatment on DNA synthesis over
4 hours in the liver and brain of male F344 rats 48 hours after an oral dose of 50 mg/kg AN.
DNA synthesis was decreased in the brain but not in the liver; replicative indices (i.e., the ratio
of DNA synthesis in treated animals over controls) were 0.29 and 1.30, respectively. Thus, the
carcinogenicity of AN in rat brain is not likely from cytotoxicity, followed by an increased rate
of DNA replication and leading to a greater chance of error during the rapid DNA synthesis.
UDS was demonstrated in lung (Ahmed et al.,1992a), testis (Ahmed et al., 1992b), and
gastric tissue (Abdel-Rahman et al., 1994a) of AN-treated male Sprague-Dawley rats (12/group).
Animals received a single oral dose of 46.5 mg/kg AN in saline, with or without hydroxyurea
cotreatment to block the endogenous deoxynucleotide pool. [3H]-Thymidine was administered
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0.5, 6, or 24 hours after AN dosing, and animals were sacrificed 2 hours later. The replicative
index for DNA synthesis was significantly reduced at all three time points in lung, while DNA
repair in the lung was increased by twofold at 0.5 hour and 1.6-fold at 6 hours following AN oral
treatment. Similarly, DNA synthesis was inhibited in testes at 0.5 and 24 hours after treatment
(but increased at 6 hours), whereas DNA repair increased at 1.5- and 33-fold at 0.5 and 24 hours
after treatment. For gastric tissue, DNA replicative synthesis was inhibited 6 hours after AN
administration but was rebounded and followed by a twofold increase at 24 hours. A threefold
increase in UDS was observed at 24 hours after dosing.
On the other hand, Butterworth et al. (1992) followed the incorporations of
[3H]-thymidine into the hepatocytes isolated from male F344 rats gavaged with either a single
dose of 75 mg/kg AN or five daily doses of 60 mg/kg AN. Single-dosed animals were sacrificed
2 or 12 hours after dosing. Multiple-dosed animals were sacrificed 4 hours after the last dose.
Hepatocytes were isolated and plated on cover slips and incubated with [3H]-thymidine.
Autoradiography was used to detect UDS. No sign of AN-induced UDS was found in
hepatocytes from exposed rats. In addition, no UDS was found in the in vivo spermatocyte DNA
repair assay with AN, using cells isolated from the seminiferous tubules of the same treated rats
(Butterworth et al., 1992). The difference in results from Butterworth et al. (1992) and Hogy and
Guengerich (1986) regarding UDS in rat liver may be due to differences in methodology in that
incorporation of [3H]-thymidine actually took place in vitro in the study by Butterworth et al.
(1992).
AN also produced negative results for dominant lethal assay in male F344 rats (Working
et al., 1987). In this assay, groups of 50 male F344 rats received AN by gavage at 0 or
60 mg/kg-day in 0.9% saline for 5 days. AN exposure in males had no effect on the incidence of
pre- or postimplantation losses, indicating a negative result to germ cells.
Mice
When the alkaline comet assay was used to detect DNA lesions, Sekihashi et al. (2002)
demonstrated DNA damage in the forestomach, colon, bladder, lung, and brain of male ddY
mice treated with 20 mg/kg AN i.p. DNA damage was not detected in the liver, kidney, or bone
marrow.
Sharief et al. (1986) determined that AN caused a slight increase on SCE frequencies in
bone marrow cells of male C57B1/6 mice (four/dose group). An increase in SCE frequency (2 x
control) was observed in the only surviving mouse at the 45 mg/kg-dose group. No increase in
SCE frequencies was observed in mice administered a single dose of up to 30 mg/kg AN
intraperitoneally. Higher doses were lethal to most of the animals. However, Fahmy (1999)
reported AN-induced SCEs in bone marrow cells of male Swiss mice (five/group) treated with
7.5 mg/kg or 10 mg/kg AN i.p. 8 hours following BrdU treatment and with colchicines 2 hours
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prior to sacrifice. The lowest dose of 5 mg/kg i.p. produced no significant effect on SCE
frequency.
Earlier CA studies in mice have been largely negative. Rabello-Gay and Ahmed (1980)
showed that AN did not produce increases in CAs in bone marrow cells of male Swiss mice
(six/group) when given orally for 4, 15, or 30 days at doses up to 21 mg/kg-day or by i.p.
injection at doses up to 20 mg/kg-day for the same duration. No increase in the incidence of
CAs compared with controls was observed in bone marrow cells of NMRI mice that were
injected intraperitoneally with 20 or 30 mg/kg AN (Leonard et al., 1981). No increase in CAs in
bone marrow cells and spermatogonia was observed in ICR mice treated with 20 or 100 mg/m3
AN for 5 days (Zhurkov et al., 1983). However, Fahmy (1999) reported AN-induced CAs in
mouse spermatocytes after single oral doses of 15.5 or 31 mg/kg or three or five successive oral
doses of 7.75 mg/kg (1/8 LDso) in male Swiss mice. In addition, Fahmy (1999) reported AN-
induced CAs in mouse bone marrow cells and spleen cells after a single oral dose of 7.75 mg/kg
or three or five successive doses of 7.75 mg/kg. The aberrations were mainly of chromatid type
(gaps, breaks, fragments, and deletions), with metaphases carrying only one aberration being
dominant.
Leonard et al. (1981) reported AN did not induce MN in polychromatic erythrocytes of
male NMRI mice injected intraperitoneally with 20 or 30 mg/kg AN. Morita et al. (1997)
demonstrated a marginal, but statistically significant, increases in MN in bone marrow
polychromatic erythrocytes but not in peripheral blood when AN at doses of 5.6-45 mg/kg was
administered to male CD-I mice (five/group) via i.p. administration. Oral or i.v. injection
yielded negative results (Morita et al., 1997).
Treatment with AN produced negative results for dominant lethal assays in male mice.
The dominant lethal assay was used to detect CAs in meiotic and postmeiotic male germ cells.
Groups of five male NMRI mice were injected intraperitoneally with 0 or 30 mg/kg AN in saline
or with isopropyl methanesulfonate (positive control) and then mated to untreated females (three
per male) for 5 weeks (Leonard et al., 1981). Females were replaced after 7, 14, 21, and 28 days
and the uterine contents were examined 17 days after mating. No evidence for dominant lethal
effects was observed.
4.5.2.3. Short-term Tests: Bacteria, Fungi, Drosophila, Others
There are a large number of reports of short-term genotoxicity test results on AN that
have been made available through the auspices of the International Programme on Chemical
Safety (IPCS). The IPCS coordinated the investigation of eight organic carcinogens known to be
either inactive or difficult to detect in the Salmonella assay, including AN, benzene,
diethylhexylphthalate, diethylstilbestrol, hexamethylphosphoramide (HMPA), PB, safrole, and
o-toluidine, as well as two noncarcinogens in rodent bioassays (benzoin and caprolactam). Most
of the available short-term genotoxicity tests were employed, and the work was carried out at
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some of the major research and testing laboratories throughout the world. The purpose of this
endeavor was to evaluate the efficacy of these tests, to evaluate the strengths and weaknesses of
such tests, and to identify the assay systems to complement the widely used Salmonella assay.
All of the results from these studies have been compiled in a 750-page collection (Ashby et al.,
1985), and a 56-page synopsis of the results is available on the internet (TPCS, 1985). The
conclusion from these evaluations was that AN, along with HMPA, o-toluidine, and safrole,
belonged to a group of genotoxins that were detected by most of the eukaryotic assays studied
and could easily be found to be nonmutagenic in the Salmonella assay because of protocol
deficiencies associated with the overall metabolic capacity of the assay system. While several of
these studies have been cited in this section, an overview of the findings as they pertain to AN
follows.
AN, at concentrations not overtly toxic in a given assay, was found in 42 of 68 tests to
positively cause genotoxicity in bacteria, fungi, Drosophila melanogaster (mutation, gene
reversion, mitotic crossing over, aneuploidy), and mammalian cell culture assay systems, both
human and animal (single-strand breaks, UDS, CAs, SCEs, MN, and transformation). Among
the eight carcinogens studied, AN gave the most positive results, inducing genotoxicity
responses in 62% of all tests (42 out of 68 tests; 25 were negative and 1 was questionable). That
number of positive responses rose to 81% when a more stringent selection of assays was applied
(IPCS, 1985). Other studies, covering all assay types used in that collaborative study, are listed
in the two subsections on short-term assays (Sections 4.5.2.3 and 4.5.2.4).
4.5.2.3.1. Bacterial tests. A number of research groups have examined the capacity of AN to
induce gene reversion in the Ames test. One of the first published papers was Milvy and Wolff
(1977), which reported positive results of AN on Salmonella typhimurium strains TA 1535,
1538, and 1978 in the presence of S9 and an NADPH generating system. The first strain is
sensitive to base substitution mutagens, and the latter two strains are sensitive to frameshift
mutagens. Negative results were obtained when metabolic activation was excluded from the
system. Gene mutation by exposure of bacteria to an atmosphere containing AN was found to
occur at concentrations as low as 57 ppm (equivalent to 2 uL of AN), lower than exposure in
solution or spotting AN to a "lawn" of bacteria on a plate. This could be because AN vaporized
readily in solution such that the actual exposure concentrations in solution or by spotting were
uncertain. The findings of Milvy and Wolff (1977) were criticized by Venitt (1978) in a letter to
the editor. While not disagreeing with the overall conclusion, Venitt (1978) considered the study
authors to have calculated mutation frequency incorrectly from the data. Venitt (1978)
confirmed AN to be mutagenic in TA 1535 but not in the frameshift strains TA 1538 and 1978.
Among other reports of AN activity in the Ames test, Lijinsky and Andrews (1980) found
AN at doses of 100-1,000 ug per plate to be positive in S. typhimurium TA 1535 in the presence
of S9 but negative in TA 98, 100, 1537, and 1538 in the plate incorporation assay. Similarly,
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Brams et al. (1987) reported AN (50-750 (J,g/L) to be negative for gene reversion in TA 97, 98,
and 100 using plate incorporation assay, irrespective of metabolic activation. The study authors
accounted for these negative results with inadequate experimental conditions. Previously, they
had demonstrated that AN was mutagenic in TA 1530, 1535, and 1950 when 0.2% gaseous AN
was injected into the dessicator where the plates were incubated for 1 hour in the presence of S9
(the plate agar contained 200 ug AN/plate) (de Meester et al., 1978). A lower mutagenic activity
was also detected with strains reversed by frameshift mutation (TA 98, 1978, 100), and assays
conducted by the plate incorporation method gave negative results (de Meester et al., 1978; Dow
Chemical Co., 1976).
Jung et al. (1992) provided data from three laboratories that examined the ability of AN
to induce gene reversion in TA 102 in the plate incorporation assay, with uniformly negative
results. Hakura et al. (2005) reported that AN induced dose-dependent increases in the number
of revertants per plate in strain TA 100 exposed to concentrations ranging from 806 to 12,100 ug
per plate in the presence of rat or human liver S9 preparations, but the maximum response at
12,100 ug per plate was slightly less than twofold higher than the number of revertants per plate
observed in the negative controls.
Although the methodologies of the experiments may have been different, the positive
finding of mutagenicity by de Meester et al. (1978) for AN in TA 98 and 100 is in agreement
with data from Khudoley et al. (1987) that AN (concentration not provided) was positive in
TA 98 and 100 (with two- to fivefold increase in frequency of induced mutants), irrespective of
the presence of S9. Zhurkov et al. (1983) reported AN to dose-dependently induce mutations in
S. typhimurium TA 1535 but not in TA 1538. The presence of complete S9 fraction was required
for this effect, but a 9,000 x g microsomal supernatant without cofactors gave inconsistent
results. The concentrations tested were between 0.1 and 10,000 ug per dish. The highest
concentration was overtly toxic to the bacteria.
Other tests of the mutagenicity of AN in bacterial systems were negative. For example,
Nakamura et al. (1987) employed the SOS test to evaluate the capacity of AN to induce
expression of the umu gene in S. typhimurium TA 1535/pSK1002. This strain contains a
umuC-lacZ fused gene, such that a forward mutation results in umu gene expression and the
transcription of the lac operon and can be demonstrated phenotypically by a twofold increase in
p-galactosidase activity. AN, at concentrations up to 2,820 ug/mL, and seven other known
mutagens gave negative results in this system. Similarly, AN was negative in the SOS
chromotest in Escherichia coli PQ37 (Brams et al., 1987). It should be noted that only 4 out of
14 compounds that were positive in the Ames assay were positive in the SOS chromotest kit
used. Thus, some technical issues were responsible for the poor performance of the test kit
(Brams et al., 1987).
Venitt et al. (1977) tested AN for mutagenicity in a gene reversion assay (2-3 days at
37°C), using the tryptophan-dependent E. coli WP2 series of bacteria as indicator organisms.
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Doses of 75 and 150 umol per plate of AN produced a dose-related increase in the number of
revertant colonies (trp~ —> trp+) compared with untreated bacteria in WP2 (which is DNA repair
proficient), WP2 uvrA (which lacks excision repair), and WP2 uvrApolA (which lacks both
excision repair and DNA polymerase 1) without a need for S9 fraction. The study authors
confirmed these results by using a simplified "fluctuation" assay in which they obtained a dose-
dependent increase in mutation rate induced by 0.4-2 mM AN (20- to 40-fold lower than the
levels in plate test) in E. coli WP2. An exponential dose response was seen at lower
concentrations, 0.1-0.4 mM, with mutant WP2 uvrApolA. The effective mutagenic
concentration range for AN was lowered by an order of magnitude when an error-prone DNA-
repair plasmid, pKMlOl, was introduced into E. coli WP2. The study authors concluded from
these results that AN caused non-excisable mis-repair DNA damage that ultimately gave rise to
DNA strand breaks. Venitt et al. (1977) hypothesized that AN might react with thymine residues
in DNA since AN has been shown to cyanoethylate ring N atoms of minor tRNA nucleosides
and ribothymidine and thymidine (Ofengand, 1967).
Lambotte-Vandepaer et al. (1985, 1981, 1980) collected urine from male Wistar rats
(two/group) and NMRI mice (five/group) that had been administered a single dose of AN
(30 mg/kg intraperitoneally). The urine samples were evaluated for potential induction of gene
reversion in S. typhimurium TA 1530. Positive results were obtained in rats and mice in the
absence of S9. Mutagenic activity in urine was abolished by the presence of S9 in rats and
decreased in mice, a finding that suggests that the mutagenic agent in urine can be inactivated
metabolically. Pretreatment with PB (induces CYP450 monooxygenase), CoCb (inhibits
CYP450 monooxygenase), and DEM (depletes GSH), before AN treatment, slightly decreased
the mutagenic response in urine from mice and completely abolished the response in urine from
rats. However, the study authors were unable to identify the genotoxic compound in urine.
4.5.2.3.2. Fungi. Available studies that employed fungi in short-term assays on the
mutagenicity/genotoxicity of AN produced mostly positive results. AN induced mitotic gene
conversion in both stationary-phase and log-phase cultures at the his4 and tips loci of
Saccharomyces cerevisiae JD1, in the presence of metabolic activation by S9. Negative results
were obtained without metabolic activation (Brooks et al., 1985; Shell Oil, 1984a). AN did not
induce chromosome loss in S. cerevisiae D61.M (Whittaker et al., 1990) but elicited respiratory
deficiency, reflecting antimitochondrial activity.
4.5.2.3.3. Drosophila. A range of in vivo experimental systems used the fruit fly,
D. melanogaster, to examine the mutagenicity/genotoxicity of AN. Drosophila can biotransform
certain procarcinogens to their reactive metabolites and are used in short-term tests for
identifying carcinogens and in studies on the mechanism of mutagenesis of chemicals (Vogel et
al., 1999).
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Osgood et al. (1991) used the Drosophila ZESTE system to monitor the potential for AN
to induce sex chromosome aneuploidy, following inhalation exposure of adult females to
2.7 ppm for up to 70 minutes. AN induced chromosome loss after exposure for 50 and
70 minutes. AN was mostly nontoxic at the tested dose, with only 13% killed after a 70-minute
exposure. Similarly, in a Drosophila somatic recombination and mutation assay, Drosophila
larvae were exposed to 5-20 mM AN in water for 9-11 days, and hatching females were scored
for twin mosaic spots and single mosaic light spots in their eyes (Vogel, 1985). These genetic
markers might arise from many types of genetic alterations. Mitotic recombination and
chromosome breakage would result in mosaic twin spots, whereas deletions and gene mutations
would give rise to mosaic single light spots. AN was positive for the induction of mosaic single
spots at 5 mM (LCso concentration was 10 mM) and negative for twin spots. Since the
classification of single spots or twin spots might be subjected to personal bias, the total of twin
and single spots was also reported, and AN gave positive result.
AN was shown to be mutagenic by having marginally positive effects in somatic
mutation/recombination assay on wing spots of Drosophila, with gas exposure of larva to 0.5-
1 uL/1,150 mL for 0.5 or 1 hour (Wiirgler et al.,1985). AN gave negative results in a sex-linked
recessive lethal mutation test on postmeiotic and meiotic germ cells of male/), melanogaster,
exposed either by feeding of 420 ppm AN or injected with 3,500 ppm AN (Foureman et al.,
1994).
The in vitro effect of AN on taxol-purified microtubules from Drosophila and mouse
brain was evaluated by Sehgal et al. (1990). (Taxol promoted the formation and stability of
microtubules.) Microtubules assist in the movement of chromosomes in both mitosis and
meiosis. Polymerization and depolymerization of microtubules occur in cell division to separate
the chromosomes from the metaphase plate during anaphase. In this study, the assembly and
disassembly of microtubules was monitored spectrophotometrically in vitro. Previous results
from in vivo assays monitoring induced sex chromosome aneuploidy indicated that effective
aneuploidogens affected microtubule assembly. When taxol-purified D. melanogaster
microtubule was incubated at 37°C to allow polymerization, addition of 5 or 50 mM AN resulted
in 28 and 64% inhibition of microtubule assembly, respectively. When taxol-purified mouse
brain microtubules were incubated with 5 or 50 mM AN, 74 and 96% inhibitions of microtubule
assembly were observed, respectively. Thus, these results indicated that AN was an
aneuploidogen in vitro. On the other hand, taxol significantly affects microtubule
depolymerization assay, probably by stabilizing the formed microtubules. None of the tested
aneuploidogens, including colchicine, promoted disassembly to taxol-purified microtubules.
4.5.2.3.4. Other short-term tests. Yates et al. (1994) reported on the ability of CEO to induce
single- and double-strand DNA breaks in supercoiled DNA plasmid pBR322 DNA. Supercoiled
DNA (1 ug) was incubated for 3 hours at 37°C with >50 mM CEO and then subjected to agarose
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gel electrophoresis. The study authors reported that CEO non-enzymatically induced DNA
strand breaks in a dose- and time-dependent manner, but detailed data were not provided. Peter
et al. (1983b) found that AN did not induce strand breaks in SV40 phage DNA in vitro, whereas
synthetic glycidonitrile (i.e., CEO) was effective in the same system (both agents were incubated
with DNA at 1 mmol/L for 17 hours in the dark, at 37°C, in a buffered solution without any
enzyme addition).
4.5.2.4. Mammalian Cell Short-term Tests
4.5.2.4.1. Mutations. A number of research groups have examined the ability of AN to induce
forward mutations in human lymphoblast cell lines and the well-known mouse lymphoma
LSI78Y Tk+/~ system.
Crespi et al. (1985) used two human lymphoblast cell lines, Tk6 (which does not contain
CYP450 activities) and AHH-1 (which is metabolically competent) to assess AN mutagenicity at
the thymidine kinase (Tk+/~) locus and the hprt locus, respectively. Tk6 cell cultures were treated
with 0-40 ug/mL AN for 3 hours with and without externally added metabolic activation (rat
liver S9). On the third day after treatment, the cultures were plated in a selective medium
containing 2 ug/mL trifluorothymidine. After incubation for 12 days, the plates were scored for
the presence of mutant colonies. Similarly, AHH-1 cell cultures were treated with AN for
28 hours, and the cultures were plated on the 6th and 7th day after treatment in a selective medium
containing 0.6 ug/mL 6-TG. AN induced dose-dependent mutations at the Tk+/~ locus in Tk6
cells in the presence, but not in the absence, of S9. AN induced mutations at the hprt locus in
AHH-1 cells. The lowest AN concentrations that were mutagenic in these test systems were
40 ug/mL with Tk6 cells +S9 and 25 ug/mL with AHH-1 cells.
Similarly, Recio and Skopek (1988a, b) assessed the mutagenicity of AN and its epoxide
metabolite, CEO, at the Tk+/~ locus in Tk6 human lymphoblast cells in the presence and absence
of rat liver S9. In the presence of S9, 2-hour incubations with 1.4 mM AN induced mutations at
the Tk+/~ locus as demonstrated by the presence of mutant clones when plated in trifluoro-
thymidine selection medium after treatment, but, in the absence of S9, no mutagenic activity was
observed over the concentration range of 0.4-1.5 mM (Recio and Skopek, 1988b). In contrast,
2-hour incubation with CEO at concentrations as low as 100 and 150 uM induced a mutagenic
response (without metabolic activation) at the Tk+ ~ locus (Recio and Skopek, 1988b). On a
molar basis, CEO was as mutagenic in this system as the well-known mutagen, ethyl
methanesulfonate (Recio and Skopek, 1988a).
Two classes of CEO-induced TK~~ mutant phenotypes were identified that differed in
their growth rates: Tkn with normal growth rate and Tks with slower growth rates. Southern blot
analysis of DNA of these two classes indicated that the phenotypes differed genotypically (Recio
and Skopek, 1988b). Ninety-six percent (25/26) of Tks mutants had lost a 14.8 kb DNA
fragment corresponding to the active Tk allele, whereas only 8% (1/12) of CEO induced Tkn
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mutant clones and 22% (2/9) of spontaneous Tkn mutant clones had lost the 14.8 kb fragment
(Recio and Skopek, 1988a, b). Recio and Skopek (1988a) suggested that Tks mutants resulted
from large-scale DNA structural alterations involving the active Tk allele. CEO induced
predominantly Tkn mutants. Southern blot analysis of CEO-induced Tkn mutants indicated the
majority of these mutants were below the detection limit of <2 kb. Thus, CEO-induced
alterations are relatively small DNA alterations. Recio and Skopek (1988a) suggested that
CEO-induced Tkn mutants resulted from point mutations or small insertion/deletions that
occurred during the replication or repair of CEO-modified DNA. Karyotypic analysis on two
Tkn mutants and 16 Tks mutants indicated that the majority of Tk mutants were not accompanied
by abnormalities of the chromosome on which the Jit gene resides (chromosome 17).
CEO also induced mutations at the hprt locus in Tk6 cells (Recio and Skopek, 1988a).
Characterization of the hprt mutations by cDNA sequencing analysis indicated that several hprt
mutations were formed. A major (8/14) type of CEO-induced mutation was the specific loss of
exons from the coding region of hprt. Remaining mutants (6/14) were single base substitutions
(point mutation) resulting from amino acid changes (A:T base pairs and G:C base pairs).
The mutagenicity of AN was evaluated in the mouse lymphoma L5178Y Tk+ ~ forward
mutation assay by a number of laboratories. Oberly et al. (1996) demonstrated that AN
(activated with S9) was mutagenic at 40 ug/mL in this assay because of producing a more than
twofold increase in mutant frequency when compared to the mutant frequency of the solvent
controls. Earlier results in the mouse lymphoma L5178Y Tk+/- system from several studies all
pointed to the capacity of AN at 12.5-200 ug/mL to induce forward mutations at the 7ft+~with or
without S9 (Lee and Webber, 1985; Myhr et al., 1985; Amacher and Turner, 1985; Rudd, 1983).
Garner and Campbell (1985) also reported that AN induced mutations to ouabain and 6-TG
resistance in the mouse lymphoma L5178 Y cells in the presence of S9. However, negative
results were obtained by Styles et al. (1985) in the mouse lymphoma L5178Y Tk+ + cell line
(Na+/K+ ATPase locus, ouabain was used for mutant selection) and Tk+ ~ cell line
(trifluorothymidine was used for mutant selection). Using P388F mouse lymphoma 7ft+~cell line
in the presence of S9, Anderson and Cross (1985) also obtained forward mutations to 5-iodo-
2-deoxyuridine resistance with AN. However, negative results were obtained by Lee and
Webber (1985) in Chinese hamster V79/HGPRT assay in which AN did not induce 8-azaguanine
resistance mutation with or without S9.
4.5.2.4.2. Other DNA effects
UDS
AN, at concentrations up to 2.5 mg/mL, was negative for induction of UDS in HeLa cells
with or without the presence of S9, using the scintillometric method (Martin and Campbell,
1985). However, an increase in UDS, as measured by uptake of [3H]-thymidine, was reported in
cultured human lymphocytes treated with 5 x 10"1 M AN and S9 (Perocco et al., 1982). Negative
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results were reported for the ability of AN to induce DNA repair synthesis (measured by
incorporation of [3H]-thymidine and autoradiographic techniques) in rat hepatocyte primary
culture (Probst and Hill, 1985; Williams et al., 1985). However, no UDS, as measured by
autoradiography, was observed in any of the cultures treated with the eight tested carcinogens
(Probst and Hill, 1985). Similarly, only one of the eight carcinogens tested by Williams et al.
(1985) induced DNA repair synthesis as determined by autoradiography of [3H]-thymidine
incorporation (AN was negative in this assay). Thus, IPCS (1985) concluded the rat hepatocyte
autoradiographic UDS assay was an insensitive assay for determination of genotoxicity of these
chemicals and should be avoided for use as a complement to the Ames assay.
Butterworth et al. (1992) used incorporation of [3H]-thymidine and autoradiographic
techniques to study the effect of AN and CEO on unscheduled DNA repair in vitro in rat
hepatocytes, and human mammary epithelial cells. AN and CEO were negative for the induction
of DNA repair in hepatocytes in vitro. There was some indication of a statistically insignificant
response at 0.1 mM CEO. However, CEO was toxic to the hepatocyte culture at 1 mM, the next
higher tested concentration. As noted previously, IPCS (1985) has determined that rat
hepatocyte autoradiographic UDS assay is insensitive for genotoxicity testing of AN and a group
of seven other carcinogens. However, CEO, but not AN, was positive for UDS in human
mammary epithelial cells in vitro (Butterworth et al., 1992).
DNA strand breaks
DNA single-strand breaks were measured by the alkaline elution method in the following
studies. DNA single-strand breaks were induced in [14C]-thymidine-labeled cultured adult
human bronchial epithelial cells treated with 200 or 500 ug/mL AN for 20 hours (Chang et al.,
1990). These concentrations were below the cytotoxic concentration of 600 ug/mL AN. DNA
single-strand breaks were also reported in cultured rat hepatocytes treated with 65.8 ug/mL AN
for 3 hours (Bradley, 1985). Higher concentrations of 197 or 658 ug/mL AN resulted in
cytotoxicity. In another study, AN induced DNA single-strand breaks in cultured Chinese
hamster ovary (CHO) cells treated with 3.7 x 103 or 5.3 x 104 ug/mL AN (7 x 10"2 or 1 x 10"1 M
AN) with or without S9 mix for 1 hour (Douglas et al., 1985). These concentrations were above
the cytotoxic concentrations of 5.31 and 53.1 ug/mL AN (10~5 and 10"4 M) with and without
S9 mix, respectively. Thus, the observed DNA strand break effect was relatively weak in CHO
cells. Lakhanisky and Hendrickx (1985) reported that AN (concentrations not reported) did not
induce DNA strand breaks in cultured CHO cells with or without S9. Therefore, cultured CHO
cells may not be a sensitive assay system to test for DNA strand breaks induced by AN when
compared with other cell cultures. On the other hand, DNA strand breaks was reported in SHE
cells treated with AN. Parent and Casto (1979) observed incubation of [3H]-thymidine-labeled
primary SHE cells with 200 or 400 ug/mL AN for 18 hours caused a shift in the sedimentation
pattern of the labeled cellular DNA when subjected to alkaline sucrose gradient.
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Induction ofp53 andp21w 1 proteins
The tumor suppressor protein p53 is a key molecule induced after DNA damage and by
conditions of cellular oxidative stress (Donehower and Bradley, 1999). Expression of p53
results either in cell cycle arrest or apoptosis (Janus et al., 1999). Yang and Duerksen-Hughes
(1998) proposed the measurement of p53 protein induction to identify genotoxic carcinogens.
The cyclin-dependent kinase inhibitor p21WAF1 protein is an important down-stream effector of
p53-induced cell cycle arrest. Expression of p21WAF1 induces cell cycle arrest either in the Gl, S,
or G2 phase, enabling DNA repair (Binkova et al., 2000).
In vitro exposure of human embryonic lung fibroblasts with 0.3 to 1.0 mM AN for
24 hours resulted in the induction of both p53 and p21WAF1 protein as determined by the ELISA
assay (Rossner et al., 2002). A change in the shape of cells from an elongated shape to a round
one was also observed. Cells treated with >2.5 mM AN changed their morphology after 4 hours
of treatment.
4.5.2.4.3. Cytogenic effects. AN induced SCE in cultured adult human bronchial epithelial cells
treated with noncytotoxic concentrations of 150 or 300 ug/mL AN for 20 hours (Chang et al.,
1990). An increase in frequency of SCE was observed in human lymphocytes from two different
donors incubated for 1 hour with 5 x 10"4M AN and S9 mix. No increase in SCE was observed
without the S9 mix metabolizing system. (Perocco et al., 1982). AN was negative for the
induction of SCEs in CHO cells (Ved Brat and Williams, 1982). However, when CHO cells
were cocultured with freshly isolated rat hepatocytes, Ved Brat and Williams (1982) observed
that AN at 10"4 M produced a greater than twofold increase in SCEs in the CHO cells, suggesting
that the rat hepatocytes metabolized AN to its reactive metabolite, which was then transported
into the CHO cells. Other cytogenetic findings in CHO cells included positive results for the
induction of SCEs at 2 mM AN with S9 (Natarajan et al., 1985) and CA at 4 mM AN with or
without S9, while Douglas et al. (1985) reported that AN at 10"1 M induced the formation of MN.
AN was reported to induce CAs in Chinese hamster lung (CHL) fibroblasts in culture without
metabolic activation at nontoxic concentrations of 12.5 ug/mL (Ishidate and Sofuni, 1985). AN
induced structural CAs in Chinese hamster liver fibroblast cell line (CH1-L) at the lowest
concentration of 2.5 ug/mL and higher (Danford, 1985).
Sasaki et al. (1980) found that 0.0053 mg/mL AN induced chromosome breaks in a
pseudodiploid Chinese hamster cell line (Don-6). AN was negative at concentrations up to
10 ug/mL for the induction of CAs, SCEs, and polyploidy in cultured epithelial-like cells from
rat liver (RL4 cell line) (Priston and Dean, 1985; Shell Oil Co., 1984b). Mangir et al. (1991)
reported that CHO cells treated with AN demonstrated a growth and RNA synthesis rate that is
similar to that for agents that cause damage to nuclear DNA in cells. Growth and RNA synthesis
of CHO cells was inhibited with 0.001% (v/v) AN and completely inhibited at 0.005% (v/v) AN.
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Kodama et al. (1989) conducted cytogenetic analyses on eight spontaneous and eight
CEO-induced Tks mutant clones in Tk6 human lymphoblastoid cultures that had lost the 14.8 kb
polymorphic band corresponding to the active Tk allele. These CEO-induced Tk'1' mutants were
reported in the Recio and Skopek (1988a, b) studies. No chromosomal abnormalities were found
in the eight spontaneous mutants. On the other hand, a visible abnormality on chromosome 17
was found in one of the CEO-induced tks mutants and was marked by duplication of the long arm
of chromosome 17, with break points at ql 1 and q21. The latter break point was close to the
Tk locus, suggesting the observed aberration might be associated with Tk ~ phenotype.
4.5.2.4.4. Transformation assays. Parent and Casto (1979) studied the capacity of AN to
induce transformations in primary SHE cells in culture by monitoring the incidence of
microscopically observed foci of morphologically transformed cells. In two experiments, SHE
cell cultures were exposed to 25-200 ug/mL AN for 18 hours, after which AN was removed and
cultures were subsequently inoculated with 200 focus-forming units of simian adenovirus SA7
and incubated for 3 hours. Colonies of surviving cells were counted after 8 days, and virus-
transformed foci were counted after 21 days. Pretreatment of cells with AN prior to viral
inoculation resulted in only slight enhancement of 1.8-fold in SA7 foci. In another experiment,
SHE cells were treated with the same concentrations of AN 5 hours after viral inoculation. This
resulted in an enhancement ratio that was markedly increased (8.9 at AN concentration of 200
ug/mL vs. 1.0 in controls).
When SHE cells were treated for 6 days with 12-100 ug/mL AN without added
SA7 virus, foci of morphologically transformed cells were observed at 50 ug/mL AN (two
foci/six dishes) and 100 ug/mL AN (three foci/nine dishes).
The ability of AN to induce transformation in SHE cells was also evaluated by Barrett
and Lamb (1985). AN was considered to give a positive response according to the criterion of
inducing four or more transformed colonies per 2,000 surviving colonies. With a relative
survival of unity, the lowest concentration of AN (0.01 ug/mL) generated morphologically
transformed colonies at a rate of 4/1,149.
Lawrence and McGregor (1985) evaluated the ability of 10 potential carcinogens,
including AN, to induce morphological transformation in cultured embryonic mouse fibroblasts
(C3H/10T1/2, Clone 8) in the presence or absence of S9. Positive response was obtained at
16 ug/mL AN in the presence of S9, while responses were uniformly negative in its absence.
Banerjee and Segal (1986) studied whether AN (0-200 ug/mL) could produce in vitro
transformation of C3H/10T1/2 and NIH/3T3 mouse fibroblast cells in culture. AN was
cytotoxic at the higher concentrations, with cell survival dropping below 75% at>50 ug/mL in
C3H/10T1/2 and NIH/3T3 cells. Optimal transformation rates were obtained at AN
concentrations of 12.5 ug/mL in C3H/10T1/2 cells. AN-induced transformation was observed at
concentrations between 3 and 100 ug/mL in NIH/3T3 cells.
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Matthews et al. (1985) evaluated whether AN induced morphological transformation and
mutation to ouabain resistance (Ouar) in Balb/c-3T3 cells in culture with or without exogenous
metabolic activation. For activation transformation assay, Balb/c-3T3 cells were cocultured with
lethally X-irradiated primary F344 rat liver cells (RLCs). The RLC-3T3 cocultures were treated
with 0-25 ug/mL AN for 48 hours. The cocultures were then treated biweekly for 3 weeks with
0.05 ug/mL 12-(9-tetradecenoyl-phorbol-13-acetate beginning 1-2 days after completion of AN
treatment. For nonactivation transformation assay, 3T3 cell cultures were incubated with 0-
20 ug/mL AN for 72 hours. In the activation transformation assay, significant increase in
relative transformation activity was found in RLC-3T3 coculture treated with 8.8 ug/mL AN
(noncytotoxic concentration). Relative cell survival was only 22%, and no significant increase in
transformation activity was found at 16.7 ug/mL. No significant increase in transformation
activity was observed in AN-treated 3T3 cell cultures without RLCs.
For the Ouar mutation assay, 3T3 cell cultures were treated with 0-150 ug/mL AN with
or without S9 mix for 4 and 24 hours, respectively. After the treatment, AN was removed and
the cultures were refed and maintained for 5-6 days for expression and selection, using 2 mM
cardiac glycoside ouabain. The appearance of Quar variants would indicate a mutation arose in
the gene controlling the synthesis of cell membrane Na+/K+ ATPase (Corsaro and Migeon,
1978). Significant increase in relative Ouar frequency was observed at 50 ug/mL AN with S9
activation.
Yuan and Wong (1991) used a nonfocus transfection-transformation assay to study the
capacity of the oxidative metabolite CEO to bring about functional changes in a plasmid that
would be indicative of a compound-induced mutation. A new plasmid that had been constructed
by ligating a human c-HA-ras-1 protooncogene to a pSV2neo mammalian vector was reacted
with CEO in vitro and then transfected into NIH3T3 cells. Cells were selected for neomycin
resistance and/or abnormal growth characteristics, the latter serving to discriminate between
colonies arising from ras mutations and those from cells that were not transfected (or that were
transfected with nonplasmid DNA). Although CEO-modified ras gave rise to two neomycin
resistant clones, they were probably not indicative of a ras mutation because their normal growth
rate and monolayer density were similar to negative control. Southern blot analysis of
transformant DNA also supported this conclusion. For example, when anti-benzo(a)pyrene-
7,8-dihydrodiol-9,10-epoxide transformant DNA was examined in this system as a positive
control, a fragment of 411 base pairs was revealed, indicating a ras mutation at codon 11 or 12.
However, both CEO-derived clones and untreated control showed the WT band of 355 base
pairs.
4.5.2.4.5. Genotoxicity summary. All identified studies concerning the mutagenicity or
genotoxicity of AN are compiled in Table 4-54. The overall weight of evidence from in vitro
and in vivo studies is adequate to support mutagenicity for the AN metabolite, CEO.
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Table 4-54. Summary of studies on the mutagenicity/genotoxicity of AN
Test system
Endpoint/effect
Exposure
concentration
Exposure
duration
Result3
Reference
Humans
AN-exposed workers
n=30
AN-exposed
workers:
31 controls, 41 low,
47 intermediate
AN-exposed
polymerization
workers:
14 Maintenance
workers:
10 controls: 20
AN-exposed
workers: 47;
47 controls
AN-exposed workers
Group 1 = 39
Group 2 = 22
Unexposed controls
= 49
AN-exposed workers
(n=45)
Matched controls =
23, unexposed
controls = 33
AN-exposed workers
(n=49)
Unexposed controls
= 24
AN-exposed workers
(n=18);
controls =18
DNA strand breaks;
nondisjunction of sex
chromosomes in sperm
Increase in MN in
buccal mucosal cells
(low and intermediate
groups) and blood
lymphocytes
(intermediate group)
of AN-exposed
workers.
CAs in lymphocytes
Deletion of
mitochondrial DNA in
lymphocytes
CAs (detected by
FISH) in cultured
lymphocytes from
peripheral blood
samples
CAs in cultured
lymphocytes in
peripheral blood
samples
Induction of p5 3 and
p21WAF1 proteins in
blood plasma
CAs in lymphocytes
0.8 mg/m3
Low =
0.522 mg/m3
Intermediate =
1.998 mg/m3
NDb
0.11 ppm
Group 1 = 0.05-
0.3 mg/m3
Group 2 = 0.05-
0.7 mg/m3
0.05-0.3 mg/m3
0.05 - 0.3 mg/m3
5 ppm, reduced
to 1.5 ppm
between 1975
and 1977
(possible
exposure to other
chemicals)
2.8 yrs
Low = 1-
33 yrs
(average
15.7 yrs)
Intermediate =
1-33 yrs
(average
17.2 yrs)
ND
17.3 yrs
3 mos
3 mos
3 mos
15.3 yrs
+
+
Maintenance
workers: +
Production
workers: (+)
+
(+)
Significant
increase in the
number of
reciprocal
translocations and
relative number of
insertions.
Increase in
frequency of
aberrant cells not
significant.
—
"
Xu et al.
(2003)
Fan et al.
(2006)
Borba et al.
(1996)
Ding et al.
(2003)
Beskid et al.
(2006)
Sram et al.
(2004)
Rossner et al.
(2002)
Thiess and
Fleig (1978)
239
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Table 4-54. Summary of studies on the mutagenicity/genotoxicity of AN
Test system
Endpoint/effect
Exposure
concentration
Exposure
duration
Result3
Reference
Rats
Sprague-Dawley
(male) (4/group)
Sprague-Dawley
(male)
Sprague-Dawley
(male) n = 3
Sprague-Dawley
(male) n = 3-4/group
Sprague-Dawley
n = 4/group
Sprague-Dawley
(male) n = 4/group
Sprague-Dawley
(male) n = 3/group
Wistar (male)
F344
n= 1
F344
n=3
Wistar
Wistar, male
F344
n = 12/group
Sprague-Dawley
F344
MN
MN in bone marrow
MN in peripheral
blood
CAs in bone marrow
Binding to DNA in
stomach, brain, and
liver
Binding to testicular
DNA
Binding to lung DNA
Binding to gastric
tissue DNA
Alkylation of hepatic
DNA
Binding to liver and
brain DNA
DNA adduct formation
DNA damage in
forestomach, colon,
kidney, bladder, and
lung but not in brain or
bone marrow
Fragmentation of brain
DNA
UDS in liver but not
brain
UDS in lung
UDS in gastric tissue
UDSintestis
UDS in isolated
hepatocytes or
spermatocytes
125 mg/kg i.v.
10-40 mg/kg oral
24.5-98 mg/kg
i.v.
40 mg/kg oral
46.5 mg/kg oral
46.5 mg/kg oral
46.5 mg/kg oral
46.5 mg/kg oral
0.2 mmol i.p.
0.6 mg/kg CEO
i.p.
50 mg /kg AN
i.p. or 6 mg/kg
CEO i.p.
30 mg/kg i.p.
100 ppm AN in
drinking water
50 mg/kg gavage
46.5 mg/kg oral
46.5 mg/kg oral
46.5 mg/kg oral
a. 75 mg/kg
b. 60 mg/kg oral
Two
treatments
Single dose
Single dose
16 d
Single dose
Single dose
Single dose
Single dose
Single dose
Single dose
Single dose
Single dose
14 or 28 d
Single dose
Single dose
Single dose
Single dose
a. Single dose
b. five daily
doses
Bone marrow: +
Peripheral blood:
-
-
+
+
+
+
+
"
Liver: +
Brain: (+)
+
+
+
+
Wakata et al.
(1998)
Morita et al.
(1997)
Morita et al.
(1997)
Rabello-Gay
and Ahmed
(1980)
Farooqui and
Ahmed (1983a)
Ahmed et al.
(1992a)
Ahmed et al.
(1992b)
Abdel-Rahman
etal. (1994b)
Peter et al.
(1983a)
Hogy and
Guengerich
(1986)
Hogy (1986);
Hogy and
Guengerich
(1986)
Sekihashi et al.
(2002)
Mahalakshmi
et al. (2003)
Hogy and
Guengerich
(1986)
Ahmed et al.
(1992a)
Abdel-Rahman
etal. (1994a)
Ahmed et al.
(1992b)
Butterworth et
al. (1992)
240
DRAFT - DO NOT CITE OR QUOTE
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Table 4-54. Summary of studies on the mutagenicity/genotoxicity of AN
Test system
F344
n=50
Endpoint/effect
Dominant lethal
mutations
Exposure
concentration
60 mg/kg
Exposure
duration
5d
Result3
-
Reference
Working et al.
(1987)
Mice
ddY
C57B1/6
n = 1-4/group
Swiss
Swiss
n = 3-6/group
NMRImale
n = 4/group
C57B1/6
n = 1-4/group
NMRImale
n = 4-5/group
CD-I
CD-I
ICR
NMR1
n = 5/group
DNA damage in
forestomach, colon,
bladder, lung, and
brain
SCEs in bone marrow
SCEs in bone marrow
CAs in spermatocytes,
bone marrow, and
spleen cells
CAs in bone marrow
CAs in bone marrow
CAs in bone marrow
MN in erythrocytes
MN in bone marrow
MN in peripheral
blood
CAs in bone marrow
cells and
spermatogonia
Dominant lethal
mutation
20 mg/kg i.p.
10-45 mg/kg
7.5 or 10 mg/kg
i.p.
a. 15.5 or
3 1 mg/kg oral
b. 7.75 mg/kg
7, 14, or
21 mg/kg-doral
or 10, 15, or
20 mg/kg-d i.p.
20 or 30 mg/kg
i.p.
10-45 mg/kg
20 or 30 mg/kg
i.p.
0-45 mg/kg i.p.
0-32 mg/kg oral
0-40 mg/kg i.v.
5.6-45 mg/kg i.p.
or 10-40 mg/kg
i.v.
100 or 20 mg/m3
30 mg/kg i.p.
Single dose
Single dose
Single dose
a. Single dose
b. three or five
doses
4, 15, or 30 d
Single dose
Single dose
Single dose
Single dose
Single dose
5d
Single dose
Short-term assays — bacteria
S. typhimurium
TA 1535
TA 1535
TA 1538
TA 1978
Gene reversion (His+
revertant)
Gene reversion
Gene reversion
Gene reversion
5-20 uL AN
solution
2-300 uL AN
vapor
200 uL AN vapor
5-10 uL AN
solution
0.5 h
0.5-4 h
2h
0.5 h
+
(+) positive at
toxic dose of
45 mg/kg
+
+
-
-
-
i.p.: (+)
oral and i.v.: -
-
(-S9 / +S9)
-/+
-/+
-/+
-/+
Sekihashi et al.
(2002)
Sharief et al.
(1986)
Fahmy (1999)
Fahmy (1999)
Rabello-Gay
and Ahmed
(1980)
Leonard et al.
(1981)
Sharief et al.
(1986)
Leonard et al.
(1981)
Morita et al.
(1997)
Morita et al.
(1997)
Zhurkov et al.
(1983)
Leonard et al.
(1981)
Milvy and
Wolff (1977)
Milvy and
Wolff (1977)
Milvy and
Wolff (1977)
Milvy and
Wolff (1977)
241
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Table 4-54. Summary of studies on the mutagenicity/genotoxicity of AN
Test system
TA 1535
TA 1537, 1538, 98,
100
TA 1535
TA 1538
TA 97, 98, 100
TA102
TA98, 100, 1535,
1537, 1538
TA 1530, 1535,
1950, 1538, 100, 98,
1978
TA 98, 100
TA100
TA 1535/pSK1002
TA 1530
E. coli WP2 series
E. coli PQ37
Endpoint/effect
Gene reversion
Gene reversion
Gene reversion
Gene reversion
Gene reversion
Gene reversion
Gene reversion
Gene reversion
Gene reversion
Gene reversion
umu gene expression
(increased
p-galactosidase
activity)
Gene reversion
Gene reversion
(trp-->trp+)
SOS chromotest
Exposure
concentration
100-1,000 ug
AN plate
incorporation
assay
Plate
incorporation
assay
0.1-
1,000 |ag/dish
0.1-
10,000 |ag/dish
50-750 (ig/mL
plate
incorporation
assay
Up to
5,000 (a,g/plate
0.1-5,000 |ag/
plate
0.2% AN vapor
(200 (a,g/plate)
ND, plate
incorporation
assay
806-12,100 |ag/
plate
ND, O.lmLAN
in 2. 5 mL culture
medium
0. ImL 24-hr
urine from rats
and mice treated
with a single dose
30 mg/kg AN i.p.
(plate
incorporation
assay)
75 or 150 (omol/
plate
ND
Exposure
duration
ND
ND
ND
ND
48 h
ND
2d
Ih
ND
48 h
2h
48 h
2-3 d
2h
Result3
-/+
-/-
-/+
-/-
-/-
-/-
-/-
-/+ (weaker
response with
TA 100, 98, and
1978)
+/+
-/(+)
-/-
+
+c
-/-
Reference
Lijinsky and
Andrews
(1980)
Lijinsky and
Andrews
(1980)
Zhurkov et al.
(1983)
Zhurkov et al.
(1983)
Brams et al.
(1987)
Jung et al.
(1992)
Dow Chemical
Co. (1977)
de Meester et
al. (1978);
Dow Chemical
Co. (1976)
Khudoley et al.
(1987)
Hakura et al.
(2005)
Nakamura et
al. (1987)
Lambotte-
Vandepaer et
al. (1985, 1981,
1980)
Venitt et al.
(1977)
Brams et al.
(1987)
242
DRAFT - DO NOT CITE OR QUOTE
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Table 4-54. Summary of studies on the mutagenicity/genotoxicity of AN
Test system
Endpoint/effect
Exposure
concentration
Exposure
duration
Result3
Reference
Short-term assays — -fungi
S. cerevisiae JD 1
S. cerevisiae D61.M
Mitotic gene
conversion
Chromosome loss
250 or
500 (4,g/mL
0.8 or
1.36 mg/mL
18 h
16 h
-/+
-/-
Brooks et al.
(1985); Shell
Oil Co. (1984a)
Whittaker et al.
(1990)
Short-term assays — -fruit fly
Adult female
Drosophila ZESTE
system
Mosaic eye in
hatching females
Germ cells of male
D. melanogaster
Wing spots of
Drosophila
(mwh+l+flr +/mei-9)
D. melanogaster
ZESTE (inhibition of
taxol-purified
microtubule
assembly in vitro)
Sex chromosome loss
Somatic recombination
and mutation
Sex-linked recessive
lethal
Somatic mutation and
recombination
Aneuploidy
Inhalation
exposure to
2.7 ppm AN
Treatment of
larvae with 5-
20mMAN
Feeding:
420 ppm or
injection:
3,500 ppm
Gas exposure of
larvae to 0.5-
1 uL/l,150mL
5mM
50 min or
70 min
Single dose,
incubate for
9-1 Id
Feeding: 3 d
0.5 or 1 h
Microtubule
assembly
monitored for
80 min
+
(+)
(+)
+
Osgood et al.
(1991)
Vogel (1985)
Foureman et al.
(1994)
Wtirgler et al.
(1985)
Sehgal et al.
(1990)
Other short-term assays
Supercoiled plasmid
DNApBR322
SV40 phage DNA
CEO induced DNA
strand breaks
CEO induced DNA
strand breaks, but not
AN
50 mM CEO
incubated with
1-1 MS
supercoiled
pBR322 plasmid
DNA
ImMCEO
incubated with
5,000 dpm
[3H]-thymidine
labelled SV-40
phage DNA
3h
17 h
+
+
Yates et al.
(1994)
Peter et al.
(1983b)
243
DRAFT - DO NOT CITE OR QUOTE
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Table 4-54. Summary of studies on the mutagenicity/genotoxicity of AN
Test system
Endpoint/effect
Exposure
concentration
Exposure
duration
Result3
Reference
In vitro mammalian cell assays
Human lymphoblasts
Tk6
Human lymphoblasts
AHH-1
L5178Y mouse
lymphoma cells
L5178Y71+/+
L5178Y71+/-
P388F mouse
lymphoma tk+/~
V79/hprt
Gene-locus mutations
at the Tk (thymidine
kinase) locus (Tk+/~— >
Tk+)
Mutation at Tk locus
Mutation at Tk locus
Abnormality at
chromosome 17 in 1 of
8 CEO-induced tks
mutant clones
CEO-induced
mutations at the hprt
locus
Gene-locus mutations
at the hypoxanthine
guanine
phosphoribosyl
transferase locus
Mutation at Tk+l~ locus
Mutation at Tk+l~ locus
Mutation to ouabain or
6-TG resistance
Induction of Tk~^~
mutants
Induction of Tk~^
mutants
Induction of Tk~^~
mutants
Mutation to ouabain
resistance (Na+/K+
ATPase locus)
Mutation to
trifluorothymidine
resistance
Mutation to 5-iodo-
2-deoxyuridine
resistance
Induction of
8-azaguanine
resistance
40 ug/mL AN
1.4mMAN
100 uM and
150 uM CEO
NAb
NA
25 ug/mL AN
30 and 40 ug/mL
AN
10-40 ug/mL AN
12.5-200 ug/mL
AN
80-225 ug/mL
AN
30 nL/mL AN
5-69 ug/mL AN
12.5-100 ug/mL
AN
12.5-100 ug/mL
AN
80-160 ug/mL
AN
50-200 ug/mL
AN
3h
2h
2h
NA
NA
28 h
ND
4h
2h
2h
4h
3h
2h
2h
24-48 h
2h
-/+
-/+
+
+
+
+
+
+/+
+/+
+/+
+/+
+/+
-/+
-/-
Crespi et al.
(1985)
Recio and
Skopek (1988a,
b)
Recio and
Skopek (19883,
b)
Kodama et al.
(1989)
Recio and
Skopek
(19883)
Crespi et al.
(1985)
Oberly et al.
(1996)
Rudd (1983)
Garner and
Campbell
(1985)
Lee and
Webber (1985)
Myhr et al.
(1985)
Amacher and
Turner (1985)
Styles et al.
(1985)
Styles et al.
(1985)
Anderson and
Cross (1985)
Lee and
Webber (1985)
244
DRAFT - DO NOT CITE OR QUOTE
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Table 4-54. Summary of studies on the mutagenicity/genotoxicity of AN
Test system
HeLa cells
Human lymphocytes
Primary cultures of
F344 rat hepatocytes
Primary cultures of
F344 rat hepatocytes
Primary F344 rat
hepatocyte
Primary F344 rat
hepatocyte
Human mammary
epithelial cell
Human bronchial
epithelial cells
Rat hepatocytes
CHO cells
CHO cells
SHE cells
Human bronchial
epithelial cells
Human embryonic
lung fibroblasts
Human lymphocytes
CHO cells
CHO cells cocultured
with freshly isolated
rat hepatocytes
CHO cells
CHO cells
CHL fibroblast
CH1-L liver
fibroblast
Chinese hamster cell
line Don-6
Endpoint/effect
UDS
UDS
UDS
UDS
UDS
UDS
UDS
DNA single-strand
breaks
DNA single-strand
breaks
DNA single-strand
breaks
DNA single-strand
breaks
DNA single-strand
breaks
SCEs
Induction of p5 3 and
p21WAF1 protein
SCEs
SCEs
SCEs
SCEs
CAs
CAs
CAs
CAs
Exposure
concentration
2.5 mg/niL AN
5 x 10'1 M
0.026-53 ug/mL
AN
lO-'-lO2 ug/mL
AN
0.01-1 mM AN
0.01-0.1 mM
CEO
O.lmMCEO
200 and
500 ug/mL AN
65.8 ug/mL AN
7 x 10'2-
1 x 10'1 M AN
ND
200 or
400 ug/mL AN
150 and
300 ug/mL AN
0.3 to 1.0 mM
AN
5 x 10'4 M
10"7-10"4MAN
10'4MAN
2mMAN
4mMAN
12.5 ug/mL AN
2.5 ug/mL AN
1 x 1Q-4 M or
0.0053 mg/mL
Exposure
duration
2.5 h
4h
20 h
18-20 h
17-19 h
17-19 h
24 h
20 h
3h
Ih
ND
18 h
20 h
24 h
Ih
3h
3h
Ih
Ih
24 and 48 h
36 h
26-30 h
Result3
-/-
-/+
-
-
-
-
+
+
+
+/+
-/-
+
+
+
-/+
d
+d
-/+
+/+
+d
+
+
Reference
Martin and
Campbell
(1985)
Perocco et al.
(1982)
Probst and Hill
(1985)
Williams et al.
(1985)
Butterworth et
al. (1992)
Butterworth et
al. (1992)
Butterworth et
al. (1992)
Chang et al.
(1990)
Bradley (1985)
Douglas et al.
(1985)
Lakhanisky
and Hendrickx
(1985)
Parent and
Castro (1979)
Chang et al.
(1990)
Rossner et al.
(2002)
Perocco et al.
(1982)
Ved Brat and
Williams
(1982)
Ved Brat and
Williams
(1982)
Natarajan et al.
(1985)
Natarajan et al.
(1985)
Ishidate and
Sofuni (1985)
Danford (1985)
Sasaki et al.
(1980)
245
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Table 4-54. Summary of studies on the mutagenicity/genotoxicity of AN
Test system
CHO cells
Rat liver (RL4) cells
CHO cells
Endpoint/effect
MN
CAs, polyploidy,
SCEs
Inhibition of cell
growth and RNA
synthesis
Exposure
concentration
10"1 M AN
1.25, 2.5, 5.0, or
10 ug/mL AN
0.001,0.002, and
0.005% (v/v) AN
Exposure
duration
Ih
2h
8d
Result3
+/+
d
+d
Reference
Douglas et al.
(1985)
Priston and
Dean (1985);
Shell Oil Co.
(1984b)
Mangir et al.
(1991)
In vitro mammalian cell transformation
SHE cells
SHE cells
SHE cells
Mouse fibroblasts
NIH/3T3 cells
Mouse fibroblasts
C3H/10T1/2 cells
Mouse fibroblasts
C3H/10T1/2 cells
Balb/c-3T3 cells
Balb/c-3T3 cells
NIH 3T3 cells
Cell transformation
Enhancement of viral
transformation
Cell transformation
Cell transformation
Cell transformation
Cell transformation
Cell transformation
Ouabain-resistant
mutants
Cell transformation
after transfection with
CEO modified ras
DNA
50 or 100 ug/mL
AN
100 or
200 ug/mL AN
0.01-1 ug/mL
AN
12.5-200 jig/mL
AN
3-100 ug/mL AN
16 ug/mL AN
8.8 jig/mL AN
50 ug/mL AN
NA
6d
18 hrs before
SA7 or 5 hrs
after SA7
inoculation
7d
48 h
48 h
24 h
48 h
24 h
14 h
+
+
+
+
+
-/(+)
-/+C
-/+
Parent and
Castro (1979)
Parent and
Castro (1979)
Barett and
Lamb (1985)
Banerjee and
Segal (1986)
Banerjee and
Segal (1986)
Lawrence and
McGregor
(1985)
Matthews et al.
(1985)
Matthews et al.
(1985)
Yuan and
Wong (1991)
aNA = not applicable; ND = not determined.
b+ = Positive; - = negative; (+) = borderline positive.
°S9 activation not needed.
dS9 not included in the assay.
eWith or without coculture with lethally X-irradiated primary F344 RLCs.
In vitro evidence indicates that AN can be mutagenic, most likely through the formation
of CEO-DNA adducts. In short-term tests with bacteria, AN induced mutations in a majority of
test systems, most often requiring exogenous metabolic activation (Table 4-54). In mouse
lymphoma cell assays, AN induced mutations at the Tk locus in most assays (Table 4-54). In
human lymphoblast Tk6 cells devoid of CYP450 activity, AN induced multiple mutations at the
Tk locus only in the presence of metabolic activation (Recio and Skopek, 1988a, b; Crespi et al.,
246
DRAFT - DO NOT CITE OR QUOTE
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1985). CEO was mutagenic in Tk6 cells at 10-fold lower concentrations than AN itself and was
as mutagenic in this test system as the well-known mutagen, ethyl methanesulfonate (Recio and
Skopek, 1988a, b). CEO also induced multiple mutations at the hprt locus in Tk6 cells (Recio
and Skopek, 1988a). CEO induced UDS in cultured human mammary epithelial cells but not in
cultured rat hepatocytes (Butterworth, 1992).
Supporting evidence for the mutagenicity of AN and its metabolites comes from in vivo
studies (Table 4-54). Increases in the occurrence of MN in buccal mucosal cells and blood
lymphocytes were recently reported in AN exposed workers in China (Fan et al., 2006). Three
studies of AN workers reported genotoxic effects such as DNA strand breaks, nondisjunction of
sex chromosomes, and CAs (Beskid et al., 2006; Xu et al., 2003; Borba et al., 1996), whereas an
earlier study did not find elevated frequency of CAs in exposed workers (Thiess and Fleig,
1978). Increases in cytogenetic aberrations such as MN were found in assays of exposed rats
(Wakata et al., 1998) and a single assay of mice (Fahmy, 1999), but were not evident in other rat
and mouse assays (Morita et al., 1997; Zhurkov et al., 1983; Leonard et al., 1981; Rabello-Gay
and Ahmed, 1980). Comet assays found DNA damage in forestomach, colon, bladder, lung, and
brain in mice, following single i.p. injections of 20 mg/kg AN, and in forestomach, colon,
kidney, bladder, and lung of rats injected with 30 mg/kg (Sekihashi et al., 2002).
UDS was detected by following the time course of 3H-thymidine incorporation into DNA
in lung, testis, and gastric tissue, following administration of single oral doses of 46.5 mg/kg AN
to Sprague-Dawley rats (Ahmed et al., 1996b, 1992a, b; Abdel-Rahman et al., 1994a).
Autoradiographic techniques did not detect UDS following incubation of primary cultures of
hepatocytes or spermatocytes from F344 rats given single oral doses of 75 mg/kg or five daily
doses of 60 mg/kg-day AN (Butterworth et al., 1992). Dominant lethal effects (from mutations
in germ cells) were not found in mice given single i.p. doses of 30 mg/kg AN (Leonard et al.,
1981) or rats given five oral doses of 60 mg/kg (Working et al., 1987). DNA binding by AN or
its metabolites was indicated by elevated levels of radioactivity in DNA from several tissues in
rats given single oral doses of 46.5 mg/kg radiolabeled AN (Abdel-Rahman et al., 1994b; Ahmed
et al., 1992a, b; Farooqui and Ahmed, 1983a).
In the European Union Risk Assessment Report on Acrylonitrile (EC 2004), AN was
regarded as "genotoxic or at least mutagenic, despite the recent publication of Whysner et al.
(1998a) which argues for a possible nongenotoxic mechanism for the tumour induction in
experimental animals."
4.5.2.5. Genotoxicity resulting from oxidative stress
As discussed in Section 4.5.1.2.2, studies are available that investigated the genotoxicity
of AN resulting from oxidative stress. These studies measured 8-oxodG levels in DNA as
biomarkers of oxidative DNA damage from AN exposure. 8-OxodG is mutagenic (Kamiat et al.,
1992; Moriya et al., 1991; Wood et al., 1990) and causes GC —»• TA transversions during DNA
247 DRAFT - DO NOT CITE OR QUOTE
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replication.
In vitro studies
Zhang et al. (2000) studied AN-induced morphological transformation in SHE cells.
SHE cell culture was treated in vitro for up to 7 days with 0-75 ug/mL AN in 12.5 ug/mL
increments (75 ug/mL was cytotoxic, with 50% reduction in cell colony number). After 7 days
of exposure, there was a dose-dependent increase in morphological transformation at 50, 62.5,
and 75 ug/mL, reaching a transformation frequency of 1.3% at 75 ug/mL. After a 24-hour AN
exposure, an increase in transformation was not observed at all concentrations. Levels of
8-oxodG in DNA isolated from cells incubated with 75 |ig/mL were increased to 192 and 186%
of control after 2 and 3 days, indicating an association between cell transformation and oxidative
DNA damage. However, no increase in 8-oxodG was observed after 1 or 7 days. AN-induced
morphological transformation was inhibited by cotreatment with the antioxidants a-tocopherol
(100 uM: up to 65% inhibition) and (-)-epigallocathechin-3-gallate (5 uM: up to 87%
inhibition) for 7 days. Cotreatment with antioxidants also inhibited the formation of 8-oxodG in
SHE cell DNA from treatment with 75 ug/mL AN.
In a later study by the same research group, Zhang et al. (2002) investigated the time
course of SHE cell morphological transformation in the presence of 75 ug/mL AN. The results
indicated that statistically significant increases in morphological transformation frequency could
not be observed until after 2 consecutive days of exposure; a plateau at a transformation
frequency of about 2% was reached after 4-5 days of exposure. In another experiment,
coadministration of 0.5 mmol/L ABT, a nonspecific suicidal CYP450 inhibitor, for 7 days
significantly reduced the rate of cell transformation from about 1.25% to about 0.3% (shown
graphically), demonstrating the need for metabolic activation of AN in this test system.
Zhang et al. (2002) also showed that AN (25, 50, and 75 ug/mL) increased the amount of
ROS (measured by 2,3-dihydroxybenzoic acid production) in the SHE cells after 4, 24, and
48 hours of treatment. At the same time, xanthine oxidase (which generates the superoxide
radical and hydrogen peroxide via oxidation of hypoxanthine or xanthine by oxygen) activity
was increased by 47% in SHE cells after 24 hours of treatment with 75 ug/mL AN. After
48 hours of treatment, xanthine oxidase activity was increased in both 50 and 75 ug/mL (80%)
AN groups. This increase in xanthine oxidase activity was blocked by cotreatment with 0.5 mM
ABT. On the other hand, antioxidant GSH was depleted 66-80% by all doses of AN (25, 50, and
75 ug/mL) after 4 hours of treatment and returned to control levels after 24 hours. At 48 hours, a
significant increase in GSH was observed with the 75 ug/mL group but not in other dose groups.
Antioxidant enzyme catalase activity was significantly decreased after 4 hours of treatment with
50 and 75 ug/mL AN but increased after 24 and 48 hours of treatment. Cotreatment with ABT
prevented the decrease and increase in activity of catalase at 4 and 24 hours after treatment,
248 DRAFT - DO NOT CITE OR QUOTE
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respectively. A transient decrease in SOD activity was also observed after 4 hours of treatment
with 75 ug/mL AN.
The effect of AN on catalase and SOD activities was also studied in a cell-free system
(Zhang et al., 2002). Catalase and SOD activities were determined in samples containing
purified catalase or SOD incubated with 75 ug/mL AN in the presence or absence of SHE cell
homogenate. In the absence of a metabolic source of SHE cell homogenate, no inhibition of
catalase activity was seen following incubation up to 60 minutes with AN. In the presence of
SHE cell homogenate, AN significantly decreased catalase activity in a time-dependent manner
after 10 minutes of incubation. Similarly, a significant time-dependent decrease in SOD activity
was observed following 30 minutes incubation with AN. The study authors concluded that
morphological transformation of SHE cells is caused by oxidative stress as a result of oxidative
metabolism of AN.
Kamendulis et al. (1999a) also investigated oxidative DNA damage induced by AN in
DITNC1, a rat glial astrocyte cell line, and primary rat hepatocytes exposed to AN in vitro (see
Section 4.5.1.2.2). AN was cytotoxic at concentrations >2.5 mmol/L (133 ug/mL) (as measured
by the release of LDH from the cells) to both cell lines, following >4 hours of exposure.
Concentrations of 0.01, 0.1, and 1.0 mmol/L AN (0.53, 5.3, and 53 ug/mL, respectively) caused
a dose-dependent increase in formation of 8-oxodG in astrocytes but not in hepatocytes.
Corresponding increases in ROS formation were also observed in astrocytes only. However, no
oxidative lipid damage (as evaluated by formation of MDA, a product of lipid peroxidation) was
found in either cell type following treatment with AN at all exposure concentrations or durations.
The formation of 8-oxodG in rat astrocytes was reversible. Following treatment with AN for
24 hours and removal of AN for 24 hours afterwards, 8-oxodG levels returned to control values
in all concentrations examined. Kamendulis et al. (1999a) concluded that this demonstrated
property was consistent with tumor promoting agents.
Pu et al. (2006) investigated oxidative DNA damage induced by AN in D1TNC1 rat
astrocyte cell line using the fpg-modified comet assay. Increase in oxidative DNA damage was
observed in astrocytes treated with 1 mM AN for 24 hours. No increase in oxidative DNA
damage was observed in astrocytes treated with 0.005-0.75 mM AN for 24 hours.
Jacob and Ahmed (2003b) investigated oxidative stress in cultured NHA (4631) treated
with up to 400 uM AN for 12 hours. AN was cytotoxic at concentrations >50 uM. Cell viability
was 85, 78, and 58%, respectively, at AN concentrations of 100, 200, and 400 uM. Significant
increases in measures of oxidative stress were observed at cytotoxic concentrations of 200-400
uM AN: the production of ROS was increased four- to sevenfold, whereas 8-oxodG levels were
increased more than twofold.
Esmat et al. (2007) investigated cytoxicity in rat (strain not known) primary glial cells
exposed to 0-5.0 mM AN up to 12 hours. Cell membrane integrity was evaluated by trypan blue
exclusion and LDH leakage. About 50% membrane damage in primary glial cells was observed
249 DRAFT - DO NOT CITE OR QUOTE
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in incubations containing 1.0 mM AN for 3 hours. Thus, subsequent studies on AN-induced
oxidative stress were performed using 1 mM AN for 3 hours incubation. AN increased MDA
levels (indicator of lipid peroxidation) to about ninefold compared with control incubations and
depleted GSH level to about 7% of controls, while no change in total glutathione level was
observed. AN induced CIST formation by glial cells and decreased ATP level by about 90% as
compared with the control. Pretreatment with 5 mM NAC (an acetylated precursor of cysteine
and GSH, and an antioxidant), reduced MDA level by 40% as compared with glial cells treated
with AN alone and raised GSH and total glutathione level in cell extract to about 2.5-fold of
control and AN alone treated group. Pretreatment with NAC also caused reduction of CN~
induced by AN to about 15% as compared with the AN alone treated group, and raised ATP
level to about sixfold, as compared with the AN alone treated group.
It should be noted that NAC stimulated GSH synthesis, enhanced glutathione-S-
transferase activity, and was a powerful nucleophile capable of scavenging free radicals (De
Vries and De Flora, 1993). Observed increases in GSH and total glutathione and decrease in CN"
formation with NAC pretreatment in Esmat et al. (2007) would suggest that most of the
administered AN could be detoxified via conjugation via the GSH pathway (consistent with
findings by Carerra et al., 2007), as less AN was available for oxidation to CN". Esmat et al.
(2007) concluded that AN toxicity was at least partly mediated by oxidative stress.
In vivo studies
As described in Section 4.5.1.2.2, Jiang et al. (1998) reported increases in measures of
oxidative stress in the brain cortices but not the livers of male Sprague-Dawley rats exposed to
AN in drinking water at concentrations of 50-200 ppm for up to 90 days. Increases in ROS and
oxidative DNA damage (increased levels of 8-oxodG) and concomitant decreases in antioxidant
enzymes (catalase, SOD) were observed in the brain cortices of exposed rats. Transient small
decreases in the antioxidants vitamin E and glutathione were also observed in the brain cortices
of AN-exposed rats only at 14 days. A transient increase in MDA level was observed only at the
highest dose group after 14 days but not at other time points.
Pu et al. (2009) reported oxidative DNA damage in WBC and brain of male Sprague-
Dawley rats treated with 100-200 ppm AN. Further discussion of study findings are provided in
Section 4.7.3.3.1..
In another study, Whysner et al. (1998a) examined the ability of AN to induce oxidative
DNA damage in the brain, liver, and forestomach of rats by exposing male Sprague-Dawley and
F344 rats up to 300 ppm AN in drinking water for 21 days (see Section 4.5.1.2.2). As shown in
Table 4-55, elevated levels of 8-oxodG were found in DNA from the brain and liver of Sprague-
Dawley rats. However, 8-oxodG levels in the forestomach of exposed Sprague-Dawley rats and
the brain of exposed F344 rats were not statistically significantly elevated compared with
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controls. Significant increase in 8-oxodG levels was found in the liver of exposed Sprague-
Dawley rats, although the liver is not a target organ for carcinogenicity in adult rats.
Table 4-55. Formation of 8-oxodG in DNA from tissues of male Sprague-
Dawley and F344 rats exposed to AN in drinking water for 21 days
Dose group
Control
3 ppm
30ppm
300 ppm
Formation of 8-oxodG (mol/105 mol dG)
Sprague-Dawley rats
Liver
0.67 ± 0.22
0.72 ± 0.06
0.95±0.19a
0.96±0.15a
Forestomach
0.68 ±0.14
1.77 ±0.55
1.59 ±0.30
1.44± 1.22
Brain
0.62 ±0.08
0.86 ±0.41
1.35±0.49a
1.29±0.10a
F344 rats
Brain
0.79 ±0.37
1.07 ±0.41
1.03 ±0.38
1.06±0.48b
aSignificantly different from controls (p < 0.05) as calculated by the study authors.
bExposure was at 100 ppm.
Source: Whysneretal. (1998a).
Whysner et al. (1998a) also exposed male Sprague-Dawley rats to 100 ppm AN in
drinking water for up to 94 days. (This dose was carcinogenic in Sprague-Dawley rats in a
chronic study [Johannsen and Levinskas, 2002a; Biodynamics, 1980b].) The rats were divided
into four groups and were exposed to distilled water, 100 ppm AN, 5 mg MNU (a DNA-reactive
carcinogen that produces glial cell tumors in rats) per week, or 100 ppm AN plus 5 mg MNU per
week. Levels of 8-oxodG in brain and liver of rats exposed to 100 ppm AN were significantly
greater than those in controls after 10 days (1.31 ± 0.52 in brains of exposed rats vs. 0.65 ±
0.22 mol per 105 mol dGin controls and 0.70 ± 0.20 in livers of exposed rats vs. 0.49 ± 0.10 mol
per 105 mol dG in controls). Administration of 5 mg/kg MNU alone did not increase the level of
8-oxodG in the brain of treated rats but increased 8-oxodG in the liver after 10 days. However,
coadministration of 100 ppm AN and MNU increased 8-oxodG level in the brain after 31 days
and 94 days when compared with the MNU-only group.
Whysner et al. (1998a) suggested that AN-induced tumors may be produced by a mode of
action involving 8-oxodG. However, several findings in this study did not support this proposed
mode of action. First, no significant increase in 8-oxodG levels were found in the brain DNA of
F344 rats exposed to AN in the 21-day study whereas AN was carcinogenic to F344 rats in a
chronic drinking water study (Johannsen and Levinskas, 2002b). Second, although a significant
increase in 8-oxodG levels was found in the brain DNA of Sprague-Dawley rats exposed to AN
for 21 days, no dose-dependent increase was observed above 30 ppm, which was not the dose
producing the highest occurrence of tumors in the chronic bioassay. Third, no increase in
8-oxodG levels was found in the forestomach DNA of exposed rats. The forestomach was a
target organ for AN carcinogenicity. Finally, increase in 8-oxodG levels was found in liver
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DNA of exposed rats. The liver was not a target organ for AN carcinogen!city in adult rats.
Therefore, there was no association between 8-oxodG levels and tumorigenicity in target organs.
Results of mutagenicity/genotoxicity studies of AN are summarized in Table 4-56.
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Table 4-56. Summary of studies on the mutagenicity or genotoxicity
resulting from oxidative stress of AN
Test system
Endpoint/effect
Exposure
concentration
Exposure
duration
Result
Reference
In vitro mammalian cell assays
SHE cells
SHE cells
SHE cells
DITNClratglial
astrocytes
D1TNC1 rat
astrocytes
Rat hepatocytes
NHAs
NHAs
Cell
transformation
8-oxodGinDNA
ROS
8-oxodGinDNA
DNA damage in
fpg-comet assay
8-oxodGinDNA
8-oxodGinDNA
ROS
50-75 ug/mL AN
75 ug/mL AN
25-75 ug/mL AN
0.01-1 mM AN
(0.53-53 ug/mL)
ImM
0.01-1 mM AN
200-400 uM AN
200-400 uM AN
7d
Increased after
2 and 3 d but not
after 1 or 7 d
4-48 h
4 or 24 h
24 h
4 or 24 h
12 h
12 h
+
(+)
+
+a
+
-
+
+
Zhang et al.
(2000)
Zhang et al.
(2000)
Zhang et al.
(2002)
Kamendulis et al.
(1999a)
Pu et al. (2006)
Kamendulis et al.
(1999a)
Jacob and Ahmed
(2003b)
Jacob and Ahmed
(2003b)
In vivo studies in rats
Male Sprague-
Dawley rats
Male Sprague-
Dawley rats
Male Sprague-
Dawley rats
Male Sprague-
Dawley rats
Male Sprague-
Dawley rats
Male Sprague-
Dawley rats
Male F344 rats
Male Sprague-
Dawley rats
8-oxodG in brain
cortex DNA
8-oxodG in liver
DNA
8-oxodG in WBC
and brain DNA
8-oxodG in brain
DNA
8-oxodG in liver
DNA
8-oxodG in
forestomach DNA
8-oxodG in brain
DNA
8-oxodG in brain
DNA
0-200 ppm AN in
drinking water
0-200 ppm AN in
drinking water
100 or 200 ppm
AN in drinking
water
30 or 300 ppm
AN in drinking
water
30 or 300 ppm
AN in drinking
water
0-300 ppm AN in
drinking water
0-100 ppm AN in
drinking water
100 ppm AN in
drinking water
100 and 200 ppm:
+ after 14-90 d
50 ppm: + after
28 and 90 d
14-90 d
28 days
21 d
21 d
21 d
21 d
3-94 d
+
-
+
+
+
-
-
+
Jiang etal. (1998)
Jiang etal. (1998)
Pu et al. (2009)
Whysner et al.
(1998a)
Whysner et al.
(1998a)
Whysner et al.
(1998a)
Whysner et al.
(1998a)
Whysner et al.
(1998a)
aThe formation of 8-oxodG was reversible. Following treatment with AN for 24 hrs and removal of AN for 24 hrs
afterwards, 8-oxodG levels returned to control values.
+ = Positive;- = negative; (+) = borderline positive.
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4.6. SYNTHESIS OF MAJOR NONCANCER EFFECTS
4.6.1. Oral
No studies are available regarding noncancer health effects in humans following acute,
subchronic, or chronic oral exposure to AN. The major noncancer findings from repeat-dose oral
toxicity studies of AN in experimental animals are summarized in Table 4-57.
Table 4-57. Summary of noncancer oral toxicity study findings for AN
Reference
Species
Johannsen and
Levinskas
(2002b);
Biodynamics
(1980c)
F344 rat
Johannsen and
Levinskas (2002a);
Biodynamics
(1980a)
Spr ague-Daw ley
rat
Johannsen and
Levinskas (2002a);
Biodynamics
(1980b)
Sprague-Dawley
rat
Exposure conditions,
mg/kg-d
0,0.1,0.3,0.8,2.5,8.4
(M)
0,0.1,0.4, 1.3,3.7, 10.9
(F)
DW, 2 yrs; sacrifices at 6,
12, and 18 mos and
termination (99 wks or
699-706 d[F]; 107 wks or
770-777 d [M])
0, 0.09, 8.0 (M)
0,0.15, 10.7 (F)
DW, 22 mos (M); 19 mos
(F); sacrifices at 6, 12, and
18 mos and termination
0,0.1, 10(M + F)
Gavage, 7 d/wk for 20
mos; sacrifices at 6, 12,
and 18 mos and
termination
NOAEL
(mg/kg-d)
O.l(M)
0.1 (F)
2.5 (M)
1.3 (F)
2.5 (M)
3.7 (F)
0.4 (F)
2.5 (M)
ND(M)
0.15(F)
8.0 (M)
0.15(F)
0.09 (M)
0.15(F)
0.1 (M + F)
0.1
0.1
LOAEL
(mg/kg-d)
0.3a (M)
0.4a (F)
8.4b(M)
3.7b(F)
8.4a(M)
10.9a(F)
1.3b(F)
8.4b(M)
0.09b(M)
10.7a (F)
ND(M)
10.7a (F)
8.0b (M)
10.7a (F)
10a (M + F)
10a
10a
Effect
Forestomach squamous cell hyperplasia
and hyperkeratosis, increased incidence
Decreased survival after 18 mos
Decreased BW (>10% compared with
control)
Increase in serum alkaline phosphatase
(F only)
Increase in epidermal inclusion cysts
(M only)
Forestomach squamous cell
hyperplasia, increased severity
Renal transitional cell hyperplasia
(F only)
Decreased survival at 10 mos (M + F);
decreased BW (10/8% [M/F])
Forestomach squamous cell
hyperplasia, increased severity
Renal transitional cell hyperplasia at
some but not all sacrifices
Decreased survival after 14 mos (M +
F); decreased BW (6-13% compared
with controls [M only]).
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Table 4-57. Summary of noncancer oral toxicity study findings for AN
Reference
Species
Quast (2002);
Quast et al.
(1980a)
Spf ague-Daw ley
rat
NTP (2001)
B6C3F] mouse
NTP (2001)
B6C3Fj mouse
Tandon et al.
(1988)
CD-I mouse
Friedman and
Bellies (2002);
Litton Bionetics
(1992)
Spr ague-Daw ley
rat
Exposure conditions,
mg/kg-d
0, 3.4, 8.5, 21.3 (M)
0, 4.4, 10.8, 25.0 (F)
DW, 2 yrs
0,2.5, 10, 20(M + F)
Gavage, 5 d/wk for 2 yrs
Adjusted doses: 0,1.8,
7.1, 14.3 mg/kg-d (M + F)
0, 5, 10, 20, 40, 60 (M +
F)
Gavage, 5 d/wk for 14
wks.
Adjusted doses: 0, 3.6,
7.1, 14.3,28.6, and 42.9
(M + F)
0, 1, 10 (M only)
Gavage, for 60 d
0, 11, 37 (M)
0, 20, 40 (F)
DW, three-generation
reproduction study
NOAEL
(mg/kg-d)
3.4 (M)
ND(F)
ND(F)
ND(M)
4.4 (F)
8.5 (M)
ND(F)
8.5 (M)
10.8 (F)
1.8 (M)
7.1 (F)
1.8 (M)
14.3 (F)
ND(F)
7.1 (M + F)
28.6 (M)
14.3 (F)
14.3 (M + F)
1
ND
ND
LOAEL
(mg/kg-d)
8.5b(M)
4.4b(F)
4.4b(F)
3.4b(M)
10.8b (F)
21.3b(M)
4.4b (F)
21.3b(M)
25.0b (F)
7.1b(M)
14.3b(F)
7.1b(M)
ND(F)
1.8b(F)
14.3a(M + F)
ND(M)
28.6a (F)
28.6b (M + F)
10b
llb(M)
20b (F)
llb(M)
20b(F)
Effect
Forestomach squamous cell hyperplasia
and hyperkeratosis, increased incidence
Gliosis in brain with or without
perivascular cuffing in females; not
significant in males
Chronic nephropathy
Decreased survival after 300 (F) or 480
(M)d
Decreased BW (>10% compared with
control); clinical signs of nervous
system dysfunction
Forestomach squamous cell hyperplasia
and hyperkeratosis, increased incidence
Increased incidence of Harderian gland
hyperplasia
Increased incidence of ovarian cysts or
ovarian atrophy
Decreased survival after 15 wks
Forestomach inflammation, ulceration,
and epithelial hyperplasia, increased
incidence in F only
No histopathology data for 60 rng/kg
groups
Increased mortality
45% decreased sperm count; 40% of
seminiferous tubules examined
degenerated
Decreased viability for Fib pups;
no changes in fertility index or
pregnancy outcome forFla, Fib, F2a,
F2b, F3a, orF3b generations
LOAEL for decreased lactation index
in Fib parents for F2a pups; small
deficits in postnatal pup weights (10-
40%, variable across generations) and
postnatal survival (about 10%, variable
across generations) in higher dosed
groups
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Table 4-57. Summary of noncancer oral toxicity study findings for AN
Reference
Species
Dow Chemical
(1976b)
Sprague-Dawley
rat
Gagnaire et al.
(1998)
Spr ague-Daw ley
rat
Rongzhu et al.
(2007)
Spr ague-Daw ley
rat
Szabo etal,
(1984)
Spr ague-Daw ley
rats
Szabo etal. (1984)
Spr ague-Daw ley
rats
Quast etal. (1975)
Beagle dog
Exposure conditions,
mg/kg-d
0, 10, 25, 65 (F only)
Gavage CDs 6-15
0, 12.5, 25, 50 (M only)
Gavage, 5 d/wk for 12
wks
0,4, 13.5 (M only)
DWforO, 4, 8, 12 wks
0, 0.2, 4, 20, and 100
mg/kg-day (F only)
DW for up to 60 days
0, 0.2, 4, 20 or 100
mg/kg-day (F only)
Daily gavage for up to 60
days
0, 10, 16, 17 (M)
0, 8, 17, 18 (F)
DW for 6 months
NOAEL
(mg/kg-d)
10
10
25
ND
4
ND
10 (M)
8(F)
LOAEL
(mg/kg-d)
25
25
50a
4b
20
0.2
16 (M)
17 (F)
Effect
Maternal effects: 3/33 with
forestomach hyperplasia
Fetal effects: 9% increase in litters
with any malformation
SCV decreased beginning wk 6
(-7.5%), -17.8% by wk 12;
-10.6% after 8 wks recovery; decreased
BW after 4 wks (-17% at wk 12); hind
limb weakness in 5/1 1 surviving high-
dose rats after wk 9
Neurobehavioral alterations, as
indicated by decreased motor
coordination, increased training
duration, head twitching, trembling,
circling, backwards pedaling, and
decreased home-cage activities
Enlarged kidneys, increase in regional
hyperplasia of the gastric mucosa
Adrenocortical hyperplasia
Early mortality, histopathological
lesions of the esophagus and tongue
aSignificantly different from controls (p < 0.01).
bSignificantly different from controls (p < 0.05).
DW = drinking water; F = female; M = male; ND = not determined
Forestomach lesions (epithelial hyperplasia and hyperkeratosis) were the most
consistently observed noncancer effect associated with chronic oral exposure of rats and mice to
AN and were associated with the lowest LOAELs in AN-exposed rats (Table 4-57). The lowest
LOAELs were 0.09 and 10.7 mg/kg-day for increased severity of forestomach lesions in male
and female Sprague-Dawley rats exposed by drinking water (Johannsen and Levinskas, 2002a;
Biodynamics, 1980a), 0.3 and 0.4 mg/kg-day for increased incidence of forestomach lesions in
male and female F344 rats exposed in drinking water (Johannsen and Levinskas, 2002b;
Biodynamics, 1980c), and 8.5 and 4.4 mg/kg-day for increased incidence of forestomach lesions
in male and female Sprague-Dawley rats exposed in drinking water (Quast, 2002; Quast et al.,
1980a). Increased incidences of animals with hyperplasia or hyperkeratosis of the forestomach
were also observed in B6C3Fi mice exposed by gavage (5 days/week) for 2 years to 10 mg/kg-
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day (males) or 20 mg/kg-day (females) (NTP, 2001), female B6C3Fi mice exposed to 40 mg/kg-
day by gavage (5 days/week) for 14 weeks (NTP, 2001), and pregnant Sprague-Dawley rats
exposed to 25 mg/kg-day by gavage on GDs 6-15 (Murray et al., 1978). Other effects observed
in orally exposed animals generally were observed at higher doses (Table 4-57).
Kidney effects included increased incidence of renal transitional cell hyperplasia in
Sprague-Dawley rats exposed by gavage or in drinking water to about 10 mg/kg-day but not
0.1 mg/kg-day (Johannsen and Levinskas, 2002a; Biodynamics, 1980a) and increased incidence
of chronic nephropathy in Sprague-Dawley rats exposed to drinking water doses of 3.4 mg/kg-
day (males) or 10.8 mg/kg-day (females) (Quast 2002; Quast et al., 1980a).
Decreased survival was observed after 18 months in F344 rats exposed to drinking water
at 8.4 mg/kg-day (males) and 3.7 mg/kg-day (females) (Johnannsen and Levinskas, 2002b) and
in Sprague-Dawley rats exposed to drinking water at 21.2 mg/kg-day (males) and 4.4 mg/kg-day
(females) (Quast, 2002). In another drinking water study of Sprague-Dawley rats, decreased
survival was observed at 8.0 mg/kg-day (males) and 10.7 mg/kg-day (females).
Increased incidences of ovarian lesions (cysts or atrophy) were observed in female mice
exposed to gavage doses of 2.5, 10, or 20 mg/kg-day (NTP, 2001), but exposure-related ovarian
lesions were not observed in the chronic rat bioassays. In a reproductive toxicity study in
Sprague-Dawley rats exposed to 0, 100, or 500 ppm AN in drinking water (11 or 37 mg/kg-day
[males]; 20 or 40 mg/kg-day [females]), small deficits in postnatal pup weights and pup survival
without effects on fertility or pregnancy success were observed in three generations exposed to
either 100 or 500 ppm (Friedman and Beliles, 2002; Litton Bionetics, 1992) (see Table 4-58).
Other reproductive effects observed in animals were a reduction of epididymal sperm counts and
degeneration of seminiferous tubules in CD-I mice exposed to 10 mg/kg-day by gavage for
60 days (Tandon et al., 1988), but no exposure-related lesions in male reproductive organs were
found in B6C3Fi mice exposed by gavage to up to 40 mg/kg-day for 14 weeks or up to
20 mg/kg-day for 2 years (NTP, 2001). Likewise, no effects on sperm motility parameters were
found in male B6C3Fi mice exposed to up to 20 mg/kg-day for 14 weeks (NTP, 2001).
The only oral developmental toxicity study (Dow Chemical, 1976b) identified 10 and
25 mg/kg-day as a NOAEL and LOAEL, respectively, for maternal effects (increased incidence
of forestomach hyperplasia) and developmental effects (increased incidence of litters with any
malformations). The study involved daily gavage exposure on GDs 6-15.
Neurological effects in animals associated with chronic oral exposure to AN were
observed in: (1) a report of gross clinical signs of neurological impairment in about 10% of
Sprague-Dawley rats (paralysis, head tilt, circling, and seizures) exposed to drinking water
concentrations of 500 ppm (providing doses of about 65 and 74 mg/kg-day for males and
females) but not in rats exposed to 100 ppm (13 and 15 mg/kg-day, males and females) (Bigner
et al., 1986); (2) decreases (8-15%) in SCV in tail nerves and hind-limb weakness in male
Sprague-Dawley rats exposed by gavage to 50 mg/kg-day (5 days/week) for 6-12 weeks but not
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in rats exposed to 25 mg/kg-day (Gagnaire et al., 1998); and (3) neurobehavioral alterations in
male Sprague-Dawley rats exposed to 4 and 13.46 mg/kg-day AN for 4, 8, or 12 weeks (head
twitching, trembling, circling, backwards pedaling, decreased home-cage activities, decreased
motor coordination, and learning and memory) (Rongshu et al., 2007). These studies indicated
that neurological impairment in AN-exposed rats occurred at higher administered doses than
doses associated with hyperplasia and hyperkeratosis in forestomach squamous epithelial cells.
4.6.2. Inhalation
Acute inhalation exposure to AN can cause blood chemistry changes indicative of slight
liver damage and a range of subjective symptoms (including dizziness, headache, chest tightness,
feebleness, hyperactive knee jerk, sore throat, dyspnea, vomiting, abdominal pain, fainting, and
congestion of the pharynx (Chen et al., 1999). Less frequently reported symptoms include
numbness of limbs, convulsion, rapid heart rate, cough, hoarseness, coma, and abnormal liver
function. AN poisoning victims are generally treated with antidotes for cyanide poisoning and
pure oxygen to overcome respiratory distress caused by damage to the lung (Chen et al., 1999).
Table 4-58 summarizes the effects of AN as reported in epidemiological investigations of
populations occupationally exposed to AN. The LOAELs and NOAELs in Table 4-58 represent
mean or midpoint values of the range of reported exposure concentrations. The major noncancer
findings from repeat-dose inhalation toxicity studies of AN in experimental animals are
summarized in Table 4-59.
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Table 4-58. Noncancer effects observed in epidemiology studies of worker
cohorts exposed to AN
Reference/
Study subjects
Exposure
Characteristics
NOAEL
(ppm)
LOAEL
(ppm)
Effect
Symptoms and clinical chemistry
Sakuraietal. (1978)
acrylic fiber workers
(n = 102; 62 controls),
males
Mutoetal. (1992)
acrylic fiber workers
(n= 157; 537 controls),
males
Kaneko and Omae
(1992)
acrylic fiber workers
(n = 504; 249 controls)
Chen etal. (2000)
acrylic fiber workers
(n=224, 180 males, 44
females; 224 controls)
Average 10-12 yrs
exposure
Average 17 yrs
exposure
Average 5.6, 7.0,
8.6 yrs exposure;
mean AN
concentrations 1.8,
7.4, 14.1 ppm in 3
groups of factories
Average 13 yrs
exposure
ND
0.19
ND
ND
4.2
1.13
1.8
0.48
Increased incidence of palpable liver,
reddening of the conjunctiva and pharynx,
skin rashes compared with unexposed
controls, but not statistically significant.
Statistically significantly increased
prevalence of subjective symptoms (e.g.,
heaviness of stomach, poor memory,
irritability); no increases in physical signs or
abnormal values in urinary, hematological,
liver function, or blood pressure variables
Statistically significantly increased
prevalence of subjective symptoms (e.g.,
headaches, tongue trouble, choking lump in
chest, fatigue) in workers from all three
groups of factories
Statistically significantly increased
prevalence of subjective symptoms (e.g.,
headache, dizziness, poor memory, choking
feeling in chest, loss of appetite)
Neurological effects
Lu et al. (2005a)
AN -monomer workers:
(n=81; 68 males, 13
females); AN-acrylic
fiber workers (n=94; 67
males, 27 females); 174
controls (130 males, 44
females)
>1 yr exposure
(average duration
not available)
ND
0.11
Small, but statistically significant, deficits in
tests of neurobehavior in monomer workers
and fiber workers (geometric mean air
concentrations of 0. 1 1 ppm for monomer
workers and 0.91 ppm for fiber workers)
Reproductive and developmental effects
Dong et al. (2000b)
Acrylic fiber workers
(n=548 males, 391
females; 496 male and
427 female controls)
Li (2000)
AN manufacturing
workers (n= 379; 511
controls), female
Xu et al. (2003)
AN production workers
(n = 30; 30 controls),
males
Average 1 1 and
10.4 yrs exposure
Average 14 yrs
exposure
Mean 2. 8 yrs
exposure
ND
ND
ND
3.6
7.5
0.37
Statistically significantly increased
prevalence of adverse reproductive outcomes
(increased stillbirths [2.7 vs. 0.7%], birth
defects [2.13 vs. 0.48%], and premature
deliveries [8.2 vs. 3.9%]) in female workers
Statistically significantly increased
prevalence of adverse reproductive outcomes
(sterility [2.6 vs. 0.8%], pregnancy
complications [20.8 vs. 7.1%], premature
deliveries [11.6 vs. 4.7%], and congenital
defects [25.4 vs. 4.2%]) in female exposed
workers
Statistically significant decrease in sperm
density and sperm number and increase in
DNA strand breakage and sex chromosome
aneuploidy in sperm cells in exposed workers
ND = cannot be determined
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Table 4-59. Summary of noncancer inhalation toxicity study findings for AN
Reference/
Study subjects
Exposure conditions
NOAEL
(ppm)
LOAEL
(ppm)
Effect
Chronic toxicity studies
Quast et al.
(1980b)
Sprague-Dawley
rats (M + F,
100/sex/group)
Gagnaire et al.
(1998)
Sprague-Dawley
rats (M only,
12/group)
0, 20, 80 ppm, 6 hrs/d,
5 d/wk for 2 yrs
0, 25, 50, 100 ppm,
6 hrs/d, 5 d/wk for 24
wks
ND
ND
20
25
Statistically significant increase in
incidence of lesions in nasal epithelia at 20
ppm in males and females; focal necrosis in
liver of females. At 80 ppm, other lesions
occurred at increased incidences — gliosis
and perivascular cuffing in brain in both
sexes, hepatic necrosis in females, focal
nephrosis and thyroid cysts in males, and
hyperplasia in nonglandular stomach.
Statistically significant deficits (5%
decreased compared with controls) in
sensory conduction velocity of the tail
nerve.
Reproductive and developmental studies
Haskell Laboratory
(1992a)
Pregnant Sprague-
Dawley rats
(30/group)
Saillenfait et al.
(1993)
Pregnant Sprague-
Dawley rats
(20-2 I/group)
Nemec et al.
(2008)
Crl:CD (SD) rats
(M + F,
25/sex/group)
two-generation
reproductive study
0, 40, or 80 ppm, 6 hrs/d
on CDs 6-15
0, 12, 25, 50, 100 ppm,
6 hrs/d on CDs 6-20
FO: 0, 5, 15, 45 ppm
6 hrs/d, 7 d/wk for 10
weeks via whole-body
inhalation
Fl : exposed in utero, via
milk through nursing
during PNDs 0 to 28,
then exposed to 0, 5, 15,
or 45 ppm AN for 6
hrs/d, 7 d/wk for 10
weeks via whole-body
inhalation
ND
40
12
12
15
ND
40
80
25
25
45
5
Maternal weight gain (GD6-15) decreased
by >20% compared with controls.
6/35 litters with any malformation (short
tail, short trunk, missing vertebrae, missing
ribs, or anteriorly displaced ovaries) vs. 1/33
in controls.
Statistically significantly decreased maternal
weight gain compared with controls.
Statistically significantly decreased (>5%)
fetal BW compared with controls. No
exposure-related increased incidences of
litters with fetal anomalies.
FO generation: histological changes in nasal
tissues
Fl generation: histological changes in nasal
tissues
ND = cannot be determined
In the cross-sectional studies of AN-exposed workers, an increased prevalence compared
with unexposed workers of subjective symptoms such as dizziness, headache, chest tightness,
and poor memory was seen, indicating respiratory irritation and neurological effects. Average
workplace air concentrations associated with these subjective symptoms were 1.13 ppm (Muto et
al., 1992), 1.8 ppm (Kaneko and Omae, 1992), and 0.48 ppm (Chen et al., 2000) (see Table 4-
58). No statistically significant increases in the prevalence of subjective symptoms were found
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in a group of acrylic fiber workers whose average workplace air concentration was 0.19 ppm
(Muto et al., 1992). The studies by Muto et al. (1992) and Sakurai et al. (1978) included clinical
physical examinations, but no statistically significant increases in the prevalence of physical
signs (such as reddened conjunctiva or pharynx) or abnormal values in clinical chemistry
variables (including activities of liver enzymes) were found in exposed workers (more details of
results from these studies can be found in Section 4.1.2.2.2).
Exposure levels associated nasal lesions (and lesions at other sites) in Sprague-Dawley
rats were higher than the workplace air concentrations associated with adverse effects in AN-
exposed workers (Table 4-59). The lowest exposure level in the 2-year inhalation bioassay, 20
ppm, produced increased incidence of lesions in nasal epithelia (hyperplasia of mucus-secreting
cells in males and flattening of the respiratory epithelium in females). At 80 ppm, further
increases in incidences of nasal lesions of a wider variety were observed as well as statistically
significant increases in the incidences of histopathological lesions at other sites, including gliosis
and perivascular cuffing in the brain of males and females, focal nephrosis and thyroid cysts in
males, hepatic necrosis in females, and hyperplasia and hyperkeratosis of the nonglandular
portion of the stomach in both sexes combined (Quast et al., 1980b). In the two-generation
reproductive study of inhaled AN vapors in Crl:CD (SD) rats (Nemec et al., 2008), statistically
significant increases in the incidence of nasal lesions (respiratory/transitional epithelial
hyperplasia, subacute inflammation, squamous metaplasia, and/or degeneration of the olfactory
epithelium) were observed in FO males and females at 45 ppm. Increases in nasal lesions were
observed in Fl males at 5 ppm and the increase in incidence was statistically significant in Fl
males and females at 15 ppm.
Statistically significant deficits in several neurobehavioral tests were measured in
exposed workers in a Chinese acrylic fiber manufacturing plant with mean workplace air
concentrations of 0.11 ppm (range 0.00-1.70 ppm) and 0.91 ppm (range 0.00-8.34 ppm) in two
different process areas (Lu et al., 2005a). Deficits in exposed workers compared with
nonexposed workers were noted in a profile of mood states test (20-68% higher for negative
moods such as anger and confusion), a simple reaction time test of attention and response speed
(10-16% deficits), and the backward sequence of the digit span test of auditory memory (21-
24% deficits). The neurologic findings observed in humans are supported by observations of
deficits in sensory nerve conduction in the tail nerve in Sprague-Dawley rats repeatedly exposed
by inhalation to 25 ppm AN (Gagnaire et al., 1998).
Reproductive and developmental effects in relation to AN exposure have been seen in
occupational exposure studies in men and women (Table 4-58) and in animal studies (Table 4-
59). In AN exposed workers, statistically significant increases in the prevalence of adverse
reproductive outcomes were associated with mean workplace air concentrations 3.6 ppm (Dong
et al., 2000a) and 7.5 ppm (Li, 2000), indicating that reproductive effects from occupational
exposure may occur at higher exposure levels than those associated with mild neurobehavioral
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effects. However, a statistically significant decrease in sperm density and number and
statistically significant increase in DNA strand breakage and sex chromosome aneuploidy in
sperm cells were reported in workers exposed to 0.37 ppm AN with an average 2.8 years of
exposure (Xu et al., 2003). In rats exposed to 80 ppm, 6 hours/day on GDs 6-15, a statistically
significantly increased incidence of litters with any malformation (missing vertebrae, missing
ribs, or anteriorly displaced ovaries) was observed (Haskell Laboratory, 1992a). In another
study, inhalation exposure of Sprague-Dawley rats to AN for 6 hours/day on GDs 6-20 resulted
in statistically significant decreases in fetal weight gain per litter, compared with controls, at
exposure concentrations of 25 ppm (5% decrease), 50 ppm (8% decrease), or 100 ppm (15%
decrease) (Saillenfait et al., 1993). In a two-generation reproductive study of inhaled AN vapors
(Nemac et al., 2008), a decrease in sperm motility and progressive sperm motility (up to 9.5%)
was observed in FO and Fl males at 45 and 90 ppm. The decrease was statistically significant at
90 ppm. In addition, exposure-dependent increase in normalized anogenital distance on PND 1
was found in Fl males. Body weight at acquisition of balanopreputial separation for Fl males
was statistically significantly decreased at 45 ppm and 90 ppm. A statistically significant delay
in acquisition of sexual developmental landmark (vaginal patency) was observed in Fl females
at 90 ppm.
In summary, human and animal studies provide evidence of neurotoxicity from AN
exposure. This evidence includes an increased prevalence of subjective symptoms such as
headaches and memory impairments (Muto et al., 1992; Kaneko and Omae, 1992; Chen et al.,
2000), deficits in several neurobehavioral tests (Lu et al., 2005a) in workers exposed to AN, and
deficits in sensory nerve conduction in rats (Gagnaire et al., 1998). The LOAEL for these effects
in the studies in humans are approximately 10 to 100 times lower than the LOAEL in the study
in rats. Reproductive effects were also demonstrated in human and animal studies at exposures
that were similar to or higher than those producing neurological effects.
4.6.3. Mode-of-Action Information
The precise modes of action whereby AN induces noncancer effects are unknown.
However, a general understanding of the processes by which AN is metabolized within the body
allows some conclusions to be drawn about the range of processes that might be involved in
bringing about one or more of its toxic responses. Relevant metabolic processes are likely to
include partitioning between detoxification and oxidative activation sub-pathways, conversion of
AN to one or more toxic metabolites, depletion of GSH, the association of AN metabolism with
the onset of oxidative stress, and the ability of CEO, the reactive metabolite of AN, to covalently
bind to macromolecules, such as proteins. In addition, other processes related to the parent
compound AN, such as its cholinomimetic effects, may also be involved.
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4.6.3.1. GIEffects
The relationship between AN metabolism and GI hemorrhage in rats was suggested by
Ghanayem and Ahmed (1983). GI bleeding was observed 3 hours after a single dose of
50 mg/kg AN was administered to Sprague-Dawley rats orally or subcutaneously, with no
significant difference in the amount of GI blood loss resulting from either route of
administration. Thus, AN-induced GI bleeding was not a result of direct irritation of AN on the
GI tract.
Pretreatment of rats with CYP450 enzyme inducer Aroclor 1254 increased blood loss by
240% (Ghanayem and Ahmed, 1983). In contrast, pretreatment of rats with CYP450 inhibitors
cobalt chloride or SKF 525 A prior to AN administration produced significant decreases in blood
loss of 10 and 40%, respectively. Pretreatment of rats with DEM, a known depletor of GSH,
prior to AN administration produced no significant change in GI bleeding. In addition,
administration of a sublethal dose (6 mg/kg s.c.) of KCN did not induce GI bleeding when
compared with controls. Therefore, Ghanayem and Ahmed (1983) concluded that metabolic
activation of AN by CYP450 to a reactive metabolite other than cyanide (probably CEO) was a
prerequisite for AN to induce gastric hemorrhage.
Ahmed et al. (1996a) showed irreversibly bound AN-derived radioactivity in intestinal
mucosa following a single i.v. injection of 2-[14C]-AN to male F344 rats. A recent study by
Jacob and Ahmed (2003a) also demonstrated that AN and/or its metabolites accumulated and
covalently interacted in GI mucosa of male F344 rats treated either by i.v. or orally with
2-[14C]-AN. These studies supported the hypothesis that AN-induced injury of the GI mucosa is
not due to direct irritation by AN but by metabolic activation and macromolecular interaction of
AN metabolite in these tissues.
Ghanayem et al. (1985) also studied the mechanism of AN-induced gastric mucosal
necrosis in the glandular stomach in male Sprague-Dawley rats. Subcutaneous administration of
40 or 50 mg/kg AN caused a significant decrease in hepatic and gastric GSH concentration
3 hours after treatment, and induced gastric necrosis. Pretreatment of rats with various metabolic
modulators (CYP450 monooxygenase and GSH) before administration showed that there was a
significant inverse relationship between gastric GSH concentration and AN-induced gastric
erosions. Pretreatment of rats with sulfhydryl-containing compounds (cysteine or cysteamine)
protected against AN-induced gastric necrosis and blocked the depletion of gastric GSH.
In addition, AN-induced gastric erosions could be prevented by pretreatment with
atropine, a muscarinic receptor blocker, suggesting the involvement of muscarinic receptors in
the AN-induced gastric mucosal necrosis (Ghanayem et al., 1985). Activation of acetylcholine
muscarinic receptors is known to increase gastric acid secretion and cause gastric erosions.
Because muscarinic receptors are known to contain sulfhydryl groups in their active site (Ikeda
et al., 1980; Aronstam et al., 1978), Ghanayem et al. (1985) hypothesized that AN inactivated
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critical sulfhydryl groups and caused gastric erosions by locally modulating muscarinic
acetylcholine receptors in the stomach.
4.6.3.2. Neurological Effects
Increased prevalence of subjective symptoms of neurological effects was associated with
average workplace air concentrations of 1.13 ppm (Muto et al., 1992), 1.8 ppm (Kaneko and
Omae, 1992), and 0.48 ppm (Chen et al., 2000), and small deficits in performance in a battery of
neurobehavioral tests were observed in workers from factories with average workplace air
concentrations of 0.11 and 0.91 ppm (Lu et al., 2005a). In a case of acute severe accidental AN
poisoning (AN concentration at 62 mg/m3) of a worker (Fei and Xu, 2006), impairment of the
CNS, including diffused damage to the cerebral cortex layer (pyramidal system) and damage to
the subcortical layer (extrapyramidal system), was observed. In addition to poisoning symptoms
(dizziness, headache, difficulty breathing, confusion, convulsion, etc.), cerebral focal damage
was also detected in the patient, suggesting Parkinson's syndrome (static tremor, muscle rigidity,
increased muscle tension and lead-pipe rigidity, slow motor activity). The clinical symptoms
were related to those induced by adverse effects on the cholinergic system and dopaminergic
system.
Neurotoxicological effects of AN in animals included cholinomimetic effects on the
peripheral and central muscarinic systems in Sprague-Dawley rats after administration of
nonlethal oral doses of 20, 40, or 80 mg/kg (Ghanayem et al., 1991) and the development of
brain lesions in Sprague-Dawley rats (gliosis and perivascular cuffing) following chronic
inhalation exposure to 80 ppm (Quast et al., 1980b) or chronic drinking water exposure to
4.4 mg/kg-day (females only).
The neurobehavioral effect of AN was suggested to be related to changes in brain
monoamine neurotransmitter levels (Lu et al., 2005b). In a 12-week drinking water study of
male Sprague-Dawley rats (Lu et al., 2005b), dopamine levels were decreased by 76 and 46% in
rats exposed to 50 ppm AN, in the striatum and cerebellum, respectively. Serotonin levels were
decreased by 38 and 41% in the striatum and cerebellum, respectively, for rats exposed to
50 ppm AN. In the case of acute severe AN poisoning, the development of Parkinson's
syndrome and other CNS impairment in the exposed worker would suggest involvement of the
cholinergic system and dopaminergic system (Fei and Xu, 2006).
The cholinomimetic effects were thought to be due to an effect of AN on muscarinic
receptors, since atropine sulfate (which blocks both central and peripheral muscarinic receptors)
protected animals against these effects (Ghanayem et al., 1991). In an earlier study on
AN-induced gastric mucosal necrosis (Ghanayem et al., 1985), it was reported that pretreatment
with atropine and sulfhydryl-containing chemicals protected against such lesions. Since
muscarinic receptors contain sulfhydryls in their active sites (Aronstam et al., 1978) and
sulfhydryl-depleting chemicals are known to potentiate chemically induced activation of
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muscarinic receptors (Hedlund and Bartfai, 1979), Ghanayem et al. (1991) speculated that
depletion and/or inactivation of endogenous sulfhydryls by AN and/or its metabolite may cause
configurational changes of muscarinic receptor binding affinity that, in turn, lead to the
development of acetylcholine-like (cholinomimetic) toxic effects. It is unlikely that these
cholinomimetic effects are due to inhibition of acetylcholinesterase activity. Rajendran and
Muthu (1981) reported that AN did not inhibit the activity of acetylcholinesterase. In addition,
Satayavivad et al. (1998) reported that AN had no effect on decreased motor activity induced by
physostigmine (an inhibitor of acetylcholinesterase). Thus, Satayavivad et al. (1998) proposed
that the cholinomimetic effect of AN might be mediated by the release of acetylcholine from
nerve endings.
In addition, Ghanayem et al. (1991) proposed that lipid peroxidation may at least partly
be involved in AN-induced cholinergic overstimulation. In noting that AN enhanced lipid
peroxidation and inhibited Na+,K+-ATPase in RBCs in vitro, Farooqui et al. (1990) speculated
that disruption of the lipid microenvironment in membranes by either or both of these processes
might impact the muscarinic receptor function and induce cholinergic overstimulation.
For acute CNS effects, the signs were similar to those produced by cyanide. Ghanayem
et al. (1991) proposed that the free cyanide liberated from AN during its metabolism may
contribute to these effects. It is well established that cyanide causes CNS dysfunction by
inhibition of cellular respiration via inactivation of tissue cytochrome c oxidase, which is the
terminal electron acceptor in cellular energy production (Klaassen, 2001). Intraperitoneal
injection of 30 mg/kg AN in Chinese hamsters decreased cerebral succinate dehydrogenase and
cytochrome oxidase activities (Zitting et al., 1981). These authors suggested that the observed
biochemical effects were likely due to the formation of cyanide from AN.
The neurotoxicity of AN may also result from the covalent binding of AN or its
metabolites to enzymes. There is abundant evidence that important proteins bearing cysteine
residues (e.g., enzymes such as GSTM1) can bind AN, thereby possibly impairing their functions
and creating metabolic imbalances leading to toxicity. For example, AN was shown to
covalently bind to the important glycolytic enzyme GAPDH in vitro (Campian et al., 2002). AN
specifically targeted and bound to cysteine 149 in the active center of this enzyme, causing
irreversible inhibition of its activity. This suggested that AN might impair glycolytic ATP
production in vivo. Campian et al. (2002) speculated that the combination of glycolytic
impairment with inhibition of mitochondrial ATP synthesis by cyanide released from AN could
result in metabolic arrest. However, Campian et al. (2008) demonstrated in male Sprague-
Dawley rats that acute lethality of AN was not due to brain metabolic arrest (see Section
4.5.1.1.3).
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4.6.3.3. Reproductive/Developmental Effects
As discussed in Section 4.5.2.1, lower sperm density was noted in workers exposed to
0.8 mg/m3 AN compared with controls of approximate age range from the general population
(75 x !06/mL vs. 140 x 106/mL) (Xu et al., 2003). DNA strand breakage was also detected in
AN-exposed workers using single-cell gel electrophoresis, with the rate of comet sperm higher in
the exposed workers than in the control (28.7 vs. 15%). The frequency of sex chromosome
disomy was 0.69% in exposed groups and was higher than 0.35% in the control group. There
were also significant differences in the frequencies of XX-, YY-, and XY-bearing sperm between
exposed and control groups.
AN-induced effects on the male reproductive system were observed in a study in which
CD-I mice treated for 60 days with gavage doses of 10 mg/kg-day produced degeneration of the
seminiferous tubules and altered testicular activities of several enzymes (SDH, acid phosphatase,
LDH, and p-glucuronidase) (Tandon et al., 1988). Exposure-related increases in the incidence of
lesions in male reproductive organs were not observed in 14-week or 2-year gavage bioassays
with B6C3Fi mice (NTP, 2001) or in the 2-year bioassays with Sprague-Dawley or F344 rats
(see Table 4-58).
In a two-generation reproductive study of inhaled AN vapors in Crl:CD (SD) rats (Nemec
et al., 2008), a decrease in sperm motility was observed in FO and Fl males at 45 and 90 ppm
(the decrease was statistically significant at 90 ppm). An exposure-related increase in
normalized anogenital distance on PND 1 was observed in Fl males, and was statistically
significant at 45 ppm and 90 ppm. Slight delays in the acquisition of sexual developmental
landmarks and lower body weights on the day of acquisition were found in Fl males in the 45-
and 90-ppm groups, and in Fl females in the 90-ppm group.
The potential male reproductive effect of AN may involve the distribution and
metabolism of AN to CEO in the testis and interaction of CEO with tissue protein and DNA.
Radiolabeled AN distributed to the rat testis after oral and i.v. administration (Ahmed et al.,
1996a; Young et al., 1977). The testis has the capability to bioactivate AN (Abdel-Aziz et al.,
1997). Thus, CEO can either be formed in the liver and transported to the testis or be formed in
the testis in situ. AN has been reported to interact with testicular DNA in rats treated with a
single 46.5 mg/kg oral dose (Ahmed et al., 1992b). Covalent binding of [2,3-14C]-AN-derived
radioactivity to testicular DNA was maximal at 0.5 hours following administration, while DNA
synthesis in testicular tissue was decreased (80% of control). In addition, testicular DNA repair
was increased 1.5-fold at 0.5 hour and more than threefold at 24 hours after treatment (Ahmed et
al., 1992b). Alkylating agents have the potential to produce infertility via destruction of dividing
primary spermatogonia (Heinrichs and Juchau, 1980). Thus, any effect of AN on male
reproductive tissue may be due to its interference with testicular DNA synthesis and repair
processes. The consequence may be reproductive abnormalities as well as impact on altered
heritability in offspring.
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No mechanistic studies are available to elucidate the mode of action whereby AN
induced ovarian atrophy and cysts in female mice chronically exposed by gavage to doses as low
as 2.5 mg/kg-day (NTP, 2001).
The mild developmental effects associated with gestational exposure to 80 ppm AN by
inhalation or 65 mg/kg-day by gavage (Murray et al., 1978) may be associated with the release of
cyanide during maternal metabolism of AN. Concurrent administration of thiosulfate, a cyanide
antagonist, was shown to protect against the malformations induced by i.p. injection of 80 mg/kg
AN in hamsters (Willhite et al., 1981). However, Saillenfait et al. (1993) tested for relative
developmental toxicities of eight aliphatic mononitriles and proposed that factors other than
cyanide liberation from the nitrile may be involved, since teratogenicity of inhaled aliphatic
mononitriles in rats could not be predicted based on the presence of a vinyl moiety in their
molecular structure.
4.6.3.4. Hematological Effects
Farooqui and Ahmed (1983b) conducted work to elucidate the mechanism of the
hematological effects of AN. Their findings pointed to substantial AN-induced covalent binding
of CEO to RBCs, GSH depletion, increase in rate of RBC metabolism, and increase in the
formation of two metabolic intermediates, ATP and 2,3-diphosphoglycerate, that regulate the
oxygen dissociation curve. These authors suggested that chronic exposure to AN may lead to
methemoglobinemia, damage to RBC membranes, and impaired delivery of oxygen to the
tissues.
Oxidative stress and lipid peroxidation were also suggested as additional factors aiding in
the destruction of RBCs, evidenced by a significant decrease of Na+,K+-ATPase activity in
isolated RBC membranes (Farooqui et al., 1990).
AN-Hb adducts have been measured and used as a marker of exposure in humans. AN
can bind to amino acid residues other than cysteine. Thus, MacNeela et al. (1992) identified the
N-terminal cyanoethyl-valine adduct of Hb (CEVal), the formation of which may have
implications for the efficiency of oxygen transport to the tissues.
4.6.3.5. Immunological Effects
Zabrodskii et al. (2000) suggested a potential mechanism of action for the AN-induced
DTH. They found that, in mice, AN reduced the number of esterase-positive splenocytes, the
number of antibody-producing cells, and the inflammatory response induced by injection of
SRBCs in the paws of animals. Treatment of the animals with an esterase activity-restoring drug
restored the paw-response completely but not the numbers of immune-competent cells.
However, a combination of the esterase-restoring drug and a cyanide-trapping drug restored
immune function completely. Therefore, the study authors concluded that the immunotoxic
effect of AN was due to combined inhibition of esterase and cytochrome c oxidase as activities.
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4.6.3.6. Covalent Binding to Sulfhydryl Groups
The capacity of AN to bind to proteins may be an important determinant of its
toxicological effects. AN has been shown to have high affinity for cysteine residues on proteins
and polypeptides. There is abundant evidence that AN will bind to the cysteine-bearing
tripeptide, GSH, even without the contribution of enzymes such as GST. As discussed in
Section 3, the formation of 2-(cyanoethyl)glutathione and the appearance of
2-(cyanoethyl)cysteine and N-acetyl 2-(cyanoethyl)cysteine in the urine have been taken as an
indication of the ready interaction of AN and GSH. However, the presence of excess AN can
cause GSH to become depleted. This will tend to channel AN into an oxidative reaction with
CYP2E1 and result in formation of CEO and other products. Perturbing the balance between
detoxification of AN with GSH and oxidation by CYP2E1 (for example, by blockade of
CYP2E1) may facilitate the binding of AN to other cysteine-bearing proteins when GSH levels
are low.
AN also has been shown to bind to the cysteine 186 residue of the enzyme CAIII in rat
liver in vivo. Nerland et al. (2003) pointed out that this enzyme may play a role in protecting
cells from oxidative stress. AN binding to this component, with possible conformational
changes in tertiary structure and functionality, might abolish this protective ability and increase
cell vulnerability to oxidative stress. In a similar study, Nerland et al. (2001) demonstrated that
AN can bind also with high selectivity to cysteine 86 of GSTM1. However, in this case, the
binding did not exert an effect on the catalytic activity of the enzyme.
4.7. EVALUATION OF CARCINOGENICITY
4.7.1. Summary of Overall Weight of Evidence
Following EPA 's Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a), AN is
"likely to be carcinogenic to humans," based predominantly on consistent results showing that
lifetime inhalation or oral exposure caused a statistically significantly increased incidence of
tumors at multiple tissue sites in rats and mice. The most consistently observed tissue sites with
tumors were the brain, forestomach, Zymbal gland in the ear canal in rats, and forestomach and
ocular Harderian gland in mice. Acrylonitrile exposure was also assocated with increased
incidences of tumors of the mammary gland, intestine, and tongue in rats. In addition, rats
exposed during gestation and throughout adulthood showed higher incidences of brain tumors,
Zymbal gland tumors, extrahepatic angiosarcomas, and hepatomors than did rats with exposure
throughout adulthood only.
In humans, the most extensive data available is for lung cancer. The largest and best-
designed epidemiologic study (Blair et al., 1998) reported a RR of 1.2 (95% CI 0.9-1.6) for lung
cancer. However, the association seen in the full study population would not be expected to
correctly represent the association seen in the higher risk groups within a population with
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varying levels of exposure. In the Blair et al. (1998) study, workers with the longest duration
and highest exposures to AN had a two-fold increased risk of dying from lung cancer (i.e., RR
2.1, 95% CI 1.2-3.8). The observations of Blair et al. (1998) are supported by the patterns seen
in analyses stratified by exposure level, latency period or age group in numerous, but not all,
other studies (see Table 4-18).
U.S. EPA's Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a) indicate that
for tumors occurring at a site other than the initial point of contact, the weight of evidence for
carcinogenic potential may apply to all routes of exposure that have not been adequately tested at
sufficient doses. An exception occurs when there is convincing information (e.g., toxicokinetic
data) that absorption does not occur by other routes. For AN, systemic tumors were observed in
rats and mice following oral and inhalation exposure. No animal cancer bioassay data following
dermal exposure to AN are available. Based on the observance of systemic tumors following
oral and inhalation exposure, and in the absence of information to indicate otherwise, it is
assumed that an internal dose will be achieved regardless of the route of exposure. Therefore,
AN is considered "likely to be carcinogenic to humans" by all routes of exposure.
The hypothesized mode of action for AN-induced tumors is through a mutagenic mode of
action involving DNA modification by the reactive metabolite, CEO. The mutagenic mode of
action for AN-induced tumors is considered to be relevant to humans. Other modes of action,
including oxidative stress and inhibition of intercellular communication, may also contribute.
4.7.2. Synthesis of Human, Animal, and Other Supporting Evidence
Section 4.1.2.2 presents an historical perspective and evaluation of the epidemiologic
studies of possible relationships between occupational exposure to AN and elevated risks for
cancer. The early studies were of small cohorts with limited follow-up periods and limited
assessment of actual exposures. Elevations for incidence and death from lung cancer and
incidence of prostate cancer were sufficiently consistent among the early studies to provoke the
conduct of several additional studies over a 20-year period. The later studies had increased
power due to increased number of workers, longer periods of follow-up, increased sophistication
and quantification of exposure assessments, or inclusion of information on smoking habits.
Small excess risks for lung and prostate cancer were identified in a few of these studies, but
consistently and statistically significantly elevated risks were not observed across studies (see
Tables 4-3 and 4-9). Other studies reported small excess risks for other types of cancer (e.g.,
bladder, colon, and brain cancers), but these findings were even less consistent across studies
than the findings for lung and prostate cancer (see Tables 4-8, 4-11, 4-12 and 4-13).
The most informative study in the later generation of studies was the large, well-
documented study by Blair et al. (1998). This study examined a large cohort and attempted to
adjust for known problems with earlier studies by quantifying exposures, estimating the effect of
smoking, and using an internal control group of unexposed workers. The AN-exposed cohort as
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a whole experienced fewer deaths from lung cancer than those expected from the experience of
the general U.S. population (probably reflecting a healthy worker effect), but when the exposed
workers were grouped into quintiles of cumulative exposure, the lung cancer rate in the highest
quintile of exposure (>8 ppm-years) for those who were followed for 20 or more years was about
two times the rate in the unexposed group of workers (RR = 2.1; 95% CI = 1.2-3.8).
Conclusions regarding the association between AN exposure and other cancer types (stomach,
brain, breast) are limited due to the small number of site-specific cancer deaths (Blair et al.,
1998).
As concluded in Section 4.1.2.2, there is no strong or consistent evidence from the body
of epidemiologic studies that mortality from any type of cancer is elevated in groups
occupationally exposed to AN at the levels that have been measured in workplaces that used this
chemical. However, these epidemiology studies indicate a possible association between
occupational exposure to AN and increased risk of lung cancer.
In rat and mouse bioassays, AN has been demonstrated to be a multiple-site carcinogen.
Chronic oral exposure to AN induced tumors in the brain or spinal cord, Zymbal gland in the ear
canal, the forestomach, and, to a lesser degree and less consistently, the female mammary gland,
tongue, and intestine in several oral bioassays with F344 rats (Johanssen and Levinskas, 2002b;
Biodynamics, 1980c) and Sprague-Dawley rats (Johannsen and Levinskas, 2002a; Quast, 2002;
Biodynamics, 1980a, c; Quast et al., 1980a). In addition, lifetime inhalation cancer bioassays
with Sprague-Dawley rats found exposure-related increased incidences of brain tumors, Zymbal
gland tumors, intestinal tumors, malignant mammary gland tumors, and tongue tumors (Dow
Chemical Co., 1992a; Maltoni et al., 1988, 1977; Quast et al., 1980b). Strong evidence exists for
dose-response relationships for the carcinogenic responses in the CNS, Zymbal gland, and the
forestomach of rats, especially in the lifetime drinking water studies with multiple exposure
levels. The lowest drinking water concentrations associated with significant increased incidence
of forestomach and brain tumors were 3 and 30 ppm, respectively, corresponding to daily doses
of about 0.3 and 2.5 mg/kg-day in F344 rats (Johanssen and Levinskas, 2002b; Biodynamics,
1980c).
With chronic inhalation exposure of Sprague-Dawley rats, increased incidences of brain
tumors occurred at air concentrations of 20 and 80 ppm, whereas the incidences of Zymbal gland
tumors, mammary gland adenocarcinomas and intestinal tumors were increased only at the
80 ppm level (Quast et al., 1980b). In mice, a single gavage lifetime bioassay identified the
forestomach and the Harderian gland as sites of tumor development, but elevated incidences of
brain, Zymbal gland, or mammary gland tumors were not found (NTP, 2001). A significant
increase in lung tumors was found in the female mid-dose group of exposed mice but not in the
female high-dose group or in any of the male exposed groups. NTP (2001) concluded that the
evidence for carcinogenicity in the mouse lung was equivocal; this conclusion is consistent with
the small magnitude of the increase, lack of a monotonic dose-response relationship, and lack of
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a demonstrated carcinogenic response in exposed male mice. No consistent evidence was found
for carcinogenicity in the rat lung or the prostate or bladder tumors in rats or mice.
The possible human relevance of the rodent carcinogenic responses to AN is not fully
understood. The brain is the only organ for which there is a direct human counterpart among
rodent organs showing strong carcinogenic responses to AN, but the forestomach squamous
epithelium and Zymbal gland have analogous tissues in humans. Although humans do not have
a forestomach, the human oral cavity and upper two-thirds of the esophagus are lined with
squamous epithelial cells morphologically similar to those in which tumors develop in rats, with
the exception that the rat cells are keratinized and the human cells are not (Cohen, 2004; Wester
and Kroes, 1988). Likewise, although humans do not have a Zymbal gland, a sebaceous gland in
the ear canal, they do have sebaceous glands. In contrast, humans and other primates do not
have tissue analogous to the Harderian gland, an ocular gland in rodents that secretes lipids and
porphyrins (Cohen, 2004; Sheldon, 1994; Albert et al., 1986).
It has been suggested that most genotoxic forestomach carcinogens appear to act through
a mutagenic mode of action (IARC, 2003). Given that mutagenic modes of carcinogenic action
are plausible for AN or its metabolites (see Section 4.6.3), formation of tumors in these organs
may be indicative of a more generic carcinogenic hazard (Cohen, 2004). Chemicals with a
mutagenic mode of action are frequently observed to cause cancers in many sites in one species,
as well as to have different sites of tumor formation in different species. IARC (2003) concluded
that multi-site carcinogens that induce forestomach tumors and are genotoxic are likely relevant
to human carcinogenesis. As stated in EPA's Guidelines for Carcinogen Risk Assessment (U.S.
EPA, 2005a), site concordance is not always assumed between animals and humans. Therefore,
it is reasonable to consider the tumor responses in these rodent organs as indicators of AN-
induced carcinogenicity in humans.
The National Academy of Sciences (NAS) (2008) in its Science and Decisions:
Advancing Risk Assessment, stated on page 143 that:
.. .the target organ in a rodent species, such as the forestomach or Zymbal gland,
may not have an exact human counterpart. However, the presence of carcinogenic action
in tissues for which there is no correspondence in humans or that may be regulated
differently in humans does not mean that the toxicity or tumor finding in animals is
irrelevant. That the rodent tissue is sensitive to the toxicant signifies that the toxicant
MO As operate in a mammalian system that has characteristics in common with similar or
even not obviously related tissues in humans or human subpopulations. Because
epidemiologic studies are often limited in their ability to explore outcomes related to
workplace or environmental exposures, it is typically impossible to rule out the relevance
of an effect seen in a particular rodent tissue unless there is detailed mechanistic
information on why humans would not be affected (IARC, 2006). The finding that the
high sensitivity of the rat Zymbal gland to benzene tumorigenesis occurs via an MOA
(clastogenesis) similar to that which produces benzene-induced bone marrow toxicity and
cancer in humans (Angelosanto et al., 1996) is an indication that a tissue that is specific to
the rat can still provide important hazard and potency information related to human risk.
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In general, tissues that are responsive to a toxicant should be considered relevant to
human risk assessment unless mechanistic information demonstrates that the processes
occurring in the tissues could not occur in humans.
Therefore, in the absence of data to indicate otherwise, EPA considers the tumors in
rodents (e.g., forestomach and Zymbal gland tumors) to be relevant to humans.
4.7.3. Mode-of-Action Information
The U.S. EPA (2005a) Guidelines for Carcinogen Risk Assessment defines mode of
action as a sequence of key events and processes, starting with the interaction of an agent with a
cell, proceeding through operational and anatomical changes, and resulting in cancer formation.
Mode of action is distinct from "mechanism of action," a term that implies greater understanding
and description of events, including those at the molecular level. Examples of possible modes of
carcinogenic action include mutagenic, mitogenic, anti-apoptotic (inhibition of programmed cell
death), cytotoxic with reparative cell proliferation, and immunologic suppression.
AN has been demonstrated to be a multiple-site carcinogen in rats and mice, inducing
tumors in the forestomach, brain, and Zymbal gland in two strains of rats following chronic oral
exposure; in the forestomach and Harderian gland in a mouse strain following chronic oral
exposure; and in the forestomach, brain, Zymbal gland, intestinal tract, and tongue in one rat
strain following chronic inhalation exposure. Carcinogenic responses have also been reported
less consistently across studies in the female mammary gland and small intestine.
Several hypothesized modes of action by which AN causes cancer in the brain of rats
have been investigated to varying degrees and are discussed within the context of the modified
Hill criteria of causality as recommended in the Guidelines for Carcinogen Risk Assessment
(U.S. EPA, 2005a). These discussions are followed by a discussion of the possible mode of
action involved in the development of forestomach tumors in rats. Modes of action by which
AN may induce tumors in the lung, liver, Zymbal gland, Harderian gland, mammary gland,
intestine, and tongue have received little or no investigation.
4.7.3.1. Hypothesized Mode of Action for Brain Tumors: Mutagenic Mode-of-Action
Key Events
This mode of action hypothesizes that AN is first activated by CYP2E1 (mostly in the
liver but also at other sites, such as intestinal mucosa or squamous epithelium of the
forestomach) to its reactive metabolite, CEO. CEO is then distributed to target organs (e.g., rat
brain) where it reacts with DNA, forming DNA adducts. DNA adduct formation results in
genetic damage, especially the formation of point mutations. Other types of DNA damage by
CEO are also observed in vivo, including DNA strand breaks, SCEs, and MN formation.
Mutagenicity is a biologically plausible mechanism for tumor induction.
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AN
^ CYP2E1
CEO
4,
Interaction with DNA
Gene mutation for tumor initiation
Following these key events, tumor growth may be promoted by any one or combination
of a number of cell-signaling pathways, leading to enhanced cell proliferation or inhibition of
programmed cell death.
4.7.3.2. Experimental Support for the Hypothesized Mode of Action
4.7.3.2.1. Strength, consistency, specificity of association. The evidence that supports a
mutagenic mode of action is strong and consistent (as summarized in Table 4-54). In addition to
positive findings in blood lymphocytes, buccal mucosal cells, and sperm in five epidemiologic
studies, DNA alkylation by AN was found in numerous tissues in rats or mice (brain, liver,
testes, forestomach, colon, kidney, bladder, and lung) treated with a single dose of AN. AN or
its reactive metabolite CEO yielded positive results in in vitro mutation assays using bacteria,
fungi, and insects, as well as animal and human cell cultures. The mutagenicity/genotoxicity is
specific and occurs in the absence of cytotoxicity or other overt toxicity. Although the temporal
relationship of adduct formation and mutagenicity with carcinogen! city has not been adequately
explored, these effects are seen in short-term assays (before tumor formation). Dose-response
concordance is observed between mutagenic doses in vivo and tumorigenic doses in rats and
mice. A mutagenic mode of action also comports with notions of biological plausibility and
coherence because AN is metabolized to an epoxide intermediate. Such agents are generally
capable of forming DNA adducts, which in turn have the potential to cause genetic damage,
including mutations, and mutagenicity, in its turn, is a well-established cause of carcinogenicity.
This chain of key events is consistent with the current understanding of the biology of cancer.
The possible association between a mutagenic action of CEO, the epoxide metabolite of
AN, and brain tumors in rats is supported by the detection of CEO in rat brains after oral
administration of 10 mg/kg AN to rats (Kedderis et al., 1993b), reported binding of CEO to brain
DNA after oral administration of [2,3-14C]-AN to rats (Farooqui and Ahmed, 1983a), results
from studies of in vitro DNA reactivity, positive results from in vitro tests of mutagenicity and
cytogenetic effects, and positive results from in vivo studies of DNA damage and repair assays
and other genotoxic endpoints in rats and mice. Studies that demonstrate the mutagenicity of AN
in occupational exposed workers are also available.
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In the following subsections, experimental evidence supporting the mutagenic mode of
action of AN using different test systems is summarized.
In vitro DNA binding
AN itself reacts very slowly with DNA in vitro at very high, nonphysiological
concentrations (>1 M) (Solomon et al., 1984), but CEO forms different adducts more rapidly in
vitro with DNA (Solomon et al., 1993; Yates et al., 1993) or nucleotides (Yates et al., 1994) (see
Section 4.5.1.2.1). When calf thymus DNA was incubated with CEO for 3 hours (Solomon et
al., 1993), the main adducts formed included N7-(oxoethyl)guanine, N3-(2-hydroxy-2-carboxy-
ethyl)deoxyuridine, and smaller amounts of adenine and thymine adducts. Yates et al. (1993)
also identified the formation of N3-(2-cyano-2-hydroxylethyl)deoxythymidine when CEO was
incubated with calf thymus DNA in vitro.
Mutations in bacteria
In short-term tests with bacteria, AN induced mutations in a majority of test systems,
often requiring the presence of exogenous metabolic systems (Table 4-54). AN induced
mutation in S. typhimurium strains TA 1530, 1535, 1538, 1937, and 1950 with strong responses
when AN was tested in the vapor phase in the presence of S9. AN induced a weak response for
TA 98, 100, and 1978. Thus, AN induced gene reversion mainly by base substitution and
induced lower mutagenic activity with strains reversed by frameshift mutation. AN also
produced a dose-related increase in the number of revertant colonies compared with untreated
bacteria in E. coll WP2 (which is DNA repair proficient), WP2uvrA (which lacks excision
repair), and WP2 uvrApolA (which lacks both excision repair and DNA polymerase 1) without a
need for S9 fraction (Venitt et al., 1977).
Mutations in fungi and Drosophila
As shown in Table 4-54, AN induced mitotic gene conversion in S. cerevisiae JD1
(Brooks et al., 1985). AN also induced sex chromosome loss in the adult female Drosophila
ZESTE system (Osgood et al., 1991), as well as somatic recombination and mutation in hatching
Drosophila with exposure in the larvae stage (Vogel, 1985; Wiirgler et al., 1985).
Mutations in mammalian cell culture
AN also induced mutations in mammalian cells in vitro. In the mouse lymphoma cell
assay, AN induced forward mutations at the Tk+ ~ locus in most assays (Table 4-54). In human
lymphoblastoid Tk6 cells devoid of CYP450 activity, AN induced mutations at the Tk locus only
in the presence of an exogenous S9 metabolic system (Crespi et al., 1985), and CEO was
effective at 10-fold lower concentrations than AN (Recio and Skopek, 1988a, b).
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Characterization of the Tk mutants in human lymphoblastoid Tk6 cultures by Recio and
Skopek (1988b) identified two classes of CEO-induced TK~~ mutant phenotypes that differed in
their growth rates. CEO induced predominantly Tkn mutant with normal growth rate and Tks
with slower growth rate. Southern blot analysis of CEO-induced Tkn mutants indicated that the
majority of these mutants were below the detection limit of <2 kb. Thus, CEO-induced
alterations are relatively small DNA alterations. Recio and Skopek (1988a) suggested that CEO
induced Tkn mutants resulted from point mutations or small insertions/deletions that occurred
during the replication or repair of CEO-modified DNA. Kodama et al. (1989) conducted
cytogenetic analyses of eight CEO-induced Tks mutant clones reported in Recio and Skopek
(1988a, b). A visible abnormality on chromosome 17 was found in one of the CEO-induced Tks
mutants and was marked by duplication of the long arm of chromosome 17, with break points at
ql 1 and q21. The latter break point was close to the Tk locus, suggesting that the observed
aberration might be associated with r&~~phenotype.
CEO also induced mutations at the hprt locus in Tk6 cells (Recio and Skopek, 1988a).
Characterization of the hprt mutations by cDNA sequencing analysis indicated that several hprt
mutations were formed. The majority of CEO-induced mutations were the specific loss of exons
from the coding region of hprt. Remaining mutations were single base substitutions (point
mutation) resulting from amino acid changes (A:T base pairs and G:C base pairs).
Cytogenetic effects in vitro
Additional evidence of mutagenicity included AN-induced cytogenetic changes, such as
SCEs or CAs in a majority of assays with CHO or CHL cells (Natarajan et al., 1985; VedBrat
and Williams, 1982; Ishidate et al., 1981) and human lymphocytes in vitro (Perocco et al., 1982)
but not in epithelial-like cells from rat liver (RL4 cell line) (see Table 4-54). AN also induced
DNA single-strand breaks in rat hepatocytes (Bradley, 1985) and CHO cells (Douglas et al.,
1985). Of note is that AN induced SCEs and DNA single-strand breaks in human bronchial
epithelial cells in culture without S9 (Chang et al., 1990), indicating that human bronchial
epithelial cells have the metabolic capabilities to activate AN to CEO.
DNA repair in vitro
In DNA repair assays, neither AN nor CEO induced UDS in cultured rat hepatocytes
(Butterworth et al., 1992; Probst and Hill, 1985; Williams et al., 1985) (as measured by
incorporation of [3H]-thymidine and autoradiographic techniques) or in HeLa cells with or
without the presence of S9 (using liquid scintillation spectrometry) (Martin and Campbell, 1985).
However, IPCS (1985) concluded that the rat hepatocyte autoradiographic UDS assay was too
insensitive for determination of genotoxicity of the eight tested carcinogens, including AN. On
the other hand, CEO, but not AN, induced UDS in human mammary epithelial cells in vitro
(Butterworth et al., 1992).
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DNA repair in vivo
Unscheduled DNA repair activity was detected by following the time course of
[3H]-thymidine incorporation into DNA isolated from lung (Ahmed et al., 1992a), testis (Ahmed
et al., 1992b), and gastric tissue (Ahmed et al., 1996b; Abdel-Rahman et al., 1994a) after
exposure of Sprague-Dawley rats to single oral doses of 46.5 mg/kg AN. In a study by Hogy and
Guengerich (1986), UDS was found in the liver of male F344 rats administered 50 mg/kg AN or
6 mg/kg CEO i.p. but was not found in the brain. The absence of detected UDS in the brain
could reflect the absence of DNA damage. Alternatively, this absence could reflect differences
in the repair rate of CEO-DNA adducts and could account for the observed target organ
specificity of AN in rat brain but not liver. Since the difference in DNA repair rates in brain and
liver is well known (Kleihues et al., 1977), the second explanation may be more sound. In
another study, UDS activity was not detected by autoradiographic techniques following
incubation of primary cultures of hepatocytes or spermatocytes from F344 rats given single oral
doses of 75 mg/kg or five daily doses of 60 mg/kg-day with [3H]-thymidine (Butterworth et al.,
1992). This difference in results from those by Hogy and Guengerich (1986) was probably due
to differences in methodology.
DNA binding in vivo
Several studies examining the amounts of radioactivity in DNA fractions of tissues
following acute in vivo exposure of rats to radiolabeled AN provide evidence of in vivo DNA
reactivity of AN or its metabolites. Maximal amounts of radioactivity covalently bound to
hydroxyapatite-purified DNA from the liver, stomach, and brain showed the following order
24 hours after male Sprague-Dawley rats were given single oral doses of 46.5 mg/kg
[2,3-14C]-AN: brain (-120 pmol AN equivalent/mg DNA) > stomach (-80 pmol/mg) > liver
(-25 pmol/mg) (Farooqui and Ahmed, 1983a). Other studies from the same group of
investigators found elevated covalent binding of radioactivity in gastric, testicular, and lung
DNA from similarly exposed rats (Abdel-Rahman et al., 1994b; Ahmed et al., 1992a, b). The
results from these studies provided some evidence of associations between covalent DNA
binding following acute exposure and sites of tumor development following chronic exposure.
The methods to isolate DNA in these studies may not have been stringent enough to exclude
covalent binding of radiolabel to proteins (Whysner et al., 1998b; Geiger et al., 1983); however,
the degree to which these alternative processes may have contributed to the measured amounts of
radioactivity in the DNA fractions is unknown. When 0.6 mg/kg [2,3-14C]-CEO was
administered to one F344 rat i.p., covalent binding to both liver and brain protein was found, but
no covalent binding to both liver and brain nucleic acids could be detected at the level of
0.3 alkylations per 106 base (Hogy and Guengerich, 1986). However, the N7-(2-oxoethyl)
guanine adduct was detected in liver DNA of F344 rats treated with 6 mg/kg CEO or 50 mg/kg
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AN i.p. in the same experiment. Thus, covalent binding of CEO to DNA had to occur. No DNA
binding was detected in the Hogy and Guengerich (1986) study, presumably due to the small
DNA sample from one rat, low CEO concentration (0.6 vs. 6 mg/kg CEO), stringent DNA
isolation procedure, and overcorrection of protein binding.
DNA adducts in vivo
Limited attempts to detect CEO-DNA adducts in brain tissues following acute in vivo
exposure protocols have been largely unsuccessful, but in vitro studies indicate that the
formation of other DNA adducts from AN and its metabolites are possible (Yates et al., 1994;
Solomon et al., 1993; Hogy and Guengerich, 1986; Solomon et al., 1984). N7-(2-oxoethyl)-
guanine, a CEO-DNA adduct formed in vitro following incubation of calf thymus DNA with
CEO, was detected in DNA isolated from the livers of male F344 rats, following single i.p.
administration of doses of 50 mg/kg AN or 6 mg/kg CEO, but was found only equivocally at
detection limit in DNA isolated from the brains of exposed rats (Hogy and Guengerich, 1986).
In another study, 7-(2-cyanoethyl)guanine and O6-(2-cyanoethyl)deoxyguanosine adducts were
not detected in DNA isolated from the liver or brain of male F344 rats given s.c. or i.v. injections
of 50 or 100 mg/kg AN (Prokopczyk et al., 1988). The DNA samples were analyzed for the
presence of the two adducts using HPLC with fluorescence detection. The detection limits were
20 umol/mol guanine for 7-(2-cyanoethyl)guanine and 15 umol/mol guanine for O6-(2-cyano-
ethyl)deoxyguanosine. These detection limits may be not sensitive enough. More importantly,
these two DNA adducts were not formed from incubation of CEO with calf thymus DNA in
vitro. Hence, the limitations of these studies were that the studies were not looking for all the
major adducts formed in in vitro incubation of CEO with DNA.
As discussed previously, the main adducts found after 3 hours of incubation of CEO with
calf thymus DNA were N7-(oxoethyl)guanine, N3-(2-hydroxy-2-carboxyethyl)deoxyuridine
(Solomon et al., 1993), andN3-(2-cyano-2-hydroxylethyl)deoxythymidine (Yates et al., 1993).
Yet only N7-(oxoethyl)guanine has been measured in available studies. Thus, it is highly likely
that the actual adducts formed from interaction of CEO with brain DNA have not yet been
looked for. It is also likely that the analytical methods used to measure DNA adducts in
available studies were not sensitive enough for their detection. In addition, DNA adducts were
measured only in single-dose studies not repeated-dose studies.
DNA damage in rats and mice
Evidence of DNA damage after AN exposure in rats and mice is available. Comet assays
showed DNA damage in various tissues in both rats and mice exposed to AN (Sekihashi et al.,
2002). Single i.p. injections of 20 mg/kg AN induced DNA damage in forestomach, bladder, and
brain but not in colon, liver, kidney, or bone marrow of ddY mice, whereas single doses of
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30 mg/kg induced DNA damage in forestomach, colon, kidney, bladder, and lung but not in
brain or bone marrow of Wistar rats (Sekihashi et al., 2002).
Increased frequencies of MN in bone marrow were found in Sprague-Dawley rats,
following i.v. injection of 98 or 124 mg/kg AN (Wakata et al., 1998) but were not found in
Sprague-Dawley rats following administration of oral doses up to 40 mg/kg (Morita et al., 1997)
or in male CD-I mice following administration of oral, i.p., or i.v. doses up to 45 mg/kg (Morita
et al., 1997). Increases in CAs in bone marrow cells were not found in Swiss albino mice
exposed to oral doses up to 20 mg/kg-day for 4, 15, or 30 days (Rabello-Gay and Ahmed, 1980);
in Sprague-Dawley rats given 16 daily doses of 40 mg/kg-day (Rabello-Gay and Ahmed, 1980);
in NMRI mice given single i.p. doses of 30 mg/kg (Leonard et al., 1981); or in ICR mice
exposed in inhalation chambers to 20 or 100 mg/m3 for 5 days (Zhurkov et al., 1983). In
contrast, Fahmy (1999) reported that increased CAs occurred in spermatocytes of Swiss mice
following single oral doses of 15.5 or 31 mg/kg AN or five daily doses of 7.75 mg/kg and in
spleen cells and bone marrow cells after a single oral dose of 7.75 mg/kg.
Dominant lethal mutations in rats and mice
Dominant lethal mutations were not increased by treating male NMRI mice with single
i.p. doses of 30 mg/kg AN (Leonard et al., 1981) or male F344 rats with five oral doses of
60 mg/kg-day AN (Working et al., 1987), indicating no AN-induced germ cell mutations.
Chromosomal mutations in humans
There is evidence of AN-induced mutagenicity in occupationally exposed workers. Fan
et al. (2006) reported significant increase in the occurrence of MN in buccal mucosal cells of
workers exposed to 0.52 or 1.99 mg/m3 AN for an average duration of 15.7-17.2 years and
significant increase in MN in peripheral blood lymphocytes in workers exposed to 1.99 mg/m3
AN for an average of 17.2 years. The workers in the Fan et al. (2006) study were exposed to
higher concentrations than those in a previous study by Sram et al. (2004), in which the exposure
concentration range was 0.05-0.3 mg/m3 AN, thus explaining the negative results reported in
Sram et al. (2004) or in other studies where exposure concentrations were not reported.
Evidence of CAs in AN-exposed workers was also reported by Borba et al. (1996), and DNA
strand breakage and sex chromosome aneuploidy in the sperm of AN-exposed workers were
reported by Xu et al. (2003). In addition, Beskid et al. (2006) reported an increase in the number
of reciprocal translocations and the relative number of insertions in the chromosomes of cultured
lymphocytes of AN-exposed male workers.
Therefore, the overall in vitro and in vivo evidence in support of the mutagenicity of AN
is strong, consistent, and specific.
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4.7.3.2.2. Dose-response concordance. Chronic exposures of rats to AN in drinking water
concentrations >30 ppm (>2.5 mg/kg-day) or air concentrations >20 ppm were associated with
significantly increased incidences of brain tumors. No published studies are available that have
measured CEO-DNA adducts or other endpoints pertinent to mutagenicity in brain tissues
following acute-, subchronic-, or chronic-duration oral or inhalation exposures of rats to AN.
The available in vivo DNA adduct studies were conducted before the Solomon et al. (1993) and
Yates et al. (1993) studies that reported on DNA adducts formed from CEO in vitro. Hence,
these DNA adduct studies were not looking for the adducts formed in vitro from CEO. The
available studies that looked for CEO-DNA adducts in brain or other tissues involved single i.p.
(Hogy and Guengerich, 1986), s.c., or i.v. (Prokopczyk et al., 1988) administration protocols at
dose levels (50 or 100 mg/kg-day AN) that were higher than the chronic oral doses associated
with brain tumors in rats (0.3 to 40 mg/kg-day). However, no repeated-dose studies have been
conducted for detection of DNA adducts. The single oral dose of 46.5 mg/kg used in studies that
demonstrated UDS in the lung, gastric tissue, and testis of treated Sprague-Dawley rats (Abdel-
Rahman et al., 1994b; Ahmed et al., 1992a, b) was comparable to the high dose of 40 mg/kg-day
used in studies by Friedman and Beliles (2002), providing evidence that the dose that caused
DNA repair was carcinogenic.
A study in mice by Fahmy (1999) showed that increased CAs occurred in spermatocytes
of Swiss mice following single oral doses of 15.5 or 31 mg/kg AN or five daily doses of
7.75 mg/kg, and in spleen cells and bone marrow cells after a single oral dose of 7.75 mg/kg.
Fahmy (1999) also reported increases in SCEs in bone marrow cells of male Swiss mice after a
single i.p. dose of 7.5 or 10 mg/kg of AN. These doses were all in the range of, or comparable
to, the tumorigenic doses in B6C3Fi mice of 2.5-20 mg/kg-day in a 2-year bioassay.
The single doses employed by Sekihashi et al. (2002) that demonstrated DNA damage in
forestomach, colon, kidney, bladder, and lung of rats treated with 30 mg/kg i.p. and in colon,
bladder, lung, and brain of male ddY mice treated with 20 mg/kg i.p. were also within the range
of tumorigenic doses in the 2-year bioassay.
Therefore, the doses that demonstrated DNA damage or chromosome mutations in single-
dose studies are in concordance with tumorigenic doses in chronic bioassays.
4.7.3.2.3. Temporal relationships. Currently available examinations of DNA damage,
chromosome mutations, SCEs, UDS, or CEO-DNA adducts in brain or other tissues are
restricted to single-dose acute administration protocols. Studies designed to examine temporal
relationships of key events, such as the presence of CEO-DNA adducts in brain tissue, are not
available. Nevertheless, results from the rat bioassays indicated that most AN tumors occur after
12-14 months of exposure or longer. Thus, the observed mutagenic effects of AN occurred
before tumor formation.
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4.7.3.2.4. Biological plausibility and coherence. The hypothesis that the key primary event in
AN induction of brain tumors is the formation of CEO-DNA adducts that leads to mutations that
initiate tumor formation is plausible based on in vitro and in vivo evidence for the mutagenicity
of the AN metabolite, CEO. However, the available data do not establish that this is the only
mode of action by which AN may induce brain tumors in rats. Notably lacking in the database
are studies designed to detect a range of DNA adducts or mutations associated with tumor
initiation in brain tissues following prolonged oral or inhalation exposures at levels that induced
brain tumors in the chronic rat bioassays. In vitro studies by Solomon et al. (1993) revealed that
CEO interacted with calf thymus DNA and, in addition to N7-(oxoethyl)guanine, formed N3-(2-
hydroxy-2-carboxyethyl)deoxyuridine from an initial cytosine adduct. Other adenine and
thymine adducts were also formed. Yates et al. (1993) demonstrated that CEO reacted calf
thymus DNA formed N3-(2-cyano-2-hydroxylethyl)-deoxythymidine. Yet, only
N7-(oxoethyl)guanine has been measured in exposed rats in available studies (Hogy, 1986).
The available data do not provide an explanation of why AN induces brain tumors in
F344 and Sprague-Dawley rats but not in B6C3Fi mice. Kedderis et al. (1993b) reported that
when male F344 rats and male B6C3Fi mice were administered 10 mg/kg AN in water by
gavage, higher CEO concentrations were found in blood and brains of rats than in mice (13%
higher in blood, 23% higher in brain). Higher CEO concentrations in rat brain may at least
partially explain why tumors were only found in the brains of exposed rats but not in mice. In
addition, the clearance of CEO in mice was more rapid than in rats (Roberts et al., 1991). It may
also be that mice are more efficient in repairing DNA damage since DNA damage has been
detected in the brain of ddY mice exposed to 20 mg/kg AN, i.p. (Sekihashi et al., 2002). It has
been noted that the B6C3Fi mouse is generally insensitive to chemically induced neurogenesis in
NTP carcinogenesis bioassays (Radovsky & Mahler, 1999).
The difference in susceptibility to AN induction of brain tumors between mice and rats is
similar to that observed for glycidol, an aliphatic epoxide structurally similar to CEO (Irwin et
al., 1996). Glycidol is a direct acting alkylating agent, which induces tumors at a variety of sites
in both rats and mice. However, although significant induction of gliomas was observed in both
sexes of F344 rats after glycidol treatment, no brain tumors were induced in B6C3Fi mice.
Ethylene oxide also shows a similar pattern of turnorigenesis, inducing brain tumors in both
sexes of exposed rats (Garmin et al., 1985; 1986), but no brain tumors in exposed mice (NTP,
1987). Thus, induction of brain tumors in rats but not in mice by known genotoxic carcinogens
appears to reflect primarily a species difference in inherent susceptibility to brain tumorigenesis.
Data for another chemical, ethylene oxide, that causes brain tumors (gliomas) in rats
show a different accumulation pattern for N7-(2-oxoethyl)guanine adducts than that observed
following i.p. administration of AN or CEO. In rats exposed to 500 ppm ethylene oxide for
1 day, higher levels of N7-(2-oxoethyl)guanine adducts were detected in DNA from brain than in
DNA from liver (Walker et al., 1990), whereas N7-(2-oxoethyl)guanine adducts were detected in
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brain at detection limit and only at low levels in the liver of rats given single i.p. injections of
50 mg/kg AN or 6 mg/kg CEO (Hogy and Guengerich, 1986). Whysner et al. (1998b) suggested
that these and other results indicated that glioma formation from chronic exposure to AN may
not involve the formation of N7-(2-oxoethyl)guanine adducts; however, several alternative
explanations are possible. The lack of unequivocal detection of CEO-DNA adducts in rat brain
may indicate that the detection limits of the methods used were not sensitive enough. More
sensitive methods of detection, such as liquid chromatography-mass spectrometry (Poirier,
2004), have not been used in AN-induced DNA adduct studies. Alternatively, the stringent
methods used to isolate purified DNA (without associated proteins) in the experiments by Hogy
and Guengerich (1986) may have caused the loss of adducts or inhibited the recovery of
adducted DNA (Meek et al., 2003). Another possibility is that N7-(2-oxoethyl)guanine adduct
may not be involved in mutations leading to AN-induced brain tumors, and other, as yet
uninvestigated, DNA adducts that were reported by Solomon et al. (1993) and Yates et al. (1993)
may be involved. Still another possibility is that CEO-DNA adducts and resultant mutations in
target tissues may occur only after prolonged exposure to AN. Notably absent from the available
mode-of-action database are experiments designed to detect a range of possible DNA adducts in
target tissues following repeated oral or inhalation exposure at exposure levels producing tumors
in the chronic bioassays.
Guengerich et al. (1986) discussed the findings in which AN was metabolized by liver
microsomes but not brain microsomes to form CEO. N7-(2-oxoethyl)guanine DNA adducts were
detected in liver but not the brain, yet AN induced tumors in the rat brain with chronic exposure
but not in the adult rat liver. In addition, AN-induced UDS was demonstrated in rat liver but not
brain (Hogy and Guengerich, 1986). Guengerich et al. (1986) proposed that AN was
metabolized in the liver to CEO. Since CEO formed by liver microsomes from AN has a half-
life of about 2 hours in neutral buffer, it can be transported easily via blood from liver to the
brain. Although DNA adducts have not yet been detected unequivocally in rat brains, CEO has
been measured in rat brains (Kedderis et al., 1993b) and shown to bind to brain DNA (Farooqui
and Ahmed, 1983b). Liver cells are efficient in repairing DNA damage via UDS while brain
cells do not have this capability, and may explain why the rat brain but not the adult rat liver is a
target tissue of AN carcinogenicity.
Meek et al. (2003) noted that there are several aspects of the development of AN-induced
tumors that are characteristic of tumors induced by compounds or metabolites that directly
interact with DNA. These comparisons add support to the evidence of a mutagenic mode of
carcinogenic action.
• Tumors are systemic and occur at multiple sites. Exposure-related increased incidences
were found for forestomach tumors, CNS tumors, and Zymbal gland tumors in chronic
oral exposure bioassays with F344 rats and Sprague-Dawley rats (which also showed an
exposure-related increased incidence of tongue tumors); for forestomach and Harderian
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gland tumors in a chronic oral exposure bioassay with B6C3Fi mice; for brain tumors,
Zymbal gland tumors, intestinal tumors, and tongue tumors in a chronic inhalation
exposure bioassay with Sprague-Dawley rats; and for brain tumors, Zymbal gland
tumors, hepatomas, and extrahepatic angiosarcomas in a chronic inhalation bioassay with
Sprague-Dawley rats exposed during gestation and extending throughout adulthood.
Particularly noteworthy is that Zymbal gland tumors in rats commonly occur with
carcinogens that are also mutagens. Of the 27 chemicals associated with site-specific
tumor induction in this gland found in NTP database, 23 were mutagenic in the
Salmonella assay. The NTP database indicated that these chemicals were multisite
carcinogens. Melnick (2002) observed that epoxide-forming chemicals usually induced
Zymbal gland and brain tumors in rats; lung, liver, Harderian gland, and circulatory
systems in mice; and mammary gland and forestomach tumors in both species. Hence,
AN induced tumors at sites consistent with other DNA reactive, epoxide-forming
chemicals.
Tumors sometimes occur at nontoxic doses or concentrations. Elevated incidences for
CNS tumors occurred in Sprague-Dawley (Quast, 2002; Quast et al., 1980a) and F344
(Johannsen and Levinskas, 2002b; Biodynamics, 1980c) rats chronically exposed to
drinking water concentrations (30 or 35 ppm) that did not induce elevated incidences of
nonneoplastic CNS lesions in interim sacrifices at 6, 12, or 18 months.
Tumors developed as early as 7-12 months following AN exposure. In a chronic drinking
water bioassay, CNS tumors were observed in Sprague-Dawley female rats treated with
300 ppm AN as early as 0-6 months (1/1 [0-6 months], 5/13 [7-12 months], 14/23 [13-
18 months], and 11/11 [19-24 months]) (Quast, 2002). Nearly all tumors in females rats
treated with 35 or 100 ppm AN and male rats from each treatment group were detected
after 13 months of exposure (Quast, 2002). The time of first detection of brain tumors
was approximately 16 months in a chronic drinking water bioassay in F344 rats (481 days
for males and 495 days for females) (Johannsen and Levinskas, 2002b; Biodynamics,
1980c). In a three-generation reproductive toxicity study, incidences of brain tumors
were 0/19, 1/20, and 2/24 in FO breeding females treated with 1, 100, or 500 ppm AN,
respectively, in drinking water for 48 weeks (approximately 12 months). In the other
generations, incidences of brain tumors were 0/20, 1/19, and 4/17 for the Fl breeding
females and 0/20, 1/20, and 1/20 for the F2 breeding females under the same exposure
conditions (Friedman and Beliles, 2002). The available evidence indicates that most AN-
induced brain tumors require at least a half-lifetime duration of exposure to develop, but
tumors in AN-exposed animals have been observed following shorter exposure durations,
specifically in female Sprague-Dawley rats exposed to 300 ppm in drinking water.
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• The ratio of benign to malignant tumors is small. The brain tumors noted in AN-exposed
Sprague-Dawley or F344 rats were astrocytomas, which are malignant tumors. Most of
the Zymbal gland tumors were also malignant tumors.
4.7.3.2.5. Human relevance. The postulated key events, the metabolism of AN to the DNA-
reactive compound, CEO, and the alteration of the genetic material leading to tumor-inducing
mutations, are both possible in humans. The metabolic scheme of AN in rats and humans is
similar. Humans are known to be able to activate AN to its reactive metabolite, CEO. As
discussed in Section 4.5.2.1, studies that demonstrate the mutagenicity of AN in humans are
available. There is evidence for mutagenicity of AN in exposed humans. Thus, the mutagenic
mode of action of AN-induced carcinogenicity is considered to be relevant to humans.
4.7.3.3. Other Possible Modes of Action
Other modes of action may contribute, along with the mutagenic mode of action, to
tumorigenesis. These additional modes of action are evaluated in the following subsections. The
results of these evaluations indicate that these modes of action are not likely to be principal
modes of action or to contribute to the carcinogenicity of AN in a significant manner.
4.7.3.3.1. Oxidative stress
Key events
This mode of action hypothesizes that ROS are generated either directly from the oxidant
or are indirectly produced via activation of endogenous sources when the oxidant or its
metabolite is distributed to the target organ. Oxidative stress is induced. These free radical ROS
can interact with DNA and produce DNA damage leading to gene mutation for tumor initiation.
ROS can also interact with lipids via lipid peroxidation, resulting in cell damage.
One of the most prevalent biomarkers of oxidative DNA damage is 8-oxodG. This DNA
lesion has been found to produce mutations involving GC^TA transversions due to base
mispairing and AT^CG transversions due to misincorporation during DNA synthesis (Cheng et
al., 1992). G-C base pairs provide a common target for activating point mutations (e.g., in both
p53 and retinoblastoma tumor suppressor genes and in the ras family of oncogenes). Thus,
G-C base pairs in both tumor suppressor genes and oncogenes may represent a vulnerable target
for mutation by oxidative stress (Guyton and Kensler, 1993). The induction of base changes in
the DNA sequence of these genes may be the basis for tumor initiation by the oxidant.
Another role that oxidants may play in carcinogenesis is tumor promotion. ROS
generating systems are known to possess some of the biochemical actions of tumor promoters,
such as promoting a rapid and sustained decrease in antioxidant defenses, including SOD,
catalase, and glutathione peroxidase activities (O'Connell et al., 1986; Slaga et al., 1981).
Tumor growth may also be promoted by oxidative stress via modification of gene expression
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through induction of gene transcription factors (e.g., NF#B or transcription factor protein [AP-1])
or change in DNA methylation status by ROS. Signal transduction pathways, including
AP-1 and NF>B, are known to be activated by ROS, and they lead to the transcription of genes
involved in cell growth regulatory pathway. Oxidative DNA damage can also result in DNA
hypomethylation by interfering with the ability of methyltransferases to interact with DNA,
allowing the expression of normally quiescent genes and promoting tumor growth. ROS can
also induce the release of calcium from intracellular stores, resulting in the activation of kinases,
including PKC (Larsson and Cerrutti, 1989), which is known to regulate many intracellular
processes, including those related to growth and differentiation. Hence, any one or a
combination of these events may lead to enhanced cell proliferation or inhibition of programmed
cell death.
The plausibility of an oxidative stress-related mutagenicity mode of action for AN-
induced tumors is evaluated in the following discussion.
Strength, consistency, specificity of association
Experimental data for oxidative stress-related mutagenicity of AN are discussed in
Section 4.5.2.5 and summarized in Table 4-56. These studies are only briefly discussed here.
In vitro studies
There is some in vitro experimental evidence that supports an oxidative stress-related
mode of action for AN. When SHE cells were treated in vitro with 0-75 ug/mL AN in 12.5
ug/mL increments, there was a dose-dependent increase in morphological transformation at 50,
62.5, and 75 ug/mL after 7 days of exposure (Zhang et al., 2000). Levels of 8-oxodG isolated
from cells incubated with 75 ug/mL AN were increased to 192 and 186% of control after 2 and 3
days, respectively; however, no increase in 8-oxodG was observed after 1 or 7 days.
AN-induced oxidative stress in SHE cells was confirmed in a later study by Zhang et al.
(2002). AN at 25, 50, or 75 ug/mL increased the amount of ROS in SHE cells after 4, 24, and
48 hours of treatment and increased xanthine oxidase activity 24 and 48 hours after treatment
with 75 ug/mL AN. AN also caused temporal changes in GSH levels and antioxidant enzyme
catalase and SOD activities. The involvement of CYP450 metabolism of AN in the production
of oxidative stress was indicated by the observation that inclusion of a nonspecific suicidal
inhibitor of CYP450 enzyme, ABT (0.5 mM), in the medium resulted in a significant reduction
(about 77%) in the cell transformation activity of 75 ug/mL AN (Zhang et al., 2002).
In another study, when cultured rat astrocytes were exposed for 4 or 24 hours to 0.01, 0.1,
or 1 mM AN, up to a 3.9-fold increase in 8-oxodG was found in cellular DNA of the rat
astrocytes (Kamendulis et al., 1999a). No increase in 8-oxodG was found in rat hepatocytes
exposed to 0.01, 0.1, or 1 mM AN for 4 or 24 hours (Kamendulis et al., 1999a). Pu et al. (2006)
also reported a 3-fold increase in oxidative DNA damage (as measured by the fpg-modified
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comet assay) in cultured D1TNC1 rat astrocytes treated with 1 mM AN for 24 hours. When
NHAs were treated with 200-400 uM AN for 12 hours, a four- to sevenfold increase in the
generation of ROS and a greater than twofold increase in 8-oxodG were observed (Jacob and
Ahmed, 2003b).
In vivo studies
Three studies in rats are available that investigated oxidative DNA damage in the brain
after exposure to AN in drinking water (Pu et al., 2009; Jiang et al., 1998; Whysner et al.,
1998a). A two- to threefold increase in 8-oxodG levels was found in cellular DNA from brain
cortex of Sprague-Dawley rats exposed to 50, 100, or 200 ppm AN in drinking water for 28 or 90
days (Jiang et al., 1998). Levels of 8-oxodG in DNA increased with increasing exposure levels
in this study. At 90 days, levels of 8-oxodG were more than twofold higher in DNA from 50
ppm rat brain cortex compared with controls (Jiang et al., 1998). No increased levels of 8-
oxodG were found in the liver DNA of exposed rats at all dose levels following 14, 28, or 90
days of exposure (Jiang et al., 1998). The liver is not a target organ for AN-induced
carcinogenicity in adult rats.
In a follow-up study, Pu et al. (2009) reported an increase in 8-oxodG levels in brain and
WBC DNA of male Sprague-Dawley rats exposed to 100 or 200 ppm AN in drinking water for
28 days. A dose-dependent increase in DNA damage in the brain and WBCs of the rats as
detected by the fpg-modified comet assay was also reported. Cotreatment with 200 ppm NAC
blocked the increase in 8-oxodG levels and DNA damage in both brain and WBCs. Pu et al.
(2009) concluded that AN induced oxidative DNA damage in the treated rats, and that the
antioxidant action of NAC prevented the oxidative damage when coadministered with AN.
Certain issues associated with the study design and analytical methods used by Pu et al.
(2009) influence the interpretation of the reported findings. As a means to measure oxidative
DNA damage, the fpg-modified comet assay has limited specificity; the assay has been shown to
detect oxidative and alkalative DNA damage (Smith et al., 2006; Speit et al., 2004). The fpg-
modified comet assay has been shown to be especially sensitive for the detection of DNA
damage by N-7 guanine alkylation, which was responsible for the observed DNA damage by
alkylating agents methylmethanesulfonate (MMS) and ethylmethanesulfonate (EMS) (Speit et
al., 2004). N-7 guanine adduct was detected in liver and brain (at the limit of detection) of AN-
treated rats. Therefore, the DNA damage observed by Pu et al. (2009) using the fpg-modified
comet assay may be a combination of oxidative DNA damage and damage resulting from N-7
guanine alkylation. Additionally, NAC was used by Pu et al. (2009) to demonstrate the effect of
an antioxidant to protect against observed DNA damage. NAC is a precursor of GSH. GSH
conjugates with AN to form N-acetyl-S-(2-cyanoethyl)cysteine, which can then be excreted in
the urine, thereby reducing DNA damage by reducing the availability of AN for oxidation to
CEO. The protective effect of NAC on AN toxicity was demonstrated by Carrera et al. (2007)
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(see Section 4.5.1.1.4). Therefore, the protective effect observed by Pu et al. (2009) with the
coadministration of NAC with AN may reflect increased detoxification of AN rather than the
antioxidant action of NAC.
In another study, levels of 8-oxodG in brain nuclear DNA were increased by about
twofold in Sprague-Dawley rats exposed to 30 or 300 ppm AN in drinking water for 21 days and
by about 1.5-fold in rats exposed to 100 ppm AN for 94 days (Whysner et al., 1998a). However,
no significant increase in 8-oxodG levels was found in F344 rats exposed to 10, 30, or 100 ppm
for 21 days (Whysner et al., 1998a). (A statistically insignificant increase of about 30% was
found in the 3-100 ppm AN dose groups.) Levels of 8-oxodG in liver DNA were increased by
about 1.4-fold in Sprague-Dawley rats exposed to 30 or 300 ppm for 21 days and by about
1.3- and 2-fold following exposure to 100 ppm for 10 and 94 days, respectively (Whysner et al.,
1998a). Levels of 8-oxodG in liver DNA were not measured in F344 rats in this study. No
significant increase in 8-oxodG levels in DNA of forestomach (a target organ of AN
carcinogenicity) were found in Sprague-Dawley rats exposed to 3-300 ppm AN (see
Table 4-55).
Several inconsistencies can be found in the results of Jiang et al. (1998) and Whysner et
al. (1998a). While a dose-related increase in 8-oxodG levels in brain cortex DNA of Sprague-
Dawley rats exposed up to 200 ppm AN was reported by Jiang et al. (1998), a twofold increase
in 8-oxodG levels in brain DNA compared with controls was found for Sprague-Dawley rats
exposed to either 30 or 300 ppm in the study by Whysner et al. (1998a). Thus, there was no
increase in 8-oxodG level with a 10-fold increase in exposure concentration in Sprague-Dawley
rats (Whysner et al., 1998a). While 8-oxodG levels were measured via different sample
preparation methods in these two studies (i.e., 8-oxodG level was measured in nuclear DNA in
whole brain in Whysner et al. [1998a] and in cellular DNA from brain cortex in Jiang et al.
[1998]), 8-oxodG level measured in Sprague-Dawley rats exposed to 3 or 30 ppm AN for 21
days in Whysner et al. (1998a) was comparable to that measured in Sprague-Dawley rats
exposed to 5 or 50 ppm AN for 28 days in Jiang et al. (1998) (0.86/105 and 1.35/105 dG vs.
1.8/105 and 2.5/105 dG). Thus, inconsistencies in the two studies regarding 8-oxodG levels in rat
brain are not likely due to differences in sample preparation.
Inconsistencies were also found in 8-oxodG levels in livers in the two rat studies. Jiang
et al. (1998) reported no increase in 8-oxodG levels in liver DNA of exposed rats, but Whysner
et al. (1998a) reported a 1.4-fold increase in 8-oxodG levels in liver DNA of Sprague-Dawley
rats exposed to 30 or 300 ppm for 21 days and a 1.3- and twofold increase following exposure to
100 ppm AN for 10 and 94 days, respectively. Moreover, the increase in 8-oxodG levels in brain
DNA was not much higher than that in liver DNA (see Table 4-56). The rat brain is a target
organ for AN-induced carcinogenicity but not adult rat liver. In addition, no significant increase
in 8-oxodG levels in forestomach DNA was found. Thus, no specificity is indicated regarding
increased 8-oxodG levels (oxidative DNA damage) in target organ DNA and tumor formation.
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Inconsistencies were also found regarding AN-induced disruption of antioxidant defense.
While Jiang et al. (1998) reported increases in ROS and concomitant persistent decreases in
antioxidant enzyme catalase activity in the brain cortex of Sprague-Dawley rats exposed to 50-
200 ppm AN in drinking water after 14, 28, and 90 days, as well as decrease in SOD activity and
GSH level in all dose groups after 14 days of treatment, Whysner et al. (1998a) reported no
changes in catalase and glutathione peroxidase activities and GSH levels in the brains of
Sprague-Dawley rats treated with 3, 30, or 300 ppm AN in drinking water for 21 days. Whysner
et al. (1998a) did report a dose-related increase in cysteine levels in the brains of Sprague-
Dawley rats exposed to AN for 21 days, and the increase was significant for the 300 ppm dose
groups. However, no changes in GSH and cysteine levels were found in the brains of F344 rats
exposed to 0, 1,3, 10, 30, or 300 ppm AN in the drinking water for 21 days (Whysner et al.,
1998a). Since F344 rats exposed to these dose levels of AN also developed brain tumors in
chronic bioassay, the formation of these tumors cannot be explained by oxidative stress resulting
from disruption of antioxidant defense. Whysner et al. (1998a) also found no changes in
cytochrome oxidase activities in the brain mitochondria of both exposed Sprague-Dawley rats
and F344 rats. Cyanide, a metabolite of AN, is a noncompetitive inhibitor of cytochrome
oxidase. No change in cytochrome oxidase activity indicated that no metabolic hypoxia occurred
in brain mitochondria as a result of inhibition of the enzyme by cyanide. Therefore, cyanide-
induced metabolic hypoxia did not appear to be involved in the mechanism of ROS generation
by AN.
Moreover, Whysner et al. (1998a) reported no changes in TEARS in the brains of all
groups of AN-exposed Sprague-Dawley rats, indicating the absence of lipid peroxidation,
another biomarker of oxidative stress and oxidative lipid damage. Jiang et al. (1998) reported
significant increase in MDA only in the brain cortex of rats exposed to 200 ppm AN for 14 days
and not in other dose groups at 14 days. No increase was found in all dose groups at 28 and
90 days. Since lipid peroxidation is another biomarker of oxidative stress, there is no strong
evidence for occurrence of significant oxidative stress in the brain of AN-exposed rats.
In addition, Chantara et al. (2006) demonstrated that AN induced ERK activation via
PKC in SK-N-SH neuroblastoma cells. However, oxidative stress was found not to be involved
in AN-induced ERK1/2 activation, which played a crucial role in cell proliferation and tumor
progression. Thus, the potential tumor promotion effect of AN has not been related to oxidative
stress.
Dose-response concordance
Levels of 8-oxodG in DNA from brain cortex of Sprague-Dawley rats (Jiang et al., 1998)
or rat astrocytes (Kamendulis et al., 1999a) increased with increasing in vivo (50, 100, 200 ppm)
or in vitro (0.01, 0.1, 1.0 mM) exposure levels. However, 8-oxodG levels in DNA from brain of
F344 rats exposed to 1-100 ppm AN were not significantly increased. In addition, 8-oxodG
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levels in DNA from brain of F344 rats did not show a dose-response relationship (see Table 4-
55), and no correlation can be found between 8-oxodG levels in brain DNA of these rats and
brain tumor incidence of F344 rats exposed to the same concentration of AN in drinking water
for 2 years (Table 4-59). Thus, experimental data do not support the hypothesis that brain
tumors from F344 rats are the result of oxidative DNA damage.
Table 4-59. 8-OxodG in brain DNA and brain tumor incidence in male F344
rats exposed to AN in drinking water
Dose group (ppm)
0
1
3
10
30
100
8-oxodG (mol/105 mol dG)a
0.79 ±0.37
0.84 ±0.25
1.07 ±0.41
1.04 ±0.30
1.03 ±0.38
1.06 ±0.48
Incidence of brain astrocytomasb
2/160
2/80
1/78
2/80
10/79
21/76
aData are from 21-d drinking water study by Whysner et al. (1998a).
bData are from 2-yr drinking water study by Biodynamics (1980b) and Johannsen and Levinskas (2002b). The
denominators for incidence of brain astrocytomas excluded rats from the 6- and 12-mo interim sacrifices and rats
that died before the appearance of the first tumor for this site.
Moreover, although Whysner et al. (1998a) demonstrated a significant increase in levels
of 8-oxodG in brain DNA of Sprague-Dawley rats exposed to AN for 21 days, no correlation can
be found between 8-oxodG levels in brain and tumor incidence in the 2-year bioassay
(Table 4-60). Therefore, dose-response data on Sprague-Dawley rats from Whysner et al.
(1998a) did not support oxidative DNA damage as the mode of action for AN-induced brain
tumor formation.
Table 4-60. 8-OxodG in brain DNA and brain tumor incidence in male
Sprague-Dawley rats exposed to AN in drinking water
Dose group (ppm)
0
1
3
10
30
100
8-oxodG (mol/105 mol dG)a
0.62 ±0.08
0.86 ±0.41
1.35 ±0.49
ND
ND
1.29 ±0.10
Incidence of brain astrocytomasb
1/80
ND
ND
8/47
19/48
23/48
Data are from 21-d drinking water study by Whysner et al. (1998a).
bData are from 2-yr drinking water study by Quast (2002).
ND = cannot be determined
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Temporal relationships
The detection of increased 8-oxodG levels in brains of Sprague-Dawley rats exposed for
subchronic durations (see Table 4-60) (Jiang et al., 1998; Whysner et al., 1998a) is temporally
consistent with oxidative DNA damage being a plausible key precursor event in the development
of later-appearing tumors. However, no significant increase in 8-oxodG levels in the brain of
exposed F344 rats was found (see Table 4-59).
Biological plausibility and coherence
The demonstration of AN-induced oxidative DNA damage in rat brain cortexes following
subchronic-duration exposure to AN at dose levels producing brain tumors with chronic
exposure would have provided support for the involvement of an oxidative stress-related mode
of action in AN carcinogenicity. However, no increase in oxidative DNA damage was found in
F344 rats exposed to AN in drinking water. Brain tumors were found in F344 rats at similar
frequencies as Sprague-Dawley rats in chronic bioassays. Thus, brain tumors in F344 rats cannot
be explained by oxidative DNA damage, and the predicted greater sensitivity of Sprague-Dawley
rats vs. F344 rats based on 8-oxodG levels measured in short-term studies is not reflected in the
cancer bioassays. In addition, the presence of increased oxidative DNA damage in the livers of
Sprague-Dawley rats exposed to AN in drinking water in one study (Whysner et al., 1998a) but
not in another (Jiang et al., 1998) also raised questions regarding a significant association
between oxidative DNA damage and tumor formation.
The origin of oxidative stress is unclear. Proposed molecular mechanisms that may be
involved in AN induction of oxidative stress in the brain (or other toxicity targets) include direct
generation of free radicals by AN (or its metabolites), stimulation by AN or its metabolites of
systems that generate free radicals, binding of AN to free radical scavengers (e.g., GSH, vitamins
C or E) and depletion of stores of these antioxidants, inhibition of the expression or activities of
antioxidant enzymes, and interference of mitochondrial respiratory electron flow via cyanide
inhibition of cytochrome c oxidase (Zhang et al., 2002; Jiang et al., 1998). As discussed
previously, some of these potential mechanisms have been ruled out.
As with the mutagenic mode of action, there are no data currently available to indicate
how or if this oxidative stress-related mode of action may explain the occurrence of brain tumors
in rats but not in mice, following chronic exposure to AN.
Human relevance
The key metabolic step involved in oxidative stress, i.e., oxidation of AN via the
CYP2E1 pathway, and antioxidant stores (e.g., GSH) are known to occur in humans. This
hypothetical mode of action, if it occurs, is considered to be relevant to humans.
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Conclusion
While plausible, this postulated mode of action is not supported by in vivo studies in rats.
Studies on oxidative DNA damage or oxidative lipid damage in the brain of exposed rats do not
provide sufficient evidence to support this mode of action. In addition, Chantara et al. (2006)
demonstrated AN-induced ERK activation via PKC in SK-N-SH neuroblastoma cells. Oxidative
stress was found to be not involved in AN-induced ERK1/2 activation, which plays a crucial role
in cell proliferation and tumor progression. Therefore, while this mode of action may play a
role, it is not likely to be the principal mode of action for AN-induced carcinogenicity.
4.7.3.3.2. Other modes of action for brain tumors
Key events
Other possible modes of action by AN or its metabolites to directly or indirectly alter the
expression of genes (either at the level of translation, post-translation, or protein activity),
leading to the loss of control of cell growth and the ultimate promotion of initiated cells into
brain tumors include stimulation of cell proliferation (mitogenic), cytotoxicity with subsequent
reparative cell proliferation (cytotoxic), inhibition of programmed cell death (anti-apoptotic), and
inhibition of GJIC. Studies specifically designed to examine these possible modes of action are
restricted to a study of GJIC in rat astrocytes (Kamendulis et al., 1999b).
Strength, consistency, and specificity of association
No studies are available that examine cellular proliferation indices in brain cells
following in vivo exposure to tumor-producing doses of AN. In in vitro studies with rat
astrocytes, indices of cytolethality after 24 hours of exposure occurred at higher AN
concentrations (2.5, 5.0, and 10.0 mM) than increased SoxodG levels in DNA (0.01, 0.1, and
1.0 mM) (Kamendulis et al., 1999a). The available data do not support a prominent role for
cytotoxic or mitogenic modes of action in AN-induced rat brain tumors.
Inhibition of GJIC has been shown to correlate with tumor promotion activity (i.e., the
loss of control of growth) and to be induced by various chemical agents thought to operate via
other modes of carcinogenic action, such as phorbol esters and PB (Kamendulis et al., 1999b).
Exposure of rat astrocytes to 0.01, 0.1, or 1 mM AN for 4-48 hours statistically significantly
inhibited GJIC compared with controls (Kamendulis et al., 1999b). The inhibition was reversible
and prevented by the presence of an antioxidant, a-tocopherol, or a precursor for the synthesis of
glutathione, OTC, in the culture medium. The results are consistent with the involvement of
AN-induced oxidative stress in the inhibition of GJIC. The inhibitory concentrations were the
same as those that produced oxidative DNA damage in a companion experiment (Kamendulis et
al., 1999a). The specificity of the inhibitory response to rat astrocytes was demonstrated by the
lack of inhibition of GJIC in rat hepatocytes exposed to 0.01, 0.1, or 1.0 mM AN (Kamendulis et
al., 1999b).
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Dose-response concordance
The inhibition of GJIC in rat astrocytes increased with increasing sublethal
concentrations in the range of 0.01-1 mM (Kamendulis et al., 1999b). A concordance of these
concentrations to dose levels associated with brain tumors is not available.
Temporal relationships
The acute nature of the observed inhibition of GJIC is consistent with the hypothesis that
this action may be one of a number of precursor events involved in tumor promotion.
Biological plausibility and coherence
Other potential modes of action for AN-induced brain tumors have not been adequately
studied. The involvement of inhibition of GJIC by oxidative stress induced by AN or its
metabolites is possible, based on the limited evidence with rat astrocytes (Kamendulis et al.,
1999b). However, the available histopathology data from interim sacrifices of the chronic rat
bioassays do not support the involvement of a mode of action involving brain cell cytotoxicity
followed by reparative cell proliferation.
4.7.3.3.3. Conclusions about modes of action for brain tumors. Data gaps exist in the current
understanding of the mode of action for carcinogenicity of AN. However, there is experimental
evidence to support mutagenicity as the principal mode of action. Key events are the generation
of DNA damage by the AN metabolite, CEO, and interaction with DNA. There is in vitro and in
vivo evidence to support the occurrence of key events in the brain following AN exposure.
Other modes of action may contribute, along with mutagenesis, to tumorigenesis. For example,
AN-induced ERK activation via PKC may play a role in cell proliferation and tumor progression.
However, limited experimental evidence does not support these modes of action as alternatives.
Available data are inadequate to establish an oxidative stress mode of action for AN-induced
carcinogenicity.
The available data do not provide an explanation for AN induction of brain tumors in
F344 and Sprague-Dawley rats but not in B6C3Fi mice. However, it should be noted that
generally mice are much less susceptible than rats in developing brain tumors resulting from
exposure to chemical carcinogens (Rice and Wilbourn, 2000; Radovsky and Mahler, 1999).
Hence, this species difference in response is not limited to AN alone.
Possible differences between rats and humans in distribution of AN or its metabolites to
the brain, susceptibility to oxidative stress, or repair of DNA damage have not been investigated.
Identified differences between rats and humans in AN disposition are restricted to the finding of
higher rates of metabolic oxidation of AN in human vs. rat hepatic microsomes, presumably due
to a more active EH in humans (Kedderis and Batra, 1993; Kedderis et al., 1993c). The
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relevance of this apparent difference in AN metabolism to possible species differences in
susceptibility to AN carcinogenicity in the brain is not understood. Within the framework of the
hypothesized mutagenic mode of action, the balance between the formation of CEO by CYP2E1
and its hydrolysis by EH is thought to be important in determining the levels of CEO that might
be available to bind DNA or GSH. Experimental comparison of this balance in human and rat
brain tissues is not available.
In summary, AN is proposed to induce brain tumors via a mutagenic mode of action.
4.7.3.4. Possible Modes of Action for Forestomach Tumors
There is evidence to suggest that, once delivered to forestomach epithelial cells, AN or its
metabolites shift the normal balance between cell proliferation and apoptosis, leading to
hyperplasia and eventually, with continued exposure and sustained net cellular proliferation, to
tumor formation. Sustained increased cellular proliferation is viewed as the key precursor event
in this hypothesized mode of carcinogenic action. Exposure of male F344 rats to gavage doses
of about 12 and 23 mg/kg-day AN for 6 weeks produced minimal to mild hyperplasia and
hyperkeratinization of the squamous mucosa of the forestomach but not the epithelium of the
glandular stomach or the liver (Ghanayem et al., 1997). The early induction of hyperplasia by
AN doses that resulted in forestomach tumors in chronic rat bioassays is temporally consistent
with the involvement of sustained cell proliferation in the development of these tumors. A dose-
related increase in cell proliferation (as determined by BrdU incorporation into DNA) was
observed in the forestomach epithelium, but no increases were found in the glandular stomach or
the liver, which are not targets of AN toxicity or carcinogenicity from chronic exposure to AN
during adulthood. At the high-dose level, an increase in apoptotic cells was found. These results
suggest that AN stimulation of cellular proliferation overcame the apparent stimulation of
apoptosis at the higher dose, since hyperplasia was observed. Whether or not the stimulation of
cellular proliferation was due to a reparative response to cytotoxicity or to a mitogenic action is
uncertain. Acute administration of a higher dose of AN (50 mg/kg) caused gastric mucosal
necrosis in rats, which was shown to involve CYP-mediated metabolism of AN (Ghanayem et
al., 1985; Ghanayem and Ahmed, 1983). Whether or not these findings relate to the stimulation
of cellular proliferation and development of hyperplasia at the lower doses is uncertain.
Possible mutagenic modes of action for forestomach tumors, such as those investigated
for brain tumors, have not been investigated. However, DNA damage, as detected by the comet
assay, was reported in the stomach of rats and mice exposed by i.p. injection (Sekihashi et al.,
2002). No significant oxidative DNA damage (levels of 8-oxodG) was measured in the
forestomach of rats exposed to 3, 30, or 300 ppm in drinking water for 21 days (Whysner et al.,
1998a). In addition, AN was reported to bind to DNA in the stomach of rats following a single
oral dose of 46.5 mg/kg (Farooqui and Ahmed, 1983a). Given that a mutagenic MOA is
hypothesized for AN-induced brain tumors in rats, that mutagenic carcinogens usually cause
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tumors in multiple sites, and evidence of DNA damage in the forestomach of AN-treated rats,
mutagenicity is a likely mode of action for AN-induced forestomach tumors.
4.7.3.5. Possible Modes of Action for Other Tumors - Hepatoma, Mammary Gland, Lung,
Intestinal, Tongue, Zymbal Gland, andHarderian Gland Tumors
The incidence of benign or malignant mammary gland tumors was significantly increased
in female Sprague-Dawley and F344 rats exposed to AN (Quast, 2002; Johannsen and
Levinskas, 2002b). No mechanistic studies were conducted to explore the mode of action of
AN-induced mammary tumors in rats. However, DNA-reactive epoxides and epoxide-forming
chemicals commonly form mammary tumors in rats and mice (Melnick, 2002). It is assumed
that a mutagenic mode of action may be involved in the formation of AN-induced mammary
tumors in rats.
Evidence of a possible association between occupational exposure to AN and lung cancer
is found in some studies, including the best available epidemiologic study in which workers in
the highest exposure category (>8 ppm-years) with more than 20 years of employment displayed
a twofold increased risk for lung cancer compared with unexposed workers (Blair et al., 1998).
Human bronchial epithelium has CYP2E1 metabolic activities. In a short-term mutagenicity
assay, AN induced SCEs and DNA single-strand breaks in human bronchial epithelial cells
without the addition of S9 mix (Chang et al., 1990). AN also induced p53 and p21WAF1 proteins
in human embryonic lung fibroblasts (Rossner et al., 2002). Thus, a mutagenic mode of action is
supported for potential AN-induced lung cancer. In contrast, there is no convincing evidence of
lung cancer in rodents chronically exposed by the oral or inhalation routes. The possible mode
of action by which AN may induce lung tumors in humans and not in rodents has not been
investigated but may involve species differences in inhalation rates and anatomical features of
the respiratory tract.
By analogy to other mutagens that cause tumors at multiple sites in animal bioassays, it is
possible that a mutagenic mode of action may be involved in the formation of AN-induced
tumors of the intestines, tongue, liver, and Zymbal gland in rats, and Harderian gland in mice.
Although possible modes of carcinogenic action of AN at these sites have not been investigated,
a mutagenic mode of action is the most likely mode of action. CYP2E1 enzymes occur in
intestinal mucosa, and CEO can form in the intestine and bind to DNA. Moreover, as discussed
previously in Section 4.7.3.2.4, chemicals that cause Zymbal gland tumors in rats are usually
mutagens. In addition, there is no evidence of tissue specificity associated with AN that would
lead to mutagenesis in brain but not other organs. Moreover, AN appears to act systemically
with DNA damage, and tumors occur in multiple tissues of treated animals. Thus, the mutagenic
mode of action is considered relevant to all tumor sites.
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4.8. SUSCEPTIBLE POPULATIONS AND LIFE STAGES
4.8.1. Possible Childhood Susceptibility
Evidence that children may be more susceptible to the acute or chronic toxicity of AN is
restricted to a report that a young girl died after sleeping in a room fumigated with a commercial
form of AN called Ventox, whereas adults sharing the same room only experienced skin or eye
irritation (Grunske, 1949; see Section 4.1.3.1).
Animal studies that examined whether or not early-life stages of development are more
susceptible than adult stages to the induction of noncancer or cancer-related effects by AN
include a lifetime inhalation cancer bioassay in Sprague-Dawley rats (Maltoni et al., 1988); a
three-generation drinking water reproductive toxicity bioassay in Sprague-Dawley rats
(Friedman and Beliles, 2002); and a two-generation reproductive toxicity study of inhaled AN
vapors in Crl:CD (SD) rats (Nemec et al., 2008). In addition, a subacute study in Sprague-
Dawley rats (Szabo et al., 1984) also found weanling rats to be more susceptible than adult rats
to the action of AN on the adrenals (see Section 4.2.1.1). A discussion of evidence for early-life
susceptibility to cancer is provided in Section 5.4.4.3
As discussed in Sections 4.5.1 and 4.6.3, toxic effects from acute or chronic exposure to
AN have been associated with inhibition of glutathione-mediated detoxification of AN by
conjugation (e.g., glutathione depletion), transformations in the CYP2E1 metabolic pathway to
CEO and cyanide, and oxidative stress. Differences in enzymatic activities (e.g., CYP2E1, EH,
DNA repair enzymes, or antioxidant enzymes such as catalase or SOD) or pool sizes of reactive
oxygen scavengers (e.g., levels of vitamin E or C) or glutathione between early-life and adult
stages may result in life stage differences in susceptibility to AN toxicity.
For CYP2E1, immunoreactive CYP2E1 protein has been detected in human liver
microsomes as early as the second trimester (GDs 93-186) (Johnsrud et al., 2003). CYP2E1
enzyme activity increases shortly after birth but less in neonates than in older infants, children,
and adults (Johnson, 2003; Johnsrud et al., 2003; Vieira et al., 1996). However, in the fetal
brain, CYP2E1 activity is seen as early as 50 days gestation, with increasing levels seen to at
least the end of the first trimester (Brzezkinski et al., 1999). In rats, CYP2E1 protein is not
significantly expressed in fetal hepatic tissues, although an elevation of CYP2E1 mRNA was
seen within a few hours after birth, coincident with the transcriptional activation of the gene
(Borlakoglu et al., 1993). Moreover, only neonates expressed small quantities of the CYP2E1
protein, despite the large quantities of CYP2E1 mRNA found in perinatal and neonatal rats.
For EH, fetal immunoreactive enzyme content in human livers averages only 25% of that
found in adults with corresponding less enzyme activity (Cresteil et al., 1985). EH activity in
fetal liver and adrenal glands are about threefold higher than in kidney and lung with only a
weak correlation with gestational age between 10 and 25 weeks. Overall, EH activity in fetal
tissue is about 30-40% of that in adults (Pacifici and Rane, 1983b; Pacifici et al., 1983a). In
another study (Omiecinski et al., 1994), human microsomal EH (mEH) was detected in fetal liver
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as early as 7.2 weeks, although at a much lower level of mEH protein, and demonstrated a linear
increase with gestational age to a level at 77 weeks that is about half of that observed in adult
liver. However, mEH activity in the fetal lung did not correlate with increasing age. Lung mEH
activity from days 85 to 130 of gestation was maintained at consistent levels, at about the same
level as fetal hepatic EH at 53 days, and varied only by threefold. In addition, Omiecinski et al.
(1994) reported that mEH activities in liver or lung, from either fetal or adult tissues, did not
correlate with corresponding mRNA levels.
The overall balance of pertinent enzymatic activities (AN metabolizing enzymes and
antioxidant enzymes) and pool sizes of reactive oxygen scavengers and glutathione will
determine the relative susceptibility of an individual or a life stage. The relatively high activity
of CYP2E1 in the brain compared to the liver of the developing human fetus, and low EH
activity for detoxification raise concern for increased susceptibility in early life to lung and brain
tumors from AN exposure.
The importance of CYP2E1 metabolism in the acute toxicity and lethality of AN has been
demonstrated by the lack of lethality or gross signs of intoxication in CYP2El-null male mice
given single gavage doses up to 40 mg/kg, whereas all WT male mice given doses of 40 mg/kg
died within 3 hours of administration (Wang et al., 2002).
4.8.2. Possible Geriatric Susceptibility
Age-related reductions in antioxidant or glutathione pool sizes may increase
susceptibility of elderly people to the tissue-damaging actions of AN and reactive metabolites,
but specific studies of the possible increased susceptibility of elderly people or rats to AN are not
available. A 35% decrease in glutathione levels in the liver of aged F344 rats compared with
younger animals was associated with an age-related decrease in the levels and activity of
y-glutamylcysteine ligase, a key enzyme in the synthesis of glutathione (Suh et al., 2004). In
Wistar rats, the liver and kidney of 22-month-old rats showed significant decreases, compared
with 10-week-old rats, in glutathione and glutathione peroxidase and increased levels of
biomarkers of lipid peroxidation (Martin et al., 2003). In another study with F344 rats (Tian et
al., 1998), activities of several antioxidant enzymes in several tissues displayed an age-dependent
decline. Enzymatic activities showing significant decline with age included SOD in the heart,
kidney, and serum; glutathione peroxidase in the serum and kidney; and catalase activities in the
brain, liver, and kidney. These changes indicated a lower resistance to oxidative stress in older
animals.
The possible toxicological impact of an age-related decline in glutathione could be offset
by an age-related decline in CYP2E1-mediated metabolism of AN leading to reactive
metabolites (CEO), cyanide, or ROS. Many studies have reported that hepatic enzymatic
activities of CYP2E1 are lower in elderly human subjects (>65 years) compared with younger
adults (see Tanaka, 1998, for review). Decreased hepatic CYP2E1 enzyme activities have also
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been reported in aged rats. For example, hepatic CYP2E1 enzyme activities were decreased by
46% in 18-month-old rats compared with 8-month-old rats, whereas no age-related changes in
CYP2E1 mRNA or protein content were evident (Wauthier et al., 2004). Such age-related
declines in CYP2E1 enzyme activities may lead to lower tissue levels of reactive metabolites or
cyanide but also may lead to elevated levels or increased residence time of AN in aged tissues
compared with younger tissues. No definitive conclusions about the possible susceptibility of
the elderly to AN toxicity can be drawn without more specific studies designed to examine the
effects of age on susceptibility to AN toxicity.
As studied in a small number of individuals (n = 47), long-term occupational exposure to
AN increased the deletion rate in mitochondrial DNA to a level equivalent to that seen in a group
of elderly nonexposed subjects (n = 12) (Ding et al., 2003) (see Section 4.1.2.2). The study
authors suggested that AN may have an effect on the molecular processes of aging.
4.8.3. Possible Gender Differences
No reports of gender differences in susceptibility of humans to AN toxicity are available.
No consistent gender-related differences in carcinogenic responses were observed in rats or mice
following chronic oral exposure to AN, or in rats following chronic inhalation exposure. With
oral exposure, male and female groups showed similarly increased incidences of brain
astrocytomas (rats: Quast, 2002; Bigner et al., 1986; Biodynamics, 1980a, b, c; Quast et al.,
1980a), forestomach tumors (rats: Quast, 2002; Bigner et al., 1986; Biodynamics, 1980a, b, c;
Quast et al., 1980a; mice: NTP, 2001), Zymbal gland tumors (rats: Quast, 2002; Bigner et al.,
1986; Biodynamics, 1980a, b, c; Quast et al., 1980a), and Harderian gland tumors (mice: NTP,
2001). Following inhalation exposure to 80 ppm AN for 2 years, male and female exposed
groups of rats showed similarly increased incidences of brain/CNS tumors and Zymbal gland
tumors compared with the respective control groups (Quast et al., 1980b).
No marked gender-related differences in noncarcinogenic responses were observed in the
rat chronic oral toxicity study reported by Quast (2002). Groups of male and female rats were
exposed to AN in drinking water at concentrations of 0, 35, 100, or 300 ppm. At the 1-year
interim sacrifice, the only exposure-related noncancer histopathologic finding was an increase in
the incidence of forestomach squamous cell hyperplasia in rats exposed to concentrations of
100 ppm (4/10 males, 7/10 females) or 300 ppm (10/10 males, 9/10 females). At the 2-year
sacrifice, incidences of stomach lesions (nonglandular hyperplasia and/or hyperkeratosis) were
similar in male (15/80, 15/47, 44/48, and 45/80 for the control through high-concentration
groups, respectively) and female (20/80, 23/48, 41/48, and 47/48) rats at the same exposure level
(Quast, 2002). These data, however, give some indication that the forestomach epithelium of
female rats may have been slightly more susceptible than that of male rats: at the 1-year
sacrifice, 7/10 100 ppm females (70%) had lesions compared with 4/10 100 ppm males (40%); at
2 years, 23/48 35 ppm females (48%) had lesions compared with 15/47 35 ppm males (32%).
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In contrast to the lack of apparent gender differences in carcinogenic or noncarcinogenic
responses in rodents with chronic exposure, male mice appear to be more susceptible to the acute
oral toxicity of AN than female mice (Chanas et al., 2003; NTP, 2001).
In a 14-week study, groups of 10 male and 10 female mice were given 0, 5, 10, 20, 40, or
60 mg/kg AN in deionized water by gavage 5 days/week (NTP, 2001). Exposure-related
mortalities were restricted to the first week of the study and occurred in the 40- and 60-mg/kg
groups. Both male and female groups showed mortality, but only at 40 mg/kg was the incidence
of deaths higher in male mice compared with females (9/10 and 10/10, males, and 3/10 and
10/10, females, at 40 and 60 mg/kg, respectively).
In a subsequent study, groups of three to four male and three to four female WT or
CYP2El-null mice (mixed 129/Sv and C57BL) were given single gavage doses of 0, 2.5, 10, 20,
or 40 mg/kg in tap water and sacrificed 1 or 3 hours later (Chanas et al., 2003). All male WT
mice exposed to 40 mg/kg died within 3 hours, showing gross signs typical of cyanide poisoning
(rapid shallow breathing, cyanosis, trembling, and convulsions). In contrast, exposed female WT
mice showed milder gross signs of poisoning, and none died within the 3-hour period. One hour
after dose administration, concentrations of cyanide in blood were statistically significantly
elevated, compared with vehicle controls, in WT male mice given doses >2.5 mg/kg. At doses
>10 mg/kg, female WT mice also showed elevated cyanide levels in blood, compared with
controls, but cyanide levels were <50% that in male WT mice. Exposed CYP2El-null mice of
both genders showed no elevation in blood cyanide concentrations compared with vehicle
controls. Cyanide levels in brain and kidney tissues were also higher in WT males compared
with females; cyanide levels in liver and lung tissues showed less distinct differences between
male and female mice.
Expression of hepatic, renal, and pulmonary CYP2E1, soluble EH, and microsomal EH
were measured in male and female WT mice using Western blot analysis (Chanas et al., 2003).
In the liver, WT males showed greater expression of EH, both soluble and microsomal, than did
females; CYP2E1 levels were similar in males and females. In the kidney, male WT mice
showed markedly higher levels of CYP2E1 (about fourfold), moderately higher levels of soluble
EH (about twofold), and comparable levels of microsomal EH compared with female mice.
Higher CYP2E1 and soluble EH in the kidney of male mice provided explanation for higher
blood cyanide levels in the kidney and acute lethality in male mice. In the lung, no gender
differences in the expression of these enzymes were apparent. The results indicated that male
mice were more susceptible than female mice to the acute toxicity and lethality of AN and that
this difference was associated with higher blood, kidney, and brain levels of cyanide in males
shortly after dose administration. Chanas et al. (2003) noted that another possible explanation
was that female mice had greater detoxification capability of converting cyanide to thiocyanate,
as demonstrated by excretion of greater amount of thiocyanate in urine than males after repeated
administration of equal doses of AN (NTP, 2001).
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In a subsequent study, male mice (mixed 129/Sv and C57BL) showed higher blood levels
of cyanide than female mice 1 hour after gavage administration of 0.047, 0.095, 0.19, or
0.38 mmol/kg (2.5, 5, 10, or 20 mg/kg) AN (El Hadri et al., 2005). This difference between
genders was evident in WT mice and in microsomal EH-null (mEH-null) mice, but blood levels
of cyanide were lower in exposed mEH-null mice, compared with comparably exposed WT
mice. As in the previous experiments reported by Chanas et al. (2003), exposure to AN induced
no elevation of blood levels of cyanide in CYP2El-null mice of either gender. Western blot
analysis revealed no gender differences in expression of CYP2E1 in the liver of WT or mEH-
null mice or in expression of mEH in WT or CYP2El-null mice. However, expression of
soluble EH in the liver was greater in WT males, compared with females. No gender differences
in expression of this enzyme was observed in mEH-null mice or in CYP2El-null mice.
The effect of gender on the expression of CYP2E1 and EH have been investigated in
humans and rats. Gender was reported to have no influence on the level of CYP2E1 in human
liver (George et al., 1995). Sex-related patterns in the activity of hepatic EH activity was studied
in Sprague-Dawley rats (Chengelis, 1988). At week 4, epoxide activity in both male and female
rats was equivalent (3.5-4.0 nmol/minute per mg protein, or 85-100 nmol/minute per g liver).
However, there were consistent increases in activity in males from week 4 to 78, while activity in
females actually decreased, but returned to week 4 levels during the later stages (week 78-103).
EH activity in male liver peaked at week 78 (about 10 nmol/minute per mg protein, or
350 nmol/minute per g liver). There was a sharp decline in EH activity in aged male rats to
about 6 nmol/minute per mg protein, or 200 nmol/minute per g liver at 104 weeks. Cornet et al.
(1994) studied gender-related changes in microsomal and cytosolic EH activity in male and
female Brown Norway rats. At 15 weeks, microsomal EH activity was about the same for male
and female rats at 4.5 nmol/mg protein/minute. The microsomal EH activity decreased strongly
as a function of age in female rats, and in the 125-week-old females, the activity was only half of
that found in 15-week-old rats. However, there was no age-related change in males, although the
activity in the 83- and 125-week age groups was significantly lower than that in 28-week-old
males. The activity of EH activity was higher in males than in females in these aged animals. In
another study that compared hepatic microsomal EH in different strains of adult rats (170-250 g)
(Oesch et al., 1983), EH activities in females were found to be 71-88% of those in males in all
strains. Denlinger and Vesell (1989) studied the hormonal regulation on the developmental
pattern of EH in F344 rats, and found that EH activities in males increased gradually until
puberty, when activities in males rose rapidly to be from 1.5- to twofold higher than those in
females. The higher activity in males was not seen if the males were castrated 24 hours after
birth. When castrated males and females were injected with testosterone propionate (0.5 mg s.c.)
on days 1, 3, and 5 postpartum, increased mEH an cEH activities were observed at adulthood.
Thus, Denlinger and Vesell (1989) concluded that full adult expression of EH activities depends
on hormonal influences exerted neonatally.
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Gender-related differences were also observed in cytosolic EH activity of Brown Norway
rats (Cornet et al., 1994). Significant gender-related differences in the cytosolic EH activity
were found in 15-, 28-, and 83-week-old rats with the male animals showing the highest values.
In the oldest animals and in 56-week-old rats comparable cytosolic EH activities were found in
both genders.
In summary, the available data from animal studies provide no evidence of consistent or
marked gender differences in susceptibility to noncancer or cancer-related effects from chronic
exposure to AN. There is evidence that male mice are more susceptible than female mice to the
acute, cyanide-induced toxicity and lethality of AN, but whether or not this apparent gender
dimorphism extends to other species is uncertain at the present time.
4.8.4. Genetic Polymorphisms
4.8.4.1. CYP450
In humans, CYP2E1 exists in several modifications that differ in amino acid sequence
(alleles). Human variability in their susceptibility to the toxic effects of AN likely exists since
CYP2E1 activities may fluctuate between one person and another.
Microsomal CYP2E1 activities varied from 6- to 20-fold in human livers (Lucas et al.,
1993). Environmental factors, diet habits, and/or genetic factors may account for the observed
interindividual variations observed. In addition, CYP2E1 is elevated in obese overfed rats
(Salazar et al., 1988) and diabetic rats (Song et al., 1987), suggesting induction by increased
plasma levels of ketone bodies (Bellward et al., 1988). Moreover, CYP2E1 is elevated in
lymphocytes from poorly controlled insulin-dependent diabetics (Song et al., 1990). Thus, obese
and diabetic individuals may have elevated levels of CYP2E1.
Besides induction, polymorphism of the human CYP2E1 gene may have an impact on
AN metabolism in humans. Stephens et al. (1994) compared two restriction fragment length
polymorphic sites of the CYP2E1 gene (Rsa 1 and Dra 1) in 695 African-American, European-
American, and Taiwanese subjects. Rare alleles at these two loci have been associated with a
reduced risk for lung cancer in Japanese and Swedish populations. Stephens et al. (1994)
demonstrated that rare alleles (c2 and C) at the Rsa 1 and Dra 1 sites were at least twice as
frequent in Taiwanese populations (28 and 24%, respectively) compared with African-
Americans (1-8%) or European-Americans (4-11%), raising the possibility of differential
susceptibility to chemically induced cancers across ethnic groups. However, when Carriere et al.
(1996) measured the allele frequencies for Ras 1, Dra 1, and Taq 1 polymorphic sites in liver
CYP2E1 from kidney donors (n = 93) in Geneva, they failed to find a correlation between
frequencies of rare alleles and CYP2E1 activity. This implied that observed differences in
enzyme activity in humans were more likely to be the result of different levels of induction by
environmental factors or other genetic factors.
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McCarver et al. (1998) identified a 100-bp insertion mutation in the regulatory region of
the CYP2E1 gene. Associated with an elevated CYP2E1 metabolic activity, this insertion
mutation appears to be present only in obese people or persons who had recently consumed
alcohol. The incidence of this mutation was seen in 31% of 65 African-American samples but
only in 6.9% of 58 Caucasian samples (McCarver et al., 1998). If tumor formation is determined
by the activity of CYP2E1, this mutation might put affected individuals at a higher risk of cancer
from tumorigenic agents that are metabolized by CYP2E1.
Thier et al. (2002) were unable to detect any influence of six genetic CYP2E1
polymorphisms (G.1259C, A.3i6G, T_297A, 0.35!, G4804A, and T7668A) on the formation of
N-(cyanoethyl)valine Hb adducts of AN. Conversely, in a study confined to a cohort of
individuals from ethnic minorities (African-Americans and Mexican-Americans), Wu et al.
(1998) observed an increased incidence of the CYP2E1 Dral DD genotype in peripheral WBC
DNA in 126 patients with untreated lung cancer compared with 193 unaffected controls. This
could imply an etiological association between the CYP2E1 Dral polymorphism and tumor
formation in the lung. Altered toxicokinetic characteristics of CYP2E1 in persons exposed to
AN might result in some persons being more vulnerable than others to the tumorigenic effects of
the compound, with concomitant changes to the rate and amount of formation of the toxic
metabolite and associated changes in susceptible individuals.
Kim et al. (1996), investigating the differences in CYP2E1 activities between
20 Caucasian and 20 Japanese men, pointed out that the significantly lower activity of CYP2E1
in Japanese men might account for the lower rate of some cancers in Japanese compared to
Caucasian men.
4.8.4.2. Glutathione S-tramferases
Thier et al. (1999) studied the formation of Hb adducts of AN (N-[cyanoethyl]valine,
N-[methyl]valine, and N-[hydroxyethyl]valine) in a group of 59 people occupationally exposed
to the chemical. They reported their findings in relation to subjects' smoking habits and their
genetic status with respect to the GST isozymes GSTM1 and GSTT1. Included in the study was
an evaluation of smoking habits, since elevated adduct levels of AN in Hb have been reported in
smokers. There was no correlation between adduct levels and either the subjects' status of GST
isozymes or smoking habit. Thus, neither GSTM1 nor GSTT1 appears as a major AN-
metabolizing isoenzyme in humans. However, in a follow-up study with the same group of
59 workers, Thier et al. (2001) reported that polymorphism of the GSTP1 gene at codon 104 was
associated with a higher level of N-(cyanoethyl) valine adducts, while GSTM3 variants had no
effect on Hb adduct formation. Whether such polymorphisms influence health risks to AN-
exposed humans has not been evaluated. However, Zielinska et al. (2004) reported an increase
in the frequency of the GSTPlb/b genotype in children with cancer (OR = 5.7, CI = 2.4-13.8).
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4.8.4.3. EH
Two major polymorphisms of mEH have been identified in human population (Hasset et
al., 1997). One is a polymorphism in exon 3 that changes the tyrosine residue 113 (Tyrl 13) to
histidine (Hisl 13), and the other is an A^-G substitution in exon 4 that changes the histidine
residue 139 (His 139) to arginine (Argl39). In vitro expression studies demonstrated that with
the Tyrl 13His polymorphism, the corresponding mEH enzymatic activity is decreased by about
40% (Hasset et al., 1997). On the other hand, the Hisl39Arg polymorphism results in increased
enzyme activity. Population studies have demonstrated that the low activity 113His allele
correlates with an increased risk for lung cancer (Benhamou et al., 1998), colon cancer (Harrison
et al., 1999), hepatocellular carcinoma developed after aflatoxin exposure (McGlynn et al.,
1995), and chronic obstructive pulmonary disease (Smith and Harrison, 1997).
The mEH genotypes was shown to play a significant role in human sensitivity to the
genotoxic effects of exposure to 1,3-butadiene (Abdel-Rahman et al., 2003, 2001). The
carcinogenic and mutagenic effects of 1,3-butadiene are thought to be due to its epoxide
metabolites, and the hydrolytic pathway involving mEH is the main detoxification pathway for
1,3-butadiene-reactive intermediates in humans (Jackson et al., 2000). In a study of
49 nonsmoking workers from two styrene-butadiene rubber plants, the hprt gene mutation assay
was used as a biomarker of genotoxic effect of BD (Abdel-Rahman et al., 2003, 2001). Abdel-
Rahman et al. (2003, 2001) evaluated the effect of polymorphisms in both exon 3 and exon 4 of
the mEH gene as modifiers of individual susceptibility to the mutagenic response associated with
exposure to 1,3-butadiene, and found a progressive increase in hprt mutant frequency with
declining mEH activity in the high exposure group (>150 ppb). The highest frequency of hprt
mutant lymphocytes occurred in the group with the mEH low-activity genotype. Individuals
with low mEH activity had three- and twofold increases in hprt mutant frequency compared to
individuals with high and intermediate mEH activity, respectively. In the low exposure group,
there was no difference in hprt mutant frequency between high-, intermediate-, and low-activity
individuals. Although there are no studies that evaluate the role played by mEH genotypes in
human sensitivity to the toxicity of AN, since EH is involved in the hydrolysis of CEO and its
elimination, polymorphisms in exon 3 and exon 4 of the mEH gene are likely to have a role in
human susceptibility.
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5. DOSE-RESPONSE ASSESSMENTS
5.1. ORAL REFERENCE DOSE (RfD)
5.1.1. Choice of Principal Study and Critical Effect
As previously discussed in Section 4, no human studies currently exist that involve oral
exposures to AN because the primary route of AN exposure in humans is via inhalation. In
animals, two 2-year drinking water studies, one in Sprague-Dawley rats (Quast, 2002; Quast et
al., 1980a) and the other in F344 rats (Johannsen and Levinskas, 2002b; Biodynamics 1980c),
and a 2-year gavage study in B6C3Fi mice (NTP, 2001) provided the best available dose-
response data on which to base the RfD (see Table 4-58 for a summary of exposure protocols
and study results). These three studies were considered candidate principal studies for derivation
of the oral RfD for AN.
The available animal toxicity studies identify forestomach lesions (i.e., squamous cell
epithelial hyperplasia and hyperkeratosis) as the most sensitive, prevalent, and consistent
noncancer effect associated with chronic oral exposure to AN (see Table 4-58 and Figure 5-1).
Although anatomically, humans do not possess a forestomach, they have comparable squamous
cell epithelial tissues in their oral cavity and in the upper two-thirds of their esophagus (IARC,
1999). (It should be noted that in AN-treated beagles, histopathological lesions occurred in the
esophagus and tongue [Quast et al., 1978].) Moreover, the forestomach lesions observed in
animals were not likely due to the direct irritating effect of AN on gastric tissue. GI bleeding has
been observed with single s.c. or oral administration of AN in rats (Ghanayem and Ahmed,
1983) that likely resulted from the distribution of AN metabolites from blood into the GI
mucosa. Jacob and Ahmed (2003a) have also demonstrated that AN or its metabolites
accumulated and covalently interacted with the GI mucosa of F344 rats treated either orally or
intravenously with 2-[14C]-AN, supporting the theory of metabolic incorporation and
macromolecular interaction of AN or its metabolites with gastric tissue. Since the metabolic
scheme of AN in rats and humans is similar, and humans are able to activate AN to its reactive
metabolite, CEO, the forestomach lesions in rodents are considered relevant to humans.
Significantly elevated incidences of hyperplasia and hyperkeratosis in squamous
epithelium of the forestomach occurred in both male and female Sprague-Dawley rats exposed to
AN in drinking water at the two highest concentrations administered (i.e., 100 and 300 ppm)
with incidences approaching 100% at the highest dose (Quast, 2002; Quast et al., 1980a).
Female Sprague-Dawley rats also exhibited statistically significantly elevated incidences of
forestomach lesions at the lowest concentration of AN administered (i.e., 35 ppm). In F344 rats,
both males and females administered 3, 10, and 30 ppm AN in drinking water exhibited
statistically significantly elevated incidences of forestomach lesions with incidences of
approximately 20-30% (Johannsen and Levinskas, 2002a; Biodynamics, 1980c). However, in
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this same study, male and female F344 rats exposed to the lowest and highest concentrations of
AN in drinking water (i.e., 1 and 100 ppm) did not show statistically significantly elevated
incidences of forestomach lesions. In B6C3Fi mice, males exhibited statistically significantly
elevated incidences of forestomach lesions (hyperplasia or hyperkeratosis) at the two highest
doses administered (i.e., 10 and 20 mg/kg-day), while females showed statistically significantly
elevated incidences of forestomach lesions at the highest dose only (NTP, 2001).
The Johannsen and Levinskas (2002b) drinking water study in F344 rats was selected as
the principal study on which to base the RfD. This study was selected as the principal study
primarily because it employed five doses of AN, ranging from 1 to 100 ppm, and thus tested a
more complete range of doses, especially in the low-dose region, than did either the Quast (2002)
study in Sprague-Dawley rats or the NTP (2001) study in B6C3Fi mice, which both employed
three doses (i.e., 35, 100, and 300 ppm AN in drinking water in rats and 2.5, 10, and 20 mg/kg-
day in mice, respectively). In addition, Johannsen and Levinskas (2002b) and Quast (2002) are
both drinking water studies, which are preferred over the NTP (2001) gavage study in B6C3Fi
mice because drinking water exposure is more relevant to humans.
Other endpoints associated with AN exposure (i.e., chronic nephropathy, ovarian cysts,
and gliosis in the brain in rodents) were either less sensitive or were observed less consistently
across studies than forestomach lesions (see Table 4-58). Therefore, the squamous cell epithelial
hyperplasia and hyperkeratosis of the forestomach was selected as the critical effect for
derivation of the RfD.
5.1.2. Methods of Analysis—Including Models
While Johannsen and Levinskas (2002b) was selected as the principal study, dose-
response analyses using data from Quast (2002) in Sprague-Dawley rats and NTP (2001) in
B6C3Fi mice was conducted for comparison purposes. Incidences of forestomach lesions in
male and female Sprague-Dawley rats (Quast, 2002) and F344 rats (Johannsen and Levinskas,
2002b) provided four sets of dose-response data from which to derive candidate RfDs (see Table
5-1), while incidences of forestomach lesions from a chronic gavage study in male and female
B6C3Fi mice (NTP, 2001) provided an additional two sets of dose-response data (see Table 5-2).
As described further below, dose-response modeling of the incidence of forestomach
lesions was carried out using benchmark dose (BMD) modeling. For rats, in addition to
administered dose, two internal dose metrics (AN in blood and CEO in blood), as estimated by
PBPK modeling, were employed in BMD modeling. In mice, only administered dose was
employed as a dose metric for BMD modeling purposes.
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Table 5-1. Incidences of forestomach lesions (hyperplasia or
hyperkeratosis) in Sprague-Dawley and F344 rats exposed to AN in
drinking water for 2 years
Sex
Administered
concentration
(ppm in drinking
water)
Administered
dose"
(mg/kg-d)
Predicted internal dose metrics1"
AN-AUC in rat
blood
(mg/L)
CEO-AUC in rat
blood
(mg/L)
Incidence of
forestomach
lesions0
Sprague-Dawley rats
(Sources: Quast, 2002; Quast et al, 1980a)
Male
Female
0
35
100
300
0
35
100
300
0
3.4
8.5
21.3
0
4.4
10.8
25.0
0
2.06 x 10"2
5.36 x 10"2
1.46 x 10'1
0
2.37 x 10"2
6.18 x 10'2
1.56 x 10'1
0
1.83 x 10"3
4.36 x 10"3
9.70 x 10'3
0
2.07 x 10"3
4.87 x 10'3
1.01 x 10'2
15/80 (19%)
15/47 (32%)
44/48 (92%)c
45/48 (94%)c
20/80 (25%)
23/48 (48%)c
41/48 (85%)c
47/48 (98%)c
F344ratsd
(Sources: Johannsen and Levinskas, 2002b; Biodynamics, 1980c)
Male
Female
0
1
3
10
30
100
0
1
3
10
30
100
0
0.08
0.25
0.83
2.48
8.37
0
0.12
0.36
1.25
3.65
10.90
0
4.33 x 10"4
1.35 x 10'3
4.52 x 10"3
1.37 x 10'2
4.85 x 10'2
0
5.73 x 10"4
1.72 x 10'3
6.02 x 10"3
1.79 x 10"2
5.63 x 10'2
0
4.06 x 10"5
1.27 x 10'4
4.19 x 10"4
1.23 x 10'3
3.97 x 10'3
0
5.32 x 10"5
1.59 x 10'4
5.49 x 10"4
1.58 x 10"3
4.46 x 10'3
11/159(7%)
3/80 (4%)
18/75 (24%)c
13/80 (16%)c
17/80 (22%)c
9/77 (12%)
4/156 (3%)
2/80 (3%)
16/80 (20%)c
23/74 (3 1%)C
13/80 (16%)c
5/74 (7%)
""Administered doses were averages calculated by the study authors based on animal B W and drinking water
intake.
bThe EPA-modified rat physiologically based pharmacokinetic (PBPK) model of Keddaris et al. (1996) was
employed to predict a rat internal dose (i.e., either AN-AUC or CEO-AUC concentration in blood, where AUC =
area under the curve) resulting from the ingestion of the specified administered dose of AN consumed in six bolus
episodes/d.
Indicates significantly different (at/> < 0.05) from control incidence by Fisher's exact test.
Incidences for F344 rats do not include animals from the 6- and 12-mo sacrifices and were further adjusted to
exclude (from the denominators) rats that died between 0 and 12 mos in the study. Rats dying during this time
period were determined from page 6 of Appendix H and Table 1 in Biodynamics (1980c) and Table 8 in
Johannsen and Levinskas (2002b). Unscheduled deaths between 0 and 12 mos in the study occurred in two female
controls, two males at 3 ppm, three females at 10 ppm, and three males and three females at 100 ppm.
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Table 5-2. Incidences of forestomach lesions (hyperplasia or hyperkeratosis)
in male and female B6C3Fi mice administered AN via gavage for 2 years
Lesion site and type
Dose (mg/kg-d)a
0
2.5
10
20
Males
Forestomach hyperplasia or hyperkeratosis
2/50
(4%)
4/50
(8%)
10/503
(20%)
13/50b
(26%)
Females
Forestomach hyperplasia or hyperkeratosis
2/50
(4%)
2/50
(4%)
5/50
(10%)
8/50a
(16%)
""Significantly elevated above vehicle control as determined by EPA using Fisher's exact test (p < 0.05).
bSignificantly elevated above vehicle control as determined by EPA using Fisher's exact test (p < 0.01).
Source: NTP(2001).
5.1.2.1. PBPKModeling
In deriving candidate RfDs from the rat studies, internal dose metrics generated using
PBPK models developed by EPA (see Section 3.5; Appendix C) based on the rat and human
PBPK models of Kedderis et al. (1996) and Sweeney et al. (2003), respectively, were employed.
The EPA-modified PBPK models included certain realistic features, each of them leading to a
different dose metric. The primary features of interest were estimated daily average internal
concentrations of AN or CEO in blood (area under the curve [AUC] expressed on a 24-hour
basis) and estimates based on continuous versus episodic exposure, with episodic exposure more
realistically reflecting how rats (and humans) actually consume drinking water.
As indicated above, two chemical markers of internal exposure were selected (i.e., the
concentration in blood of the parent compound, AN, and its reactive metabolite, CEO). These
two internal dose metrics were evaluated under an episodic exposure pattern because rats (and
humans) consume drinking water in an episodic manner. In addition, for rats and mice, the
externally administered AN dose was also used for deriving candidate RfDs for purposes of
comparison with the candidate RfDs derived based on internal doses of AN and CEO estimated
from the PBPK model.
For this assessment, internal dose metrics were evaluated for use in cross-species
extrapolation from rats because of the following:
• The rat and human PBPK models incorporated species differences in physiological
processes influencing the disposition of AN and CEO and were developed with rat in
vivo toxicokinetic data, human in vitro metabolic data, and rat-to-human allometric
scaling (Section 3.5; Appendix C; Sweeney et al., 2003; Kedderis et al., 1996). As such,
approaches using these models were expected to provide more accurate bases for
extrapolation of dose-response relationships from rats to humans than approaches based
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on administered animal dose alone.
• For oral exposures, PBPK model predictions of AN or CEO concentrations in blood
appear to provide a more accurate basis for extrapolation than predictions based on
forestomach AN-AUC or forestomach CEO-AUC concentrations because the stomach
compartment in the PBPK model was not calibrated with measurements of AN or CEO in
the stomach epithelium, the site of the forestomach lesions. Thus, more research is
needed before a more physiologically meaningful stomach compartment can be
incorporated into the model. Blood AN concentrations predicted by the rat PBPK model
were fairly close to measured AN concentrations in rats following oral exposure, but the
model consistently predicted higher CEO concentrations in blood than was reported in
studies of orally exposed rats (see Figure 3 in Kedderis et al., 1996).
Given the uncertainties in the PBPK predictions noted above, especially for CEO, both
internal dose metrics (AN-AUC and CEO-AUC in blood) were used in deriving candidate RfDs
based on the rat data, in addition to administered dose. The concentration of AN administered in
the bioassay, in terms of mg/kg-day, was used as input into the EPA-modified rat PBPK model
of Kedderis et al. (1996) in order to predict a rat internal dose (either AN-AUC or CEO-AUC
concentration in blood) resulting from the ingestion of the total daily administered dose of AN
consumed in six bolus episodes per day. The resulting predicted AN-AUC or CEO-AUC
concentrations in rat blood are shown in Table 5-1 for male and female SD and F344 rats.
5.1.2.2. BMD Modeling
The incidences of forestomach lesions observed following 2 years of AN exposure in
male and female SD and F344 rats were modeled using AN and CEO in blood, expressed in
mg/L, as internal dose metrics. In addition, incidences of these same lesions were modeled in
male and female SD and F344 rats, as well as male and female B6C3Fi mice, employing
administered dose. For B6C3Fi mice, administered doses were multiplied by 5/7 to convert
5 day/week gavage exposures to 7 day/week continuous exposures. In all cases, all of the
dichotomous dose-response models available in EPA's BMD Software (BMDS, version 2.0)
were fit to these incidence data and BMDs and the 95% lower confidence limit on the BMD
(BMDLs) were calculated.
A benchmark response (BMR) of 10% extra risk of forestomach lesions was selected as
the response associated with the point of departure (POD) for deriving the RfD in the absence of
information regarding what level of change is considered biologically significant, and also to
facilitate a consistent basis of comparison across endpoints and assessments. It is possible that
the forestomach lesions (hyperplasia and hyperkeratosis) may progress to papilloma and
ultimately to carcinoma (Johannsen and Levinskas, 2002b), but specific data supporting this
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conclusion are not available. For purposes of comparison, BMDL values associated with a BMR
of 5% were also presented.
Because the incidence of forestomach lesions in male and female F344 rats did not
increase monotonically across all administered concentrations, none of the models exhibited
adequate fit to the data. Following a procedure of sequentially dropping the highest dose and
refitting the models, only incidence data from the three lowest concentrations (i.e., 0, 1, and 3
ppm) were ultimately used in dose-response modeling. For the same reason, the incidence data
from the highest dose group in male Sprague-Dawley rats needed to be dropped, thus only the
three lowest concentrations (i.e., 0, 35, and 100 ppm) were ultimately used in analyses.
In most cases, several models fit the data equally well (i.e., exhibited %2 goodness-of-fit
p values >0.1). Of those models exhibiting adequate fit and yielding BMDLs sufficiently close
to one another, the selected model was the one with the lowest Akaike's Information Criterion
(AIC) value, as per the EPA's Benchmark Dose Technical Guidance Document (U.S. EPA,
2000b). The AIC is a measure of the deviance of the model fit that allows for comparison across
models for a particular endpoint. BMDLio and BMDLos estimates were then derived from this
selected model. Appendix B-l provides additional details regarding the BMD modeling results
used in RfD derivation.
For the two internal dose metrics selected (AN and CEO in blood), once the BMDLio and
BMDLos estimates were derived from the selected model(s), these estimates, expressed as rat
internal AUCs (in mg/L), were then input into the EPA-modified human PBPK model of
Sweeney et al. (2003) in order to predict the human equivalent administered dose of AN that
would result in a human 24-hour blood AN-AUC or CEO-AUC equivalent to the corresponding
rat AUC, again assuming six bolus ingestion episodes per day. The resulting predicted 95%
lower bounds on the human equivalent administered dose of AN, expressed in mg/kg-day,
represent potential PODs for deriving candidate RfDs. In the case of administered dose, the
BMDLio and BMDLos estimates are already expressed as human equivalent doses (HEDs) in
mg/kg-day, and are thus used directly as potential PODs for deriving candidate RfDs. The
BMDLio and BMDLos estimates and their corresponding PODs across the three dose metrics are
presented in Tables 5-3, 5-4, and 5-5 for Sprague-Dawley rats, F344 rats, and B6C3Fi mice,
respectively.
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Table 5-3. Candidate RfDs based on BMD modeling of the incidence of forestomach lesions (hyperplasia or
hyperkeratosis) in male and female Sprague-Dawley rats exposed to AN in drinking water for 2 years
Sex
Male
Female
Endpoint
Forestomach
lesions
Forestomach
lesions
Dose metric
Administered
dose
(mg/kg-d)
Predicted AN in
blood
(mg/L)
Predicted CEO
in blood
(mg/L)
Administered
dose
(mg/kg-d)
Predicted AN in
blood
(mg/L)
Predicted CEO
in blood
(mg/L)
BMRa
5%
10%
5%
10%
5%
10%
5%
10%
5%
10%
5%
10%
BMDLb
0.72 mg/kg-d
1.27 mg/kg-d
4.32 x 10"3 mg/L
7.72 x 10'3 mg/L
3.90x 10"4mg/L
6.82 x 10'4 mg/L
0.64 mg/kg-d
1.24 mg/kg-d
1.76x 10'3mg/L
3.62 x 10"3 mg/L
2.88 x 10'4 mg/L
5.58 x 10"4 mg/L
PODC
(mg/kg-d)
0.72
1.27
1.27
2.17
3.86 x 10"2
6.75 x 10'2
0.64
1.24
5.35 x 10'1
1.07
2.84 x 10'2
5.51 x 10"2
UF
100
100
30
30
30
30
100
100
30
30
30
30
Candidate RfDd
(mg/kg-d)
7.20 x 10"3
1.27 x 10'2
4.23 x 10"2
7.23 x 10'2
1.29 x 10"3
2.25 x 10'3
6.39 x 10"3
1.24 x 10'2
1.78 x 10'2
3.57 x 10"2
9.48 x 10'4
1.84 x 10"3
aBMR refers to the 95% lower confidence limit on the administered or PBPK-predicted internal dose in the rat associated with a 5 or 10% extra risk for the incidence of
forestomach lesions.
bAll dichotomous models in EPA's BMDS (version 2.0) were fit to the incidence of forestomach lesions (hyperplasia or hyperkeratosis) in Sprague-Dawley rats using the
data presented in Table 5-1. For BMD modeling, three different dose metrics were employed: (1) administered animal dose expressed in mg/kg-d, (2) AN in blood
(predicted) expressed in mg/L, and (3) CEO in blood (predicted) expressed in mg/L. Adequate fit of a model was achieved if the %2 goodness-of-fit statistic yielded ap-
value >0.1. Of those models exhibiting adequate fit and yielding BMDLs that were sufficiently close, the selected model was the model with the lowest AIC value, as
per the EPA's Benchmark Dose Technical Guidance Document (U.S. EPA, 2000b). BMDL10 and BMDL05 estimates were derived from the selected model. Appendix
B-l provides additional details regarding these BMD modeling results.
Tor administered dose, the POD is the BMDL05 or BMDL10 based on BMD modeling using administered animal dose as the dose metric. For the internal dose metrics
(AN and CEO in blood), the PODs are PBPK model-derived human equivalent administered doses of AN that would result in a human 24-hr blood AN-AUC or CEO-
AUC equivalent to the corresponding rat BMDL05 or BMDL10 values, assuming AN ingestion in six bolus episodes/d.
dRfD = POD/UF.
Sources: Quast (2002); Quast et al. (1980a).
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Table 5-4. Candidate RfDs based on BMD modeling of the incidence of forestomach lesions (hyperplasia or
hyperkeratosis) in male and female F344 rats exposed to AN in drinking water for 2 years
Sex
Male
Female
Endpoint
Forestomach
lesions
Forestomach
lesions
Dose metric
Administered
dose (mg/kg-d)
Predicted AN in
blood (mg/L)
Predicted CEO
in blood (mg/L)
Administered
dose (mg/kg-d)
Predicted AN in
blood (mg/L)
Predicted CEO
in blood (mg/L)
BMRa
5%
10%
5%
10%
5%
10%
5%
10%
5%
10%
5%
10%
BMDLb
9.97 x 10"2 mg/kg-d
0.151 mg/kg-d
5.39 x 10"4 mg/L
8.14x 10'4mg/L
5.06 x 10"5 mg/L
7.65 x 10'5 mg/L
0.1 17 mg/kg-d
0.209 mg/kg-d
5.58 x 10"4 mg/L
9.97 x 10'4 mg/L
5.17x 10"5mg/L
9.23 x 10'5 mg/L
PODC
(mg/kg-d)
9.97 x 10"2
0.151
1.66 x 10"1
2.50 x 10'1
5.00 x 10"3
7.56 x 10'3
0.117
0.209
1.72 x 10"1
3.06 x 10'1
5.11 x 10"3
9.12x 10'3
UF
100
100
30
30
30
30
100
100
30
30
30
30
Candidate RfDd
(mg/kg-d)
9.97 x 10"4
1.51 x 10'3
5.55 x 10"3
8.35 x 10'3
1.67 x 10"4
2.52 x 10'4
1.17x 10"3
2.09 x 10'3
5.74 x 10"3
1.02 x 10'2
1.70 x 10"4
3.04 x 10'4
aBMR refers to the 95% lower confidence limit on the administered or PBPK-predicted internal dose in the rat associated with a 5 or 10% extra risk for the incidence of
forestomach lesions.
bAll dichotomous models in EPA's BMDS (version 2.0) were fit to the incidence of forestomach lesions (hyperplasia or hyperkeratosis) in F344 rats using the data
presented in Table 5-1. For BMD modeling, three different dose metrics were employed: (1) administered animal dose expressed in mg/kg-d, (2) AN in blood
(predicted) expressed in mg/L, and (3) CEO in blood (predicted) expressed in mg/L. Adequate fit of a model was achieved if the %2 goodness-of-fit statistic yielded a
p-value >0.1. Of those models exhibiting adequate fit and yielding BMDLs that were sufficiently close, the selected model was the model with the lowest AIC value, as
per the EPA's Benchmark Dose Technical Guidance Document (U.S. EPA, 2000b). BMDL10 and BMDL05 estimates were derived from the selected model. Appendix
B-l provides additional details regarding these BMD modeling results.
Tor administered dose, the POD is the BMDL05 or BMDL10 based on BMD modeling using administered animal dose as the dose metric. For the internal dose metrics
(AN and CEO in blood), the PODs are PBPK model-derived human equivalent administered doses of AN that would result in a human 24-hr blood AN-AUC or CEO-
AUC equivalent to the corresponding rat BMDL05 or BMDL10 values, assuming AN ingestion in six bolus episodes/d.
dRfD = POD/UF.
Sources: Johannsen and Levinskas (2002b); Biodynamics (1980c).
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Table 5-5. Candidate RfDs based on BMD modeling of the incidence of forestomach lesions (hyperplasia or
hyperkeratosis) in male and female B6C3Fi mice exposed to AN via gavage for 2 years
Sex
Male
Female
Endpoint
Forestomach
lesions
Forestomach
lesions
Dose metric
Administered
dose
(mg/kg-d)
Administered
dose
(mg/kg-d)
BMRa
5%
10%
5%
10%
BMDLb
(mg/kg-d)
1.43
3.01
3.02
6.20
PODC
(mg/kg-d)
1.43
3.01
3.02
6.20
UF
100
100
100
100
Candidate RfDd
(mg/kg-d)
1.43 x 10'2
3.01 x 10"2
3.02 x 10"2
6.20 x 10'2
aBMR refers to the 95% lower confidence limit on the administered dose in the mouse associated with either a 5 or 10% extra risk for the incidence of forestomach
lesions.
bAll dichotomous models in EPA's BMDS (version 2.0) were fit to the incidence of forestomach lesions (hyperplasia or hyperkeratosis) in B6C3FJ mice using the data
presented in Table 5-2. For BMD modeling, administered animal dose, expressed in mg/kg-d, was employed. Adequate fit of a model was achieved if the %2 goodness-
of-fit statistic yielded a^-value >0.1. Of those models exhibiting adequate fit and yielding BMDLs that were sufficiently close, the selected model was the model with
the lowest AIC value, as per the EPA's Benchmark Dose Technical Guidance Document (U.S. EPA, 2000b). BMDL10 and BMDL05 estimates were derived from the
selected model. Appendix B-l provides additional details regarding these BMD modeling results.
°The POD is the BMDL05 or BMDL10 based on BMD modeling using administered animal dose as the dose metric. No internal dose metrics were employed for mice
because of the absence of a PBPK model for this species.
dRfD = POD/UF.
Source: NTP(2001).
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5.1.3. RfD Derivation—Including Application of Uncertainty Factors (UFs)
The following uncertainty factors were applied to the candidate PODs to derive candidate
RfD values.
Animal to human extrapolation. In Tables 5-3 and 5-4, the candidate PODs based on the
internal dose metrics (i.e., CEO-AUC in blood and AN-AUC in blood) were divided by an UF of
3 (i.e., 10°5) to account for uncertainty associated with extrapolating from rats to humans. A
factor of 3 was chosen instead of a factor of 10 because toxicokinetic differences were largely
addressed by the application of the rat and human PBPK models. The applied factor of 3 was
selected to account for any remaining toxicokinetic uncertainties in the dosimetric extrapolation,
and possible differences in the response of rat and human target tissues to AN or its metabolites
(i.e., toxicodynamic differences).
The candidate PODs based on administered dose in Tables 5-3, 5-4, and 5-5 were divided
by an UF of 10 to account for uncertainty associated with extrapolating from rodents to humans
because information was unavailable to quantitatively assess toxicokinetic or toxicodynamic
differences between animals and humans.
Human variation. An UF of 10 was used to account for potentially sensitive human
subpopulations in the absence of information on the variability of response to AN in the human
population. Information is unavailable to assess human-to-human variability in AN
toxicokinetics and toxicodynamics.
Subchronic to chronic extrapolation. An UF to account for extrapolation from
subchronic to chronic exposure was not necessary because data from chronic oral studies in rats
and mice were used to derive the candidate PODs.
LOAEL to NOAEL extrapolation. An UF to account for LOAEL to NO AEL
extrapolation was not applied because the current approach is to address this extrapolation as one
of the considerations in selecting a BMR for BMD modeling. In this case, a BMR of a 10%
extra risk of forestomach lesions was selected under an assumption that it represents a minimal
biologically significant change.
Database deficiencies. A database deficiency UF was not used in the development of the
RfD. Although studies of health effects in humans exposed to AN by the oral route are not
available, the animal oral toxicity database is particularly robust. As discussed in Section 4.6.1,
there are nine rat toxicity and cancer bioassays, one toxicity and cancer bioassay with B6C3Fi
mice, a three-generation (46-week) developmental/reproductive toxicity study with Sprague-
Dawley rats, a 12-week gavage study of nerve conduction velocities in male Sprague-Dawley
rats, a 14-week gavage toxicity bioassay in B6C3Fi mice, a developmental toxicity study in
Sprague-Dawley rats exposed by gavage during GDs 6-15, and a 90-day study of oxidative
stress indicators in the brain and liver of F344 rats. These animal data identify forestomach
lesions as the most sensitive, prevalent, and consistent noncancer effect in animals associated
with repeated oral exposure to AN. In addition, the data for forestomach lesions in rats and mice
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provide an adequate characterization of the dose-response relationship for the development of
these lesions with chronic exposure, which, coupled with the application of the PBPK models for
cross-species dosimetric extrapolation, adds to the confidence in the estimation of a chronic oral
RfD.
For the purpose of this assessment, CEO-AUC in blood is considered to be the most
appropriate dose metric to use in deriving an RfD. CEO is believed to be the reactive metabolite
most likely responsible for the noncancer (and cancer) effects observed following AN exposure.
Results from acute exposure studies in rats indicate that CEO plays a key role in the mode of
action by which AN elicits stomach lesions. Research has shown that pretreatment with
inhibitors of CYP2E1 inhibited the development of GI ulceration, following acute exposure of
rats to AN, and doses of KCN equivalent to toxic doses of AN did not induce GI bleeding in a
similar manner to AN (Ghanayem et al., 1985; Ghanayem and Ahmed, 1983). For oral
exposures, then, the use of CEO-AUC as the internal dose metric for cross-species extrapolation
is recommended. Therefore, while the PBPK model does not predict CEO in blood as accurately
as it predicts AN in blood, CEO in blood provides the most biologically relevant dose metric for
modeling the incidence of forestomach lesions and for extrapolating from orally exposed rats to
humans.
In comparing the candidate RfDs for the preferred dose metric of predicted CEO in blood
across Tables 5-3 and 5-4, the candidate RfDs based on the forestomach lesion incidence data in
F344 rats (Johannsen and Levinskas, 2002b; Biodynamics, 1980c) are about an order of
magnitude lower than those candidate RfDs based on forestomach lesion incidence data in
Sprague-Dawley rats (Quast, 2002; Quast et al., 1980a). This comparison indicates that the F344
rats are more sensitive to AN exposure than Sprague-Dawley rats. Moreover, these same
candidate RfDs in F344 rats are approximately two orders of magnitude lower than the candidate
RfDs based on administered dose in B6C3Fi mice (NTP, 2001). Therefore, for the critical
endpoint (i.e., forestomach lesions), the F344 rat is the most sensitive species and strain.
Consequently, the RfD based on data from Johannsen and Levinskas (2002b) in male and female
F344 rats is 3 x 10"4 mg/kg-day (i.e., 2.52 x 10"4 and 3.04 x 10"4 mg/kg-day, respectively).
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5.1.4. Data Array for Oral Noncancer Endpoints
LOAELs based on selected animal studies presented in Table 4-58 are arrayed for
comparison in Figure 5-1, and provide a perspective on the RfD developed in the previous
section from data in F344 rats (Johannsen and Levinskas, 2002b). Figure 5-1 should be
interpreted with caution, however, because the LOAELs across studies are not necessarily
comparable due to the lack of any indication regarding the confidence in the data sets from
which the LOAELs were derived. In addition, the nature, severity, and incidence of effects at a
LOAEL are also likely to vary. For example, the incidence of forestomach squamous cell
hyperplasia in male and female F344 rats at the LOAEL were both 17%, while in Sprague-
Dawley rats, the incidences of the same lesion at the LOAEL in males and females were 73 and
23%, respectively.
The predominant noncancer effect of chronic oral exposure to AN is hyperplasia and
hyperkeratosis of squamous cell epithelial tissue in the forestomach. The LOAELs for this
endpoint are lower than those observed for chronic nephropathy, gliosis, ovarian cysts, or
developmental effects. Therefore, the RfD based on hyperplasia and hyperkeratosis of squamous
cell epithelial tissue in the forestomach should be protective of other effects resulting from oral
exposure to AN.
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30 i
25 -
20 -
15 -
0)
(A
O
Q
10 -
5 -
D NOAEL
A LOAEL
Verticle lines
represent the
range of doses
tested in a given
study.
A
A
(a) Johannsen & Levinskas, 2002b
(b) Quast, 2002
(c) Johannsen & Levinskas, 2002a
(d) NTP, 2001
A
A
A
A
A
A
A
Forestomach/ Forestomach/ Forestomach/ Forestomach/ Fores to mac h/ Forestomach/ Forestomach/ Forestomach/ Chronic Chronic
squamous cell squamous cell squamous cell squamous cell squamous cell squamous cell squamous cell squamous cell nephropathy nephropathy
hyperplasia, hyperplasia, hyperplasia, SD hyperplasia, SD hyperplasia, SD hyperplasia, SD hyperplasia, hyperplasia in (severe) in SD (minimal) in SD
F344 male rats F344 female rats male rats (b) female rats (b) male rats (c) female rats (c) B6C3F1 male B6C3F1 female male rats (b) female rats (b)
(a) (a) mice (d) mice (d)
Ovarian cysts in Gliosis in the
B6C3F1 female brain of SD
mice (d) female rats (b)
Figure 5-1. Exposure-response array for noncancer endpoints across target organs following oral exposure to AN in animals.
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5.1.5. Previous RfD Assessment
No RfD was derived in the previous IRIS assessment of AN.
5.2. INHALATION REFERENCE CONCENTRATION (RfC)
5.2.1. Choice of Principal Study and Critical Effect
As discussed in Sections 4.1.2.2.2 and 4.6, results from several studies of AN-exposed
workers (Chen et al., 2000; Kaneko and Omae, 1992; Muto et al., 1992; Sakurai et al., 1978) and
a study of the performance on a battery of neurobehavioral tests by exposed workers (Lu et al.,
2005a) identified increased prevalence of subjective neurological symptoms (e.g., headache,
poor memory, and irritability) and small performance deficits in neurobehavioral tests as the
critical effects from chronic occupational inhalation exposure to AN. An increased prevalence of
subjective symptoms was associated with average workplace air concentrations of 1.13 ppm
(Muto et al., 1992), 1.8 ppm (Kaneko and Omae, 1992), and 0.48 ppm (Chen et al., 2000).
Deficits in the neurobehavioral tests were associated with average workplace air concentrations
of 0.11 ppm for a group of workers designated as monomer workers and 0.91 ppm for a group of
workers designated as fiber workers (Lu et al., 2005a).
Cross-sectional epidemiologic surveys of reproductive outcomes in AN-exposed workers
found increased prevalence of adverse reproductive outcomes associated with somewhat higher
average workplace air concentrations of 3.6 ppm (Dong et al., 2000a) and 7.5 ppm (Li, 2000).
Adverse outcomes with statistically significantly increased prevalence compared with unexposed
workers included the following:
• Premature deliveries—8.2% in exposed females vs. 3.9% in controls (Dong et al.,
2000b); and 11.6% in exposed females vs. 4.7% in controls (Li, 2000)
• Stillbirths—2.7% in exposed females vs. 1.1% in controls (Dong et al., 2000b)
• Sterility— 2.6% in exposed females vs. 0.8% in controls (Li, 2000)
• Birth defects—21.3 per 1000 live births in exposed females vs. 4.8 per 1000 live births
in controls (Dong et al., 2000b); 25.4 per 1000 live births in exposed females vs. 4.2 per
1000 live births in controls (Li, 2000)
• Pregnancy complications—20.8% in exposed females vs. 7.1% in controls (Li, 2000)
Exposure levels associated with adverse effects in the available animal inhalation toxicity
studies were higher than the workplace air concentrations associated with adverse effects in AN-
exposed workers (see Tables 4-59 and 4-60). With repeated inhalation exposure to AN, effects
noted with the lowest exposure levels were as follows:
• Increased incidence (58% in males, 61% in females) of nasal epithelial lesions in
Sprague-Dawley rats exposed for 2 years to 20 ppm AN (Quast et al., 1980b)
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• Deficits in sensory nerve conduction in the tail nerve of Sprague-Dawley rats exposed for
12 weeks to AN concentrations >25 ppm, starting at 5% at 25 ppm (Gagnaire et al., 1998)
• Increased incidence of litters with any malformation (short tail, short trunk, missing
vertebrae, or anteriorly displaced ovaries) in Sprague-Dawley rats exposed to AN
concentrations of 80 ppm (by 15%) on GDs 6-15 (Murray et al., 1978)
• Decreased fetal weight gain (5%) per litter in Sprague-Dawley rats exposed to AN
concentrations >25 ppm on GDs 6-20 (Saillenfait et al., 1993).
The cross-sectional study of neurobehavioral performance in acrylic fiber workers by Lu
et al. (2005a) was selected as the principal study for deriving the RfC because it is the best
available study that identified neurobehavioral effects in workers exposed to AN. Previous
occupational studies by Kaneko and Omae (1992) and Muto et al. (1992) reported subjective
neurological symptoms (e.g., poor memory and irritability) in exposed workers. Lu et al.
(2005a) utilized the WHO-recommended NCTB administered by trained physicians to evaluate
these neurobehavioral effects systematically. Hence, the results were more reliable when
compared with those based on self reporting. In addition, neurobehavioral effects were also
reported in AN-treated rats (Rongzhu et al., 2007; Ghanayem et al. 1991). Confounding by other
workplace exposures is not considered likely.
The distribution of employment duration for the monomer workers was 1-10 years, 23%;
11-20 years, 42%; and >20 years, 35%. For fiber workers, the distribution was 1-10 years,
47%; 11-20 years, 23%; and >20 years, 30%. Geometric mean workplace AN air concentrations
were 0.11 ppm for the monomer operations areas (range 0-1.70 ppm based on 390 stationary air
samples collected between 1997 and 1999) and 0.91 ppm for the acrylic fiber operations areas
(range 0.00-8.34 ppm based on 570 samples). For monomer workers, the following statistically
significant deficits, compared with unexposed controls, were measured:
• 41-68% higher scores for negative moods (i.e., anger, confusion, depression, fatigue, and
tension) in the Profile of Mood States Test.
• 16% longer times in the Simple Reaction Time Test of attention and visual response
speed.
• 21% lower scores in the backward sequence of the Digit Span Test of auditory memory.
• 4% lower scores in the Benton Visual Retention Test, a measure of visual perception and
memory.
• 14% lower scores in the Pursuit Aiming II Test, a measure of motor steadiness.
Statistically significant deficits were not found in the Santa Ana Test for manual dexterity
or in the Digital Symbol Test for perceptual motor speed. Fiber workers showed deficits of a
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similar magnitude in many of the same tests (20-44% higher in Profile of Mood States Test,
10% longer in the Simple Reaction Time Test, 24% lower score in the backward sequence of the
Digit Span Test, 4% lower scores in the Benton Visual Retention Test, and 10% lower scores in
the Pursuit Aiming II Test), with the exception that scores in the forward sequence of the Digit
Span Test were significantly better than those of unexposed workers.
Lu et al. (2005a) reported that air in these workplaces also presented potential exposures
to cyanide in the monomer operations areas and methyl methacrylate in the fiber areas, but
measurements of these chemicals in air samples were not made. According to Dr. Lu (email
from Dr. Rongzhu Lu, Department of Preventive Medicine, College of Medicine, Jiangsu
University, China, to Dr. Diana Wong, U.S. EPA, dated 5/15/2008), cyanide is one of the
byproducts in the production of AN by oxidation of ammonia and propylene, and this byproduct
is recycled as raw material to produce sodium cyanide. Because its concentration is too low to
be detected, there are no workplace monitoring data. Therefore, the concentration of cyanide in
the monomer plant should be low compared with AN. Methyl methacrylate was generally used
as a minor second monomer in the production of acrylic fiber. According to Dr. Lu (email from
Dr. Rongzhu Lu, Department of Preventive Medicine, College of Medicine, Jiangsu University,
China, to Dr. Diana Wong, U.S. EPA, dated 5/20/2008), the ratio of AN to methyl methacrylate
to methylene succinic acid (third monomer) should be approximately 90-94 to 5-8 to 0.3-2.
Since cyanide and methyl methacrylate occurred only at trace levels, if at all, they were not
considered to be confounding exposures. Therefore, potential exposure to methyl methacrylate
and cyanide was determined not to be a significant limitation in using Lu et al. (2005a) to
identify neurological effects as a potential health hazard from occupational exposure to AN and
derive an RfC for chronic inhalation exposure to AN. Both groups of workers showed deficits,
and the results from this study are consistent with the increased prevalence of subjective
symptoms in other studies of AN-exposed workers.
An RfC based on the results from the chronic inhalation bioassay with Sprague-Dawley
rats (Quast et al., 1980b) was also derived for comparison purposes. As discussed in
Section 4.6.2, statistically significant increased incidence of inflammatory and degenerative
nasal lesions (i.e., hyperplasia of mucus-secreting cells in males and flattening of respiratory
epithelium in females) occurred in rats exposed to the lowest level of AN in this two-year
bioassay, 20 ppm (6 hours/day, 5 days/week), and represent the critical effects in animals
exposed to AN chronically by inhalation. At the higher exposure level, 80 ppm, other nasal
lesions with elevated incidences were suppurative rhinitis and focal erosion of the mucous lining
in females and hyperplasia of respiratory epithelium in males and females. Other lesions with
elevated incidences at the 80-ppm exposure level were gliosis and perivascular cuffing in the
brain of males and females, focal nephrosis and thyroid cysts in males, and hepatic necrosis in
females (which was also elevated at 20 ppm).
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5.2.2. Methods of Analysis
A NOAEL/LOAEL approach to the human data was used to derive the RfC. The
performance deficits measured in monomer and fiber workers were judged to be adverse and the
geometric mean of the range of air concentrations measured for the monomer work areas,
0.11 ppm (0.24 mg/m3), identified as the LOAEL, was selected as the POD for deriving the RfC.
For the animal data, a BMD approach was used. Dose-response models available in
EPA's BMDS (version 1.3.2) were fit to incidence data for flattening of the respiratory
epithelium in female rats and hyperplasia of mucus-secreting cells in male rats. Prior to
modeling, animal exposure data were converted to human equivalent concentrations (HECs)
using U.S. EPA (1994) methods for extrathoracic respiratory effects from a category 1 gas
(Table 5-6). A BMR of 10% extra risk was selected as the POD for deriving the RfC based on
both biological and statistical considerations. Biologically, the endpoints selected on which to
derive the RfC (i.e., flattening of the respiratory epithelium in female rats and hyperplasia of
mucus-secreting cells in male rats) are relatively benign. A BMR of 10% extra risk was selected
under the assumption that it represents a minimal biologically significant change. A BMR of
10% was also selected to facilitate a consistent basis of comparison across endpoints and
assessments. Benchmark concentrations (BMCs) and the 95% lower bounds on the BMCs
(BMCLios) for the best-fitting models are shown in Table 5-6. Potential PODs for the animal-
based RfC are the BMCLios of 0.082 and 0.059 mg/m3 for nasal effects in male and female rats,
respectively. More detailed information on these BMD modeling results is presented in
Appendix B-2.
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Table 5-6. Results of dose-response analyses of incidence data for selected
nasal lesions in male and female Sprague-Dawley rats exposed by inhalation
to AN for 2 years
Nasal lesion
Hyperplasia of mucus-
secreting cells in males
Flattening of respiratory
epithelium in females
Administered
concentration
(mg/m3)
0
43.4
173.6
0
43.4
173.6
HEC
(rng/rn3)3
0
2.1
8.5
0
2.1
8.5
Lesion
incidence
0/11(0%)
7/12 (58%)d
8/10 (80%)d
1/11(9.1%)
7/10 (70%)d
8/10 (80%)d
BMC10
(mg/m3)b
0.187
0.162
BMCL10
(mg/m3)b
0.082
0.059
Candidate
RfC
(mg/m3)c
3 x 10'3
2 x 10'3
aHEC as per U.S. EPA (1994) methods for a category 1 gas producing an upper respiratory effect.
Sample calculation: 43.4 mg/m3 x 6h/24h x 5d/7d x RGDRET = 2.1 mg/m3, where RGDRET = 0.275 = [VE/SAET] rat
-^ [VE/SAET] human; VE = minute volume = 0.281 L/min rat, 13.8 L/min human; SAET = extrathoracic surface area =
15 cm2 rat, 200 cm2 human.
^MCio and BMCL10 refer to the BMD model-predicted air concentration and its 95% lower confidence limit,
associated with a 10% extra risk for having nonneoplastic nasal lesions. BMCi0s and BMCL10s are estimated from
the best-fitting model among those fit to the data. More detailed information on the BMD modeling results is
presented in Appendix B-2.
cRfC = BMCL10/UF, where the UF is 30.
dStatistically significantly different from control value as reported by Quast et al., 1980b.
Source: Quast et al. (1980b).
5.2.3. RfC Derivation—Including Application of Uncertainty Factors (UFs)
The LOAEL of 0.11 ppm (0.24 mg/m3) from Lu et al. (2005a) for statistically significant
performance deficits in neurobehavioral tests of mood, attention and speed, auditory memory,
visual perception and memory, and motor steadiness in humans occupationally exposed to AN
via inhalation was used for derivation of the RfC. Since the LOAEL was from an occupational
study, the adjusted LOAEL for continuous exposure was obtained by multiplying the study
LOAEL by a factor of 0.36 (5 days/7 days x 10 m3/day - 20 m3/day). This adjusted LOAEL
(LOAELADj) of 0.086 mg/m3 was divided by a composite UF of 100 to derive an RfC of 9 x 10"4
mg/m3. This two-step calculation is illustrated below:
1. LOAELADJ = 0.24 mg/m3 x 0.36 = 0.086 mg/m3
2. RfC = 0.086 mg/m3 - 100 = 0.00086 mg/m3 or 0.9 ug/m3
The following UFs were applied in the calculation of the RfC: 10 for consideration of
intraspecies (human) variability, and 10 for extrapolation from a LOAEL to NOAEL. The
composite UF = 10 x 10 = 100.
Animal to human extrapolation. An UF for animal to human extrapolation was not
applied because the RfC was derived from data collected from humans.
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Human variation. An UF of 10 was used to account for potentially sensitive human
subpopulations in the absence of information on the variability of response to AN in the human
population. Information is unavailable to assess human-to-human variability in AN
toxicokinetics and toxicodynamics.
Subchronic to chronic extrapolation. An UF for study duration was not applied because
the principal study was of chronic duration.
LOAEL to NOAEL extrapolation. An UF of 10 was applied to extrapolate from a
LOAEL to a NOAEL because the POD for derivation of the RfC was a LOAEL.
Database deficiencies. A database UF was not applied because the database for AN is
robust. The inhalation database includes eight occupational exposure studies that evaluated the
noncancer health effects of AN on workers exposed via inhalation. Three of these studies
evaluated reproductive endpoints in AN-exposed workers. The database also includes one
chronic inhalation toxicity study in male and female Sprague-Dawley rats; one two-generation
reproductive toxicity study of inhaled AN vapors in Crl:CD(SD) rats (Nemec et al., 2008); one
24-week nerve conduction velocity study in male rats; and 2 developmental studies in rats
exposed from GD 6-15 or GD 6-20.
Comparative animal-based RfCs of 3 x 10"3 mg/m3 (or 3 ug/m3) and 2 x 10"3 mg/m3 (or
2 ug/m3) were derived by dividing the BMDLioS of 0.082 mg/m3 and 0.059 mg/m3 for nasal
lesions in male and female rats, respectively, by UFs of 30 (3 for extrapolating from rats to
humans using the default U.S. EPA [1994] dosimetric adjustment and 10 to protect sensitive
human subpopulations). These candidate RfCs are displayed in Table 5-6. Extrapolating the
results from animal toxicity studies to derive an RfC has inherently greater uncertainty than
using the results from the cross-sectional studies of health effects in human workers; however,
the human-based and animal-based RfCs differ by about twofold.
For this assessment, the human-based RfC of 9 x 10"4 mg/m3 or 0.9 ug/m3 is the
recommended reference value.
5.2.4. Data Array for Inhalation Noncancer Endpoints
LOAELs based on selected studies in human workers included in Table 4-59 are
summarized in Table 5-7 and provide perspective on the RfC derived from Lu et al. (2005a).
The LOAELs in Table 5-7 should be interpreted with caution because the LOAELs across
studies are not necessarily comparable, due to inherent limitations in NOAEL/LOAEL
determinations (U.S. EPA, 2000b), nor is the confidence in the data sets from which the
LOAELs were derived the same. The nature, severity, and incidence of effects occurring at a
LOAEL are likely to vary. The text in Sections 4.1.2.2.2 and 5.2.1 should be consulted for a
more complete understanding of the issues associated with each data set and the rationale for the
selection of the critical effect and principal study used to derive the RfC. The most sensitive
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endpoint is the neurobehavioral effects identified in Lu et al. (2005a) and this endpoint provides
the basis for the derivation of the RfC.
Table 5-7. Comparison of LOAELs for noncancer effects in human workers
following inhalation exposure to AN
Reference
Kaneko and Omae, 1992
Mutoetal., 1992
Chen et al., 2000
Lu et al., 2005a
Xu et al., 2003
Dongetal.,2000b
Li, 2000
Endpoint
Subjective symptoms in males
Subjective symptoms in males
Subjective symptoms in males and females
Neurobehavioral deficits in males and females
Decrease in sperm density and number and sex
chromosome aneuploidy in males
Adverse reproductive outcomes in males and
females
Adverse reproductive outcomes in females
LOAEL (ppm)
1.8
1.1
0.5
0.1
0.4
3.6
7.5
5.2.5. Previous RfC Assessment
Previously, EPA derived an RfC of 2 x 10~3 mg/m3 from the low exposure level (i.e.,
20 ppm AN, duration-adjusted to a LOAELHEC of 1.9 mg/m3) in the Quast et al. (1980b) animal
study, which was identified as a LOAEL for the onset of hyperplasia of the mucus-secreting
cells. The LOAELnEc was divided by a combined UF of 1,000 to derive the RfC. This
composite UF was made up of 10 for intraspecies variability; 3 for interspecies variability, where
dosimetric adjustments had already been applied to account for the toxicokinetic component of
this area of uncertainty; 3 for extrapolation from a minimally adverse LOAEL to a NOAEL; and
10 for database deficiencies. The latter UF was applied because of the lack of an inhalation
bioassay in a second species and the absence of reproductive data by the inhalation route where
an oral study existed that showed reproductive effects.
New pertinent information, available since the previous RfC was developed, includes:
(1) several cross-sectional health examinations and surveys of subjective symptoms and
reproductive outcomes in AN-exposed workers; (2) a published cross-sectional study of
performance in a battery of neurobehavioral tests by AN-exposed workers (Lu et al., 2005a) (the
principal study for the current RfC); (3) toxicokinetic information and the development of PBPK
models for AN; and (4) two inhalation developmental toxicity studies in rats.
5.3. UNCERTAINTIES IN THE ORAL REFERENCE DOSE AND INHALATION
REFERENCE CONCENTRATION
The following discussion identifies uncertainties associated with the RfD or RfC for AN.
As presented earlier in Sections 5.1.2 and 5.1.3 for the RfD and Sections 5.2.2 and 5.2.3 for the
RfC, the UF approach, following EPA methodology for RfC and RfD development (U.S. EPA,
2002, 1994), was applied to a POD. For the RfD, the POD was determined as the BMDL05
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(using CEO concentration in blood as the dose metric) estimated from rats and subsequently
converted to a HED. For the RfC, the POD (a LOAEL) was derived from an epidemiologic
study that examined neurobehavioral effects. The POD divided by a set of factors to account for
uncertainties including extrapolation from responses observed in animal bioassays to humans,
extrapolation from a LOAEL to an estimate of the NOAEL, to account for a diverse population
of varying susceptibilities, and to account for database deficiencies. These extrapolations are
carried out with assumptions instead of data on AN, given the limitations in experimental AN
data to inform individual steps.
Selection of principal study and critical effect for reference value determination
Hyperplasia and hyperkeratosis of squamous epithelium of the forestomach was selected
as the critical effect for the RfD. This effect is the most prevalent, consistent, and most sensitive
effect in rats and mice. Since GI bleeding occurs after s.c. injection of AN in rats, this effect is
probably not due to local irritation on gastric tissues, but is likely due to binding of CEO to GI
mucosa. Although humans do not possess a forestomach, humans do have comparable
squamous cell epithelial tissues in their oral cavity and the upper two-thirds of their esophagus
(IARC, 1999). Thus, there is little uncertainty that this effect is relevant to humans.
For derivation of the RfC, both a 2-year rat inhalation study and epidemiologic studies
demonstrating neurological effects are available. To reduce uncertainty in extrapolating from
animals to humans, epidemiologic studies are preferred. Neurobehavioral effects of AN-exposed
workers was selected as the critical effect since this effect was observed in several occupational
studies with workers chronically exposed to AN via the inhalation route. The primary limitation
of the selected principal study (Lu et al., 2005a) identified by the study authors was the extent of
exposure data, with exposure measures based on area sampling during 1997 to 1999; no
contemporaneous personal monitoring data was available.
Animal to human extrapolation
No human oral exposure studies are available for derivation of the RfD. For derivation of
the RfD, extrapolating dose-response data from animals to humans is a source of uncertainty. A
PBPK model, which has its own associated uncertainties, was used to address toxicokinetic
differences between animals and humans. Uncertainties of the PBPK model are discussed as
part of the overall uncertainty discussion (Section 5.4.4.5), with quantitative details given in
Appendix D. Residual uncertainties pertaining to unknown interspecies differences in
pharmacodynamics were addressed by application of an UF of 3.
A human occupational exposure study (Lu et al., 2005a) was used for derivation of the
RfC, eliminating uncertainty associated with extrapolation from animals to humans. An RfC
was also derived from a two-year inhalation study of rats based on increased incidence of
322 DRAFT- DO NOT CITE OR QUOTE
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inflammatory and degenerative nasal lesions. This alternative RfC was approximately twofold
higher than the RfC derived from Lu et al. (2005a).
Dose-response modeling
BMD modeling was used to estimate the POD for the RfD. While models with better
biological support may exist, the selected models provided adequate mathematical fits to the
experimental data sets. BMD modeling has advantages over a POD based on a NOAEL or
LOAEL because NOAELs/LOAELs are a reflection of the particular exposure concentration or
dose at which a study was conducted, they do not make use of the dose-response curve, and they
do not address the variability of the study population. NOAELs and LOAELs also are less
amenable to quantitative uncertainty analysis.
The RfC was based on a LOAEL identified from an occupational epidemiology study.
As stated above, there are several reasons to prefer a POD obtained from BMD modeling.
However, the available data only supported a LOAEL. An UF to address LOAEL to NOAEL
extrapolation was applied.
Intrahuman variability
Heterogeneity among humans is another source of uncertainty. Uncertainty related to
human variation needs consideration, also, in extrapolation from a small subset of presumably
healthy humans (i.e., workers) to a larger, more diverse general population. Available data from
animal studies provide no evidence of gender differences in susceptibility to toxicity of AN,
although no data are available regarding possible gender differences in susceptibility. Human
genetic polymorphisms in CYP2E1 activities likely contribute to variability in human
susceptibility to the toxic effects of AN (see Section 4.8.4.1). An UF of 10 was used to account
for human variability for derivation of the RfD and RfC. A factor of 10 has been found to be
generally sufficient to account for human variability in response to chemical exposure (Renwick
and Lazarus, 1998).
5.4. CANCER ASSESSMENT
5.4.1. Choice of Study/Data—with Rationale and Justification
As previously discussed in Section 4.1.2.2, evidence of a possible association between
exposure to AN and cancer in humans has been found in some studies. The best available
occupational epidemiologic study (Blair et al., 1998) reported that workers exposed to AN via
inhalation in the highest cumulative exposure category (i.e., >8 ppm-years) with more than
20 years of employment displayed a twofold increased risk of death from lung cancer compared
with unexposed workers (RR = 2.1, 95% CI = 1.2-3.8). To date, Blair et al. (1998) is the largest
cohort study to assess the relationship between AN and cancer, following 25,460 workers in
eight AN-producing facilities, with two-thirds of this cohort having a follow-up period of over
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20 years. The large sample size and the long follow-up time in this study provide a good
opportunity to detect any substantial elevation of case-specific cancer deaths. The findings from
Blair et al. (1998) are not inconsistent with other studies evaluating carcinogenic endpoints from
AN exposure. However, Blair et al. (1998) addressed known problems with earlier studies by
quantifying exposures, examining potential confounding from smoking, and employing an
internal control group of unexposed workers. And although only 5% of the cohort had died by
the time of analysis, this study has a large number of observed deaths that can be used to assess
the relationship between AN-exposure and cancer based on mortality. Also, Blair et al. (1998)
utilized a detailed job-exposure matrix as part of the exposure assessment. For these reasons,
data from Blair et al. (1998) were chosen to derive an inhalation unit risk (IUR) for AN.
Additionally, the use of epidemiology data for the derivation of the IUR, versus using data from
an animal bioassay, reduces uncertainty inherent in animal to human extrapolation.
Sections 4.6.1 and 4.6.2 summarized the current animal data on AN indicating that it is a
multiple-site carcinogen in chronic oral and inhalation bioassays with rats and mice. The best
available animal studies for evaluating the dose-response relationship between AN exposures
and forestomach, CNS, Zymbal gland, tongue, and mammary gland tumors were two chronic
drinking water studies, one with Sprague-Dawley rats exposed to 0, 35, 100, or 300 ppm AN in
drinking water for 2 years (Quast, 2002; Quast et al., 1980a) and the other with F344 rats
exposed to 0, 1,3, 10, 30, or 100 ppm AN in drinking water for 2 years (Johannsen and
Levinskas, 2002b; Biodynamics, 1980c). In addition, data from another bioassay with Sprague-
Dawley rats exposed to AN in drinking water at concentrations of 0, 1, or 100 ppm for 2 years
were also considered (Johannsen and Levinskas, 2002a; Biodynamics, 1980a). In the absence of
human studies demonstrating the carcinogenicity of chronic oral exposure to AN, developing an
oral cancer slope factor (CSF) from these animal studies is reasonable, especially because the
application of the AN PBPK models described in Section 3 can decrease the toxicokinetic
uncertainty in the interspecies extrapolation and in the mode of action of carcinogenicity of AN
in rodents versus humans.
Animal data from the only available chronic inhalation cancer bioassay with multiple
exposure levels (Dow Chemical Co., 1992a; Quast et al., 1980b) were also used to develop an
IUR to compare with the one derived from AN-exposed workers (Blair et al., 1998). In this
animal study, male and female Sprague-Dawley rats were exposed to 0, 20, or 80 ppm AN in air,
6 hours/day, 5 days/week, for 2 years. At 80 ppm, significantly increased incidences of
astrocytomas and glial cell proliferation and Zymbal gland tumors in males and females,
malignant mammary gland tumors (adenocarcinomas) in females, as well as intestinal and tongue
tumors in males, were found. At 20 ppm, male and female rats showed increased incidences of
astrocytomas and glial cell proliferation and Zymbal gland tumors. IUR estimates were derived
based on dose-response data for these tumors using the AN PBPK models previously described to
dosimetrically extrapolate from rats to humans.
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5.4.2. Dose-Response Data
5.4.2.1. Human Occupational Data
An analysis of the lung cancer mortality and AN exposure data from the Blair et al.
(1998) study was conducted to derive an IUR estimate for AN. This analysis used the approach
described by Starr et al. (2004) in which the risk of death from lung cancer in AN-exposed
workers was characterized using a semi-parametric Cox regression model with time-dependent
covariates. The Cox regression model has several advantages in that it allows inclusion of
individual exposure histories and utilizes internal controls, thus avoiding confounding by the
"healthy worker" effect. In contrast to the analysis done by Starr et al. (2004), the analysis
conducted for this assessment included the entire cohort (not just white male workers), and the
final model only had cumulative exposure as a covariate. Starr et al. (2004) included a second
covariate in their final model, plant of employment. This analysis is further described in
Appendix B-7.
5.4.2.2. Rat Oral Data
Incidence data for forestomach, CNS, Zymbal gland, tongue, and mammary gland tumors
in Sprague-Dawley and F344 rats exposed to AN in drinking water for 2 years were used to
develop site-specific oral CSFs for AN (Tables 5-8 through 5-10). Tumor incidences in F344
rats were adjusted to exclude animals dying before the first appearance of each tumor type,
which ranged from day 419 to 495. No early mortality adjustments were made to the cumulative
tumor incidences from the Sprague-Dawley bioassay (Quast, 2002) because CNS and Zymbal
gland tumors were seen in some high-dose female rats as early as 7 months (210 days) after
exposure began, and no differences in survival were observed across the dose groups during the
first 6 months of the study.
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Table 5-8. Incidence of CNS tumors in Sprague-Dawley and F344 rats
exposed to AN in drinking water for 2 years
Sprague-Dawley rats
(Quast, 2002; Quast et al, 1980a)a
Sex
Male
Female
AN drinking water
concentration (ppm)
0
35
100
300
0
35
100
300
Glial cell proliferation1"
0/80
4/47°
3/48
7/48c
0/80
3/48
3/48
7/48c
Astrocytomas
1/80
8/47c
19/48
23/48c
1/80
17/48C
22/48c
24/48c
Overall CNS tumor
1/80
12/47C
22/48c
30/48C
1/80
20/48C
25/48c
31/48C
F344 rats
(Johannsen and Levinskas, 2002b; Biodynamics, 1980c)d
Sex
Male
Female
AN drinking water
concentration (ppm)
0
1
o
J
10
30
100
0
1
3
10
30
100
Glial cell proliferation1"
Brain
0/160
0/80
0/78
0/80
0/79
0/76
0/157
0/80
0/80
0/77
0/80
0/76
Spinal cord
0/156
0/79
0/70
0/78
0/79
0/70
0/155
0/78
0/79
0/72
0/87
0/69
Astrocytomas
Brain
2/160
2/80
1/78
2/80
10/79C
21/76C
1/157
1/80
2/80
4/75
6/80c
23/76c
Spinal cord
1/156
0/79
0/70
0/78
0/79
4/70c
0/155
0/78
0/79
1/72
0/77
1/69
Overall CNS
tumors
3/160
2/80
1/78
2/80
10/79C
25/76c
1/157
1/80
2/80
5/75
6/80c
24/76c
"Incidence denominators were calculated from the total number of animals examined from the beginning of the
study. Numerators for CNS tumor incidences (glial cell proliferation, astrocytomas, or overall incidence of glial
cell proliferation or astrocytomas) for these Sprague-Dawley rats were reported as combined brain and spinal cord
lesions by Quast et al. (1980a).
bGlial cell proliferation is a smaller-sized lesion, either focal or multifocal, than astrocytomas, suggestive of an early
tumor.
Statistically significantly different from controls (p < 0.05) as calculated by the study authors.
dThe denominators for incidences in these F344 rats exclude rats from the 6- and 12-mo sacrifices and unscheduled
deaths prior to the 12-mo sacrifice. Numerators for the overall incidences were the number of rats with astrocytomas
in brain or spinal cord. Reviews of summaries of individual animal pathology reports for this study (Appendix H,
Biodynamics, 1980c) indicated that five of the seven F344 rats showing spinal cord astrocytomas also showed a
brain astrocytoma; thus, the number was not always as great as the sum of the numerators for the incidences of these
lesions in the two tissues (e.g., 21/76 male 100-ppm brain tissues had astrocytomas and 4/70 male 100-ppm spinal
cord tissues had astrocytomas, but all four rats with astrocytomas also had brain astrocytomas). Because the
response to AN was predominately in brain tissue and a few spinal cord tissue samples were missing in each
exposure group, denominators for the overall incidences were taken as the number of rats examined for brain lesions
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Table 5-9. Incidence of mammary gland tumors in F344 and Sprague-
Dawley rats exposed to AN in drinking water for 2 years
Sex
AN drinking water
concentration
(ppm)
Benign mammary gland
tumors"
Malignant
mammary
gland tumorsb
Benign and/or
malignant mammary
gland tumors
F344 rats
(Johannsen and Levinskas, 2002b; Biodynamics, 1980c)°'d
Male
Female
0
1
3
10
30
100
0
1
3
10
30
100
No statistically significantly elevated incidences of mammary gland
tumors were found in exposed groups compared with controls.
12/156
5/80
6/80
8/79
9/80
9/73
3/156
4/80
0/80
1/78
3/80
6/73e
14/156
8/80
6/80
9/80
12/80
14/73e
Sprague-Dawley rats
(Quast, 2002; Quast et al., 1980a)f
Male
Female
0
35
100
300
0
35
100
300
Not
reported
52/80
35/48
33/48
22/48
Not
reported
1/80
1/48
3/48
10/48e
Not
reported
58/80
42/48e
42/48e
35/48
"Incidence includes fibroadenomas for F344 rats and fibroadenomas/adenofibromas/adenomas for Sprague-Dawley
rats.
blncidence includes adenocarcinomas and carcinomas.
°The denominators for tumor incidences in F344 rats excluded rats from the 6- and 12-mo sacrifices and rats that
died prior to 12 mos. Mammary gland tumor incidences are for animals scheduled for the 18-mo and terminal
sacrifices. Microscopic examinations were only conducted on mammary glands showing gross signs of tumors—
the inclusion of all rats living for >52 wks in the denominators assumes that rats without gross signs of tumors were
also without microscopic neoplastic changes.
dAnimals with multiple tumor types within a tissue were counted only once.
Statistically significantly different from controls (atp < 0.05) via Fisher's exact test.
Denominators were calculated from the total number of animals examined from the beginning of the study.
Incidences were reported as total number of rats with benign-only, malignant-only, or benign and/or malignant
tumors.
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Table 5-10. Tumor incidences Sprague-Dawley and F344 in rats exposed to
AN in drinking water for 2 years
Sex
AN drinking
water
concentration
(ppm)
Forestomach
tumors
CNS
tumors
Zymbal gland
tumors
Tongue
tumors
Intestinal
tumors
Mammary gland
tumors
Sprague-Dawley rats
(Quast, 2002; Quast et al, 1980a)
Male
Female
0
35
100
300
0
35
100
300
0/80
2/47
23/48a
39/48a
1/80
1/48
12/483
30/483
1/80
12/47a
22/48a
30/48a
1/80
20/48a
25/48a
31/48a
3/80
4/47
3/48
16/48a
1/80
5/48a
9/48a
18/48a
1/80
2/47
4/48
5/48a
0/80
1/48
2/48
12/48a
Not
reported
0/80
1/48
4/48a
4/48a
Not reported
58/80
42/48a
42/48a
35/48
F344 rats
(Johannsen and Levinskas, 2002b; Biodynamics, 1980c)b
Male
Female
0
1
3
10
30
100
0
1
o
J
10
30
100
0/159
1/80
4/78a
3/80a
4/80a
1/77
1/157
1/80
2/79
2/77
4/80a
2/75
3/160
2/80
1/78
2/80
10/79C
25116°
1/157
1/80
2/80
5/75
6/80c
24/76c
1/147
1/76
0/73
0/67
2/7 lc
14/68C
0/157
0/73
0/73
0/70
2/73c
8/62c
Not examined
Not examined
3/159
2/80
2/78
2/80
0/80
2/77
14/156
8/80
6/80
9/80
12/80
14/73a
"Significantly different from controls (p < 0.05).
bThe denominators for tumor incidences in F344 rats excluded rats from the 6- and 12-mo sacrifices and
unscheduled deaths prior to the 12-mo sacrifice. Numerators for the incidences of CNS tumors were derived by
adding the number of rats with brain or spinal astrocytomas; denominators were taken as the greater of the number
of rats examined for brain or spinal cord lesions after the 12-mo sacrifice. Numerators for Zymbal gland tumor
incidences included squamous cell papillomas and carcinomas designated to occur in the ear canal. Mammary
gland tumor incidences are for fibroadenomas and adenocarcinomas in animals sacrificed or found dead after
12 mos. Tongues were not routinely histopathologically examined for tumors in this bioassay.
Significantly different from controls (p < 0.01).
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As previously noted, a second two-year AN drinking water study employing Sprague-
Dawley rats has been conducted (Johannsen and Levinskas, 2002a; Biodynamics, 1980a). To
determine whether these data should also be employed in the cancer assessment, an analysis was
performed to evaluate the statistical validity of pooling tumor incidence data from the two
Sprague-Dawley rat chronic AN drinking water studies. The first study (Quast, 2002; Quast et
al., 1980a) exposed animals to AN drinking water concentrations of 0, 35, 100, and 300 ppm,
while the second study (Johannsen and Levinskas, 2002a) employed AN drinking water
concentrations of 0, 1, and 100 ppm. The dichotomous multistage model in BMDS was fit to the
tumor incidence data from three sites (i.e., forestomach, CNS, and Zymbal gland) in each sex
across both studies, using administered animal dose expressed in mg/kg-day. A statistical test
described by Stiteler et al. (1993), which employs a maximum likelihood ratio statistic
distributed as a %2, was then used to test the null hypothesis that the two data sets are compatible
with a common dose-response model. If the null hypothesis is not rejected, this provides
evidence that the results from the two studies may be pooled.
As discussed in more detail in Appendix B-5, the statistical tests indicated that some, but
not all, of the data sets from the two studies were consistent with a common dose-response
model. More specifically, the results of this analysis showed that forestomach and Zymbal gland
tumors in both male and female Sprague-Dawley rats were not compatible with a common dose-
response model, while CNS tumors in male and female Sprague-Dawley rats were compatible
with a common dose-response model. Because of these conflicting results, it was decided that
the results from the two Sprague-Dawley rat drinking water studies would not be pooled.
Therefore, the final dose-response analysis for deriving the oral slope factor for AN focused on
the two rat drinking water studies containing the most dose groups (i.e., the Sprague-Dawley rat
bioassay reported by Quast [2002] and the F344 rat bioassay reported by Johannsen and
Levinskas [2002b]).
5.4.2.3. Rat Inhalation Data
Incidence data for intestinal, CNS, Zymbal gland, tongue, and mammary gland tumors in
Sprague-Dawley rats exposed to AN via inhalation were used for deriving site-specific lURs for
AN (Table 5-11). With the exception of one male in the 80 ppm exposure group that died with a
CNS tumor after 7-12 months on study, all of the remaining tumors occurred in rats that died or
were sacrificed after at least 12 months of AN exposure. Denominators for the incidences in
Table 5-11 excluded animals that died without a tumor before 12 months on study because these
animals were not exposed long enough to be at risk for tumor development. Although a
statistically significantly elevated incidence in tongue tumors was observed in male rats at 80
ppm, tongues from only 14 of the males in the 20 ppm exposure group were examined. No data
on the incidence of tongue tumors in female rats were presented in the original study report by
Quast etal. (1980b).
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Table 5-11. Incidences of intestinal, CNS, Zymbal gland, tongue, and
mammary gland tumors in Sprague-Dawley rats exposed to AN via
inhalation for 2 years
Sex
Male
Female
AN air
concentration
(ppm)
0
20
80
0
20
80
Intestinal
tumors"
4/96
3/93
17/82C
-
-
-
CNS
tumorsa'b
0/96
4/93
22/82c
0/93
8/99c
20/89C
Zymbal gland
tumors"
2/96
4/93
11/82C
0/93
1/98
11/89C
Tongue
tumors"
1/95
0/14
7/82c
-
-
-
Mammary gland
adenocarcinomas"
-
-
-
9/93
8/98
20/99C
Tor all incidence data, the denominators excluded rats dying earlier than 12 mos in the study. These data were
ascertained from Tables 22, 25, 31, 34, and 35 in the original study report by Quast et al. (1980b).
blncidences for CNS tumors (brain and spinal cord) in Sprague-Dawley rats listed in this table, as reported by
Quast et al. (1980b), include both glial cell proliferation and astrocytomas.
Statistically significantly different from controls (p < 0.05) as calculated by the study authors.
Sources: Dow Chemical (1992a); Quast et al. (1980b).
5.4.3. Dose-Response Modeling
The EPA Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a) recommend that
the method used to characterize and quantify cancer risk from a chemical is determined by what
is known about the mode of action of the carcinogen and the shape of the cancer dose-response
curve. The dose response is assumed to be linear in the low-dose range when evidence supports
a mutagenic mode of action because of DNA reactivity, or if another mode of action that is
anticipated to be linear is applicable. A linear low-dose extrapolation approach was used to
estimate human carcinogenic risk associated with both inhalation and oral exposure to AN in
light of AN's mutagenic mode of carcinogenic action.
5.4.3.1. Human Occupational Data
A Cox regression model was developed based on an analysis of the Blair et al. (1998)
lung cancer mortality data for AN-exposed workers. This model yielded a regression coefficient
(ft) estimate of 1.2 x 10~3 with an associated estimated standard error of 2.47 x 10~3 and a
corresponding/? value of 0.61. The 95% upper confidence limit (UCL) on the parameter
estimate, ft, was 5.3 x 10~3. The parameter estimate, ft, in this model was about threefold greater
than the same parameter estimate in the final model developed by Starr et al. (2004), but the
UCL was similar. The Cox regression model described above was used to estimate an AN
exposure concentration (EC) and its associated 95% lower confidence limit (LEG), associated
9 _
with a 10" (1%) risk of dying from lung cancer at age 80 (ECoi and LECoi, respectively).
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Conversions of occupational exposures to continuous environmental exposures were performed
to account for differences in the number of days exposed per year (240 vs. 365 days) and in the
amount of air inhaled during an 8-hour workday versus a 24-hour day (10 and 20 m3/day,
respectively). The resulting AN exposure estimates (by age 80) were ECoi = 0.992 ppm (or 2.2
mg/m3) and LECoi = 0.238 ppm (or 0.524 mg/m3). The IUR estimate was then derived by linear
r\ -I
extrapolation from the LECoi. The corresponding unit risk was calculated to be 4.2 x 10" ppm"
9 o
or 2 x 10" per mg/m . As previously noted, Appendix B-7 provides more details on this Cox
regression analysis.
5.4.3.2. Rat Oral Data
Within each rat strain, sex, and tumor site, the multistage model, employing EPA's
BMDS (version 1.4.1), was fit to the tumor incidence data from the bioassays in Sprague-
Dawley (Quast, 2002) and F344 (Johannsen and Levinskas, 2002b) rats shown in Table 5-10
using two internal dose metrics, CEO concentration in blood (AUC/24 hours) and AN
concentration in blood (AUC/24 hours). As already discussed, CEO is a DNA-reactive epoxide
metabolite thought to play a key role in the carcinogenic mode of action of AN. The
EPA-modified PBPK model employed consistently predicted higher CEO concentrations in
blood and brain than were reported in studies of orally exposed rats (see Figures 3 and 4 in
Kedderis et al., 1996). In contrast, AN concentrations in blood, brain, and liver predicted by the
same PBPK model were fairly close to measured AN concentrations in rats following oral
exposure. Ultimately, CEO levels in blood were chosen as the internal dose metric of choice for
the oral cancer dose-response assessment because CEO is believed to be the most biologically
relevant dose metric, and it is also consistent with the dose metric employed in derivation of the
RfD. As with the RfD, CEO concentration in blood represents a reasonable internal dose metric
for extrapolating from orally exposed rats to humans.
The AN concentrations in drinking water (in ppm) employed in the Quast (2002) and
Johannsen and Levinskas (2002b) rat studies were converted to AN administered doses (in
mg/kg-day) by the study authors, using water intake data recorded during the study. These
administered doses of AN were then used as input into the EPA-modified rat PBPK model of
Kedderis et al. (1996) in order to predict a rat internal dose (either CEO-AUC or AN-AUC
concentration in blood) resulting from the ingestion of a total daily dose of AN equivalent to the
administered dose consumed in six bolus episodes per day that reflect the daily drinking water
consumption pattern of rats. The resulting predicted CEO-AUC or AN-AUC concentrations in
rat blood were then employed in dose-response modeling. Table 5-12 displays the relationship
between AN drinking water concentrations (in ppm), administered animal doses (in mg/kg-day),
and the two internal dose metrics (i.e., CEO in blood and AN in blood, both expressed in mg/L)
predicted from the PBPK model.
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Table 5-12. Four different dose metrics, two external and two internal, based
on doses employed in studies of Sprague-Dawley and F344 rats exposed to
AN in drinking water for 2 years
Species, strain
(reference)
Rat,
Sprague-Dawley
(Quast, 2002)
Rat, F344
(Johannsen and
Levinskas, 2002b)
Sex
Male
Female
Male
Female
External dose
Concentration in
drinking water
(ppm)
0
35
100
300
0
35
100
300
0
1
3
10
30
100
0
1
3
10
30
100
Administered
doseb
(mg/kg-d)
0
3.4
8.5
21.3
0
4.4
10.8
25.0
0
0.08
0.25
0.83
2.48
8.37
0
0.12
0.36
1.25
3.65
10.90
Predicted internal dose"
CEO in blood
(mg/L)
0
1.83 x 10"3
4.36 x 10'3
9.70 x 10"3
0
2.07 x 10"3
4.87 x 10'3
1.01 x 10"2
0
4.06 x 10'5
1.27 x 10'4
4.19 x 10"4
1.23 x 10'3
3.97 x 10"3
0
5.32 x 10"5
1.59 x 10'4
5.49 x 10"4
1.58 x 10'3
4.46 x 10"3
AN in blood
(mg/L)
0
2.06 x 10"2
5.36 x 10'2
1.46 x 10"1
0
2.37 x 10"2
6.18 x 10'2
1.56 x 10"1
0
4.33 x 10'4
1.35 x 10'3
4.52 x 10"3
1.37 x 10'2
4.85 x 10"2
0
5.73 x 10"4
1.72 x 10'3
6.02 x 10"3
1.79 x 10'2
5.63 x 10"2
aThe EPA-modified rat PBPK model of Keddaris et al. (1996) was employed to predict a rat internal dose (i.e.,
either AN-AUC or CEO-AUC concentration in blood) resulting from the ingestion of an administered dose of AN
consumed in six bolus episodes/d.
bAdministered doses were averages calculated by the study authors based on animal B W and drinking water intake.
Employing the internal dose metrics in Table 5-12 (i.e., CEO in blood and AN in blood),
successive stages of the multistage model, starting with stage 1 and ending with the stage equal
to the number of dose groups minus one, were fit to the tumor incidence data at a particular site
for each rat strain and sex. Then, all stages of the multistage model that did not show a
significant lack of fit (i.e.,/? > 0.1) were compared using AIC. The stage of the multistage model
with the lowest AIC was selected as the "best-fit" model. For most tumor sites, the one-stage
model exhibited the best fit.
A BMR of 10% extra risk was selected for all tumor sites, consistent with the Guidelines
for Carcinogen Risk Assessment (U.S. EPA, 2005a), which recommend identifying the POD near
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the lower end of the observed data. Using the best-fit model, the resulting BMDi0s and
BMDLioS were estimated for each tumor site within each rat strain and sex. A summary of the
results of this BMD modeling is shown in Table 5-13. Additional details regarding dose-
response modeling used in the derivation of the oral CSFs are provided in Appendix B-3.
Table 5-13. BMD modeling results using tumor incidence data from male
and female Sprague-Dawley and F344 rat studies in which animals were
exposed to AN in drinking water for 2 years
Species, strain
(reference)
Rat, Sprague-Dawley
(Quast, 2002)
Sex
Male
Female
BMD modeling results"
Tumor site
Forestomach
CNS
Zymbal gland
Tongue
Forestomach
CNS
Zymbal gland
Tongue
Mammary gland
Dose metric
CEO
AN
CEO
AN
CEO
AN
CEO
AN
CEO
AN
CEO
AN
CEO
AN
CEO
AN
CEO
AN
BMD10
(mg/L)
1.87 x 10'3
2.26 x 10'2
8.87 x 10'4
8.82 x 10'3
5.46 x 10'3
8.94 x 10'2
8.78 x 10"3
1.29 x 10"1
3.29 x 10"3
4.22 x 10"2
5.79 x 10"4
7.12x 10"3
2.40 x 10'3
3.41 x 10'2
6.70 x 10'3
9.67 x 10'2
5.50 x 10'4
7.05 x 10'3
BMDL10
(mg/L)
1.44 x 10'3
1.70 x 10'2
7.16x 10'4
6.64 x 10'3
3.15 x 10'3
4.26 x 10'2
4.90 x 10"3
6.97 x 10"2
2.38 x 10"3
2.49 x 10"2
4.51 x 10"4
5.55 x 10"3
1.78 x 10'3
2.52 x 10'2
4.74 x 10'3
6.10x 10'2
2.98 x 10'4
3.77 x 10'3
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Table 5-13. BMD modeling results using tumor incidence data from male
and female Sprague-Dawley and F344 rat studies in which animals were
exposed to AN in drinking water for 2 years
Species, strain
(reference)
Rat, F344
(Johannsen and
Levinskas, 2002b)
Sex
Male
Female
BMD modeling results"
Tumor site
Forestomach
CNS
Zymbal gland
Forestomach
CNS
Zymbal gland
Mammary gland
Dose metric
CEO
AN
CEO
AN
CEO
AN
CEO
AN
CEO
AN
CEO
AN
CEO
AN
BMD10
(mg/L)
6.03 x lO'4
6.48 x lO'3
1.16x lO'3
1.37 x lO'2
2.73 x 10"3
3.31 x 10"2
3.65 x 10"3
4.13 x 10"2
1.39 x 10"3
1.70 x 10"2
3.78 x 10'3
5.41 x 10'2
3.58 x 10'3
4.51 x 10'2
BMDL10
(mg/L)
3.55 x 10'4
3.81 x 10'3
8.74 x 10'4
1.03 x 10'2
2.19 x 10"3
2.59x 10"2
1.69 x 10"3
1.90 x 10"2
1.05 x 10"3
1.28 x 10"2
2.97 x 10'3
3.35 x 10'2
1.97 x 10'3
2.45 x 10'2
"The multistage model in EPA's BMDS (version 1.4.1) was fit to each set of tumor incidence data from the
Sprague-Dawley and F344 rat bioassays, as shown in Table 5-10, using the two internal dose metrics, CEO in blood
and AN in blood, expressed in mg/L. An adequate fit of the multistage model was achieved if the %2 goodness-of-
fit statistic yielded/) > 0.1. In the case of CNS tumors in Sprague-Dawley female rats, an adequate fit of the
multistage model to the data could not be achieved; therefore, the best fitting of the other models available in
BMDS (assessed by AIC), the log-logistic model, was used. Appendix B-3 provides further details on these BMD
modeling results.
After completion of the dose-response modeling, the BMDLios estimated for each tumor
site within each rat strain and sex for each internal dose metric were input into the EPA-modified
human PBPK model of Sweeney et al. (2003) in order to predict the human equivalent
administered dose of AN that corresponds to the estimated BMDLio, assuming six bolus
ingestion episodes of AN per day. The resulting predicted human equivalent administered dose
of AN, expressed in mg/kg-day, are shown in the third column of Table 5-14 for the internal
dose metric CEO in blood and Table 5-15 for the internal dose metric AN in blood. Finally, for
each rat strain, sex, and tumor site, the site-specific oral CSFs shown in the last column of Table
5-14 (based on CEO concentration in blood) and Table 5-15 (based on AN concentration in
blood) were derived by dividing the BMR (i.e., 10% or 0.1) by the human equivalent
administered dose of AN (in mg/kg-day), displayed in the third column of Tables 5-13 and 5-14.
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Table 5-14. Site-specific oral CSFs for AN based on BMD modeling of
tumor incidence data in rats and predicted CEO levels in blood
(AUC/24 hours) of rats and humans assuming episodic exposure to AN
Rat strain, gender
Tumor site
Rat BMDL10a
(mg/L)
BMDL10/HEDb
(mg/kg-d)
CSFC
(mg/kg-d)1
Sprague-Dawley malesA
Forestomach
CNS
Zymbal gland
Tongue
1.44 x 10"3
7.16x 10"4
3.15 x 10"3
4.90 x 10"3
0.142
0.071
0.312
0.486
0.704
1.410
0.320
0.206
Sprague-Dawley females'*
Forestomach
CNS
Zymbal gland
Tongue
Mammary gland
2.38 x 10"3
4.51 x 10"4
1.78 x 10"3
4.74 x 10"3
2.98 x 10"4
0.236
0.045
0.176
0.470
0.029
0.424
2.222
0.567
0.213
3.390
F344 males'
Forestomach
CNS
Zymbal gland
3.55 x 10"4
8.74 x 10"4
2.19 x 10"3
0.035
0.086
0.217
2.850
1.160
0.462
F344 females'
Forestomach
CNS
Zymbal gland
Mammary gland
1.69 x lO'3
1.05 x 10'3
2.97 x 10'3
1.97 x 10"3
0.167
0.104
0.294
0.195
0.599
0.963
0.340
0.514
aRat BMDL10 refers to the estimated 95% lower confidence limit on the internal dose of CEO in blood in
the rat associated with a 10% extra risk for the incidence of tumors at the specified site in the associated
strain and sex. This value is taken from the last column of Table 5-13.
bPBPK model-derived HED of AN that would result in a human 24-hr blood CEO-AUC equivalent to the
rat CEO-AUC presented in the previous column of the table, assuming AN ingestion in six bolus
episodes/d.
CCSF = BMR/BMDL10/HED or 0.1/BMDL10/HED.
dBased on data from Quast (2002).
eBased on data from Johannsen and Levinskas (2002b).
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Table 5-15. Site-specific oral CSFs for AN based on BMD modeling of
tumor incidence data in rats and predicted AN levels in blood
(AUC/24 hours) of rats and humans assuming episodic exposure to AN
Rat strain, gender
Tumor site
Rat BMDIV
(mg/L)
BMDL10/HEDb
(mg/kg-d)
CSFC
(mg/kg-d)1
Sprague-Dawley malesA
Forestomach
CNS
Zymbal gland
Tongue
1.70 x lO'2
6.64 x lO'3
4.26 x 10"2
6.97 x 10"2
4.30
1.89
8.59
11.85
0.023
0.053
0.012
0.008
Sprague-Dawley Females'*
Forestomach
CNS
Zymbal gland
Tongue
Mammary gland
2.49 x lO'2
5.55 x lO'3
2.52 x lO'2
6.10 x 10"2
3.77 x 10"3
5.82
1.60
5.87
10.90
1.11
0.017
0.062
0.017
0.009
0.090
F344 males"
Forestomach
CNS
Zymbal gland
3.81 x 10'3
1.03 x 10'2
2.59 x 10'2
1.12
2.82
5.99
0.089
0.036
0.017
F344 females*
Forestomach
CNS
Zymbal gland
Mammary gland
1.90 x 10"2
1.28 x 10'2
3.35 x 10'2
2.45 x 10'2
4.71
3.38
7.26
5.75
0.021
0.030
0.014
0.017
aRat BMDLio refers to the estimated 95% lower confidence limit on the internal dose of AN in blood in the
rat associated with a 10% extra risk for the incidence of tumors at the specified site in the associated strain
and sex. This value is taken from the last column of Table 5-13.
bPBPK model-derived HED of AN that would result in a human 24-hr blood AN-AUC equivalent to the
rat AN-AUC presented in the previous column of the table, assuming AN ingestion in six bolus
episodes/d.
CCSF = BMR/BMDL10/HED or 0.1/BMDL10/HED.
dBased on data from Quast (2002).
eBased on data from Johannsen and Levinskas (2002b).
In comparing Tables 5-13 and 5-14, the CEO-based site-specific oral CSF estimates in
Table 5-14 are higher than those based on AN blood levels in Table 5-15. This occurs because
the VmaxC/Km for AN oxidation to CEO is estimated to be about 10 times higher in humans than
in the rat. AN enzymatic GSH conjugation is estimated to be 1.5 times higher in the rat, but the
2n -order removal constant for non-enzymatic reaction of AN and GSH is assumed to be the
same and the GSH tissue levels are approximately the same (and lower in the larger tissue
groups). The result, for example, with a steady oral "infusion" of 30 mg/kg-day, is that AN is
removed somewhat faster overall, leading to a PBPK-predicted steady-state blood level of 0.114
mg/L AN in humans vs. 0.151 mg/L in the rat, but a much higher portion of this goes to CEO in
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the human yielding a CEO blood level of 0.297 vs. 0.013 mg/L in the rat. Thus, the AN:CEO
ratio is 11.4 in the rat versus 0.385 in the human, and so the CEO-based risk estimates are
approximately 50 times higher than the AN-based estimates.
For comparative purposes, in Table 5-16, the BMDLio estimates generated from the
BMD modeling of the incidence of tumors in male and female Sprague-Dawley and F344 rats
using administered animal dose (modeling results not shown) were converted to human
equivalent administered doses of AN using BW scaling to the % power. Then, as in Tables 5-13
and 5-14, oral CSFs were derived by dividing the BMR (i.e., 10% or 0.1) by the human
equivalent administered dose of AN (in mg/kg-day).
Table 5-16. Site-specific oral CSFs for AN based on BMD modeling of
tumor incidence data in rats and BW scaling to the 3/4 power to convert from
rat to human administered doses
Rat strain, gender
Tumor site
Rat BMDL10a
(mg/kg-d)
BMDL10/HEDb
(mg/kg-d)
CSFC
(mg/kg-d)1
Sprague-Dawley males'1
Forestomach
CNS
Zymbal gland
Tongue
2.76
1.48
5.17
10.4
0.812
0.436
1.52
3.04
0.123
0.229
0.066
0.033
Sprague-Dawley females'*
Forestomach
CNS
Zymbal gland
Tongue
Mammary gland
4.81
0.99
4.19
8.30
0.66
1.27
0.262
1.11
2.19
0.174
0.079
0.382
0.090
0.046
0.575
F344 males'
Forestomach
CNS
Zymbal gland
0.70
1.81
3.43
0.193
0.493
0.932
0.517
0.203
0.107
F344 females'
Forestomach
CNS
Zymbal gland
Mammary gland
3.89
2.52
6.64
4.77
0.926
0.602
1.59
1.14
0.108
0.166
0.063
0.088
aRat BMDL10 is the estimated 95% lower confidence limit on the administered dose of AN in the rat
associated with a 10% extra risk for the incidence of tumors at the specified site in the associated strain
and sex. This value is generated from the "best-fit" dose-response model in BMDS (version 1.4.1).
bThe HED of AN equal to the rat BMDL10 in the previous column converted through use of BW scaling to
the 3/4 power.
CCSF = BMR/BMDL10/HED or 0.1/BMDL10/HED.
dBased on data from Quast (2002).
eBased on data from Johannsen and Levinskas (2002b).
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5.4.3.3. Rat Inhalation Data
Within each sex and for each tumor site, the multistage model, employing EPA's BMDS
(version 1.4.1), was fit to the tumor incidence data from the AN inhalation bioassay in Sprague-
Dawley rats (Quast et al., 1980b) shown in Table 5-11 using the internal dose metric, CEO
concentration in blood (AUC/24 hours). In contrast to the approach used for oral exposure, only
one internal dose metric was selected on which to base site-specific lURs (i.e., CEO in blood),
because, in contrast to oral exposure, the EPA-modified PBPK model adequately predicted
measured blood and brain concentrations of CEO in rats exposed to AN via inhalation (Kedderis
et al., 1996). Furthermore, as mentioned previously, CEO is the DNA-reactive metabolite
thought to be key in the carcinogenic mode of action of AN.
Prior to dose-response modeling, the AN concentrations in air (in ppm) administered in
the Quast et al. (1980b) rat study were input into the EPA-modified rat PBPK model of Kedderis
et al. (1996) in order to predict a rat internal dose (CEO-AUC concentration in blood) resulting
from inhalation exposure to the administered air concentration of AN. The resulting predicted
CEO-AUC concentrations in rat blood were then employed in dose-response modeling.
Table 5-17 displays the relationship between AN concentrations in air (expressed in ppm) and
the internal dose metric, CEO in blood (expressed in mg/L), predicted from the PBPK model.
Table 5-17. Two different dose metrics, one external and one internal, based
on administered air concentrations of AN employed in a 2-year bioassay in
Sprague-Dawley rats
Species, strain
Rat, Sprague-Dawley
Sex
Male
Female
AN concentration in air
(ppm)
0
20
80
0
20
80
Predicted CEO concentration
in blood"
(mg/L)
0
2.11 x 10'3
8.20 x 10'3
0
2.18 x 10"3
8.24 x 10"3
aSee Table 5-12.
After completion of the dose-response modeling, the BMCios and BMCLios estimated
within each sex for each tumor site using CEO in blood as the dose metric from Table 5-18 were
input into the EPA-modified human PBPK model of Sweeney et al. (2003) in order to predict the
HEC of AN in air that corresponds to the estimated BMCio and BMCLio in animals. The
resulting predicted 95% lower bounds of the HECs (BMCLi0/HEc) of AN in air, expressed in
mg/m3, are shown in the third column of Table 5-19 for the internal dose metric CEO in blood.
Finally, for each sex and tumor site, the site-specific lURs shown in the last column of
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Table 5-19 (based on CEO concentration in blood) were derived by dividing the BMR (i.e., 10%
or 0.1) by the BMCLi0/HEC.
For purposes of comparison, a similar analysis was performed using external AN
concentration in air as the dose metric. The results of this analysis are presented in Table 5-20.
Table 5-18. BMD modeling results using tumor incidence data from male
and female Sprague-Dawley rats exposed to AN via inhalation for 2 years
and CEO concentration in blood predicted from an EPA-modified PBPK
model
Strain, species
(reference)
Rat, Sprague-
Dawley
(Quast et al.,
1980b)
Sex
Male
Female
BMD modeling results
Tumor site
Intestine
CNS
Zymbal gland
Tongue
CNS
Zymbal gland
Mammary gland
Dose metric
CEO
CEO
CEO
CEO
CEO
CEO
CEO
BMC10
(mg/L)
6.06 x 10"3
3.14x 10"3
7.26 x 10"3
9.48 x lO'3
3.21 x lO'3
7.90 x 10'3
7.31 x 10'3
BMCL10
(mg/L)
4.47 x 10"3
2.31 x 10"3
4.40 x 10"3
6.39 x 10'3
2.39 x 10'3
5.09 x 10'3
4.33 x 10'3
aThe multistage model in EPA'sBMDS (version 1.4.1) was fit to each set of tumor incidence data from the
Sprague-Dawley rat bioassay, as shown in Table 5-11, using the internal dose metric, CEO in blood, expressed in
mg/L. An adequate fit of the multistage model was achieved if the % goodness-of-fit statistic yielded/? > 0.1.
Appendix B-4 provides further details on these BMD modeling results.
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Table 5-19. Site-specific lURs for AN based on BMD modeling of tumor
incidence data in Sprague-Dawley rats and PBPK modeling of CEO levels in
blood (AUC/24 hours) of rats and humans
Gender
Tumor site
Rat BMCL10a
(mg/L)
BMCL/io/HEc
(mg/m3)
IURC
(mg/m3)1
Males'1
Intestine
CNS
Zymbal gland
Tongue
4.47 x 10"3
2.31 x lO'3
4.40 x 10"3
6.39 x 10"3
6.00
3.10
5.91
8.58
0.017
0.032
0.017
0.012
Females*
CNS
Zymbal gland
Mammary gland
2.39 x lO'3
5.09 x 10"3
4.33 x 10'3
3.21
6.83
5.81
0.031
0.015
0.017
aRat BMCL10 refers to the estimated 95% lower confidence limit on the concentration of CEO in blood of
the rat associated with a 10% extra risk for the incidence of tumors at the specified site based on BMD
modeling. This value is taken from the last column of Table 5-18.
'"PBPK model-derived HEC of AN in air that would result in a human 24-hr blood CEO-AUC
concentration equivalent to the rat BMCL10 in the previous column of the table assuming continuous
exposure to AN. The human PBPK model employed did not contain the human in vitro to in vivo
modifying factor for CEO hydrolysis proposed by Sweeney et al. (2003).
CIUR = BMR/BMCL10/HEc or 0. l/BMCL10/HEc.
dBased on data from Quast et al. (1980b).
Table 5-20. Site-specific lURs for AN based on BMD modeling of tumor
incidence data in Sprague-Dawley rats exposed to AN via inhalation
Gender
Tumor site
Rat BMCL10a
(ppm)
BMCL/io/HEc
(mg/m3)
IURC
(mg/m3)1
Males'1
Intestine
CNS
Zymbal gland
Tongue
42.68
22.23
42.53
59.41
93.58
48.74
93.25
130.26
1.07 x 10'3
2.05 x 10"3
1.07 x 10"3
7.68 x 10'4
Females'1
CNS
Zymbal gland
Mammary gland
22.89
48.74
37.82
50.19
106.87
82.92
1.99 x 10"3
9.36 x 10'4
1.21 x 10'3
aRat BMCL10 refers to the estimated 95% lower confidence limit on the concentration of AN in air
associated with a 10% extra risk for the incidence of tumors at the specified site based on BMD modeling.
bThe HEC of AN in air equivalent to the rat BMCL10 in the previous column of the table generated through
use of the following conversion equation: mg/m3 = [ppm x molecular weight]/24.20, where molecular
weight = 53.06.
CIUR = BMR/BMCL10/HEc or 0.1/BMCL10/HEc.
dBased on data from Quast et al. (1980b).
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5.4.4. Oral Slope Factor and Inhalation Unit Risk
5.4.4.1. OralCSFs
Because AN has been demonstrated to produce tumors at multiple sites in rats, the
estimation of risk based on only one tumor site may underestimate the overall carcinogenic
potential of AN. Under the assumption that AN causes tumors by a mutagenic mode of action,
estimates of the composite risk of etiologically distinct tumor types in each rat strain/sex
combination considered in Section 5.4.3.2 were derived employing the Markov Chain Monte
Carlo approach described in Appendix B-6. This approach is consistent with the
recommendations of the NRC (1994) and the Guidelines for Carcinogen Risk Assessment (U.S.
EPA, 2005a), which recommends "adding risk estimates derived from different tumor sites
(NRC, 1994) as an option for "how best to represent the human cancer risk." This composite
risk is associated with the potential for developing tumors, not at all of the sites, but at any
combination of the sites observed in male or female rats. For this analysis, the etiologically
distinct tumor sites associated with AN exposure were assumed to be statistically independent.
This assumption cannot currently be verified, but if not correct could lead to an overestimate of
risk. NRC (1994) concluded that a general assumption of statistical independence of tumors
within animals was not likely to introduce substantial error in assessing carcinogenic potency
from rodent bioassay data.
Table 5-21 presents the estimated human equivalent oral CSFs for both site-specific risks
and composite risks from multiple tumor types from each rat strain and sex exposed to AN via
drinking water based on the internal dose metric, CEO concentration in blood (AUC over
24 hours).
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Table 5-21. Estimated human equivalent oral CSFs for composite risk based
on tumor incidence in AN-exposed rats and predicted CEO-AUC levels in
blood
Rat strain and gender
Tumor sites included
Site-specific oral CSFa
(mg/kg-d)1
Oral CSF for composite risk
across sites
(mg/kg-d)1
Sprague-Dawley maleb
Forestomach
CNS
Zymbal gland
Tongue
0.704
1.410
0.320
0.206
2.8
Sprague-Dawley female*
Forestomach
CNS
Zymbal gland
Tongue
Mammary gland
F344malec
Forestomach
CNS
Zymbal gland
0.424
2.222
0.567
0.213
3.390
2.850
1.160
0.462
5.0
3.1
F344femalec
Forestomach
CNS
Zymbal gland
Mammary gland
0.599
0.963
0.340
0.514
1.9
aThese site-specific oral CSFs are taken from the last column of Table 5-14.
bBased on data from Quast (2002).
°Based on data from Johannsen and Levinskas (2002b).
In the study by Quast (2002), male Sprague-Dawley rats showed statistically significantly
increased incidences of tumors at four sites (i.e., forestomach, CNS, Zymbal gland, and tongue),
while females showed statistically significantly increased incidences of tumors at five sites (i.e.,
forestomach, CNS, Zymbal gland, tongue, and mammary gland), following chronic oral
exposure to AN. The composite oral CSF based on tumor incidences in male Sprague-Dawley
rats chronically exposed to AN (using CEO concentration in blood) is estimated to be 2.8 per
mg/kg-day, about twice as high as the estimated slope factor resulting from the most sensitive
single site in this strain and sex. For female Sprague-Dawley rats, the composite oral CSF
(based on CEO concentration in blood) was 5.0 per mg/kg-day, approximately 50% higher than
the estimated slope resulting from the most sensitive single site in this strain and sex.
In the study by Johannsen and Levinskas (2002b), male F344 rats showed statistically
significant increased incidences of tumors at three sites (i.e., forestomach, CNS, and Zymbal
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gland), while females showed statistically significant increased incidences of tumors at four sites
(i.e., forestomach, CNS, Zymbal gland, and mammary gland), following oral exposure to AN in
drinking water for 2 years. Again, under the assumption that the tumors at these sites are
statistically independent, the composite oral CSF in F344 males (based on CEO concentration in
blood) was 3.1 per mg/kg-day, slightly higher than the estimated slope factor of 2.8 per mg/kg-
day based on tumors in the forestomach alone, the most sensitive single site in this strain and
sex. For female F344 rats, the composite oral CSF (based on CEO concentration in blood) was
1.9 per mg/kg-day, about twice the estimated slope factor of 0.96 per mg/kg-day from tumors of
the CNS, the most sensitive single site in this strain and sex.
The CSFs obtained from male and female Sprague-Dawley rats and F344 rats ranged
from 2 to 5 (mg/kg-day)"1. While female Sprague-Dawley rats had the highest CSF due to
increase in mammary gland tumor risk, the increase in mammary gland tumor risk in female
F344 rats was not as high. There is no information as to which rat strain may be most similar to
humans. Both strains of rats developed the same tumors in response to AN exposure. For the
purpose of this assessment, the CSF of 5 per mg/kg-day is recommended for use in humans
because it is the value obtained from the most sensitive species, strain, and sex (i.e., female
Sprague-Dawley rats). This CSF should not be used with exposures greater than 0.04 mg/kg-day
(the lowest POD supporting the composite risk) because above this level the CSF cannot be
expected to be an adequate approximation to the dose-response relationship. The fitted dose-
response relationship and pharmacokinetic models should be used to estimate risk above this
exposure level. See Section 5.6 regarding the application of age-dependent adjustment factors
(ADAFs).
5.4.4.2. Inhalation Unit Risk
Employing human data, the ECoi and LECoi, defined as the AN exposure concentration
and its associated 95% lower confidence limit, respectively, that result in a 1% increase in the
risk of dying from lung cancer at age 80 were estimated from the Cox regression model derived
from the Blair et al. (1998) occupational epidemiology study. These ECoi and LECoi values
were 0.992 and 0.238 ppm, respectively (or 2,187 and 524 ug/m3, respectively). From the
LECoi, an IUR of 4.2 x 10"2 ppm"1 or 2 x 10"2 (mg/m3)"1 was derived.
For inhalation exposures to AN in animals, as with oral exposures, estimates of composite
risk addressing multiple tumor sites (within strain and sex) based on BMD modeling results from
a chronic inhalation rodent bioassay in Sprague-Dawley rats (Dow Chemical 1992a; Quast et al.,
1980b) were generated. These estimates were derived employing the same Markov Chain Monte
Carlo approach described in Appendix B-6 for oral exposures.
Table 5-22 presents the human equivalent lURs for both site-specific risks and composite
risks from multiple tumor types observed in each sex of Sprague-Dawley rat exposed to AN
using the internal dose metric, CEO concentration in blood (AUC for 24 hours).
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Table 5-22. Estimated human equivalent composite lURs based on tumor
incidence in AN-exposed rats and predicted CEO-AUC levels in blood
Rat strain and gender
Tumor sites included
Site-specific IURa
(mg/m3)-1
Composite IUR
(mg/m3)-1
Sprague-Dawley maleb
Intestine
CNS
Zymbal gland
Tongue
0.017
0.032
0.017
0.012
6.8 x 10'2
Sprague-Dawley femaleb
CNS
Zymbal gland
Mammary gland
0.031
0.015
0.017
5.7 x 10'2
"These site-specific lURs are taken from the last column of Table 5-19.
bBased on data from Quast et al. (1980b).
From the inhalation bioassay, male rats showed statistically significant elevated tumor
incidences at four sites (i.e., intestine, CNS, Zymbal gland, and tongue), while female rats
showed statistically significant elevations at three tumor sites (i.e., CNS, Zymbal gland, and
mammary gland), when exposed to AN via inhalation for 2 years. Under the assumption that
tumors at these sites are statistically independent, the estimated composite IUR in male rats was
9 T
6.8 x 10" per mg/m , approximately 2 times higher than the site-specific IUR estimate of 3.2 x
9 T
10" per mg/m based on the most sensitive single site in males (i.e., CNS). For female rats, the
composite IUR was 5.7 x 10"2 per mg/m3, about 2 times higher than the site-specific estimate of
9 T
3.1 x 10" per mg/m from the most sensitive single site in females (i.e., CNS). These unit risks
should not be used with exposures greater than 3 mg/m3 (the lowest POD supporting the
composite risk), because above this level the unit risk cannot be expected to be an adequate
approximation of the dose-response relationship. The fitted dose-response relationship and
pharmacokinetic models should be used to estimate risk above this exposure level.
The IUR recommended for use in estimating cancer risks associated with chronic
9 T
inhalation exposures to AN during adult stages of development is 2 x 10" per mg/m (or
2 x 10"5 per ug/m3)—the value based on human data (i.e., the occupational epidemiology study
by Blair et al., 1998). The derivation of this IUR is described in Section 5.4.3.1 and Appendix
B-7. The IUR derived from epidemiologic data is chosen over the alternative IUR value derived
from animal inhalation bioassay data because the use of human study data eliminates the
uncertainty inherent in animal to human extrapolation.
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5.4.4.3. Application of Age-Dependent Adjustment Factors (ADAFs)
AN is determined to be carcinogenic by a mutagenic mode of action (see Section 4.7.3.1),
which raises concern for increased early-life susceptibility to cancer. Consistent with this
possibility, two studies in Sprague-Dawley rats provide some evidence of increased cancer
susceptibility associated with chronic AN exposure that begins in early periods of
development—a chronic-duration inhalation cancer bioassay (Maltoni et al., 1988) and a three-
generation drinking water reproductive toxicity study (Friedman and Beliles, 2002). According
to the Supplemental Guidance for Assessing Susceptibility from Early-Life Exposure to
Carcinogens ^Supplemental Guidance") (U.S. EPA, 2005b), individuals exposed during early
life to carcinogens with a mutagenic mode of action are assumed to be at increased risk for
cancer. In these situations, the Supplemental Guidance recommends that age dependent
adjustment factors (ADAFs) be applied in estimating cancer risk. The guidance further
recommends that when data are available for a susceptible lifestage, those data should be used
directly to evaluate risks for that chemical and that lifestage.
In the Maltoni et al. (1988) study conducted at the Ramazzini Institute, pregnant Sprague-
Dawley rats were exposed to 60 ppm AN by inhalation for 104 weeks (see Section 4.2.2.2.2 for a
complete study description). In addition, offspring of these dams were exposed to AN starting at
GD 12 for a total of 104 weeks. Female rats exposed from GD 12 had a higher incidence of
malignant mammary tumors, encephalic gliomas, and extrahepatic angiosarcomas than female
rats exposed as adults. The drinking water bioassay by Friedman and Beliles (2002) involved
48-week exposures of each of three generations of female breeder Sprague-Dawley rats (see
Section 4.2.1.2.9 for a complete study description). Compared with the 500-ppm FO breeders
that were exposed starting in adulthood, there was an increase in Zymbal gland and brain tumor
incidence in the 500-ppm Fib female rats exposed starting in utero. Increases in tumor incidence
in the F2 generation were not statistically significantly different from FO breeders.
The Maltoni et al. (1988) and Friedman and Beliles (2002) studies were considered for
use in deriving chemical-specific ADAFs for early-life exposure to AN. The Maltoni et al.
(1988) tumor results were not used to develop chemical-specific ADAFs because of concerns
raised by a memorandum from NTP (Malarkey et al., 2010) that discussed differences of opinion
between NTP scientists and the Ramazzini Institute in the diagnoses of certain cancers reported
in a methanol study. While these data are considered qualitatively as support for early-life
susceptibility, EPA decided not to rely on these data for quantitative purposes. The Friedman
and Beliles (2002) study was also not considered further for the derivation of chemical-specific
ADAFs. This study had several limitations, including a small number of animals per treatment
group (20) and, therefore, limited power to detect tumor increases, and a lower tumor response in
the F2b generation than the Fib generation. Therefore, although limited data are available
supporting early-life susceptibility to carcinogenicity from AN exposure, the available
information was not considered suitable for developing data-specific ADAFs. Accordingly, it is
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recommended that default ADAFs be applied to cancer risk values (OSF and IUR) for AN.
The Supplemental Guidance establishes default ADAFs for three specific age groups.
These ADAFs and their corresponding age groups are: 10 for exposed individuals <2 years; 3 for
exposed individuals 2 to <16 years; and 1 for exposed individuals >16 years. The 10- and 3-fold
adjustments are combined with age-specific exposure estimates when estimating cancer risks
from early life (<16 years of age) exposures to AN. Example calculations for estimating cancer
risks based on age at exposure are provided in Section 6 of the Supplemental Guidance (U.S.
EPA, 2005b).
5.4.4.3.1. Application of ADAFs in Oral Exposure Scenarios
To illustrate the use of ADAFs in oral exposure scenarios, sample calculations are
presented for three exposure duration scenarios. These scenarios include full lifetime exposure
(assuming a 70-year lifespan) and two 30-year exposures from ages 0-30 years and ages 20-50
years. An average daily dose of 0.0001 mg/kg-day AN was assumed for each scenario. The
dose of 0.0001 mg/kg-day is used here for illustrative purposes only to demonstrate how to apply
ADAFs. In practice, actual exposure information specific to the situation under consideration
should be used.
Table 5-23 lists the four factors (ADAFs, OSF, assumed dose, and duration adjustment)
that are needed to calculate the partial cancer risk based on the early age-specific group. The
cancer risk for each age group is the product of the four factors in columns 2 to 5. For example,
the cancer risk following daily oral exposure to AN in the age group 0 to <2 years is the product
of the values in columns 2 to 5 or 10 x 5 x 0.0001 x 2/70 = 1.4 x 10~4. The risks listed in the last
column of Table 5-23 are summed to obtain an estimate of the total risk. Thus, a 70-year
(lifetime) risk estimate associated with an average daily dose of 0.0001 mg/kg-day AN is 8.3 x
10"4, which is adjusted for early-life susceptibility and assumes a 70-year lifetime and constant
exposure rate across age groups.
Table 5-23. Application of ADAFs to AN cancer risk following a lifetime
(70-year) oral exposure
Age group
(years)
0-<2
2-<16
>16
ADAF
10
3
1
Slope factor
(per mg/kg-day)
5
5
5
Average daily dose
(mg/kg-day)
0.0001
0.0001
0.0001
Duration
adjustment
2 years/70 years
14 years/70 years
54 years/70 years
Total risk
Cancer Risk
for Specific
Exposure
Duration
Scenarios
1.4 x 1Q-4
3.0 x 1Q-4
3.9 x 1Q-4
8.3 x 1Q-4
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In calculating the cancer risk for a 30-year exposure to AN at a constant average daily
dose of 0.0001 mg/kg-day from ages 0-30 years, the duration adjustments would be 2/70, 14/70,
and 14/70. The risks for the three age groups would be 1.4 x 10'4, 3.0 x 10'4, and 1.0 x 10'4,
which would result in a total risk estimate of 5.4 x 10"4.
In calculating the cancer risk for a 30-year exposure to AN at the same average daily dose
from ages 20-50 years, the duration adjustments would be 0/70, 0/70, and 30/70. The partial
risks for the three age groups would be 0, 0, and 2.1 x io~4, which would result in a total risk
estimate of 2.1 x 10"4.
5.4.4.3.2. Application of ADAFs in Inhalation Exposure Scenarios
To illustrate the use of ADAFs in inhalation exposure scenarios, sample calculations are
presented for three scenarios involving inhalation exposure. These scenarios include full lifetime
exposure (assuming a 70-year lifespan) and two 30-year exposures from ages 0-30 years and
ages 20-50 years. A constant exposure concentration of 1 |ig/m3 AN was assumed for each
scenario. The exposure concentration of 1 |ig/m3 is used here for illustrative purposes only to
demonstrate how to apply ADAFs. In practice, actual exposure information specific to the
situation under consideration should be used.
Similar to the oral exposure scenarios presented in Section 5.4.4.3.1, Table 5-24 lists the
four factors (ADAFs, unit risk, assumed exposure concentration, and duration adjustment) that
are needed to calculate the partial cancer risk based on the early age-specific group. The cancer
risk for each age group is the product of the four factors in columns 2 to 5. For example, the
cancer risk following daily inhalation exposure to AN in the age group 0 to <2 years is the
product of the values in columns 2 to 5 or 10 x (2 x icr5) x 1 x 2/70 = 5.7 x 10"6. The risks listed
in the last column of Table 5-24 are summed to obtain an estimate of the total risk. Thus, a 70-
year (lifetime) risk estimate for continuous exposure to 1 |ig/m3 AN is 3.3 x 10"5, which is
adjusted for early-life susceptibility and assumes a 70-year lifetime and constant exposure across
age groups.
Table 5-24. Application of ADAFs to AN cancer risk following a lifetime
(70-year) inhalation exposure
Age group
(years)
0-<2
2-<16
>16
ADAF
10
3
1
Unit risk
(per fig/m3)
2 x lO'5
2 x 10'5
2 x lO'5
Exposure
concentration
Oig/m3)
1
1
1
Duration
adjustment
2 years/70 years
14 years/70 years
54 years/70 years
Total risk
Cancer Risk for
Specific Exposure
Durations
5.7 x 10'6
1.2 x 10'5
1.5 x 10'5
3.3 x 10'5
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In calculating the cancer risk for a 30-year exposure to AN at a constant exposure
concentration of 1 |ig/m3 from ages 0-30 years, the duration adjustments would be 2/70, 14/70,
and 14/70. The risks for the three age groups would be 5.7 x 10'6, 1.2 x 10'5, and 4.0 x 10'6,
which would result in a total risk estimate of 2.2 x 10"5.
In calculating the cancer risk for a 30-year constant exposure to AN at an exposure
concentration of 1 |ig/m3 from ages 20-50 years, the duration adjustments would be 0/70, 0/70,
and 30/70. The partial risks for the three groups would be 0, 0, and 8.6 x 10"6, which would
result in a total risk estimate of 8.6 x 10"6.
5.4.5. Uncertainties in Cancer Risk Values
Risk estimates have inherent uncertainties. This subsection discusses the
uncertainties that may be associated with cancer risk values for oral and inhalation cancer
assessments.
5.4.5.1. Oral Cancer Assessment
Uncertainties related to the oral cancer assessment are discussed below and summarized
in Table 5-25 and Figure 5-3.
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Table 5-25. Summary of uncertainty in the AN oral cancer risk assessment
Consideration/
approach
Low-dose
extrapolation
procedure
PBPK model
Dose metric
Statistical
uncertainty at
POD
Bioassay
Species/gender
combination
Human population
variability in
metabolism and
response/sensitive
subpopulations
Early-life
susceptibility
Impact on cancer risk
estimate
The selected model does not
represent all possible models
one might fit, and other
models could conceivably be
selected to yield more
extreme results consistent
with the observed data, both
higher and lower than those
included in this assessment.
Oral slope factor based on
PBPK modeling and internal
CEO levels is about sixfold
higher than slope factor
based on the default
approach.
Alternatives could decrease
oral slope factor (e.g., use of
AN-AUC instead of CEO-
AUC could lower slope
factor by 30-fold).
Oral slope factor will be
reduced 1.3 -fold if BMD
used as POD instead of
BMDL.
Oral slope factor would be
the same if study on F344
rats were used (Johannsen
and Levinskas, 2002b).
Human risk could decrease
or increase, depending on
relative sensitivity.
Low-dose risk increase to an
unknown extent.
Lifetime cancer risk
increased by 1.6-fold.
Decision
Multistage model to
determine POD,
linear low-dose
extrapolation from
POD.
EPA-revised model
was used.
CEO-AUC was
used.
BMDL (approach
for calculating
reasonable upper
bound).
Quast (2002) study
on Sprague-Dawley
rats was used.
Oral slope factor
derived from study
on female Sprague-
Dawley rats.
Considered
qualitatively.
Application of
ADAFs of 10 (for
individuals <2
years), 3 (for
individuals 2 to <16
years), and 1 (for
individuals > 16
years).
Justification
Available mode of action data support
mutagenicity as the key mode of action and
the application of the low-dose linear
extrapolation approach. EPA's Guidelines
for Carcinogen Risk Assessment (U.S. EPA,
2005a): mutagens "are assessed with a
linear approach." Mutagenic mode of action
functions systemically at multiple tumor
sites.
The revised PBPK model includes EH
activity in rats and provides better estimates
of internal dose.
CEO is the reactive metabolite that binds to
DNA and initiates tumor formation. AN is
not the causal agent for cancer.
Limited size of bioassay results in sampling
variability; BMDL is lower 95% confidence
limit of BMD.
Quast (2002) has separate interim sacrifice
groups and examined more endpoints than
the Biodynamics (1980c) data sets.
It was assumed that humans are as sensitive
as the most sensitive rodent gender/species
tested; true correspondence is unknown.
The carcinogenic response occurs across
species. Generally, direct site concordance
is not assumed; consistent with this view,
some rodent tumors are not found in
humans (e.g., Zymbal gland tumors,
Harderian gland tumors) and rat and mouse
tumor types also differ.
Variability in human susceptibility likely
exists due to differences in microsomal
CYP2E1 activities and possibly GST
activity.
Individuals exposed during early life to
carcinogens with a mutagenic MOA are
assumed to be at increased risk for cancer
(U.S. EPA, 2005b). In the absence of
adequate chemical-specific data to evaluate
differences in age-specific susceptibility,
EPA's Supplemental Guidance
recommends the application of ADAFs.
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5'
S 4>
o
ro
u_
icer Slope
ig/kg-day)~1
CO
<5 £.
o
I 2.
1 •
2.8
+
0.40
5.0
1.03
4CSF-CEO in Blood
• CSF - Administered Dose
3.1
.*
(J.wa
1.9
*
0.36
•
Male Sprague-Dawley
Rats
Female Sprague-
Dawley Rats
Male F344 Rats
Female F344 Rats
Sex/Strain/Species
Figure 5-2. Comparison of composite oral CSFs derived from tumor
incidence data in four different sex/strain/species of rats exposed chronically
to AN. For each sex/strain/species combination, two different dose metrics
were employed: (1) CEO concentration in blood, and (2) human equivalent
administered dose.
Choice of study
Several bioassays by the oral route are available. The 2-year drinking water studies of
Sprague-Dawley rats (Quast, 2002; Quast et al., 1980a) and F344 rats (Johannsen and Levinskas,
2002a; Biodynamics, 1980a) were selected as principal studies. These two studies have
sufficient numbers of animals, investigate a thorough set of endpoints, and are considered well
conducted. The tumor types observed and incidence of treatment-related tumors were similar in
these two studies. Moreover, cancer risk estimates from these two studies differed only by a
factor of 1.5. Similarities in these results increase the level of confidence of estimated cancer
risk. The Quast (2002) study of Sprague-Dawley rats investigated more endpoints and is
therefore a slightly stronger study.
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PBPK model
Use of the PBPK model to extrapolate from rodents to humans also introduces some
uncertainty. An extensive quantitative analysis of the PBPK model uncertainty is presented in
Appendix D. Briefly, PBPK models are computational tools used to predict chemical/drug
disposition and are comprised of three distinct types of information: physiological,
physicochemical, and biochemical. The physiological data are independent of chemical-specific
data, and describe such parameters as organ volumes and blood flows. Physicochemical
parameters are chemical-specific and specify parameters, such as PCs or permeability.
Biochemical parameters define the rates of chemical transformation or binding. Because the
physiological data are considered to be well-characterized, analysis focused on uncertainties in
the chemical-specific parameters employed in the PBPK model used to predict AN dosimetry in
humans, which is adapted from that of Sweeney et al. (2003). (Only a small number of
(metabolic) parameters were changed in the EPA's adaptation.) Uncertainty analysis and
characterization of the human model lead to the following conclusions.
• Blood:air and tissue:air PCs as measured directly (in rodents) vs. estimated by a
computational tool were compared, and the impact on model predictions of peak AN and
peak CEO levels in brain and blood was found to be less than 30% (difference between
predictions using the alternate sets of PCs).
• Comparison of model predictions to the limited human data of Jakubowski et al. (1987)
shows that the model over-predicts the inhalation respiratory retention measured
(predicted ~ 70%; measured 44-58%). This over-prediction may be due to the fact that
the model does not fully describe the exposure apparatus (which can introduce additional
airway "dead-space") or that the model does not describe gas absorption/desorption in the
conducting airways that can reduce uptake rates.
• Once absorbed, the model predicts that 8.4% of AN is converted to CEO and then
hydrolyzed, while 30% is converted to CEO and then conjugated with GSH. These
values bracket the observation of 16.3% of retained AN being excreted in urine as "CEO"
(after acid extraction). Thus, this limited observation is in line with model predictions
subsequent to AN predictions, though the exact level of agreement or error cannot be
determined due to the distinction between observed quantity (CEO in urine) and what the
model predicts (CEO metabolic rates).
• The PBPK model over-predicts CEO levels in the rat at early time-points after i.v.
injection (Figure 3-3) and in blood and brain at all time points measured after oral
exposure (Figure 3-5a). Kedderis et al. (1996) suggest that the overestimation of CEO by
the rat model at the early time points (which also occurs in the EPA's revision) may be
due to an intrahepatic first pass effect, as occurs with other epoxides formed in situ from
their parent olefins (Filser and Bolt, 1984). However, this explanation is unlikely and a
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more plausible explanation is that the model does not account for time-dependence in
GSH levels. Inclusion of GSH dynamics would be a much more significant and intensive
change to the model structure than single variation implemented here (addition of EH to
the rat model with parameter re-estimation) and so has not been considered.
• Parametric sensitivity analysis of human predictions showed a shift in importance
between the EH and GSH pathways for CEO elimination from being approximately equal
in the implementation of Sweeney et al. (2003) to lower significance for EH (now low
but significant) and high significance for GSH with the revised parameters. The
oxidation of AN to CEO has a significant effect on AN predictions but only a small effect
on CEO in both model versions, reflecting the high dependence of CEO concentrations
on CEO metabolic removal. Not surprisingly, brain:tissue PCs significantly affect
predictions in brain tissue, and the rate constant for oral absorption affects peak AN
concentrations in blood, but not the AUG. While the human parameters are largely
derived from human in vitro data, which gives a higher confidence than would occur had
they only been extrapolated from rats, the high dependence of CEO concentrations on
GSH conjugation rates, together with the hypothesis above regarding the lack of GSH
dynamics in the model, point to that pathway description in particular as being the
greatest source of quantitative uncertainty.
• Estimated coefficients of variation for predicted concentrations in brain and blood after
inhalation exposure are -0.6-0.7 for AN and 0.9-1.2 for CEO; after oral exposure these
are 0.8-0.9 for AN and 0.7-1.0 for CEO. Based on these values, the human model
predictions are expected to be accurate to within a factor of approximately 3, which is the
standard assumption for pharmacokinetic variability among humans.
It can be noted that the oral CSFs are larger when based on blood levels of CEO versus
blood levels of AN. The Vmaxc/Km for AN oxidation in humans is estimated to be about 10 times
higher than in the rat. AN enzymatic GSH conjugation is estimated to be 1.5 times higher in the
rat, but the 2nd-order removal constant is the same and the GSH tissue levels are approximately
the same (and lower in the larger tissue groups). The result, for example, with a steady oral
"infusion" of 30 mg/kg-day, is that AN is removed somewhat faster overall, leading to steady-
state blood level of 0.114 mg/L in the human vs. 0.151 mg/L in the rat. However, a much higher
portion of this goes to CEO in the human yielding a CEO blood level of 0.297 mg/L vs.
0.013 mg/L in the rat. Thus, the CEO:AN ratio is 0.088 in the rat and 2.6 in the human, so that
the CEO-based risk estimate is higher than the AN-based estimate.
Dose metric
AN is activated by CYP2E1 into its reactive metabolite, CEO, which binds to DNA and
initiates tumor formation. Therefore, AUC CEO concentration in blood was selected as the
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internal dosimetric for PBPK modeling. AN is not anticipated to be the causal agent for
carcinogenesis. Uncertainty in the risk estimate related to the dose metric is primarily associated
with PBPK model estimation.
Statistical uncertainty at the POD
Parameter uncertainty within the chosen model reflects the sample size of the cancer
bioassay. For the multistage model applied to this data set, there is a relatively small degree of
uncertainty at the BMDLio (the POD for linear low-dose extrapolation) which is approximately
1.4-fold lower than the BMDio (Appendix B, Table B-37). The highest value was selected in
order to provide a reasonable upper-bound risk estimate.
With regard to the composite risk estimate, under the assumption of independence of the
tumor type/site considered, no additional uncertainty is added to the estimated POD. Each
composite estimate is a statistically rigorous restatement of the statistical uncertainty associated
with the risk estimates derived from the individual sites. The only assumption in the combining
tumors approach is independence of tumors. This assumption is consistent with NRC (1994)
recommendations.
Choice of low dose extrapolation approach
The mode of action is a key consideration in deciding how risks should be estimated for
low-dose exposure. The mode of action for cancer effects of AN is discussed extensively in
Section 4.7.3.1 and is determined to be mainly due to mutagenicity. Other modes of action are
plausible, but the evidence suggests they may not be key to the carcinogenicity of AN. The
pattern of tumors is consistent with DNA-reactive chemicals. When the mode of action is
determined to be mutagenicity, a linear approach is used to estimate low-exposure risk, in
accordance with EPA's Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a).
Choice of species/gender
No human cancer epidemiology study via the oral route is available. Cancer risk is
estimated for both male and female rats, and cancer risk for the strain and gender with the
highest risk, Sprague-Dawley females, was selected. Mammary gland tumors were observed in
female rats of both studies. Although a cancer bioassay in B6C3Fi mice is also available, the
exposure was via gavage and no PBPK model is available for mice. It was assumed that humans
are as sensitive overall as rats, although the true correspondence is unknown. Site concordance
is not assumed, nor is it necessary (U.S. EPA, 2005a).
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Human population variability
Heterogeneity among humans is another source of uncertainty. Human genetic
polymorphisms in CYP2E1 activities likely contribute to variability in susceptibility to the toxic
effects of AN (see Section 4.8.4.1).
5.4.5.2. Inhalation Cancer Assessment
Uncertainties related to the IUR assessment are discussed below and summarized in
Table 5-26 and Figure 5-4.
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Table 5-26. Summary of uncertainty in the AN inhalation cancer risk
assessment
Consideration/
approach
Impact on cancer risk
estimate"
Decision
Justification
Choice of study
Cancer risk estimate could
increase because only lung
cancer deaths evaluated.
Lung cancer mortality and
exposure data from Blair et
al. (1998) study were used to
derive IUR.
IUR derived from best available
cancer epidemiological study of
AN-exposed workers has fewer
inherent uncertainties than IUR
derived from rat study (Quast et al.,
1980b).
Low-dose
extrapolation
procedure
The selected model does not
represent all possible
models one might fit, and
other models could
conceivably be selected to
yield more extreme results
consistent with the observed
data, both higher and lower
than those included in this
assessment.
Low-dose linear.
EPA's Guidelines for Carcinogen
Risk Assessment (U.S. EPA, 2005a):
mutagens "are assessed with a linear
approach."
Statistical
uncertainty at
POD
IUR would be reduced
fourfold if BMD used as
POD instead of BMDL.
BMDL, approach for
calculating reasonable upper
bound.
Limited size of study results in
sampling variability; BMCL is
lower 95% confidence limit on
ECoi.
Choice of
species/gender
IUR estimated from
exposed male and female
workers.
IUR estimated from entire
cohort was used.
No apparent gender difference
between male and female workers;
responses in male and female rats
were also similar.
Human population
variability in
metabolism and
response/sensitive
subpopulations
Low-dose risk for general
population can increase to
an unknown extent due to
underestimation of human
variability
Semi-parametric Cox
regression model with time-
dependent covariates was
used. The Cox model allows
inclusion of individual
exposure histories and
utilizes internal controls, thus
avoiding confounding by the
healthy worker effect.
Human variability in susceptibility
likely exists due to differences in
microsomal CYP2E1 activities, and
possibly GST activity.
Early-life
susceptibility
Lifetime cancer risk
increased by 1.6-fold.
Application of ADAFs of 10
(for individuals <2 years), 3
(for individuals 2 to <16
years), and 1 (for individuals
>16 years).
Individuals exposed during early
life to carcinogens with a mutagenic
MOA are assumed to be at
increased risk for cancer (U.S. EPA,
2005b). In the absence of adequate
chemical-specific data to evaluate
differences in age-specific
susceptibility, EPA's Supplemental
Guidance recommends the
application of ADAFs.
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0.100 -i
0.090
0.080
0.070
0.060
c 5 0.050
.2 i*
+- c
2 '—
•S 0.040
0.030 -
0.020
0.010-
0.000
0.070
IT.OU40
r
* IUR -CEO in Blood
• IUR - Human Equivalent Administered Concentration
0.060
6.0035
0.020
A
V
Male Sprague-Dawley Rats Female Sprague-Dawley Rats Humans Exposed
Occupationally
Sex/Strain/Species
Figure 5-3. Comparison of composite lURs derived from: (1) tumor
incidence data in male and female Sprague-Dawley rats exposed chronically
to AN, and (2) lung cancer mortality in humans exposed to AN
occupationally. In deriving the animal-based lURs, two different dose
metrics were employed: (1) predicted CEO concentration in blood, and (2)
human equivalent administered AN concentration in air.
Choice of study
As mentioned previously, the Blair et al. (1998) cohort study is the largest cohort
assessment of the relationship between AN exposure and cancer. This study had the distinct
advantage of quantifying exposure and using an internal control group of unexposed workers, all
factors identified as shortcomings in previous studies. Information on smoking history, a
potential confounder when lung cancer is the outcome of interest, was available, but only for a
small subset within the Blair et al. (1998) cohort. Blair et al. (1998) limited their analysis to the
white male population and noted that adjustment for smoking reduced the risk of lung cancer
slightly. The data used in this assessment to derive the IUR included the entire cohort;
consequently, the smoking history data were incomplete for this analysis, leading to an area of
uncertainty surrounding this risk estimate.
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Another source of uncertainty stems from the use of only lung cancer mortality for the
derivation of the IUR. As previously described, other studies reported small and not consistent
excess risks for other types of cancer (prostate, bladder, brain cancers) associated with AN
exposure. Thus, basing the IUR on only lung cancer mortality may underestimate the
carcinogenic potential of AN. Further, as the IUR is based on mortality rather than lung cancer
incidence, there may be potential for underestimation of carcinogenic potential with this
approach. In the case of lung cancer, however, mortality is a good surrogate of lung cancer
incidence. Additional uncertainties related to the statistical analysis of these data are further
discussed in Appendix B-7. Briefly, uncertainties of the statistical approach include: (1) Cox
model that was fit to the data is not a biologically based model; (2) the estimator of the
cumulative hazard does not account for the covariate path and hence is only an approximation;
and (3) the estimate of risk is obtained using the first-order "linearized" approximation.
However, obtained results are consistent with assumptions of the first-order approximation
validity.
Other uncertainties associated with the NCI/NIOSH cohort include nondifferential
exposure misclassification, lung cancer misdiagnosis among the internal controls, the relatively
short follow-up period (reflected by the relatively small proportion of mortality within the
cohort), and the extrapolation of continuous environmental exposure from 8-hour occupational
exposure without consideration of potential recovery mechanisms between daily exposures. The
nondifferential exposure misclassification contributes to an underestimation of risk, while the
impacts of short follow-up and extrapolation to a continuous exposure scenario are unclear.
Nonetheless, the NCI/NIOSH cohort provides the most robust data set in terms of sample size
and exposure assessment for the derivation of the IUR.
Statistical uncertainty at the POD
Parameter uncertainty within the chosen model reflects the limited sample size of the
cancer bioassay. For the results relying on the cohort study, the ECoi is approximately fourfold
higher than the LECoi.
For the multistage model applied to the rat data, in support of the human-based results,
there is a reasonably small degree of uncertainty at the 10% incidence level (the POD for linear
low-dose extrapolation). Composite BMCios for overall cancer risk are approximately 1.3-fold
higher than their corresponding BMCLioS for both male and female rats.
With regard to the composite risk estimate, under the assumption of independence of the
tumor type/site considered no additional uncertainty is added to the estimated POD. Each
composite estimate is a statistically rigorous restatement of the statistical uncertainty associated
with the risk estimates derived from the individual sites. This assumption is consistent with NRC
(1994) recommendations.
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Low-dose extrapolation procedure
As discussed previously, the key mode of action for carcinogenicity of AN is determined
to be mutagenicity. A linear approach is used to estimate low-exposure risk, in accordance with
EPA's Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a). A nonlinear low-dose
extrapolation approach is not used since data do not support other modes of action at this time.
Choice of species/gender
A human study was selected for quantification of the IUR. Both male and female
workers were evaluated. A 2-year inhalation study of male and female rats was used for
comparison. Estimates of lURs from the human study and from the rat studies were similar.
lURs estimated from male and female rats were also similar, although there was a suggestion of
higher cancer risk in females associated with mammary gland tumors.
Human population variability
Heterogeneity among humans is another source of uncertainty. See the discussion of
human population variability in Section 5.4.5.1.
5.4.6. Previous Cancer Assessment
In the previous IRIS assessment completed in 1991, EPA derived CSFs derived from
each of these three studies, i.e., 4 x 10"1 (mg/kg-day)"1 (Biodynamics, 1980a), 4 x 10"1 (mg/kg-
day)"1 (Biodynamics, 1980c), and 10 x 10"1 (mg/kg-day)"1 (Quast et al., 1980a).
Also in the previous IRIS assessment, an IUR was derived based on human data. This
IUR was based on an increase in lung cancer incidence in humans occupationally exposed to
AN, as described by O'Berg (1980). A value of 6.8 x 10"5 per ug/m3 was derived by the
application of an RR model, which was adjusted for smoking and was based on a continuous
lifetime equivalent of occupational exposure to AN.
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6. MAJOR CONCLUSIONS IN THE CHARACTERIZATION OF HAZARD AND
DOSE-RESPONSE
6.1. HUMAN HAZARD POTENTIAL
AN (CASRN 107-13-1) is a colorless, flammable, and volatile liquid with a weakly pungent
onion- or garlic-like odor. It is a commercially important chemical used in the manufacture of
acrylic and modacrylic fibers, plastics (ABS and AN-styrene resins), and nitrile rubbers and as
an intermediate in the synthesis of other chemicals, such as adiponitrile and acrylamide.
Exposure to airborne AN is possible for people living in the vicinity of emission sources such as
acrylic fiber or chemical manufacturing plants or waste sites.
AN is rapidly and nearly completely absorbed, widely distributed to tissues, and
biochemically transformed into metabolites that are excreted in the urine and, to a much lesser
extent, in feces and expired air. Two major metabolic pathways for AN have been identified:
detoxification of AN by conjugation with GSH, forming the urinary metabolite N-acetyl-S-
(2-cyanoethyl)cysteine, and oxidation of AN to its epoxide metabolite, CEO, via CYP2E1. CEO
can bind to tissue macromolecules, such as proteins and DNA. CEO can be hydrolyzed to
cyanide with further transformation to thiocyanate, which is excreted in urine. Alternatively,
CEO can interact with GSH, resulting in the formation of a number of other urinary metabolites.
There are no studies directly identifying health hazards in humans following oral
exposures of any duration, but results from a robust array of studies in rats and mice identify
noncancer lesions in the gastric squamous epithelium and tumors in multiple tissues as potential
health hazards to humans exposed to AN by the oral route for chronic durations. Results from
cross-sectional epidemiologic studies of AN-exposed workers identify increased prevalences of
neurological symptoms and adverse reproductive outcomes in occupationally exposed workers as
potential health hazards from chronic inhalation exposure. Cancer is another potential human
health hazard from chronic inhalation exposure, as indicated by increased incidences of tumors at
several tissue sites in rat chronic inhalation bioassays. Studies of AN-exposed workers have
provided no strong evidence that mortality from any type of cancer is casually related to
occupational exposure, but limited evidence from the best-designed epidemiologic study found a
small, but statistically significant, increased risk for dying from lung cancer in workers with the
longest durations and highest exposures to AN.
The animal toxicity database identifies hyperplasia and hyperkeratosis of the squamous
epithelium of the forestomach as the most sensitive noncancer effects associated with repeated
oral exposure to AN. Following chronic oral exposure, these lesions have been observed in rats
at drinking water concentrations as low as 1-3 ppm (0.09-0.3 mg/kg-day) (Johannsen and
Levinskas, 2002a, b; Biodynamics, 1980a, b). Other effects have been observed in repeatedly
exposed animals, generally at higher exposure levels. These include ovarian atrophy in female
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mice exposed to doses >2.5 mg/kg-day for 2 years (NTP, 2001); chronic nephropathy in male
and female rats exposed for 2 years to 3.4 and 10.8 mg/kg-day AN, respectively (Quast et al.,
1980a); gliosis in the brain of female rats at 4.4 mg/kg-day (Quast et al., 1980a); decreased
sperm count in male mice exposed to 10 mg/kg-day for 60 days (Tandon et al., 1988); hind-limb
weakness and decreased sensory nerve conduction velocity in male rats exposed to 50 mg/kg-day
AN for 12 weeks (Gagnaire et al., 1998); and neurobehavioral effects in male rats exposed to 4
mg/kg-day AN in drinking water for 8 or 12 weeks (Rongzhu et al., 2007). No changes in
fertility index or pregnancy success were found in a three-generation study of rats exposed to
drinking water doses as high as 39 mg/kg-day (Friedman and Beliles, 2002; Litton Bionetics,
1992). Mild developmental effects were observed at 11 and 20 mg/kg-day (small deficits in
postnatal pup weight or survival) in this three-generation rat study and at 25 mg/kg-day
(increased litters with pups with missing vertebrae) but not at 10 mg/kg-day in rat fetuses
exposed on GDs 6-15 (Murray et al., 1978).
Repeated inhalation exposure to AN in the workplace has been associated with increased
prevalence of subjective neurological symptoms, such as headache, poor memory, and irritability
(Chen et al., 2000; Kaneko and Omae, 1992; Muto et al., 1992; Sakurai et al., 1978) and small
performance deficits in neurobehavioral tests of mood, attention and speed, auditory memory,
visual perception and memory, and motor steadiness (Lu et al., 2005a). Such effects have been
associated with average workplace air concentrations of 0.1 or 0.9 ppm and appear to be the most
sensitive noncancer effects from repeated inhalation exposure to AN. Adverse reproductive
outcomes, such as increased prevalences of premature deliveries, stillbirths, sterility, birth
defects, and pregnancy complications, have been associated with occupational exposure to
average workplace concentrations ranging from 3.6 to 7.5 ppm (Dong et al., 2000a; Li, 2000).
Subchronic and chronic inhalation toxicity studies in rats identified other noncancer effects at
higher exposure levels, including nasal epithelial lesions in rats exposed to concentrations of 20
or 80 ppm for 2 years (Quast et al., 1980b), decreased nerve conduction velocity and hind-limb
weakness in rats exposed to >25 ppm AN (Gagnaire et al., 1998), increased incidence of rat
fetuses with missing vertebrae, missing ribs, or anteriorly displaced ovaries following exposure
of pregnant rats to 80 ppm on GDs 6-15 (Murray et al., 1978), and decreased rat fetal weight
gain following exposure of pregnant rats to concentrations >25 ppm on GDs 6-20 (Saillenfait et
al., 1993).
The genotoxicity of AN has been evaluated in multiple systems in vitro and in vivo. In
addition to positive findings in blood lymphocytes, buccal mucosal cells, and sperm in five
epidemiologic studies, DNA alkylation by AN was found in numerous tissues in rats or mice
(brain, liver, testes, forestomach, colon, kidney, bladder, and lung) treated with a single dose of
AN. AN or its reactive metabolite, CEO, yielded positive results in in vitro mutation assays
using bacteria, fungi, and insects, as well as animal and human cell cultures. The weight of
evidence from these studies suggests that AN is mutagenic after metabolic activation to CEO.
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Following EPA Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a), AN is
"likely to be carcinogenic to humans," based predominantly on consistent results showing that
lifetime inhalation or oral exposure caused statistically significantly increased incidence of
tumors at multiple tissue sites in rats and mice. Lifetime oral exposure to AN caused increased
incidences of tumors at multiple tissue sites, including the brain, forestomach, and Zymbal gland,
in several rat studies and the forestomach and Harderian gland in a gavage study in mice (NTP,
2001). Lifetime inhalation bioassays with Sprague-Dawley rats found exposure-related increases
in the incidences of brain tumors, Zymbal gland tumors, intestinal tumors, tongue tumors, and
malignant mammary gland tumors (Dow Chemical Co., 1992a; Quast et al., 1980b). Also, there
is some evidence for an association between AN exposure and lung cancer deaths in
occupationally exposed workers (notably Blair et al., 1998).
Although data gaps still exist in the current understanding of the mode of action for
carcinogenicity of AN, there is experimental evidence to support a mutagenic mode of action as
the key mode of action for AN-induced tumors. Other modes of action may contribute, but
limited data do not appear supportive at this time. Mutagenicity via oxidative DNA damage is
plausible, but oxidative stress was not supported by experimental evidence as a key mode of
action.
6.2. DOSE-RESPONSE
6.2.1. OralRfD
The available oral toxicity studies in animals identify nonneoplastic forestomach lesions
(i.e., squamous cell epithelial hyperplasia and hyperkeratosis) as the most sensitive noncancer
effect associated with chronic oral exposure to AN. These lesions are expected to be relevant to
humans because, although humans do not possess a forestomach, they do have comparable
squamous cell epithelial tissue in their oral cavity and in the upper two-thirds of their esophagus.
A 2-year drinking water study with F344 rats (Johanssen and Levinskas, 2002b) was
selected as the principal study on which to base the RfD. This study included five drinking water
exposure levels ranging from 1 to 100 ppm, and it identified the lowest administered-dose
LOAEL of the available chronic animal toxicity studies based on an increased incidence of
forestomach lesions (i.e., LOAELs of 0.3 and 0.4 mg/kg-day for males and females, respectively,
with associated NOAELs of 0.1 mg/kg-day for both sexes).
An RfD of 3 x 10"4 mg/kg-day is derived based on incidence of nonneoplastic
forestomach lesions (hyperplasia or hyperkeratosis) in male and female F344 rats in the 2-year
drinking water study of AN by Johanssen and Levinskas (2002b). This RfD was derived using a
BMD approach and application of PBPK modeling (Section 3.5; Appendix C; Sweeney et al.,
2003;Kedderisetal., 1996).
The RfD derivation process involved first fitting all available dichotomous models in
BMDS (version 2.0) to the incidence data for male and female rats, separately, employing two
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different internal dose metrics (i.e., AN in blood and CEO in blood). These two internal dose
metrics were derived by converting administered rat doses of AN to internal rat doses (either AN
or CEO in blood) using the rat PBPK model of Kedderis et al. (1996), as modified by EPA
(Section 3.5; Appendix C). Then, for each sex, the BMDL associated with a 10% extra risk for
gastric epithelial lesions was derived based on the best-fitting model. This BMDL based on rat
internal dose was then converted to the human equivalent administered dose of AN by using the
human PBPK model of Sweeney et al. (2003) (with EPA-modified parameters; Section 3.5;
Appendix C). As discussed in more detail in Section 5.1, CEO in blood was selected as the best
available internal dose metric for cross-species extrapolation with oral exposure. The human
equivalent administered dose of AN represents a POD for noncancer effects and was then
divided by a UF of 30 (3 to account for uncertainty in extrapolating from rats to humans with
dosimetric adjustment and 10 to account for variation in response from average humans to
sensitive humans) to arrive at an RfD.
Confidence in the principal study selected for the RfD is high. The principal study,
Johannsen and Levinskas (2002b), was selected from eight chronic rat studies and one gavage
study in exposed mice. The study employed five exposure levels of AN, ranging from 1 to
100 ppm, and thus provided a relatively complete description of the dose-response relationship in
the low-dose region. Johannsen and Levinskas (2002b) identified the lowest LOAELs based on
an increased incidence of hyperplasia and hyperkeratosis in squamous epithelium of forestomach
(0.3 mg/kg-day for males and 0.4 mg/kg-day for females). Confidence in the AN database is
high. The database for ingested AN includes nine rat toxicity and cancer bioassays, one toxicity
and cancer bioassay with B6C3Fi mice, a three-generation (46-week) developmental/
reproductive toxicity study with Sprague-Dawley rats, a 12-week gavage study of nerve
conduction velocities in male Sprague-Dawley rats, a 14-week gavage toxicity bioassay in
B6C3Fi mice, a developmental toxicity study in Sprague-Dawley rats exposed by gavage during
GDs 6-15, and a 90-day study of oxidative stress indicators in the brain and liver of F344 rats. .
Overall confidence in the RfD is high reflecting these considerations.
6.2.2. Inhalation RfC
Results from several cross-sectional epidemiologic studies of AN-exposed workers
identified increased prevalence of neurological symptoms and small performance deficits in
neurobehavioral tests as the critical effects resulting from chronic inhalation exposure to AN.
The cross-sectional study of neurobehavioral performance measures in acrylic fiber workers by
Lu et al. (2005a) was selected as the principal study for RfC derivation because it identified the
lowest reliable exposure level in humans associated with adverse neurological effects.
A NOAEL/LOAEL approach was used to derive the RfC from the human data. The
average workplace AN air concentration of 0.11 ppm for workers in the monomer work areas of
the acrylic fiber plant was selected as the LOAEL or POD for RfC derivation. At the LOAEL of
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0.11 ppm (0.24 mg/m3), small, but statistically significant, performance deficits in
neurobehavioral tests of mood, attention and speed, auditory memory, visual perception and
memory, and motor steadiness were observed. The LOAEL was converted to an equivalent
continuous exposure of 0.086 mg/m3 and was divided by a composite UF of 100 (10 for
extrapolating from a LOAEL to a NOAEL and 10 to account for extrapolating from healthy
workers to sensitive humans) to arrive at an RfC for AN of 0.9 ug/m3.
As discussed in more detail in Section 5.2, comparative animal-based RfCs for AN of
3 x 10~3 mg/m3 (or 3 ug/m3) and 2 x 10"3 mg/m3 (or 2 ug/m3) were derived based on PODs from
BMD modeling of nasal lesions observed in male and female rats, respectively, exposed to AN
via inhalation for 2 years (Quast et al., 1980b). In deriving these RfCs, the PODs were divided
by a composite UF of 30 (3 for extrapolating from rats to humans using the default U.S. EPA
(1994) dosimetric adjustment and 10 to account for variation from average humans to sensitive
humans). These animal-based RfCs are consistent with the human-based value. However, the
human-based value of 0.9 ug/m3 is selected as the RfC, because extrapolating from animals has
greater associated uncertainty than extrapolating from humans.
The principal study is given medium confidence because it is the best available study
that identified neurobehavioral effects of AN in occupationally exposed workers. Lu et al.
(2005a) utilized the WHO-recommended NCTB administered by trained physicians to
systemically evaluate neurobehavioral effects, whereas previous occupational studies by Kaneko
and Omae (1992) and Muto et al. (1992) reported subjective neurological symptoms in exposed
workers. Hence, the results of Lu et al. (2005a) were more reliable when compared with those
based on self reporting. The confidence in the principal study is medium because there are
several limitations in the study. One was that the cited exposure data represented estimates of
previous exposure levels and no contemporaneous personal monitoring data were available. In
addition, the study authors could not rule out the possibility that examiner drift may have
affected the results. Moreover, the largest measures of neurobehavioral effect occurred in the
acrylic fiber workers, who had a lower average exposure level.
Confidence in the database is high. Like the oral toxicity database, the inhalation
database for AN is robust, consisting of eight occupational exposure studies that evaluated the
noncancer health effects of AN on workers exposed via inhalation. Three of these studies
evaluated reproductive endpoints in AN-exposed workers. The database also includes one
chronic inhalation toxicity study in male and female Sprague-Dawley rats; one two-generation
reproductive toxicity study of inhaled AN vapors in Crl:CD (SD) rats; one 24-week nerve
conduction velocity study in male rats, and 2 developmental studies in rats exposed from GD 6-
15 or GD 6-20. An RfC based on the results from a chronic inhalation study with Sprague-
Dawley rats was also derived for comparison. Statistically significant increased incidence of
inflammatory and degenerative nasal lesions occurred in rats exposed to the lowest level in this
2-year bioassay. The alternative RfC derived from the rat study is only about threefold higher
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than the RfC derived from the occupational exposure study. Overall confidence in the RfC is
medium, reflecting these considerations.
6.2.3. Oral Slope Factor
Incidence data for forestomach, CNS, Zymbal gland, tongue, and mammary gland tumors
in male and female Sprague-Dawley (Quast, 2002) and F344 (Johanssen and Levinskas, 2002a)
rats were used to develop site-specific oral CSFs for AN, employing EPA-modified rat and
human PBPK models for cross-species dosimetric extrapolation (Section 3.5; Appendix C;
Sweeney et al., 2003; Kedderis et al., 1996). These animal studies were selected for the
development of oral CSFs because oral human data are not available, and these studies are the
best available chronic bioassays for characterizing the dose-response relationships for these
AN-induced tumors. Weight-of-evidence evaluation following U.S. EPA (2005a) guidelines
determined that a mutagenic mode of action, most likely via the reactive AN metabolite CEO,
was the principal mode of action. Consequently, a linear low-dose extrapolation approach was
used in the development of the oral CSFs.
As discussed in more detail in Section 5.4, CEO in blood and AN in blood were both
evaluated as internal dose metrics for use in cross-species extrapolation with oral exposure. The
multistage dose-response model in BMDS (version 1.4.1) was fit to the male and female rat
tumor incidence data, using internal animal dose expressed as either CEO or AN in blood. Rat
administered doses were converted to rat internal doses (either CEO or AN in blood) by using
the rat PBPK model of Kedderis et al. (1996), as modified by EPA (Section 3.5; Appendix C).
For each of these two dose metrics, the resulting best-fit model for each endpoint was then used
to derive a 95% lower confidence limit on the dose associated with 10% extra risk (i.e., a
BMDLio). These BMDLs, based on internal rat doses, were then converted to human equivalent
administered doses of AN using the human PBPK model of Sweeney et al. (2003), as modified
by EPA (Section 3.5; Appendix C). The site-specific oral CSFs based on incidence data from
each tumor site were derived by linear extrapolation from the human equivalent administered
doses down to the origin (oral CSF = 0.1/BMDLio/HEo)- Within rat strain and sex, an CSF for the
composite risk across these tumor sites, based on CEO in blood, was then estimated by
employing the procedure described in Section 5.4.4.1.
The highest composite CSF of 5 per mg/kg-day is recommended for use in humans
because it is the value obtained from the most sensitive species, sex, and strain (i.e., incidence of
tumors in female Sprague-Dawley rats). This slope factor should not be used with exposures
greater than 0.04 mg/kg-day (the lowest POD supporting the composite risk) because above this
level, the slope factor cannot be expected to be an adequate approximation to the dose-response
relationship. The fitted dose-response relationship and pharmacokinetic models should be used
to estimate risk above this exposure level.
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Because a mutagenic mode of action for acrylonitrile carcinogenicity is sufficiently
supported in laboratory animals and relevant to humans, and in the absence of adequate chemical-
specific data to evaluate differences in susceptibility, increased early-life susceptibility is assumed.
Accordingly, early-life susceptibility factors or ADAFs should be applied to the OSF when assessing
cancer risk associated with early-life exposures (i.e., birth to 16 years) in accordance with the
Supplemental Guidance for Assessing Susceptibility from Early-Life Exposure to Carcinogens (U.S.
EPA, 2005b) (see Section 5.4.4.3).
6.2.4. Cancer Inhalation Unit Risk
An analysis of the human lung cancer mortality data from the Blair et al. (1998) cohort
study of AN-exposed workers was conducted to derive an IUR estimate for AN based on human
data. This analysis employed the approach presented by Starr et al. (2004) and is further
described in Appendix B-7. In brief, the risk of death from lung cancer in AN-exposed workers
was characterized by using a semi-parametric Cox regression model with a cumulative exposure
metric (i.e., ppm-working years) as the only time-dependent covariate. In contrast to the analysis
by Starr et al. (2004), the entire cohort was included in the analysis conducted for this assessment
(not just white male workers), and the final model only included cumulative exposure as the
covariate.
The Cox regression model was used to estimate an AN exposure level and its associated
95% lower confidence limit corresponding to a 1% risk of dying from lung cancer by age 80
(i.e., ECoi and LECoi, respectively). Conversion of occupational exposures to continuous
environmental exposures was accomplished by adjusting for differences in the amount of air
inhaled during an 8-hour workday versus a 24-hour day (10 and 20 m3/day, respectively). The
IUR estimate was derived by linear extrapolation from the LECoi. The predicted ECoi and
LECoi derived from the Cox regression model based on the Blair et al. (1998) data were 0.992
and 0.238 ppm (2,168 and 524 ug/m3) AN, respectively. From the LECoi, an IUR estimate of
0.042 per ppm (2 x 10"5 per ug/m3) was derived.
Some uncertainty is associated with this IUR estimate because of the study on which it is
based. More specifically, no adjustment for smoking was applied in the Blair et al. (1998) study.
The investigators collected smoking information on only about 10% of the exposed and
unexposed members of the cohort and found similar incidences of smoking in the two groups.
Because their statistical analysis compared exposed and unexposed groups of workers, Blair et
al. (1998) noted that the adjustment for smoking made only a slight difference in the results of
their analysis. Other uncertainties associated with the Blair et al. (1998) data set included
nondifferential exposure misclassification and a relatively short follow-up period, an average of
21 years. Additionally, the outcome of the study was lung cancer mortality, not lung cancer
incidence. Additional uncertainties related to the statistical analysis of these data are discussed
further in Appendix B-7.
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Tumor incidence data from the best available animal inhalation study (Dow Chemical
Co., 1992a) were selected for describing the dose-response relationship between intestinal, CNS,
Zymbal gland, tongue, and mammary gland tumors and AN exposure. These dose-response
relationships were used to derive animal-based lURs for AN for comparative purposes. As with
the animal-based CSF, a mutagenic mode of action, most likely via the reactive AN metabolite
CEO, was assumed. Consequently, a linear low-dose extrapolation approach was used in the
development of the animal-based lURs.
Based on the assumption that the epoxide metabolite, CEO, is critical to AN's
carcinogenic mode of action, the selected internal dose metric was CEO in blood. In contrast to
oral exposure, the EPA-modified PBPK model adequately predicted measured blood and brain
concentrations of CEO in rats exposed to AN by inhalation (Kedderis et al., 1996). The
multistage model was fit to the rat tumor incidence and CEO concentration in blood predicted by
the PBPK model of Kedderis et al. (1996), as modified by EPA (Section 3.5; Appendix C). The
best-fitting stage of the model was used to derive 95% lower bounds on rat internal blood
concentrations of CEO associated with 10% extra risk (BMCLios). The human PBPK model of
Sweeney et al. (2003) (parameters modified; Section 3.5; Appendix C), was then used to
calculate human equivalent administered concentrations of AN, corresponding to the rat
BMCLios. These human equivalent administered concentrations of AN were used as the PODs
for the IUR estimates derived via linear extrapolation down to the origin (i.e., IUR =
0.1/BMCL10/HEc).
The IUR estimates for multiple tumor sites were derived within each rat sex by
employing the procedure described in Section 5.4.4.2. The resulting composite IUR estimates
were 7 x 10"2 and 6 x 10"2 per mg/m3 (7 x 10"5 and 6 x 10"5 per ug/m3) and were derived based on
tumor incidence data from male and female Sprague-Dawley rats, respectively. These unit risks
should not be used with exposures greater than 3 mg/m3 (the lowest POD supporting the
composite risk), because above this level, the IUR cannot be expected to be an adequate
approximation of the dose-response relationship. The fitted dose-response relationship and
pharmacokinetic models should be used to estimate risk above this exposure level.
The IUR of 2 x 10"5 per ug/m3 derived from human data is chosen as the IUR for AN
over the alternative IUR value derived from animal inhalation bioassay data because the use of
human study data eliminates the uncertainty inherent in animal to human extrapolation. This
value is consistent with the alternative IUR derived from the animal data.
As described for the OSF, early-life susceptibility factors or ADAFs should be applied to
the IUR when assessing cancer risks associated with early-life exposures (i.e., birth to 16 years) in
accordance with the Supplemental Guidance for Assessing Susceptibility from Early-Life Exposure to
Carcinogens (U.S. EPA, 2005b) (see Section 5.4.4.3).
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7. REFERENCES
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3 91 DRAFT- DO NOT CITE OR QUOTE
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APPENDIX A. SUMMARY OF EXTERNAL PEER REVIEW AND PUBLIC
COMMENTS AND DISPOSITION
[page intentionally left blank]
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APPENDIX B. BENCHMARK DOSE CALCULATIONS
APPENDIX B-l. NONCANCER ORAL DOSE-RESPONSE ASSESSMENT (RfD):
BENCHMARK DOSE MODELING RESULTS EMPLOYING THE INCIDENCE OF
FORESTOMACH LESIONS (HYPERPLASIA AND HYPERKERATOSIS) IN MALE
AND FEMALE SPRAGUE-DAWLEY RATS, F344 RATS, AND B6C3Fi MICE
CHRONICALLY EXPOSED ORALLY TO AN FOR 2 YEARS
As previously discussed in Section 4, no human studies currently exist that involve oral
exposures to AN. The available animal oral toxicity data, however, identify forestomach lesions
(i.e., squamous cell epithelial hyperplasia and hyperkeratosis) as the most sensitive, prevalent,
and consistent noncancer effect associated with chronic oral exposure to AN. Therefore, this
endpoint was selected as the critical effect on which to base derivation of the RfD.
Two 2-year drinking water studies, one in Sprague-Dawley rats (Quast, 2002; Quast et
al., 1980a) and the other in F344 rats (Johannsen and Levinskas, 2002b; Biodynamics 1980c),
and a 2-year gavage study in B6C3Fi mice (NTP, 2001) provided the best available dose-
response data on which to base the RfD. Candidate RfDs for AN were derived from the
incidence of forestomach lesions in Sprague-Dawley rats, F344 rats, and B6C3Fi mice using a
BMD approach. Incidences of forestomach lesions from chronic drinking water studies in male
and female Sprague-Dawley rats (Quast, 2002) and F344 rats (Johannsen and Levinskas, 2002b)
provided four sets of dose-response data from which to derive candidate RfDs, while incidences
of forestomach lesions from a chronic gavage study in male and female B6C3Fi mice provided
an additional two sets of dose-response data. These six data sets are presented in Tables B-l (for
rats) and B-2 (for mice).
Table B-l. Incidences of forestomach lesions (hyperplasia or
hyperkeratosis) in Sprague-Dawley and F344 rats exposed to AN in drinking
water for 2 years
Sex
Administered
concentration
(ppm in drinking
water)
Administered
dose"
(mg/kg-d)
Predicted internal dose metrics1"
AN-AUC in rat
blood
(mg/L)
CEO-AUC in rat
blood
(mg/L)
Incidence of
forestomach
lesions0
Sprague-Dawley rats
(Sources: Quast, 2002; Quast et al., 1980a)
Male
Female
0
35
100
300
0
35
0
3.4
8.5
21.3
0
4.4
0
2.06 x 10'2
5.36 x 10"2
1.46 x 10'1
0
2.37 x 10"2
0
1.83 x 10'3
4.36 x 10"3
9.70 x 10'3
0
2.07 x 10"3
15/80 (19%)
15/47 (32%)
44/48 (92%)c
45/48 (94%)c
20/80 (25%)
23/48 (48%)c
B-l
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Table B-l. Incidences of forestomach lesions (hyperplasia or
hyperkeratosis) in Sprague-Dawley and F344 rats exposed to AN in drinking
water for 2 years
Sex
Administered
concentration
(ppm in drinking
water)
100
300
Administered
dose3
(mg/kg-d)
10.8
25.0
Predicted internal dose metrics1"
AN-AUC in rat
blood
(mg/L)
6.18 x 10"2
1.56 x 10'1
CEO-AUC in rat
blood
(mg/L)
4.87 x 10"3
1.01 x 10'2
Incidence of
forestomach
lesions0
41/48 (85%)c
47/48 (98%)c
¥344 mtsd
(Sources: Johannsen and Levinskas, 2002b; Biodynamics, 1980c)d
Male
Female
0
1
o
5
10
30
100
0
1
o
5
10
30
100
0
0.08
0.25
0.83
2.48
8.37
0
0.12
0.36
1.25
3.65
10.90
0
4.33 x 10"4
1.35 x 10"3
4.52 x 10'3
1.37 x 10"2
4.85 x 10'2
0
5.73 x 10'4
1.72 x 10"3
6.02 x 10'3
1.79 x 10"2
5.63 x 10'2
0
4.06 x 10"5
1.27 x 10"4
4.19x 10'4
1.23 x 10"3
3.97 x 10'3
0
5.32 x 10'5
1.59 x 10"4
5.49 x 10'4
1.58 x 10"3
4.46 x 10'3
11/159(7%)
3/80 (4%)
18/75 (24%)c
13/80 (16%)c
17/80 (22%)c
9/77 (12%)
4/156 (3%)
2/80 (3%)
16/80 (20%)c
23/74 (3 1%)C
13/80 (16%)c
5/74 (7%)
"Administered doses were averages calculated by the study authors based on animal B W and drinking water intake.
bThe EPA-modified rat PBPK model of Keddaris et al. (1996) was employed to predict a rat internal dose (i.e.,
either AN-AUC or CEO-AUC concentration in blood) resulting from the ingestion of the specified administered
dose of AN consumed in six bolus episodes/d.
Indicates significantly different (at/> < 0.05) from control incidence by Fisher's exact test performed by Syracuse
Research Corporation.
Incidences for F344 rats do not include animals from the 6- and 12-mo sacrifices and were further adjusted to
exclude (from the denominators) rats that died between 0 and 12 mos in the study. Rats dying during this time
period were determined from page 6 of Appendix H and Table 1 in Biodynamics (1980c) and Table 8 in Johannsen
and Levinskas (2002b). Unscheduled deaths between 0 and 12 mos in the study occurred in two female controls,
two males at 3 ppm, three females at 10 ppm, and three males and three females at 100 ppm.
B-2
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Table B-2. Incidences of forestomach lesions (hyperplasia or
hyperkeratosis) in male and female B6C3Fi mice administered AN via
gavage for 2 years
Lesion site and type
Dose (mg/kg-d)a
0
2.5
10
20
Males
Forestomach hyperplasia or hyperkeratosis
2/50
(4%)
4/50
(8%)
10/503
(20%)
13/50b
(26%)
Females
Forestomach hyperplasia or hyperkeratosis
2/50
(4%)
2/50
(4%)
5/50
(10%)
8/50a
(16%)
""Significantly elevated above vehicle control as determined by EPA using Fisher's exact test (p < 0.05).
bSignificantly elevated above vehicle control as determined by EPA using Fisher's exact test (p < 0.01).
Source: NTP(2001).
The incidences of forestomach lesions observed following 2 years of AN exposure in
male and female SD and F344 rats were modeled using AN and CEO in blood, expressed in
mg/L, as internal dose metrics. In addition, incidences of these same lesions were modeled in
male and female SD and F344 rats, as well as male and female B6C3Fi mice, employing
administered dose. In all cases, all of the dichotomous dose-response models available in EPA's
BMDS software (version 2.0) (i.e., the gamma, logistic, log-logistic, probit, log-probit,
multistage, Weibull, and quantal-linear models) were fit to these incidence data. Because the
incidence of forestomach lesions in male and female F344 rats did not increase monotonically
across all administered concentrations, however, only incidence data from the three lowest
concentrations (i.e., 0, 1, and 3 ppm) were used in dose-response modeling. Similarly, in male
Sprague-Dawley rats, the incidence data from the highest dose group needed to be dropped prior
9
to BMD modeling. In most cases, several models fit the data equally well (i.e., exhibited x
goodness-of-fit/> values greater than 0.1). Of those models exhibiting adequate fit, the selected
model was the one with the lowest AIC value. BMDLio and BMDLos estimates were derived
from the selected model. If more than one model shared the lowest AIC, the mean BMDLio and
BMDLos were calculated, as per the EPA's Benchmark Dose Technical Guidance Document
(U.S. EPA, 2000b).
In this appendix, results of the dose-response modeling for forestomach lesions in male
and female SD and F344 rats, and B6C3Fi mice are presented. For each species, strain, and sex,
summaries of the dose-response data (i.e., animal administered and internal doses and incidence
data) are presented (Table B-l for rats and Table B-2 for mice). Then, again for each species,
strain, and sex, Tables B-3 (Sprague-Dawley rats), B-4 (F344 rats), and B-5 (B6C3Fi mice)
summarize the results of the dose-response modeling. Each of these tables is then followed by
B-3
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the standard output from EPA's BMDS, version 2.0, for the dose-response models that resulted
in the lowest AIC for each endpoint (presented for the 10% BMR outputs).
Table B-3. Summary of the BMD modeling results based on the incidence of
forestomach lesions (hyperplasia or hyperkeratosis) in male and female
Sprague-Dawley rats exposed to AN in drinking water for 2 years
Sex
Male
Female
Endpoint
Forestomach
lesions
Forestomach
lesions
Dose metric
Estimated
administered
dose
(mg/kg-d)
Predicted AN
in blood
(mg/L)
Predicted CEO
in blood
(mg/L)
Estimated
administered
dose
(mg/kg-d)
Predicted AN
in blood
(mg/L)
Predicted CEO
in blood
(mg/L)
Selected model(s)a
two-stage multistage
(1)
two-stage multistage
(1)
two-stage multistage
(1)
gamma
logistic
log-logistic
log-probit
one-stage multistage
Weibull
gamma
logistic
log-logistic
log-probit
one-stage multistage
Weibull
gamma
logistic
log-logistic
probit
log-probit
%2
/7-value
0.18
0.24
0.12
0.38
0.45
0.98
0.74
0.31
0.31
0.32
0.23
0.98
0.70
0.42
0.28
0.51
0.69
0.87
0.33
0.86
AIC
169.45
169.01
170.09
212.77
211.31
212.03
212.14
212.40
213.07
212.96
212.00
212.03
212.18
211.79
213.19
212.45
210.72
212.06
211.86
212.06
BMDL10b
1.27
7.72 x 10 3
6.82 x 10 4
6.87 x 10'1
1.24
1.41
1.38
6.30 x 10"1
6.74 x 10'1
3.76 x 10'3
7.07 x 10"3
6.93 x 10'3
7.02 x 10"3
3.62 x 10 3
3.72 x 10'3
3.53 x 10"4
5.58 x 10"4
7.25 x 10"4
5.87 x 10"4
7.17x 10'4
BMDLosb
7.20 x 10'1
4.32 x 10 3
3.90 x 10 4
3.34 x 10'1
6.39 x 10'1
9.17 x 10"1
9.61 x 10'1
3.07 x 10"1
3.28 x 10'1
1.83 x 10'3
3.64 x 10"3
4.36 x 10'3
4.88 x 10"3
1.76 x 10 3
1.81 x 10'3
1.74 x 10"4
2.88 x 10 4
4.87 x 10"4
3.00 x 10"4
5.15 x 10'4
"All dichotomous models in EPA's BMDS (version 2.0) were fit to the incidence of forestomach lesions
(hyperplasia or hyperkeratosis) in Sprague-Dawley rats using the data presented in Table B-l. For BMD modeling,
three different dose metrics were employed: (1) administered animal dose (estimated) expressed in mg/kg-d, (2)
AN in blood (predicted) expressed in mg/L, and (3) CEO in blood (predicted) expressed in mg/L. Adequate fit of a
model was achieved if the %2 goodness-of-fit statistic yielded a^-value > 0.1. The numbers in parentheses indicate
the number of dose groups dropped in order to obtain an adequate fit, starting with the highest dose group. Of those
models exhibiting adequate fit and yielding BMDLs that were sufficiently close, the selected model (indicated in
body in the table) was the model with the lowest AIC value, as per the EPA's Benchmark Dose Technical Guidance
Document (U.S. EPA, 2000b).
bBMDL10 and BMDL05 estimates were derived from the selected model.
Sources: Quast (2002); Quast et al. (1980a).
B-4
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SD Male Rats: Forestomach lesions (administered dose)
Multistage Model with 0.95 Confidence Level
•
I
•
CO
0.8
0.6
0.4
0.2
16:2210/032008
Multistage Model. (Version: 3.0; Date: 05/16/2008)
Input Data File: C:\USEPA\BMDS2\Temp\tmpAB.(d)
Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpAB.plt
Fri Oct 03 16:22:53 2008
BMDS Model Run
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl-beta2*doseA2)]
The parameter betas are restricted to be positive
Dependent variable = Response
Independent variable = DOSE
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
B-5
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Default Initial Parameter Values
Background = 0.110075
Beta(l) = 0
Beta(2) = 0.0325391
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Beta(l)
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Background Beta(2)
Background 1 -0.37
Beta(2) -0.37 1
Parameter Estimates
Variable
Background
Beta(l)
Beta (2)
Estimate
0.170685
0
0.0278185
Std. Err.
*
*
*
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
* *
* *
* *
* - Indicates that this value is not calculated.
Analysis of Deviance Table
Model
Full model
Fitted model
Reduced model
AIC:
Log(likelihood)
-81.807
-82.7268
-119.21
169.454
# Param's Deviance Test d.f. P-value
3
2 1.83969 1 0.175
1 74.8052 2 <.0001
Dose
Goodness of Fit
Est._Prob. Expected Observed Size
Scaled
Residual
0.0000
3.4000
8.5000
ChiA2 =1.78
0.1707
0.3987
0.8889
d.f. = 1
13.655 15.000 80 0.400
18.741 15.000 47 -1.114
42.666 44.000 48 0.613
P-value = 0.1825
Benchmark Dose Computation
Specified effect =
Risk Type =
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0.95
1.94613
1.27385
2.28688
Taken together, (1.27385, 2.28688) is a 90
interval for the BMD
% two-sided confidence
B-6
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SD Male Rats: Forestomach lesions (AN in blood)
Multistage Model with 0.95 Confidence Level
•
I
•
CO
0.8
0.6
0.4
0.2
0.01
11:5010/062008
0.02 0.03
dose
0.04
0.05
Multistage Model. (Version: 3.0; Date: 05/16/2008)
Input Data File: C:\USEPA\BMDS2\Temp\tmpBB.(d)
Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpBB.plt
Mon Oct 06 11:50:19 2008
BMDS Model Run
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl-beta2*doseA2)]
The parameter betas are restricted to be positive
Dependent variable = Response
Independent variable = DOSE
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
B-7
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Default Initial Parameter Values
Background = 0.121757
Beta(l) = 0
Beta(2) = 815.045
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Beta(l)
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Background Beta(2)
Background 1 -0.36
Beta (2) -0.36 1
Variable
Background
Beta(l)
Beta (2)
Parameter Estimates
Estimate Std. Err.
0.172109 *
0 *
714.086 *
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
* *
* *
* *
* - Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
Log(likelihood)
-81.807
-82.5048
-119.21
169.01
# Param's Deviance Test d.f. P-value
3
2 1.39565 1 0.2375
1 74.8052 2 <.0001
Dose
0.0000
0.0206
0.0536
Est._Prob.
0.1721
0.3885
0.8936
Goodness of Fit
Expected Observed Size
13.769
18.261
42.892
15.000
15.000
44.000
80
47
48
Scaled
Residual
0.365
-0.976
0.519
ChiA2 =1.35
d.f. = 1
P-value = 0.2445
Benchmark Dose Computation
Specified effect =
Risk Type =
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0.95
0.0121469
0.00772037
0.0142952
Taken together, (0.00772037, 0.0142952) is a 90
interval for the BMD
% two-sided confidence
B-S
DRAFT- DO NOT CITE OR QUOTE
-------
SD Male Rats: Forestomach lesions (CEO in blood)
Multistage Model with 0.95 Confidence Level
•
I
•
CO
0.8
0.6
0.4
0.2
0.0005 0.001
12:0510/062008
0.0015 0.002 0.0025 0.003
dose
0.0035 0.004 0.0045
Multistage Model. (Version: 3.0; Date: 05/16/2008)
Input Data File: C:\USEPA\BMDS2\Temp\tmpCC.(d)
Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpCC.plt
Mon Oct 06 12:05:11 2008
BMDS Model Run
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl-beta2*doseA2)]
The parameter betas are restricted to be positive
Dependent variable = Response
Independent variable = DOSE
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
B-9
DRAFT- DO NOT CITE OR QUOTE
-------
Default Initial Parameter Values
Background = 0.0947166
Beta(l) = 0
Beta (2) = 124268
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Beta(l)
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Background Beta(2)
Background 1 -0.37
Beta(2) -0.37 1
Variable
Background
Beta(l)
Beta (2)
Parameter Estimates
Estimate Std. Err.
0.169148 *
0 *
102914 *
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
* *
* *
* *
* - Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
Log(likelihood)
-81.807
-83.0465
-119.21
170.093
# Param's Deviance Test d.f. P-value
3
2 2.47897 1 0.1154
1 74.8052 2 <.0001
Dose
0.0000
0.0018
0.0044
Est._Prob.
0.1691
0.4114
0.8825
Goodness of Fit
Expected Observed Size
13.532
19.334
42.362
15.000
15.000
44.000
80
47
48
Scaled
Residual
0.438
-1.285
0.734
ChiA2 =2.38
d.f. = 1
P-value = 0.1228
Benchmark Dose Computation
Specified effect =
Risk Type =
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0.95
0.00101182
0.000682485
0.00118691
Taken together, (0.000682485, 0.00118691) is a 90
interval for the BMD
% two-sided confidence
B-10
DRAFT- DO NOT CITE OR QUOTE
-------
SD Female Rats: Forestomach lesions (administered dose)
Logistic Model with 0.95 Confidence Level
•
I
•
CO
0.8
0.6
0.4
0.2
15
20
25
14:1010/062008
dose
Logistic Model. (Version: 2.12; Date: 05/16/2008)
Input Data File: C:\USEPA\BMDS2\Temp\tmpD2.(d)
Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpD2.plt
Mon Oct 06 14:10:15 2008
BMDS Model Run
The form of the probability function is:
P[response] = I/[1+EXP(-intercept-slope*dose)]
Dependent variable = Response
Independent variable = DOSE
Slope parameter is not restricted
Total number of observations = 4
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
Default Initial Parameter Values
background = 0 Specified
intercept = -0.808992
slope = 0.180053
B-ll
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -background
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
intercept
slope
intercept
1
-0.66
slope
-0.66
1
Variable
intercept
slope
Parameter Estimates
Estimate Std. Err.
-1.07104 0.225024
0.23795 0.0368793
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
-1.51208 -0.630006
0.165668 0.310232
Analysis of Deviance Table
Model
Full model
Fitted model
Reduced model
AIC:
Log(likelihood)
-103.017
-103.655
-152.026
211.31
# Param's Deviance Test d.f. P-value
4
2 1.27604 2 0.5283
1 98.0188 3 <.0001
Dose
0.0000
4.4000
10.8000
25.0000
Est. Prob.
0.2552
0.4940
0.8174
0.9924
Gooc
Expected
20.416
23.711
39.235
47.637
Iness of Fi1
Observed
20.000
23.000
41.000
47.000
Size
80
48
48
48
Scaled
Residual
-0.107
-0.205
0.659
-1.062
ChiA2 =1.62
d.f. = 2
P-value = 0.4456
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 1.51893
BMDL = 1.2414
B-12
DRAFT- DO NOT CITE OR QUOTE
-------
SD Female Rats: Forestomach lesions (AN in blood)
Multistage Model with 0.95 Confidence Level
•
I
•
CO
1 -
0.8 -
0.6
0.4 -
0.2
0.02
0.14
0.16
15:4510/062008
Multistage Model. (Version: 3.0; Date: 05/16/2008)
Input Data File: C:\USEPA\BMDS2\Temp\tmpED.(d)
Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpED.plt
Mon Oct 06 15:45:27 2008
BMDS Model Run
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl)]
The parameter betas are restricted to be positive
Dependent variable = Response
Independent variable = DOSE
Total number of observations = 4
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
B-13
DRAFT- DO NOT CITE OR QUOTE
-------
Default Initial Parameter Values
Background = 0.2349
Beta(l) = 23.4597
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.46
Beta(l) -0.46 1
Parameter Estimates
Variable
Background
Beta(l)
Estimate
0.237678
22.7713
Std. Err.
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
* - Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
Log(likelihood)
-103.017
-103.895
-152.026
211.789
# Param's
4
2
1
Deviance Test d.f.
1.75601
98.0188
P-value
0.4156
<.0001
Goodness of Fit
Dose
0.0000
0.0237
0.0618
0.1560
Est._Prob.
0.2377
0.5556
0.8134
0.9782
Expected
19.014
26.670
39.042
46.951
Observed
20.000
23.000
41.000
47.000
Size
80
48
48
48
Scaled
Residual
0.259
-1.066
0.725
0.048
ChiA2 =1.73
d.f. = 2
P-value = 0.4207
Benchmark Dose Computation
Specified effect =
Risk Type =
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0.95
0.0046269
0.00361858
0.00604543
Taken together, (0.00361858, 0.00604543) is a 90
interval for the BMD
% two-sided confidence
B-14
DRAFT- DO NOT CITE OR QUOTE
-------
SD Female Rats: Forestomach lesions (CEO in blood)
Logistic Model with 0.95 Confidence Level
•
I
•
CO
0.8
0.6
0.4
0.2
0.002
09:2610/072008
0.004 0.006
dose
0.008
0.01
Logistic Model. (Version: 2.12; Date: 05/16/2008)
Input Data File: C:\USEPA\BMDS2\Temp\tmpl4.(d)
Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpl4.plt
Tue Oct 07 09:26:18 2008
BMDS Model Run
The form of the probability function is:
P[response] = I/[1+EXP(-intercept-slope*dose)]
Dependent variable = Response
Independent variable = DOSE
Slope parameter is not restricted
Total number of observations = 4
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
Default Initial Parameter Values
background = 0 Specified
intercept = -0.929208
slope = 452.992
B-15
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -background
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
intercept
slope
intercept
1
-0.66
slope
-0.66
1
Variable
intercept
slope
Parameter Estimates
Estimate Std. Err.
-1.11487 0.227873
547.334 80.7582
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
-1.56149 -0.668246
389.05 705.617
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
Log(likelihood)
-103.017
-103.359
-152.026
210.718
# Param's
4
2
1
Deviance Test d.f.
0.684912
98.0188
P-value
0.71
<.0001
Dose
0.0000
0.0021
0.0049
0.0101
Est. Prob.
0.2470
0.5045
0.8250
0.9880
Gooc
Expected
19.757
24.217
39.600
47.425
Iness of Fi1
Observed
20.000
23.000
41.000
47.000
Size
80
48
48
48
Scaled
Residual
0.063
-0.351
0.532
-0.565
ChiA2 =0.73
d.f. = 2
P-value = 0.6946
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 0.000678745
BMDL = 0.000557658
B-16
DRAFT- DO NOT CITE OR QUOTE
-------
Table B-4. Summary of the BMD modeling results based on the incidence of
forestomach lesions (hyperplasia or hyperkeratosis) in male and female F344
rats exposed to AN in drinking water for 2 years
Sex
Male
Female
Endpoint
Forestomach
lesions
Forestomach
lesions
Dose metric
Estimated
administered
dose
(mg/kg-d)
Predicted AN
in blood
(mg/L)
Predicted CEO
in blood
(mg/L)
Estimated
administered
dose
(mg/kg-d)
Predicted AN
in blood
(mg/L)
Predicted CEO
in blood
(mg/L)
Selected model(s)a
gamma (3)
two -stage multistage
(3)
gamma (3)
two -stage multistage
(3)
gamma (3)
two -stage multistage
(3)
logistic (3)
two-stage multistage
(3)
probit (3)
logistic (3)
two-stage multistage
(3)
probit (3)
logistic (3)
two-stage multistage
(3)
probit (3)
x2
^j-value
0.32
0.14
0.32
0.14
0.32
0.14
0.32
0.39
0.26
0.32
0.39
0.26
0.32
0.39
0.26
AIC
193.27
194.76
193.27
194.77
193.27
194.76
141.06
140.80
141.38
141.06
140.80
141.38
141.07
140.81
141.40
BMDL10b
1.51 x 10'1
1.40 x 10"1
8.14 x 10 4
7.56 x 10"4
7.65 x 10 5
7.10x 10'5
2.31 x 10'1
2.09 x 10'1
2.18x 10"1
1.11 x 10'3
9.97 x 10 4
1.04 x 10'3
1.02 x 10"4
9.23 x 10 5
9.63 x 10'5
BMDLosb
9.97 x 10 2
8.02 x 10"2
5.39 x 10 4
4.34 x 10"4
5.06 x 10 5
4.07 x 10'5
1.54 x 10'1
1.17 x 10'1
1.40 x 10"1
7.34 x 10'4
5.58 x 10"4
6.69 x 10'4
6.78 x 10"5
5.17 x 10 5
6.19 x 10'5
aAll dichotomous models in EPA's BMDS (version 2.0) were fit to the incidence of forestomach lesions
(hyperplasia or hyperkeratosis) in F344 rats using the data presented in Table B-l. For BMD modeling, three
different dose metrics were employed: (1) administered animal dose (estimated) expressed in mg/kg-d, (2) AN in
blood (predicted) expressed in mg/L, and (3) CEO in blood (predicted) expressed in mg/L. Adequate fit of a model
was achieved if the %2 goodness-of-fit statistic yielded a^-value > 0.1. The numbers in parentheses indicate the
number of dose groups dropped in order to obtain an adequate fit, starting with the highest dose group. Of those
models exhibiting adequate fit and yielding BMDLs that were sufficiently close, the selected model was the model
with the lowest AIC value, as per the EPA's Benchmark Dose Technical Guidance Document (U.S. EPA, 2000b).
bBMDL10 and BMDL05 estimates were derived from the selected model (indicated in bold in the table).
Sources: Johannsen and Levinskas (2002b); Biodynamics (1980c).
B-17
DRAFT- DO NOT CITE OR QUOTE
-------
F344 Male Rats: Forestomach lesions (administered dose)
Gamma Multi-Hit Model with 0.95 Confidence Level
T3
s
I
•
CO
0.35
0.3
0.25
0.2
0.15
0.1
0.05
Gamma Multi-Hit
BMDL
BMD
0.05
0.1
0.15
0.2
0.25
dose
10:2710/072008
Gamma Model. (Version: 2.13; Date: 05/16/2008)
Input Data File: C:\USEPA\BMDS2\Temp\tmpCD.(d)
Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpCD.plt
Tue Oct 07 10:27:44 2008
BMDS Model Run
The form of the probability function is:
P[response]= background+(1-background)*CumGamma[slope*dose,power]
where CumGamma(.) is the cummulative Gamma distribution function
Dependent variable = Response
Independent variable = DOSE
Power parameter is restricted as power >=1
Total number of observations = 3
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
Default Initial (and Specified) Parameter Values
Background = 0.071875
B-18
DRAFT- DO NOT CITE OR QUOTE
-------
Slope =
Power =
5.59733
2.91574
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Power
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Background Slope
Background 1 -0.24
Slope -0.24 1
Parameter Estimates
Variable
Background
Slope
Power
Estimate
0.0585772
57.0862
18
Std. Err.
0.0151899
2.86678
MA
MA - Indicates that this parameter has hit a bound
implied by some inequality constraint and thus
has no standard error.
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
0.0288055 0.0883489
51.4674 62.705
Analysis of Deviance Table
Model
Full model
Fitted model
Reduced model
AIC:
Log(likelihood)
-94.1158
-94.6364
-103.388
193.273
# Param's Deviance Test d.f. P-value
3
2 1.04132 1 0.3075
1 18.5446 2 <.0001
Dose
Goodness of Fit
Est._Prob. Expected Observed Size
Scaled
Residual
0.0000
0.0800
0.2500
ChiA2 =0.97
0.0586
0.0586
0.2400
d.f. = 1
9.314 11.000 159 0.569
4.686 3.000 80 -0.803
18.000 18.000 75 0.000
P-value = 0.3250
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 0.224601
BMDL = 0.150668
B-19
DRAFT- DO NOT CITE OR QUOTE
-------
F344 Male Rats: Forestomach lesions (AN in blood))
Gamma Multi-Hit Model with 0.95 Confidence Level
•
I
•
CO
0.35
0.3
0.25
0.2
0.15
0.1
0.05
Gamma Multi-Hit
BMDL
BMD
0.0002 0.0004
14:4310/072008
0.0006 0.0008
dose
0.001
0.0012 0.0014
Gamma Model. (Version: 2.13; Date: 05/16/2008)
Input Data File: C:\USEPA\BMDS2\Temp\tmpl07.(d)
Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpl07.plt
Tue Oct 07 14:43:15 2008
BMDS Model Run
The form of the probability function is:
P[response]= background+(1-background)*CumGamma[slope*dose,power]
where CumGamma(.) is the cummulative Gamma distribution function
Dependent variable = Response
Independent variable = DOSE
Power parameter is restricted as power >=1
Total number of observations = 3
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
Default Initial (and Specified) Parameter Values
Background = 0.071875
Slope = 1039.5
Power = 2.92132
B-20
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Power
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Background Slope
Background 1 -0.24
Slope -0.24 1
Parameter Estimates
Variable
Background
Slope
Power
Estimate
0.0585772
10571.5
18
Std. Err.
0.01519
530.886
MA
MA - Indicates that this parameter has hit a bound
implied by some inequality constraint and thus
has no standard error.
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
0.0288053 0.088349
9531.01 11612
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
Log(likelihood)
-94.1158
-94.6364
-103.388
193.273
# Param's
3
2
1
Deviance Test d.f.
1.04133
18.5446
P-value
0.3075
<.0001
Goodness of Fit
Dose Est. Prob. Expected Observed Size
Scaled
Residual
0.0000
0.0004
0.0014
ChiA2 =0.97
0.0586
0.0586
0.2400
d.f. = 1
9.314 11.000 159 0.569
4.686 3.000 80 -0.803
18.000 18.000 75 0.000
P-value = 0.3250
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 0.00121285
BMDL = 0.000814398
B-21
DRAFT- DO NOT CITE OR QUOTE
-------
F344 Male Rats: Forestomach lesions (CEO in blood)
Gamma Multi-Hit Model with 0.95 Confidence Level
T3
s
I
•
CO
0.35
0.3
0.25
0.2
0.15
0.1
0.05
Gamma Multi-Hit
BMDL
BMD
2e-005
4e-005
6e-005
dose
8e-005
0.0001
0.00012
16:2910/072008
Gamma Model. (Version: 2.13; Date: 05/16/2008)
Input Data File: C:\USEPA\BMDS2\Temp\tmpll5.(d)
Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpll5.plt
Tue Oct 07 16:29:50 2008
BMDS Model Run
The form of the probability function is:
P[response]= background+(1-background)*CumGamma[slope*dose,power]
where CumGamma(.) is the cummulative Gamma distribution function
Dependent variable = Response
Independent variable = DOSE
Power parameter is restricted as power >=1
Total number of observations = 3
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
Default Initial (and Specified) Parameter Values
Background = 0.071875
Slope = 11005
Power = 2.91337
B-22
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Power
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Background Slope
Background 1 -0.24
Slope -0.24 1
Parameter Estimates
Variable
Background
Slope
Power
Estimate
0.0585772
112374
18
Std. Err.
0.0151899
5643.28
MA
MA - Indicates that this parameter has hit a bound
implied by some inequality constraint and thus
has no standard error.
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
0.0288055 0.0883489
101314 123435
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
Log(likelihood)
-94.1158
-94.6364
-103.388
193.273
# Param's
3
2
1
Deviance Test d.f.
1.04132
18.5446
P-value
0.3075
<.0001
Goodness of Fit
Dose Est. Prob. Expected Observed Size
Scaled
Residual
0.0000
0.0000
0.0001
ChiA2 =0.97
0.0586
0.0586
0.2400
d.f. = 1
9.314 11.000 159 0.569
4.686 3.000 80 -0.803
18.000 18.000 75 0.000
P-value = 0.3250
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 0.000114098
BMDL = 7.6508e-005
B-23
DRAFT- DO NOT CITE OR QUOTE
-------
F344 Female Rats: Forestomach lesions (administered dose)
•
I
•
CO
0.3
0.25
0.2
0.15
0.1
0.05
13:5510/082008
Multistage Model with 0.95 Confidence Level
Multistage
BMDL
BMD
0.05
0.1
0.15 0.2
dose
0.25
0.3
0.35
Multistage Model. (Version: 3.0; Date: 05/16/2008)
Input Data File: C:\USEPA\BMDS2\Temp\tmpA65.(d)
Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpA65.plt
Wed Oct 08 13:55:09 2008
BMDS Model Run
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl-beta2*doseA2)]
The parameter betas are restricted to be positive
Dependent variable = Response
Independent variable = DOSE
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
B-24
DRAFT- DO NOT CITE OR QUOTE
-------
Default Initial Parameter Values
Background = 0.0147377
Beta(l) = 0
Beta (2) = 1.59649
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Beta(l)
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Background Beta(2)
Background 1 -0.51
Beta(2) -0.51 1
Parameter Estimates
Variable
Background
Beta(l)
Beta (2)
Estimate
0.0216322
0
1.47012
Std. Err.
*
*
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
* - Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
Log(likelihood)
-67.9873
-68.401
-79.8392
140.802
# Param's
3
2
1
Deviance Test d.f.
0.827463
23.7038
ChiA2 =0.74
d.f. = 1
P-value = 0.3901
P-value
0.363
<.0001
Dose
0.0000
0.1200
0.3600
Est._Prob.
0.0216
0.0421
0.1914
Goodness of Fit
Expected Observed Size
3.375
3.370
15.308
4.000
2.000
16.000
156
80
80
Scaled
Residual
0.344
-0.763
0.197
Benchmark Dose Computation
Specified effect =
Risk Type =
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0.95
0.267709
0.208731
0.349349
Taken together, (0.208731, 0.349349) is a 90
interval for the BMD
% two-sided confidence
B-25
DRAFT- DO NOT CITE OR QUOTE
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F344 Female Rats: Forestomach lesions (AN in blood)
T3
s
I
•
CO
0.3
0.25
0.2
0.15
0.1
0.05
Multistage Model with 0.95 Confidence Level
Multistage
BMDL
BMD
0.0002 0.0004 0.0006 0.0008 0.001
dose
0.0012 0.0014 0.0016 0.0018
16:04 10/082008
Multistage Model. (Version: 3.0; Date: 05/16/2008)
Input Data File: C:\USEPA\BMDS2\Temp\tmpA72.(d)
Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpA72.plt
Wed Oct 08 16:04:59 2008
BMDS Model Run
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl-beta2*doseA2)]
The parameter betas are restricted to be positive
Dependent variable = Response
Independent variable = DOSE
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
B-26
DRAFT- DO NOT CITE OR QUOTE
-------
Default Initial Parameter Values
Background = 0.0147497
Beta(l) = 0
Beta(2) = 69935.2
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Beta(l)
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Background Beta(2)
Background 1 -0.51
Beta(2) -0.51 1
Variable
Background
Beta(l)
Beta (2)
Parameter Estimates
Estimate Std. Err.
0.021634 *
0 *
64407.8 *
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
* *
* *
* *
* - Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
Log(likelihood)
-67.9873
-68.4002
-79.8392
140.8
# Param's Deviance Test d.f. P-value
3
2 0.825863 1 0.3635
1 23.7038 2 <.0001
Dose
0.0000
0.0006
0.0017
Est._Prob.
0.0216
0.0421
0.1914
Goodness of Fit
Expected Observed Size
3.375
3.369
15.310
4.000
2.000
16.000
156
80
80
Scaled
Residual
0.344
-0.762
0.196
ChiA2 =0.74
d.f. = 1
P-value = 0.3905
Benchmark Dose Computation
Specified effect =
Risk Type =
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0.95
0.001279
0.000997117
0.00166904
Taken together, (0.000997117, 0.00166904) is a 90
% two-sided confidence interval for the BMD
B-27
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F344 Female Rats: Forestomach lesions (CEO in blood)
•
I
•
CO
0.3
0.25
0.2
0.15
0.1
0.05
Multistage Model with 0.95 Confidence Level
Multistage
BMDL
BMD
14:4310/092008
2e-005 4e-005 6e-005 8e-005 0.0001 0.00012 0.00014 0.00016
dose
Multistage Model. (Version: 3.0; Date: 05/16/2008)
Input Data File: C:\USEPA\BMDS2\Temp\tmpFC.(d)
Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpFC.plt
Thu Oct 09 14:43:44 2008
BMDS Model Run
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl-beta2*doseA2)]
The parameter betas are restricted to be positive
Dependent variable = Response
Independent variable = DOSE
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
B-28
DRAFT- DO NOT CITE OR QUOTE
-------
Default Initial Parameter Values
Background = 0.0146597
Beta(l) = 0
Beta(2) = 8.18648e+006
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Beta(l)
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Background Beta(2)
Background 1 -0.51
Beta(2) -0.51 1
Parameter Estimates
Variable
Background
Beta(l)
Beta (2)
Estimate
0.0216202
0
7.53198e+006
Std. Err.
*
*
*
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
* *
* *
* *
* - Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
Log(likelihood)
-67.9873
-68.4062
-79.8392
140.812
# Param's Deviance Test d.f. P-value
3
2 0.837895 1 0.36
1 23.7038 2 <.0001
Dose
0.0000
0.0001
0.0002
Est._Prob.
0.0216
0.0423
0.1913
Goodness of Fit
Expected Observed Size
3.373
3.380
15.301
4.000
2.000
16.000
156
80
80
Scaled
Residual
0.345
-0.767
0.199
ChiA2 =0.75
d.f. = 1
P-value = 0.3873
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 0.000118273
BMDL = 9.22818e-005
BMDU = 0.000154338
Taken together, (9.22818e-005, 0.000154338) is a 90
BMD
% two-sided confidence interval for the
B-29
DRAFT- DO NOT CITE OR QUOTE
-------
Table B-5. Summary of the BMD modeling results based on the incidence of
forestomach lesions (hyperplasia or hyperkeratosis) in male and female
B6C3Fi mice exposed to AN via gavage for 2 years
Sex
Male
Female
Endpoint
Forestomach
lesions
Forestomach
lesions
Dose metric
Estimated
administered dose
(mg/kg-d)
Estimated
administered dose
(mg/kg-d)
Selected model(s)a
gamma
logistic
log-logistic
probit
log-probit
one-stage
multistage
Weibull
gamma
logistic
log-logistic
probit
log-probit
one-stage
multistage
Weibull
x2
/7-value
0.80
0.37
0.87
0.42
0.33
0.80
0.80
0.74
0.87
0.75
0.90
0.81
0.92
0.74
AIC
156.45
158.01
156.30
157.74
158.23
156.45
156.45
116.17
114.33
116.17
114.28
114.48
114.24
116.18
BMDL10b
3.45
6.37
3.01
5.96
5.53
3.45
3.45
6.24
8.95
5.98
8.51
8.12
6.20
6.24
BMDL05b
1.68
3.72
1.43
3.42
3.85
1.68
1.68
3.04
5.60
2.83
5.19
5.65
3.02
3.04
aAll dichotomous models in EPA's BMDS (version 2.0) were fit to the incidence of forestomach lesions
(hyperplasia or hyperkeratosis) in B6C3F! mice using the data presented in Table B-2. For BMD modeling, animal
dose (estimated), expressed in mg/kg-d, was employed. Adequate fit of a model was achieved if the %2 goodness-
of-fit statistic yielded a^-value > 0.1. Of those models exhibiting adequate fit and yielding BMDLs that were
sufficiently close, the selected model (indicated in bold in the table) was the model with the lowest AIC value, as
per the EPA's Benchmark Dose Technical Guidance Document (U.S. EPA, 2000b).
bBMDL10 and BMDL05 estimates were derived from the selected model.
Source: NTP(2001).
B-30
DRAFT- DO NOT CITE OR QUOTE
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B6C3F! Male Mice: Forestomach lesions (administered dose)
Log-Logistic Model with 0.95 Confidence Level
T3
s
I
•
CO
0.4
0.35
0.3
0.25
0.2
0.15
0.1
0.05
Log-Logistic
BMDL
BMD
6 8
dose
10
12
14
14:02 11/142008
Logistic Model. (Version: 2.12; Date: 05/16/2008)
Input Data File: C:\USEPA\BMDS2\Temp\tmpll6.(d)
Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpll6.plt
Fri Nov 14 14:02:24 2008
BMDS Model Run
The form of the probability function is:
P[response] = background+(1-background)/[1+EXP(-intercept-slope*Log(dose))]
Dependent variable = Response
Independent variable = DOSE
Slope parameter is restricted as slope >= 1
Total number of observations = 4
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
User has chosen the log transformed model
Default Initial Parameter Values
background = 0.04
B-31
DRAFT- DO NOT CITE OR QUOTE
-------
intercept =
slope =
-3.73286
1
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -slope
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
background intercept
background 1 -0.54
intercept -0.54 1
Parameter Estimates
Variable
background
intercept
slope
Estimate
0.0421528
-3.76687
1
Std. Err.
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
* - Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
Log(likelihood)
-76.0086
-76.1489
-82.7874
156.298
# Param's
4
2
1
Deviance Test d.f.
0.280564
13.5576
P-value
0.8691
0.003574
Goodness of Fit
Dose
0.0000
1.8000
7.1000
14.3000
Est. Prob.
0.0422
0.0804
0.1772
0.2802
Expected
2.108
4.021
8.862
14.009
Observed
2.000
4.000
10.000
13.000
Size
50
50
50
50
Scaled
Residual
-0.076
-0.011
0.422
-0.318
ChiA2 =0.28
d.f. = 2
P-value = 0.8674
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 4.80493
BMDL = 3.0101
B-32
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B6C3F! Female Mice: Forestomach lesions (administered dose)
•
I
•
CO
0.3
0.25
0.2
0.15
0.1
0.05
Multistage Model with 0.95 Confidence Level
Multistage
BMDL
BMD
15:5911/142008
6 8
dose
10
12
14
Multistage Model. (Version: 3.0; Date: 05/16/2008)
Input Data File: C:\USEPA\BMDS2\Temp\tmpl3D.(d)
Gnuplot Plotting File: C:\USEPA\BMDS2\Temp\tmpl3D.plt
Fri Nov 14 15:59:08 2008
BMDS Model Run
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl)]
The parameter betas are restricted to be positive
Dependent variable = Response
Independent variable = DOSE
Total number of observations = 4
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
B-33
DRAFT- DO NOT CITE OR QUOTE
-------
Default Initial Parameter Values
Background = 0.032589
Beta(l) = 0.00986338
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.7
Beta(l) -0.7 1
Parameter Estimates
Variable
Background
Beta(l)
Estimate
0.0341442
0.00957473
Std. Err.
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
* - Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
Log(likelihood)
-55.0321
-55.1226
-58.1629
114.245
# Param's
4
2
1
Deviance Test d.f.
0.181118
6.26165
P-value
0.9134
0.09955
Goodness of Fit
Dose
0.0000
1.8000
7.1000
14.3000
Est._Prob.
0.0341
0.0506
0.0976
0.1577
Expected
1.707
2.532
4.881
7.887
Observed
2.000
2.000
5.000
8.000
Size
50
50
50
50
Scaled
Residual
0.228
-0.343
0.057
0.044
ChiA2 =0.18
d.f. = 2
P-value = 0.9162
Benchmark Dose Computation
Specified effect =
Risk Type =
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0.95
11.004
6.2019
33.7925
Taken together, (6.2019 , 33.7925) is a 90
interval for the BMD
% two-sided confidence
B-34
DRAFT- DO NOT CITE OR QUOTE
-------
APPENDIX B-2. NONCANCER INHALATION DOSE-RESPONSE ASSESSMENT
(RfC): BMD MODELING RESULTS EMPLOYING THE INCIDENCE DATA FOR
NONNEOPLASTIC NASAL LESIONS IN RATS EXPOSED TO AN BY INHALATION
FOR 2 YEARS (TABLES B-6 THROUGH B-9)
For these modeling exercises, a BMR of 10% extra risk was selected due to the limited
number of rats in each exposure group. BMC and BMCL refer to the model-predicted
concentration and its lower 95% confidence limit, respectively, associated with a 10% extra risk
for developing the lesion.
Table B-6. Incidence data for selected nasal lesions in Sprague-Dawley rats
exposed by inhalation to AN for 2 years
Nasal lesion
Hyperplasia of mucus-
secreting cells in males
Flattening of respiratory
epithelium in females
Exposure level (ppm)
0
20
80
0
20
80
Exposure level
(mg/m3)
0
43.4
173.6
0
43.4
173.6
HEC
(mg/m3)3
0
2.1
8.5
0
2.1
8.5
Incidence
0/11(0%)
7/12 (58%)b
8/10 (80%)b
1/11(9%)
7/10 (70%)b
8/10 (80%)b
aHEC as per U.S. EPA (1994b) methods for a category 1 gas producing an upper respiratory effect.
Sample calculation: 43.4 mg/m3 x 6h/24h x 5d/7d x RGDRET = 2.1 mg/m3, where RGDRET = 0.275 =
[VE/SAET] rat -^ [VE/SAET] human; VE = minute volume = 0.281 L/min rat; 13.8 L/min human; and SAET
= extrathoracic surface area = 5 cm2 rat, 200 cm2 human.
bStatistically significantly different from control value as reported by the authors.
Source: Quastetal. (1980b).
B-35
DRAFT- DO NOT CITE OR QUOTE
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Table B-7. A summary of BMDS (version 1.3.2) modeling results based on
incidence of hyperplasia of mucus-secreting cells in male Sprague-Dawley
rats exposed to AN via inhalation for 2 years
Model
Gamma
Logistic
Log-logistic
Multistage
Probit
Log-probit
Weibull
X2/7-valuea
0.34
0.02
0.94
0.34
0.01
0.37
0.34
AIC
30.27
38.05
28.43
30.27
37.99
29.98
30.27
BMC10b
(mg/m3)
0.396
1.17
0.187
0.396
1.17
0.625
0.396
BMCL10C
(mg/m3)
0.252
0.718
0.082
0.252
0.776
0.382
0.252
a/(2 />-value from the %2 test for lack of fit. Values <0.1 fail to meet conventional goodness-of-fit criteria.
bBMC10 = BMC associated with 10% extra risk for nonneoplastic nasal lesions.
°BMCL10 = 95% lower confidence limit on the BMC10 for nonneoplastic nasal lesions.
Source: Quastetal. (1980b).
B-36
DRAFT- DO NOT CITE OR QUOTE
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BMDS (version 1.3.2) model output for the best-fit model (i.e., log-logistic) based on
incidence of hyperplasia of mucus-secreting cells in male Sprague-Dawley rats exposed to
AN via inhalation for 2 years
Log-Logistic Model with 0.95 Confidence Level
0.8
I °'6
0.4
cc
0.2
0
Log-Logistic
BlvDL
BMD
0
13:3805/162007
4
dose
8
Logistic Model $Revision: 2.1 $ $Date: 2000/02/26 03:38:20 $
Input Data File: G:\ACN DOSE-RESPONSE
MODE LING\NONCANCER\INHALATION\S D_MALE_NASAL_INHALATION. (d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODE LING\NONCANCER\INHALATION\S D_MALE_NASAL_INHALATION.p11
Wed May 16 13:38:57 2007
The form of the probability function is:
P[response] = background+(1-background)/[1+EXP(-intercept-slope*Log(dose)
Total number of observations = 3
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
B-37
DRAFT- DO NOT CITE OR QUOTE
-------
Default Initial Parameter Values
background = 0
intercept = -0.581093
slope = 1
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -background -slope
have been estimated at a boundary point, or have been specified by the user, and
do not appear in the correlation matrix )
intercept
intercept 1
Parameter Estimates
Estimate
NA - Indicates that this parameter has hit a bound
implied by some inequality constraint and thus
has no standard error.
Model
Full model
Fitted model
Reduced model
AIC:
Log(likelihood) Deviance Test DF
-13.1543
-13.2153 0.121833 2
-22.7373 19.1659 2
P-value
Est. Prob.
"hi-square =
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 0.187372
BMDL = 0.0818673
B-38
DRAFT- DO NOT CITE OR QUOTE
-------
Table B-8. A summary of BMDS (version 1.3.2) modeling results based on
incidence of flattening of respiratory epithelium in female Sprague-Dawley
rats exposed to AN via inhalation for 2 years
Model
Gamma
Logistic
Log-logistic
Multistage
Probit
Log-probit
Weibull
/2/7-valuea
0.08
0.02
0.43
0.08
0.02
0.08
0.08
AIC
35.78
38.47
33.50
35.78
38.54
35.53
35.78
BMC10b
(mg/m3)
0.419
1.02
0.162
0.419
1.05
0.616
0.419
BMCL10C
(mg/m3)
0.245
0.610
0.059
0.245
0.683
0.340
0.245
a/(2 />-value from the %2 test for lack of fit. Values <0.1 fail to meet conventional goodness-of-fit criteria.
bBMCio = BMC associated with 10% extra risk for nonneoplastic nasal lesions.
°BMCL10 = 95% lower confidence limit on the BMCio for nonneoplastic nasal lesions.
Source: Quastetal. (1980b).
B-39
DRAFT- DO NOT CITE OR QUOTE
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BMDS (version 1.3.2) model output for the best-fit model (log-logistic) based on incidence
of flattening of respiratory epithelium in female Sprague-Dawley rats exposed to AN via
inhalation for 2 years
Log-Logistic Model with 0.95 Confidence Level
T3
CD
o
=8
CO
0.8
0.6
0.2
Log-Logistic
BMDL
BMD
0
16:2505/142007
4
dose
6
8
Logistic Model $Revision: 2.1 $ $Date: 2000/02/26 03:38:20 $
Input Data File: G:\ACN DOSE-RESPONSE MODELING\NONCANCER\SD_FEMALE_NASAL_INHALATION.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODE LING\NONCANCER\S D_FEMALE_NASAL_INHALATION.p11
Mon May 14 16:25:24 2007
BMDS MODEL RUN
Total number of observations = 3
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
Default Initial Parameter Values
background = 0.0909091
intercept = -0.457634
slope = 1
B-40
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -slope
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Variable
background
intercept
slope
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
Log(likelihood) Deviance Test DF
-14.4637
-14.751 0.574692 1
-21.4714 14.0155 2
Goodness of Fit
Est._Prob. Expected Observed
"hi-square =
DF = 1
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 0.161645
BMDL = 0.0593975
B-41
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APPENDIX B-3. CANCER ORAL DOSE-RESPONSE ASSESSMENT: BMD DOSE
MODELING RESULTS FOR TUMOR INCIDENCE DATA FROM RATS
CHRONICALLY EXPOSED TO AN IN DRINKING WATER
As summarized in Section 4.6.1, AN is a multisite carcinogen in chronic oral rodent
bioassays. Oral carcinogenicity studies of animals chronically exposed to AN include drinking
water studies in two strains of rats, Sprague-Dawley (Johannsen and Levinskas, 2002a; Quast,
2002; Biodynamics, 1980a; Quast et al., 1980a) and F344 (Johannsen and Levinskas, 2002b;
Biodynamics, 1980b). In Sprague-Dawley rats, significantly increased incidences of
forestomach, CNS, Zymbal gland, tongue, and mammary gland (females only) tumors were
found. In F344 rats, significantly increased incidences of forestomach, CNS, Zymbal gland, and
mammary gland (females only) tumors were found.
Two of these chronic drinking water studies were selected for dose-response modeling
using the BMD approach and derivation of oral CSFs for AN. In one study by Quast (2002),
Sprague-Dawley rats were exposed to 0, 35, 100, or 300 ppm of AN in drinking water for
2 years. In the second study by Johannsen and Levinskas (2002b), F344 rats were exposed to 0,
1, 3, 10, 30, or 100 ppm of AN in drinking water for 2 years.
In this appendix, detailed results of the dose-response modeling for each of the tumor
sites listed above are presented (Tables B-9 though B-26). For each tumor site, first a summary
of the dose-response data is presented, followed by a table summarizing the results of the dose-
response modeling. Finally, the standard output from EPA's BMDS, version 1.4.1, for the
selected dose-response model for each tumor site is presented.
In general, the multistage model was fit to all of the data sets with the BMR set at
0.1 (i.e., 10% extra risk). In fitting this model, successive stages of the multistage model,
starting with stage 1 and ending with the stage equal to the number of dose groups minus one,
were fit to the tumor incidence data at a particular site for each rat strain and sex employing the
internal dose metrics CEO in blood and AN in blood. Then, for each dose metric, all stages of
the multistage model that did not show a significant lack of fit (i.e.,/? > 0.1) were compared
using AIC. The stage of the multistage model with the lowest AIC was selected as the best-fit
model. For most tumor sites, the one-stage model exhibited the best fit. For data sets that
exhibited a significant lack of fit for all stages of the multistage model, dose groups were
dropped (starting with the highest dose group) until an adequate fit was achieved.
B-42 DRAFT- DO NOT CITE OR QUOTE
-------
Sprague-Dawlev Rats (Quasi 2002; Quast et al., 1980a)
Tumor Site: Forestomach
Table B-9. Incidence of forestomach (nonglandular) tumors in Sprague-
Dawley rats exposed to AN in drinking water for 2 years
Sex
Male
Female
Administered
animal dose
(ppm in drinking
water)
0
35
100
300
0
35
100
300
Equivalent
administered
animal dose"
(mg/kg-d)
0
3.42
8.53
21.2
0
4.36
10.8
25.0
Predicted internal dose metrics
AN-AUC in blood
(mg/L)
0
2.06 x 10'2
5.36 x 10"2
1.46 x 10'1
0
2.37 x 10"2
6.18 x 10"2
1.56 x 10'1
CEO-AUC in blood
(mg/L)
0
1.83 x 10'3
4.36 x 10"3
9.70 x 10'3
0
2.07 x 10"3
4.87 x 10"3
1.01 x 10'2
Incidence of
forestomach
tumorsb
0/80 (0%)
2/47 (4%)
23/48 (48%)c
39/48 (81%)c
1/80 (1%)
1/48 (2%)
12/48 (25%)c
30/48 (62%)c
""Administered doses were averages calculated by the study authors based on animal BW and drinking water intake.
Incidences for Sprague-Dawley rats do not include animals from the 6- and 12-mo sacrifices and were further
adjusted to exclude (from the denominators) rats that died between 0 and 12 mos in the study.
Significantly different from controls (p < 0.05) as calculated by the study authors.
Sources: Quast (2002); Quast et al. (1980a).
Table B-10. Summary of BMD modeling results based on incidence of
forestomach (nonglandular) tumors in Sprague-Dawley rats exposed to AN
in drinking water for 2 years
Dose metric
Best-fit model3
X2/7-valueb
AIC
BMD10C
BMDL10d
Males
Administered dose
CEO
AN
2°MSe
2°MSe
2°MSe
0.46
0.39
0.54
86.82
87.28
86.45
3.62 mg/kg-d
1.87x 10"3mg/L
2.26 x 10"2 mg/L
2.76 mg/kg-d
1.44x 10"3mg/L
1.70x 10"2mg/L
Females
Administered dose
CEO
AN
2°MS
2°MS
2°MS
0.17
0.52
0.13
145.97
143.50
146.45
7.76 mg/kg-d
3.29x 10"3mg/L
4.22 x 10"2 mg/L
4.81 mg/kg-d
2.38 x 10"3 mg/L
2.49 x 10"2 mg/L
aDose-response models were fit using BMDS, version 1.4.1. "2°MS" indicates a two-stage multistage model.
bp value from the %2 goodness-of-fit test. Values <0.1 indicate a significant lack of fit.
CBMD10 = BMD at 10% extra risk.
dBMDL10 = 95% lower confidence limit on the BMD at 10% extra risk.
eHighest dose dropped prior to model fitting.
B-43
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for forestomach tumors in Sprague-Dawley male rats
employing administered dose as a dose metric
Multistage Cancer Model with 0.95 Confidence Level
0.6
0.5
1 0.4
o
t5
0.3
0.2
0.1
Multistage Cancer
Linear extrapolation
BMD
4
dose
8
15:4001/232009
Multistage Cancer Model. (Version: 1.5; Date: 02/20/2007)
Input Data File: M:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\SD_MALE_FORESTOMACH_DW.(d)
Gnuplot Plotting File: M:\ACN DOSE-RESPONSE
MODE LING\CANCERX ORAL\S D_MALE_FORE S T OMACH_DW.pit
Fri Jan 23 15:40:45 2009
BMDS MODEL RUN
The form of the probability function is:
Dependent variable = Response
Independent variable = Dose
Total number of observations = 3
Total number of records with missing values
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
B-44
DRAFT- DO NOT CITE OR QUOTE
-------
Default Initial Parameter Values
Background = 0
Beta(l) = 0
Beta(2) = 0.00929613
and do not appear in the correlation matrix )
Beta (2)
1
Variable
Background
Beta(1)
Beta(2)
Indicates that this value is not calculated.
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
Model
Full model
Fitted model
Reduced model
AIC:
Log(likelihood) # Param's Deviance Test d.f.
-41.5002 3
-42.4099 1 1.81939 2
-71.7704 1 60.5403 2
Est. Prob.
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
3.62184
2.75694
4.3186
Taken together, (2.75694, 4.3186 ) is a 90
interval for the BMD
B-45
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for forestomach tumors in Sprague-Dawley male rats
employing CEO in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
0.6
0.5
1 0.4
o
t5
0.3
0.2
0.1
0
Multistage
BMDL
0 0.0005 0.001 0.0015 0.002 0.0025 0.003 0.0035 0.004 0.0045
dose
13:4609/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_MALE_FORESTOMACH_BLOOD_CEO.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_MALE_FORESTOMACH_BLOOD_CEO.plt
Thu Sep 27 13:46:37 2007
BMDS MODEL RUN
The form of the probability function is:
Dependent variable = Response
Independent variable = Dose
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
B-46
DRAFT- DO NOT CITE OR QUOTE
-------
Default Initial Parameter Values
Background = 0
Beta(l) = 0
Beta(2) = 35739.1
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Background -Beta(l)
have been estimated at a boundary point, or have been specified by the user, and
do not appear in the correlation matrix )
Beta (2)
Beta(2) 1
Parameter Estimates
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
Model Log(likelihood) # Param's Deviance Test d.f. P-value
Full model -41.5002 3
Fitted model -42.6376 1 2.27473 2 0.3207
Reduced model -71.7704 1 60.5403 2 <.0001
AIC: 87.2752
Goodness of Fit
Est._Prob. Expected Observed
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.00186692
0.00143604
0.00222594
B-47
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for forestomach tumors in Sprague-Dawley male rats
employing AN in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
0.6
0.5
S 0.4
<
c
o
cc
0.3
0.2
0.1
0
Multistage
BMDL
BMD
0.01
0.02
0.03
0.04
0.05
dose
10:3209/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_MALE_FORESTOMACH_BLOOD_AN.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_MALE_FORESTOMACH_BLOOD_AN.plt
Thu Sep 27 10:32:01 2007
BMDS MODEL RUN
The form of the probability function is:
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
Default Initial Parameter Values
B-48
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Background -Beta(l)
have been estimated at a boundary point, or have been specified by the user, and
do not appear in the correlation matrix )
Parameter Estimates
Variable
Background
Beta(1)
Beta(2)
Indicates that this value is not calculated.
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
Analysis of Deviance Table
Model Log(likelihood) # Param's Deviance Test d.f.
Full model -41.5002 3
Fitted model -42.2261 1 1.45174 2
Reduced model -71.7704 1 60.5403 2
Goodness of Fit
Est. Prob.
d.f.
Benchmark Dose Computation
0.1
Extra risk
Confidence level = 0.95
BMD =
BMDL =
BMDU =
B-49
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for forestomach tumors in Sprague-Dawley female rats
employing administered dose as a dose metric
Multistage Cancer Model with 0.95 Confidence Level
0.8
0.7
0.6
S 0.5
< 0.4
o
=0 0.3
ro
^ 0.2
0.1
0
: - - . . ' ' :
: ivuitistage uancer :
: Linear extrapolation :
^ -p /^ r"
~- <> /^^ '-
\ ^^ }
( L^-4'^'" J
; BMDL BMD ;
10
15
20
25
dose
16:1401/232009
Multistage Cancer Model. (Version: 1.5; Date: 02/20/2007)
Input Data File: M:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\SD_FEMALE_FORESTOMACH_DW.(d)
Gnuplot Plotting File: M:\ACN DOSE-RESPONSE
MODE LING\CANCERX ORAL\S D_FEMALE_FORE S T OMACH_DW.pit
Fri Jan 23 16:14:45 2009
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl-beta2*doseA2)]
Total number of observations = 4
Total number of records with missing values = 0
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
B-50
DRAFT- DO NOT CITE OR QUOTE
-------
Default Initial Parameter Values
Background = 0
Beta(l) = 0.0138175
Beta(2) = 0.00104033
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l) Beta(2)
Background 1 -0.57 0.39
Beta(l) -0.57 1 -0.93
Beta(2) 0.39 -0.93 1
Variable
Background
Beta(1)
Beta(2)
Indicates that this value is not calculated.
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
Model
Full model
Fitted model
Reduced model
AIC:
d.f. = 1
P-value = 0.1652
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
7.76379
4.81488
9.11082
Taken together, (4.81488, 9.11082) is a 90
interval for the BMD
B-51
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for forestomach tumors in Sprague-Dawley female rats
employing CEO in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
0.8
0.7
0.6
S 0.5
£>
< 0.4
a
o
=0 0.3
cc
£ 0.2
0.1
0
Multistage
BMD
0.002
0.004 0.006
dose
0.008
0.01
14:2409/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_FEMALE_FORESTOMACH_BLOOD_CEO.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_FEMALE_FORESTOMACH_BLOOD_CEO.plt
Thu Sep 27 14:24:06 2007
BMDS MODEL RUN
The form of the probability function is:
Total number of observations = 4
Total number of records with missing values = 0
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
Default Initial Parameter Values
B-52
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Beta(l)
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Background
Background 1
Beta(2) -0.48
Parameter Estimates
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
Model
Full model
Fitted model
Reduced model
AIC:
Log(likelihood) # Param's Deviance Test d.f.
-68.9836 4
-69.7499 2 1.53257 2
-110.972 1 83.9771 3
P-value
Est. Prob.
80
48
48
48
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.00328691
0.00238295
0. 0037821
% two-sided confidence
B-53
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for forestomach tumors in Sprague-Dawley female rats
employing AN in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
Affected
S_
o
cc
LJ_
0.8
0.7
0.6
0.5
0.4
0.3
0.2
0.1
0
: - - . . :
: Multistage :
[ ^^<:> '
\ /^ -L ;
: <> /-^ ;
1. §?*~ ' T J
; B|VPL BIVD i
0 0.02 0.04 0.06 0.08 0.1 0.12 0.14 0.16
dose
10:4909/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_FEMALE_FORESTOMACH_BLOOD_AN.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_FEMALE_FORESTOMACH_BLOOD_AN.plt
Thu Sep 27 10:49:49 2007
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl-beta2*doseA2)]
Total number of observations = 4
Total number of records with missing values = 0
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
B-54
DRAFT- DO NOT CITE OR QUOTE
-------
Default Initial Parameter Values
Background = 0
Beta(l) = 3.15264
Beta(2) = 20.8062
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l) Beta(2)
Background 1 -0.57 0.4
Beta(l) -0.57 1 -0.93
Beta(2) 0.4 -0.93 1
Parameter Estimates
Variable
Background
Beta(1)
Beta(2)
Indicates that this value is not calculated.
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
# Param's Deviance Test d.f.
4
3 2.47911 1
1 83.9771 3
Goodness of Fit
Est. Prob.
d.f. = 1
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0 . 0421673
0.0249138
0.055287
B-55
DRAFT- DO NOT CITE OR QUOTE
-------
Tumor Site: CNS (Quast, 2002; Quast et al., 1980a)
Table B-ll. Incidence of CNS tumors in Sprague-Dawley rats exposed to
AN in drinking water for 2 years
Sex
Male
Female
Administered
animal dose
(ppm in drinking
water)
0
35
100
300
0
35
100
300
Equivalent
administered
animal dosea
(mg/kg-d)
0
3.42
8.53
21.2
0
4.36
10.8
25.0
Predicted internal dose metrics
AN-AUC in blood
(mg/L)
0
2.06 x 10"2
5.36 x 10'2
1.46 x 10"1
0
2.37 x 10"2
6.18 x 10'2
1.56 x 10"1
CEO-AUC in blood
(mg/L)
0
1.83 x 10"3
4.36 x 10'3
9.70 x 10"3
0
2.07 x 10"3
4.87 x 10'3
1.01 x 10"2
Incidence of CNS
tumorsb
1/80 (1%)
12/47 (26%)c
22/48 (46%)c
30/48 (62%)c
1/80 (1%)
20/48 (42%)c
25/48 (52%)c
31/48(65%)c
""Administered doses were averages calculated by the study authors based on animal BW and drinking water
intake.
blncidences for Sprague-Dawley rats do not include animals from the 6- and 12-mo sacrifices and were further
adjusted to exclude (from the denominators) rats that died between 0 and 12 mos in the study.
Significantly different from controls (p < 0.05) as calculated by the study authors.
Table B-12. Summary of BMD modeling results based on incidence of CNS
tumors in Sprague-Dawley rats exposed to AN in drinking water for 2 years
Dose metric
Best-fit model3
/2/7-valueb
AIC
BMD10C
BMDL10d
Males
Administered dose
CEO
AN
1°MS
1°MS
l°MSe
0.16
0.37
0.59
201.38
199.81
134.64
1.84 mg/kg-d
8.87 x 10"4 mg/L
8.82 x 10'3 mg/L
1.48 mg/kg-d
7.16x 10"4mg/L
6.64 x 10'3 mg/L
Females
Administered dose
CEO
AN
l°MSe
l°MSe
l°MSe
0.06
0.08
0.04
149.85
149.29
150.45
1.26 mg/kg-d
5.79 x 10"4 mg/L
7.12x 10'3mg/L
0.99 mg/kg-d
4.51 x 10"4mg/L
5.55 x 10"3 mg/L
aDose-response models were fit using BMDS, version 1.4.1. "1°MS" indicates a one-stage multistage model.
V value from the %2 goodness of fit test. Values <0.1 indicate a significant lack of fit.
CBMD10 = BMD at 10% extra risk.
dBMDL10 = 95% lower confidence limit on the BMD at 10% extra risk.
eHighest dose dropped prior to model fitting.
B-56
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for CNS tumors in Sprague-Dawley male rats employing
administered dose as a dose metric
Multistage Cancer Model with 0.95 Confidence Level
0.8
0.7
0.6
1 0.5
< 0.4
o
'•6 0.3
CD
^ 0.2
0.1
0
:--.._ :
: Multistage uancer :
: Linear extrapolation :
: ^^^-^^ <> :
L ^^^^ \
1 -p ^^ J
L // -L j
- // -
[ V \
\ BMDL BMD 1
0 5 10 15 20
dose
15:4801/232009
Multistage Cancer Model. (Version: 1.5; Date: 02/20/2007)
Input Data File: M:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\SD_MALE_CNS_DW.(d)
Gnuplot Plotting File: M:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\SD_MALE_CNS_DW.plt
Fri Jan 23 15:48:14 2009
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl)]
Dependent variable = Response
Independent variable = Dose
Total number of observations = 4
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
Default Initial Parameter Values
B-57
DRAFT- DO NOT CITE OR QUOTE
-------
Background =
Beta(l) =
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.62
Beta(l) -0.62 1
Parameter Estimates
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
# Param's Deviance Test d.f.
4
2 3.51012 2
1 75.2765 3
P-value
Est. Prob.
d.f. =2
1
12
Benchmark Dose Computation
0.1
Extra risk
Confidence level = 0.95
BMD =
BMDL =
BMDU =
Multistage Cancer Slope Factor =
B-58
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for CNS tumors in Sprague-Dawley male rats employing
CEO in blood as an internal dose metric
Multistage Model with 0.95 Confidence Level
0.7
0.6
°-5
0.4
0.3
cc
^ 0.2
0.1
0
Multistage
BIVDL BlvD
0
09:11 09/252007
0.002 0.004 0.006
dose
0.008
0.01
Multistage Model. $Revision: 2.1 $ $Date: 2000/08/21 03:38:21 $
Input Data File: G:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\SD_MALE_CNS_BLOOD_CEO.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_MALE_CNS_BLOOD_CEO.plt
Tue Sep 25 09:11:57 2007
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl)]
Total number of observations = 4
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-59
DRAFT- DO NOT CITE OR QUOTE
-------
**** WARNING:
**** WARNING 0
**** WARNING 1
**** WARNING 2
**** WARNING 3
**** WARNING 4
**** WARNING 5
**** WARNING 6
**** WARNING 7
**** WARNING 8
**** WARNING 9
Completion code = -3. Optimum not found. Trying new starting point****
Completion code = -3 trying new start****
Completion code = -3 trying new start****
Completion code = -3 trying new start****
Completion code = -3 trying new start****
Completion code = -3 trying new start****
Completion code = -3 trying new start****
Completion code = -3 trying new start****
Completion code = -3 trying new start****
Completion code = -3 trying new start****
Completion code = -3 trying new start****
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.62
Beta(l) -0.62 1
Variable
Background
Beta(1)
Model Log(likelihood) Deviance Test DF
Full model -96.9359
Fitted model -97.9055 1.93933 2
Reduced model -134.574 75.2765 3
AIC:
P-value
i: 1
i: 4
1.173
1
12
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 0. 00088713
BMDL = 0.00071632
B-60
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for CNS tumors in Sprague-Dawley male rats employing AN
in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
0.6
0.5
§ 0.4
<
o
t5
CO
0.3
0.2
0.1
Multistage
BMDL
BMD
0.01
0.02 0.03
dose
0.04
0.05
10:4009/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\SD_MALE_CNS_BLOOD_AN.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_MALE_CNS_BLOOD_AN.plt
Thu Sep 27 10:40:46 2007
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl)]
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-61
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.61
Beta(l) -0.61 1
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
134. 641
# Param's Deviance Test d.f.
3
2 0.279194 1
1 44.7792 2
P-value
1
12
B-62
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for CNS tumors in Sprague-Dawley female rats employing
administered dose as a dose metric
Multistage Cancer Model with 0.95 Confidence Level
0.7
0.6
0.5
CD
g 0.3
t5
CO
0.2
0.1
0
Multistage Cancer
Linear extrapolation
BIVDL BMP
8
10
dose
16:1901/232009
Multistage Cancer Model. (Version: 1.5; Date: 02/20/2007)
Input Data File: M:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\SD_FEMALE_CNS_DW.(d)
Gnuplot Plotting File: M:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\SD_FEMALE_CNS_DW.plt
Fri Jan 23 16:19:26 2009
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl)]
Dependent variable = Response
Independent variable = Dose
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
Default Initial Parameter Values
B-63
DRAFT- DO NOT CITE OR QUOTE
-------
Background =
Beta(l) =
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.62
Beta(l) -0.62 1
Parameter Estimates
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Log(likelihood) # Param's Deviance Test d.f.
-71.2064 3
-72.9251 2 3.43738 1
-101.108 1 59.8036 2
P-value
Est. Prob.
1
20
d.f. = 1
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
1.26376
0.985061
1.66456
% two-sided confidence
B-64
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for CNS tumors in Sprague-Dawley female rats employing
CEO in blood as an internal dose metric
Multistage Cancer Model with 0.95 Confidence Level
T3
CD
0.7
0.6
0.5
0.4
I 0.3
ts
cc
it 0.2
0.1
Multistage Cancer
Linear extrapolation
PMDL
BMD
0.001
0.002
0.003
0.004
0.005
dose
09:0312/032008
Multistage Cancer Model. (Version: 1.5; Date: 02/20/2007)
Input Data File: M:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\SD_FEMALE_CNS_BLOOD_CEO.(d)
Gnuplot Plotting File: M:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_FEMALE_CNS_BLOOD_CEO.plt
Wed Dec 03 09:03:51 2008
BMDS MODEL RUN
The form of the probability function is:
Dependent variable = Response
Independent variable = Dose
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-65
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.62
Beta(l) -0.62 1
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
Log(likelihood) # Param's Deviance Test d.f.
-71.2064 3
-72.6449 2 2.87694 1
-101.108 1 59.8036 2
P-value
Est. Prob.
d.f. = 1
80
48
48
Benchmark Dose Computation
0.1
Extra risk
Confidence level = 0.95
BMD =
BMDL =
BMDU =
Multistage Cancer Slope Factor =
B-66
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for CNS tumors in Sprague-Dawley female rats employing
AN in blood as an internal dose metric
Multistage Cancer Model with 0.95 Confidence Level
T3
CD
ts
cc
0.7
0.6
0.5
0.4
0.3
0.2
0.1
Multistage Cancer
Linear extrapolation
BMDL
0.01
0.02
0.03
dose
0.04
0.05
0.06
09:1212/032008
Multistage Cancer Model. (Version: 1.5; Date: 02/20/2007)
Input Data File: M:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\SD_FEMALE_CNS_BLOOD_AN.(d)
Gnuplot Plotting File: M:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_FEMALE_CNS_BLOOD_AN.plt
Wed Dec 03 09:12:32 2008
BMDS MODEL RUN
The form of the probability function is:
Dependent variable = Response
Independent variable = Dose
Total number of observations = 3
Total number of records with missing values
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-67
DRAFT- DO NOT CITE OR QUOTE
-------
Default Initial Parameter Values
Background = 0.106975
Beta(l) = 11.0861
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.62
Beta(l) -0.62 1
Variable
Background
Beta(1)
Estimate
0.0144346
14.7901
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
Est. Prob.
d.f. = 1
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.0071237
0.00554992
0.00939265
% two-sided confidence
B-68
DRAFT- DO NOT CITE OR QUOTE
-------
Tumor Site: Zymbal Gland (Quast, 2002; Quast et al., 1980a)
Table B-13. Incidence of Zymbal gland tumors in Sprague-Dawley rats
exposed to AN in drinking water for 2 years
Sex
Male
Female
Administered
animal dose
(ppm in drinking
water)
0
35
100
300
0
35
100
300
Equivalent
administered
animal dose"
(mg/kg-d)
0
3.42
8.53
21.2
0
4.36
10.8
25.0
Predicted internal dose metrics
AN-AUC in blood
(mg/L)
0
2.06 x 10'2
5.36 x 10"2
1.46 x 10'1
0
2.37 x 10'2
6.18 x 10"2
1.56 x 10'1
CEO-AUC in blood
(mg/L)
0
1.83 x 10'3
4.36 x 10"3
9.70 x 10'3
0
2.07 x 10'3
4.87 x 10"3
1.01 x 10'2
Incidence of
Zymbal gland
tumorsb
3/80 (4%)
4/47 (9%)
3/48 (6%)
16/48 (33%)c
1/80 (1%)
5/48 (10%)c
9/48 (19%)c
18/48 (38%)c
""Administered doses were averages calculated by the study authors based on animal B W and drinking water intake.
blncidences for Sprague-Dawley rats do not include animals from the 6- and 12-mo sacrifices and were further
adjusted to exclude (from the denominators) rats that died between 0 and 12 mos in the study.
Significantly different from controls (p < 0.05) as calculated by the study authors.
Table B-14. Summary of BMD modeling results based on incidence of
Zymbal gland tumors in Sprague-Dawley rats exposed to AN in drinking
water for 2 years
Dose metric
Best-fit model3
X2/7-valueb
AIC
BMD10C
BMDL10d
Males
Administered dose
CEO
AN
2°MS
2°MS
3°MS
0.43
0.38
0.28
142.15
142.46
143.58
11. 80 mg/kg-d
5.46 x 10'3 mg/L
8.94 x 10"2 mg/L
6.29 mg/kg-d
3.15x 10-3mg/L
4.26 x 10"2 mg/L
Females
Administered dose
CEO
AN
1°MS
1°MS
1°MS
0.94
0.95
0.85
156.80
156.77
156.98
5.66 mg/kg-d
2.40 x 10"3 mg/L
3.41 x 10"2mg/L
4. 19 mg/kg-d
1.78x 10"3mg/L
2.52 x 10"2 mg/L
"Dose-response models were fit using BMDS, version 1.4.1. "1°MS" indicates a one-stage multistage model,
"2°MS" indicates a two-stage multistage model, and "3°MS" indicates a three-stage multistage model.
bp value from the %2 goodness of fit test. Values <0.1 indicate a significant lack of fit.
CBMD10 = BMD at 10% extra risk.
dBMDL10 = 95% lower confidence limit on the BMD at 10% extra risk.
B-69
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for Zymbal gland tumors in Sprague-Dawley male rats
employing administered dose as a dose metric
Multistage Cancer Model with 0.95 Confidence Level
0.5
0.4
CD
~&
O)
o
=6 0.2
co
0.1
Multistage Cancer
Linear extrapolation
BIN/PL
BMD
10
dose
15
20
16:01 01/232009
Multistage Cancer Model. (Version: 1.5; Date: 02/20/2007)
Input Data File: M:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\SD_MALE_ZYMBAL_DW.(d)
Gnuplot Plotting File: M:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\SD_MALE_ZYMBAL_DW.plt
Fri Jan 23 16:01:39 2009
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl-beta2*doseA2)]
Total number of observations = 4
Total number of records with missing values = 0
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
B-70
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Beta(l)
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Background
Background 1
Beta(2) -0.49
Parameter Estimates
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
Model Log(likelihood) # Param's Deviance Test d.f.
Full model -68.2481 4
Fitted model -69.0738 2 1.65141 2
Reduced model -80.2977 1 24.0991 3
AIC:
P-value
Est. Prob.
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
11. 8043
6.28905
15.532
% two-sided confidence
B-71
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for Zymbal gland tumors in Sprague-Dawley male rats
employing CEO in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
0.5
0.4
-22
I
o
••e 0.2
CO
0.1
Multistage
BIVDL
BMP
0.002
0.004
0.006
0.008
0.01
dose
14:0909/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\SD_MALE_ZYMBAL_BLOOD_CEO.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAIA SD_MALE_ZYMBAL_BLOOD_CEO.pit
Thu Sep 27 14:09:41 2007
BMDS MODEL RUN
The form of the probability function is:
Total number of observations = 4
Total number of records with missing values = 0
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
Default Initial Parameter Values
Background = 0.0373755
B-72
DRAFT- DO NOT CITE OR QUOTE
-------
0
Background
1
-0.51
Variable
Background
Beta(1)
Beta(2)
Indicates that this value is not calculated.
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
Model
Full model
Fitted model
Reduced model
AIC:
Log(likelihood) # Param's Deviance Test d.f.
-68.2481 4
-69.231 2 1.96583 2
-80.2977 1 24.0991 3
Est. Prob.
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.00545893
0.00315099
0.00717256
% two-sided confidence
B-73
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for Zymbal gland tumors in Sprague-Dawley male rats
employing AN in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
0.5
0.4
CD v-'-»J
<
£Z
I 0.2
CD
0.1
0
Multistage
BIN/PL
BMD
0.02 0.04 0.06 0.08
dose
0.1
0.12
0.14
10:4409/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\SD_MALE_ZYMBAL_BLOOD_AN.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_MALE_ZYMBAL_BLOOD_AN.plt
Thu Sep 27 10:44:18 2007
The form of the probability function is:
Total number of observations = 4
Total number of records with missing values = 0
Total number of parameters in model = 4
Total number of specified parameters = 0
Degree of polynomial = 3
B-74
DRAFT- DO NOT CITE OR QUOTE
-------
Background
Beta (1)
Beta (3)
Parameter Estimates
Variable
Background
Beta(1)
Beta(2)
Beta (3)
* - Indicates that this value is not calculated.
Analysis of Deviance Table
Model
Full model
Fitted model
Reduced model
AIC: 143.576
Goodness of Fit
Est._Prob. Expected Observed
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
d.f. = 1
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.0894055
0.0425525
0.117269
% two-sided confidence
B-75
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for Zymbal gland tumors in Sprague-Dawley female rats
employing administered dose as a dose metric
CD
0.5
0.4
0.3
i 0.2
0.1
Multistage Cancer Model with 0.95 Confidence Level
Multistage Cancer
Linear extrapolation
BMDL
BMD
0 5
16:2801/232009
10 15
dose
20
25
Multistage Cancer Model. (Version: 1.5; Date: 02/20/2007)
Input Data File: M:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\SD_FEMALE_ZYMBAL_DW.(d)
Gnuplot Plotting File: M:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_FEMALE_ZYMBAL_DW.plt
Fri Jan 23 16:28:50 2009
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl)]
Total number of observations = 4
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-76
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.64
Beta(l) -0.64 1
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
# Param's
4
2
1
Deviance Test d.f.
P-value
Est. Prob.
18
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
5.66068
4.19199
8.10463
% two-sided confidence
B-77
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for Zymbal gland tumors in Sprague-Dawley female rats
employing CEO in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
03
<
C
O
cc
0.5
0.4
0.3
0.2
0.1
Multistage
BIVDL
BMD
0.002
0.004 0.006
dose
0.008
0.01
14:4309/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_FEMALE_ZYMBAL_BLOOD_CEO.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_FEMALE_ZYMBAL_BLOOD_CEO.plt
Thu Sep 27 14:43:02 2007
BMDS MODEL RUN
The form of the probability function is:
Total number of observations = 4
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
Default Initial Parameter Values
B-78
DRAFT- DO NOT CITE OR QUOTE
-------
Background =
Beta(l) =
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.65
Beta(l) -0.65 1
Parameter Estimates
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
# Param's Deviance Test d.f.
4
2 0.103357 2
1 34.6128 3
P-value
Est. Prob.
18
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.00239973
0.00178178
0.00341242
% two-sided confidence
B-79
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for Zymbal gland tumors in Sprague-Dawley female rats
employing AN in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
T3
CD
0.5
0.4
0.3
i 0.2
0.1
0
Multistage
BMD
0 0.02
0.04
0.06
0.08
dose
0.1
0.12
0.14
0.16
10:5509/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_FEMALE_ZYMBAL_BLOOD_AN.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELINGXCANCERXORAIA SD_FEMALE_ZYMBAL_BLOOD_AN.pit
Thu Sep 27 10:
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl)]
Total number of observations = 4
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-80
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.63
Beta(l) -0.63 1
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
# Param's Deviance Test d.f.
4
2 0.313733 2
1 34.6128 3
P-value
Est. Prob.
d.f.
18
P-
Benchmark Dose Computation
0.1
Extra risk
Confidence level = 0.95
BMD =
BMDL =
BMDU =
B-81
DRAFT- DO NOT CITE OR QUOTE
-------
Tumor Site: Tongue (Quast, 2002; Quast et al., 1980a)
Table B-15. Incidence of tongue tumors in Sprague-Dawley rats exposed to
AN in drinking water for 2 years
Sex
Male
Female
Administered
animal dose
(ppm in drinking
water)
0
35
100
300
0
35
100
300
Equivalent
administered
animal dose"
(mg/kg-d)
0
3.42
8.53
21.2
0
4.36
10.8
25.0
Predicted internal dose metrics
AN-AUC in blood
(mg/L)
0
2.06 x 10'2
5.36 x 10"2
1.46 x 10'1
0
2.37 x 10'2
6.18 x 10"2
1.56 x 10'1
CEO-AUC in blood
(mg/L)
0
1.83 x 10'3
4.36 x 10"3
9.70 x 10'3
0
2.07 x 10'3
4.87 x 10"3
1.01 x 10'2
Incidence of
tongue tumorsb
1/80 (1%)
2/47 (4%)
4/48 (8%)
5/48 (10%)c
0/80 (0%)
1/48 (2%)
2/48 (4%)
12/48 (25%)c
"Administered doses were averages calculated by the study authors based on animal B W and drinking water intake.
blncidences for Sprague-Dawley rats do not include animals from the 6- and 12-mo sacrifices and were further
adjusted to exclude (from the denominators) rats that died between 0 and 12 mos in the study.
Significantly different from controls (p < 0.05) as calculated by the study authors.
Table B-16. Summary of BMD modeling results based on incidence of
tongue tumors in Sprague-Dawley rats exposed to AN in drinking water for
2 years
Dose metric
Best-fit model3
X2/7-valueb
AIC
BMD10C
BMDL10d
Males
Administered dose
CEO
AN
1°MS
1°MS
1°MS
0.69
0.78
0.62
91.62
91.40
91.83
18.89 mg/kg-d
8.78 x 10'3 mg/L
1.29 x 10"1 mg/L
10.35 mg/kg-d
4.90 x 10'3 mg/L
6.97 x 10"2 mg/L
Females
Administered dose
CEO
AN
3°MS
3°MS
3°MS
0.92
0.88
0.93
84.51
84.59
84.48
15.99 mg/kg-d
6.70 x 10"3 mg/L
9.67 x 10"2 mg/L
10.64 mg/kg-d
4.74 x 10"3 mg/L
6.10x 10"2mg/L
"Dose-response models were fit using BMDS, version 1.4.1. "1°MS" indicates a one-stage multistage model and
"3°MS" indicates a three-stage multistage model.
bp value from the %2 goodness of fit test. Values <0.1 indicate a significant lack of fit.
CBMD10 = BMD at 10% extra risk.
dBMDL10 = 95% lower confidence limit on the BMD at 10% extra risk.
B-82
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for tongue tumors in Sprague-Dawley male rats employing
administered dose as a dose metric
Multistage Cancer IVbdd with 0.95 Confidence Level
0.25
0.2
0.15
•B 0.1
"G
CO
0.05
Multistage Cancer
Linear extrapolation
BMP
10
dose
15
20
15:5301/232009
Multistage Cancer Model. (Version: 1.5; Date: 02/20/2007)
Input Data File: M:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\SD_MALE_TONGUE_DW.(d)
Gnuplot Plotting File: M:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\SD_MALE_TONGUE_DW.plt
Fri Jan 23 15:53:59 2009
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl)]
Total number of observations = 4
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-83
DRAFT- DO NOT CITE OR QUOTE
-------
Background =
Beta(l) =
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.69
Beta(l) -0.69 1
Parameter Estimates
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Log(likelihood) # Param's Deviance Test d.f.
-43.4536 4
-43.8112 2 0.715245 2
-46.7384 1 6.56959 3
P-value
Est. Prob.
Benchmark Dose Computation
0.1
Extra risk
Confidence level = 0.95
BMD =
BMDL =
BMDU =
B-84
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for tongue tumors in Sprague-Dawley male rats employing
CEO in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
0.25
0.2
Fraction Affected
o
O LJ.
LJ. Ul
0.05
0
f l\
. ... .
multistage
~-
-
~-
~- -
<,
^
<
^^^^^
>
PIS/DL.
^^^—~^~^^
^~-^
i
> \
:
BIVP ;
0.002
0.004 0.006
dose
0.008
0.01
14:1909/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\SD_MALE_TONGUE_BLOOD_CEO.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_MALE_TONGUE_BLOOD_CEO.plt
Thu Sep 27 14:19:57 2007
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl)]
Total number of observations = 4
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-85
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.7
Beta(l) -0.7 1
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
Log(likelihood) # Param's Deviance Test d.f.
-43.4536 4
-43.6998 2 0.492394 2
-46.7384 1 6.56959 3
P-value
Est. Prob.
Benchmark Dose Computation
0.1
Extra risk
Confidence level = 0.95
BMD =
BMDL =
BMDU =
B-86
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for tongue tumors in Sprague-Dawley male rats employing
AN in blood as an internal dose metric
Multistage MDdel with 0.95 Confidence Level
0-25 Multistage
0.2
0.15
-B 0.1
o
CO
0.05
0
BIVPL
BMD
0.02 0.04 0.06 0.08
dose
0.1
0.12
0.14
10:4709/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\SD_MALE_TONGUE_BLOOD_AN.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_MALE_TONGUE_BLOOD_AN.plt
Thu Sep 27 10:47:13 2007
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl)]
Total number of observations = 4
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-87
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.68
Beta(l) -0.68 1
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
# Param's
4
2
1
Deviance Test d.f.
P-value
Est. Prob.
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.12938
0.0696872
0.4562
% two-sided confidence
B-88
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for tongue tumors in Sprague-Dawley female rats employing
administered dose as a dose metric
Multistage Cancer IVbdd with 0.95 Confidence Level
0.4
0.35
0.3
CD
2 °-25
< 0.2
o
TO °-15
LL
0.1
0.05
0
Multistage Cancer
Linear extrapolation
BIVDL
BMD
10
15
20
25
dose
16:2601/232009
Multistage Cancer Model. (Version: 1.5; Date: 02/20/2007)
Input Data File: M:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\SD_FEMALE_TONGUE_DW.(d)
Gnuplot Plotting File: M:\ACN DOSE-RESPONSE
MODE LING\CANCERX ORAL\S D_FEMALE_T ONGUE_DW.pit
Fri Jan 23 16:26:08 2009
BMDS MODEL RUN
The form of the probability function is:
Dependent variable = Response
Independent variable = Dose
Total number of observations = 4
Total number of records with missing values
Total number of parameters in model = 4
Total number of specified parameters = 0
Degree of polynomial = 3
B-89
DRAFT- DO NOT CITE OR QUOTE
-------
Default Initial Parameter Values
Background = 0.00365935
Beta(l) = 0.00218662
Beta (2) = 0
Beta(3) = 1.46642e-005
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Background -Beta(2)
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Beta(l) Beta(3)
1 -0.93
-0.93 1
Parameter Estimates
- Indicates that this value is not calculated.
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
Model
Full model
Fitted model
Reduced model
AIC:
# Param's Deviance Test d.f.
4
2 0.172251 2
1 29.7469 3
P-value
d.f.
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
15.9932
10.6421
19.9332
B-90
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for tongue tumors in Sprague-Dawley female rats employing
CEO in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
0.4
0.35
0.3
CD
t> 0.25
1 °'2
1 0.15
^ 0.1
0.05
0
; i\
?
^
^
r
r
r ^
. ' ... '
multistage
<
^^^^
BMDL
X
^>
^/
- ^
^
H
)> -
\
BMD ;
0.002
0.004 0.006
dose
0.008
0.01
14:4909/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_FEMALE_TONGUE_BLOOD_CEO.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_FEMALE_TONGUE_BLOOD_CEO.plt
Thu Sep 27 14:49:16 2007
BMDS MODEL RUN
The form of the probability function is:
Total number of observations = 4
Total number of records with missing values = 0
Total number of parameters in model = 4
Total number of specified parameters = 0
Degree of polynomial = 3
Default Initial Parameter Values
Background = 0.0043768
Beta(l) = 2.92561
Beta(2) = 0
Beta(3) = 245868
B-91
DRAFT- DO NOT CITE OR QUOTE
-------
Beta (1)
Beta (3)
Variable Estimate Std. Err.
Background
Beta(1)
Beta(2)
Beta (3)
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
# Param's
4
2
1
Goodness of Fit
Est. Prob.
0
1
2
12
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0. 00669671
0.00473574
0.00818609
% two-sided confidence
B-92
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for tongue tumors in Sprague-Dawley female rats employing
AN in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
CD
0.4
0.35
0.3
0.25
0.2
0.15
0.1
0.05
0
i Multistage
?
^
^
^
r
r
^ <
x____— — — ~~~~~
0 0.02
' '
>
/^
^^__^
BIVPL
^~~~~^~^
,^^
- \
'-_
'-.
y -
^
BMD ;
0.04 0.06 0.08 0.1 0.12 0.14 0.16
dose
10:5709/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_FEMALE_TONGUE_BLOOD_AN.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELINGXCANCERXORAIA SD_FEMALE_TONGUE_BLOOD_AN.pit
Thu Sep 27 10:57:41 2007
BMDS MODEL RUN
The form of the probability function is:
Total number of observations = 4
Total number of records with missing values = 0
Total number of parameters in model = 4
Total number of specified parameters = 0
Degree of polynomial = 3
B-93
DRAFT- DO NOT CITE OR QUOTE
-------
Beta (1)
Beta (3)
Parameter Estimates
Variable
Background
Beta(1)
Beta(2)
Beta (3)
Indicates that this value is not calculated.
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
Model
Full model
Fitted model
Reduced model
Est. Prob.
0
1
2
12
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0 . 0966976
0.0610292
0.122886
B-94
DRAFT- DO NOT CITE OR QUOTE
-------
Tumor Site: Mammary Gland (Quast, 2002; Quast et al., 1980a)
Table B-17. Incidence of mammary gland tumors in Sprague-Dawley rats
exposed to AN in drinking water for 2 years
Sex
Female
Administered
animal dose
(ppm in drinking
water)
0
35
100
300
Equivalent
administered
animal dose"
(mg/kg-d)
0
4.36
10.8
25.0
Predicted internal dose metrics
AN-AUC in blood
(mg/L)
0
2.37 x 10'2
6.18 x 10"2
1.56 x 10'1
CEO-AUC in blood
(mg/L)
0
2.07 x 10'3
4.87 x 10"3
1.01 x 10'2
Incidence of
mammary gland
tumorsb
58/80 (72%)
42/47 (89%)c
42/48 (88%)c
35/48 (73%)
"Administered doses were averages calculated by the study authors based on animal B W and drinking water intake.
blncidences for Sprague-Dawley rats do not include animals from the 6- and 12-mo sacrifices and were further
adjusted to exclude (from the denominators) rats that died between 0 and 12 mos in the study.
Significantly different from controls (p < 0.05) as calculated by the study authors.
Table B-18. Summary of BMD modeling results based on incidence of
mammary gland tumors in Sprague-Dawley rats exposed to AN in drinking
water for 2 years
Dose metric
Best-fit model3
X2/7-valueb
AIC
BMD10C
BMDL10d
Females
Administered dose
CEO
AN
l°MSe
l°MSe
l°MSe
0.25
0.27
0.24
171.81
171.69
171.93
1.22 mg/kg-d
5.50 x 10"4 mg/L
7.05 x 10"3 mg/L
0.66 mg/kg-d
2.98 x 10'4 mg/L
3.77x 10"3mg/L
"Dose-response models were fit using BMDS, version 1.4.1. "1°MS" indicates a one-stage multistage model.
bp value from the %2 goodness of fit test. Values <0.1 indicate a significant lack of fit.
CBMD10 = BMD at 10% extra risk.
dBMDL10 = 95% lower confidence limit on the BMD at 10% extra risk.
eHighest dose dropped prior to model fitting.
B-95
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for mammary gland tumors in Sprague-Dawley female rats
employing administered dose as a dose metric
Multistage Cancer IVbdd with 0.95 Confidence Level
0.95
0.9
-o 0.85
CD
•G
»t= 0.8
-------
Default Initial Parameter Values
Background = 0.771359
Beta(l) = 0.0674842
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.58
Beta(l) -0.58 1
Variable Estimate Std. Err.
Background
Beta(1)
- Indicates that this value is not calculated.
Analysis of Deviance Table
Model Log(likelihood) # Param's Deviance Test d.f. P-value
Full model -83.2234 3
Fitted model -83.9062 2 1.36543 1
Reduced model -86.3815 1 6.31609 2
AIC:
Dose Est. Prob.
d.f. = 1
80
48
48
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 1.22462
BMDL = 0.659237
BMDU = 4.94647
% two-sided confidence
B-97 DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for mammary gland tumors in Sprague-Dawley female rats
employing CEO in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
0.95
0.9
-o 0.85
CD
0.8
0.75
cc
£ 0.7
0.65
0.6
Multistage
BIN/PL
BMD
0.001
0.002 0.003
dose
0.004
0.005
14:5309/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_FEMALE_MAMMARY_BLOOD_CEO.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_FEMALE_MAMMARY_BLOOD_CEO.plt
Thu Sep 27 14:53:05 2007
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl)]
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-98
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.58
Beta(l) -0.58 1
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
# Param's
3
2
1
Deviance Test d.f.
P-value
Est. Prob.
d.f. = 1
80
48
48
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.000549528
0.000298184
0.00214159
% two-sided confidence
B-99
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for mammary gland tumors in Sprague-Dawley female rats
employing AN in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
0.95
0.9
-o 0.85
CD
0.8
0.75
cc
£ 0.7
0.65
0.6
Multistage
BIN/PL
0.01
0.02
0.03
dose
0.04
0.05
0.06
11:0009/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_FEMALE_MAMMARY_BLOOD_AN.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\SD_FEMALE_MAMMARY_BLOOD_AN.pit
Thu Sep 27 11:00:59 2007
BMDS MODEL RUN
The form of the probability function is:
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
Default Initial Parameter Values
B-100
DRAFT- DO NOT CITE OR QUOTE
-------
Background =
Beta(l) =
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.58
Beta(l) -0.58 1
Parameter Estimates
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Log(likelihood) # Param's Deviance Test d.f.
-83.2234 3
-83.9649 2 1.48294 1
-86.3815 1 6.31609 2
P-value
Est. Prob.
ChiA2 = 1.41
d.f. = 1
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0 . 00705091
0.00376506
0.0295793
B-101
DRAFT- DO NOT CITE OR QUOTE
-------
F344 Rats (Johannsen and Levinskas, 2002b; Biodynamics, 1980b)
Tumor Site: Forestomach
Table B-19. Incidence of forestomach (nonglandular) tumors in F344 rats
exposed to AN in drinking water for 2 years
Sex
Male
Female
Administered
animal dose
(ppm in drinking
water)
0
1
3
10
30
100
0
1
o
J
10
30
100
Equivalent
administered
animal dose"
(mg/kg-d)
0
0.08
0.25
0.83
2.48
8.37
0
0.12
0.36
1.25
3.65
10.90
Predicted internal dose metrics
AN-AUC in blood
(mg/L)
0
4.33 x 10'4
1.35 x 10'3
4.52 x 10'3
1.37 x 10'2
4.85 x 10'2
0
5.73 x 10"4
1.72 x 10"j
6.02 x 10"j
1.79 x 10"2
5.63 x 10"2
CEO-AUC in blood
(mg/L)
0
4.06 x 10°
1.27 x 10'4
4.19 x 10'4
1.23 x 10'3
3.97 x 10'3
0
5.32 x 10°
1.59 x 10"4
5.49 x 10"4
1.58 x 10"J
4.46 x 10"J
Incidence of
forestomach
tumorsb
0/159 (0%)
1/80 (1%)
4/78 (5%)c
3/80 (4%)c
4/80 (5%)c
1/77 (1%)
0/157 (0%)
1/80 (1%)
2/79 (3%)
2/77 (3%)
4/80 (5%)c
2/75 (3%)
"Administered doses were averages calculated by the study authors based on animal B W and drinking water intake.
blncidences for F344 rats do not include animals from the 6- and 12-mo sacrifices and were further adjusted to
exclude (from the denominators) rats that died between 0 and 12 mos in the study.
Significantly different from controls (p < 0.05) as calculated by the study authors.
B-102
DRAFT- DO NOT CITE OR QUOTE
-------
Table B-20. Summary of BMD modeling results based on incidence of
forestomach (nonglandular) tumors in F344 rats exposed to AN in drinking
water for 2 years
Dose metric
Best-fit model3
X2/7-valueb
AIC
BMD10C
BMDL10d
Males
Administered
dose
CEO
AN
l°MSe
l°MSe
l°MSe
0.18
0.19
0.18
74.08
74.04
74.12
1.19mg/kg-d
6.03 x 10"4 mg/L
6.48 x 10'3 mg/L
0.70 mg/kg-d
3.55 x 10"4mg/L
3.81 x 10-3mg/L
Females
Administered
dose
CEO
AN
l°MSf
l°MSf
l°MSf
0.83
0.83
0.82
96.64
96.62
96.65
8.43 mg/kg-d
3.65x 10'3mg/L
4.13 x 10"2mg/L
3. 89 mg/kg-d
1.69x 10-3mg/L
1.90 x 10"2mg/L
"Dose-response models were fit using BMDS, version 1.4.1. "1°MS" indicates a one-stage multistage model.
bp value from the %2 goodness of fit test. Values <0.1 indicate a significant lack of fit.
CBMD10 = BMD at 10% extra risk.
dBMDL10 = 95% lower confidence limit on the BMD at 10% extra risk.
eTwo highest doses dropped prior to model fitting.
fHighest dose dropped prior to model fitting.
B-103
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for forestomach tumors in F344 male rats employing
administered dose as a dose metric
Multistage Cancer IVbdd with 0.95 Confidence Level
0.14
0.12
0.1
"§ 0.08
g 0.06
CD
it 0.04
0.02
0
Multistage Cancer
Linear extrapolation
BMDL
0.2
0.4
0.6
dose
0.8
1.2
14:2901/262009
Multistage Cancer Model. (Version: 1.5; Date: 02/20/2007)
Input Data File: M:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\F344_MALE_FORESTOMACH_DW.(d)
Gnuplot Plotting File: M:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_MALE_FORESTOMACH_DW.plt
Mon Jan 26 14:29:05 2009
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl)]
Total number of observations = 4
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-104
DRAFT- DO NOT CITE OR QUOTE
-------
Default Initial Parameter Values
Background = 0.0148134
Beta(l) = 0.0377128
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Background
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Beta(1)
Beta(1)
Parameter Estimates
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
Log(likelihood) # Param's Deviance Test d.f.
-33.9463 4
-36.0377 1 4.18282 3
-39.1548 1 10.417 3
of Fit
Est. Prob.
d.f. =3
P-value = 0.1836
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
1.19186
0.701326
3.81474
Taken together, (0.701326, 3.81474) is a 90
interval for the BMD
% two-sided confidence
B-105
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for forestomach tumors in F344 male rats employing CEO in
blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
0.14
0.12
0.1
l£ 0.08
g 0.06
CO
it 0.04
0.02
0
\ Multist
^
r
^
^ <
j^"""^
age
<
,^-
\ ^^^~^
^^
BIVDL
^^^^
'-_
-
\
BM
0.0001
0.0002
0.0003
dose
0.0004
0.0005
0.0006
14:5709/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_MALE_FORESTOMACH_BLOOD_CEO.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_MALE_FORESTOMACH_BLOOD_CEO.plt
Thu Sep 27 14:57:13 2007
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl)]
Total number of observations = 4
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-106
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Background
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Beta(1)
Beta(1)
Parameter Estimates
Estimate
0
174.815
- Indicates that this value is not calculated.
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
Log(likelihood) # Param's Deviance Test d.f.
-33.9463 4
-36.0209 1 4.14909 3
-39.1548 1 10.417 3
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.000602696
0.000354643
0.00191036
% two-sided confidence
B-107
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for forestomach tumors in F344 male rats employing AN in
blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
0.14
0.12
0.1
l£ 0.08
g 0.06
CO
it 0.04
0.02
0
\ Multist
^
r
^
^ <
age
<
^^-
> ^^-^^
_^^
BIVDL
^-^^
'-_
-
\
BM
0.001 0.002 0.003 0.004
dose
0.005
0.006
11:0409/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_MALE_FORESTOMACH_BLOOD_AN.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_MALE_FORESTOMACH_BLOOD_AN.plt
Thu Sep 27 11:04:04 2007
BMDS MODEL RUN
The form of the probability function is:
Total number of observations = 4
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
Default Initial Parameter Values
B-108
DRAFT- DO NOT CITE OR QUOTE
-------
Background =
Beta(l) =
Beta (1)
Beta(1)
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
Est. Prob.
Goodness of Fit
80
B-109
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for forestomach tumors in F344 female rats employing
administered dose as a dose metric
Multistage Cancer IVbdd with 0.95 Confidence Level
0.14
0.12
0.1
CD
Ijj 0.08
it
.1 O-06
CO
ul 0.04
0.02
0
?
?
r
- -,
?
? -
Multistac
Linear exti
_^ ^
X
\
T'
/
y^-
je Cancer
apolation
/
/
^^^
/
y
BIVDL
^^^^^
^^^^
*^^
^
T
-
:
BM
012345678
dose
14:4901/262009
Multistage Cancer Model. (Version: 1.5; Date: 02/20/2007)
Input Data File: M:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_FEMALE_FORESTOMACH_DW.(d)
Gnuplot Plotting File: M:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_FEMALE_FORESTOMACH_DW.plt
Mon Jan 26 14:49:27 2009
BMDS MODEL RUN
The form of the probability function is:
Dependent variable = Response
Independent variable = Dose
Total number of observations = 5
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-110
DRAFT- DO NOT CITE OR QUOTE
-------
Default Initial Parameter Values
Background = 0.0127929
Beta(l) = 0.0107517
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.58
Beta(l) -0.58 1
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
Goodness of Fit
Est. Prob.
d.f. =3
1.155
4.346
P-value = 0.8283
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
8.42727
3.88972
48.9918
Taken together, (3.88972, 48.9918) is a 90
interval for the BMD
Multistage Cancer Slope Factor =
B-lll
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for forestomach tumors in F344 female rats employing CEO
in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
0.14
0.12
0.1
CD
2 0.08
§ 0.06
ts
CO
ii 0.04
0.02
0
f Multistage
?
?
1 _
r
? -
^>
<
> ^^<
^^<
\
">
BIVPL
^^
^^^^
~^~~^
'-_
\
L
-_
BM
0.0005 0.001
0.0015 0.002 0.0025 0.003 0.0035
dose
15:0709/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_FEMALE_FORESTOMACH_BLOOD_CEO.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_FEMALE_FORESTOMACH_BLOOD_CEO.plt
Thu Sep 27 15:07:15 2007
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl)]
Total number of observations = 5
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-112
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.58
Beta(l) -0.58 1
Variable
Background
Beta(1)
- Indicates that this value is not calculated.
Analysis of Deviance Table
Model Log(likelihood) # Param's Deviance Test d.f. P-value
Full model -45.9122 5
Fitted model -46.3121 2 0.799693 3
Reduced model -48.4586 1 5.09287 4
AIC:
Est. Prob.
0 . 0000
0.0001
0 . 0101
0.0116
1.587
0. 930
1
1
157
80
-0.469
0. 072
0.0002 0.0146
0.0005 0.0257
0.0016 0.0543 4.342 4 80
d.f. =3
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 0.0036478
BMDL = 0.00168696
BMDU = 1361.26
% two-sided confidence
B-l 13 DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for forestomach tumors in F344 female rats employing AN in
blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
0.14
0.12
0.1
CD
2 0.08
§ 0.06
ts
CO
ii 0.04
0.02
0
f Multistage
?
?
r
? -
^>
\
^^
' '
^~^^
^
r
BMPL
^^^
^^^
~-^^
~-_
\
-_
BM
0.005 0.01
0.015 0.02 0.025 0.03
dose
0.035 0.04
11:3009/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_FEMALE_FORESTOMACH_BLOOD_AN.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_FEMALE_FORESTOMACH_BLOOD_AN.plt
Thu Sep 27 11:30:39 2007
BMDS MODEL RUN
The form of the probability function is:
Total number of observations = 5
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
Default Initial Parameter Values
B-114
DRAFT- DO NOT CITE OR QUOTE
-------
Background =
Beta(l) =
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.58
Beta(l) -0.58 1
Parameter Estimates
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Log(likelihood) # Param's Deviance Test d.f.
-45.9122 5
-46.3252 2 0.826038 3
-48.4586 1 5.09287 4
P-value
Prob.
80
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.0413364
0.0190324
0.242436
% two-sided confidence
B-115
DRAFT- DO NOT CITE OR QUOTE
-------
Tumor Site: CNS (Johannsen and Levinskas, 2002b; Biodynamics, 1980b)
Table B-21. Incidence of CNS tumors in F344 rats exposed to AN in
drinking water for 2 years
Sex
Male
Female
Administered
animal dose
(ppm in drinking
water)
0
1
3
10
30
100
0
1
3
10
30
100
Equivalent
administered
animal dosea
(mg/kg-d)
0
0.08
0.25
0.83
2.48
8.37
0
0.12
0.36
1.25
3.65
10.90
Predicted internal dose metrics
AN-AUC in blood
(mg/L)
0
4.33 x 10"4
1.35 x 10"J
4.52 x 10"J
1.37 x 10'2
4.85 x 10'2
0
5.73 x 10'4
1.72 x 10'J
6.02 x 10'3
1.79 x 10'2
5.63 x 10'2
CEO-AUC in blood
(mg/L)
0
4.06 x 10"'
1.27 x 10"4
4.19x 10"4
1.23 x 10"j
3.97 x 10"j
0
5.32 x lO'3
1.59 x 10'4
5.49 x 10'4
1.58 x 10'3
4.46 x 10'3
Incidence of CNS
tumorsb
0/160 (0%)
2/80 (3%)
1/78 (1%)
2/80 (3%)
10/79 (13%)c
25/76 (33%)c
1/157 (1%)
1/80 (1%)
2/80 (3%)
5/75 (7%)
6/80 (8%)c
24/76 (32%)c
""Administered doses were averages calculated by the study authors based on animal B W and drinking water intake.
blncidences for F344 rats do not include animals from the 6- and 12-mo sacrifices and were further adjusted to
exclude (from the denominators) rats that died between 0 and 12 mos in the study.
Significantly different from controls (p < 0.01) as calculated by the study authors.
Table B-22. Summary of BMD modeling results based on incidence of CNS
tumors in F344 rats exposed to AN in drinking water for 2 years
Dose metric
Best-fit model3
^/J-value13
AIC
BMD10C
BMDL10d
Males
Administered dose
CEO
AN
1°MS
1°MS
1°MS
0.73
0.70
0.74
240.57
240.75
240.44
2.41 mg/kg-d
1.16x 10"3mg/L
1.37x 10"2mg/L
1.81 mg/kg-d
8.74 x 10"4 mg/L
1.03 x 10"2mg/L
Females
Administered dose
CEO
AN
1°MS
1°MS
1°MS
0.68
0.66
0.68
222.13
222.30
222.04
3.34 mg/kg-d
1.39 x 10"3mg/L
1.70x 10-2mg/L
2.52 mg/kg-d
1.05 x 10"3mg/L
1.28x 10-2mg/L
aDose-response models were fit using BMDS, version 1.4.1. "1°MS" indicates a one-stage multistage model.
bp value from the %2 goodness of fit test. Values <0.1 indicate a significant lack of fit.
CBMD10 = BMD at 10% extra risk.
dBMDL10 = 95% lower confidence limit on the BMD at 10% extra risk.
B-116
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for CNS tumors in F344 male rats employing administered
dose as a dose metric
Multistage Cancer Model with 0.95 Confidence Level
0.4
I0'3
o 0.2
CD
LL
0.1
0
Multistage Cancer
Linear extrapolation
BMDL
BMD
1
4
dose
8
14:3401/262009
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl)]
Dependent variable = Response
Independent variable = Dose
Total number of observations = 6
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
Default Initial Parameter Values
B-117
DRAFT- DO NOT CITE OR QUOTE
-------
Background =
Beta(l) =
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.47
Beta(l) -0.47 1
Parameter Estimates
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
# Param's
Goodness of Fit
Test d.f.
P-value
Prob.
d.f. = 4
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
2.40586
1.8134
3.32187
Taken together, (1.8134 , 3.32187) is a 90
interval for the BMD
Multistage Cancer Slope Factor =
B-118
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for CNS tumors in F344 male rats employing CEO in blood as
an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
0.4
i °-3
<
o 0.2
ro
u_
0.1
Multistage
BIN/PL
BMD
0 0.0005
15:0009/272007
0.001 0.0015
0.002
dose
0.0025 0.003 0.0035 0.004
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\F344_MALE_CNS_BLOOD_CEO.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_MALE_CNS_BLOOD_CEO.plt
Thu Sep 27 15:00:12 2007
The form of the probability function is:
Total number of observations = 6
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
Default Initial Parameter Values
B-119
DRAFT- DO NOT CITE OR QUOTE
-------
Background =
Beta(l) =
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.47
Beta(l) -0.47 1
Parameter Estimates
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Log(likelihood)
-117.105
-118.375
-151.112
# Param's
Goodness of Fit
Test d.f.
P-value
Prob.
d.f. = 4
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.00115833
0.000873641
0.00159795
% two-sided confidence
B-120
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for CNS tumors in F344 male rats employing AN in blood as
an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
0.4
i °-3
<
o 0.2
ro
0.1
Multistage
BMDL
BMD
0.01
0.02 0.03
dose
0.04
0.05
11:0709/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\F344_MALE_CNS_BLOOD_AN.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_MALE_CNS_BLOOD_AN.plt
Thu Sep 27 11:07:15 2007
The form of the probability function is:
Total number of observations = 6
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
Default Initial Parameter Values
B-121
DRAFT- DO NOT CITE OR QUOTE
-------
Background =
Beta(l) =
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.46
Beta(l) -0.46 1
Parameter Estimates
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Log(likelihood)
-117.105
-118.218
-151.112
# Param's
Goodness of Fit
Test d.f.
P-value
Prob.
d.f. = 4
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.0137355
0.0103458
0.0189838
% two-sided confidence
B-122
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for CNS tumors in F344 female rats employing administered
dose as a dose metric
Multistage Cancer Model with 0.95 Confidence Level
o
t5
CO
0.4
0.3
0.2
0.1
0
Multistage Cancer
Linear extrapolation
BMPL
8
10
dose
14:5501/262009
Multistage Cancer Model. (Version: 1.5; Date: 02/20/2007)
Input Data File: M:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\F344_FEMALE_CNS_DW.(d)
Gnuplot Plotting File: M:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_FEMALE_CNS_DW.plt
Mon Jan 26 14:55:28 2009
BMDS MODEL RUN
The form of the probability function is:
Dependent variable = Response
Independent variable = Dose
Total number of observations = 6
Total number of records with missing values
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-123
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.49
Beta(l) -0.49 1
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
Deviance Test d.f.
P-value
Est. Prob.
d.f. = 4
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
3.34421
2.51554
4.64166
Taken together, (2.51554, 4.64166) is a 90
interval for the BMD
% two-sided confidence
B-124
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for CNS tumors in F344 female rats employing CEO in blood
as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
0.4
•§ 0.3
o
t5
CO
0.2
0.1
0
Multistage
BMDL
B
-------
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.5
Beta(l) -0.5 1
Variable
Background
Beta(1)
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
Deviance Test d.f.
P-value
Est. Prob.
d.f. = 4
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.0013937
0.00104985
0.00193063
B-126
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for CNS tumors in F344 female rats employing AN in blood
as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
0.4
•§ 0.3
o
t5
CO
0.2
0.1
0
Multistage
BIVP
0.01
0.02
0.03
dose
0.04
0.05
11:3409/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\F344_FEMALE_CNS_BLOOD_AN.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_FEMALE_CNS_BLOOD_AN.plt
Thu Sep 27 11:34:20 2007
The form of the probability function is:
Total number of observations = 6
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
Default Initial Parameter Values
B-127
DRAFT- DO NOT CITE OR QUOTE
-------
Background =
Beta(l) =
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.49
Beta(l) -0.49 1
Parameter Estimates
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
# Param's
Goodness of Fit
Test d.f.
P-value
Prob.
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.0170014
0.0127694
0.0236456
% two-sided confidence
B-128
DRAFT- DO NOT CITE OR QUOTE
-------
Tumor Site: Zymbal's Gland (Johannsen and Levinskas, 2002b; Biodynamics, 1980b)
Table B-23. Incidence of Zymbal gland tumors in F344 rats exposed to AN
in drinking water for 2 years
Sex
Male
Female
Administered
animal dose
(ppm in drinking
water)
0
1
o
J
10
30
100
0
1
o
J
10
30
100
Equivalent
administered
animal dose"
(mg/kg-d)
0
0.08
0.25
0.83
2.48
8.37
0
0.12
0.36
1.25
3.65
10.90
Predicted internal dose metrics
AN-AUC in blood
(mg/L)
0
4.33 x 10"4
1.35 x 10"3
4.52 x 10"3
1.37 x 10"2
4.85 x 10"2
0
5.73 x 10"4
1.72 x 10"3
6.02 x 10"3
1.79 x 10"2
5.63 x 10"2
CEO-AUC in blood
(mg/L)
0
4.06 x 10"5
1.27 x 10"4
4.19 x 10"4
1.23 x 10"3
3.97 x 10"3
0
5.32 x 10"5
1.59 x 10"4
5.49 x 10"4
1.58 x 10"3
4.46 x 10"3
Incidence of
Zymbal's gland
tumorsb
1/147 (1%)
1/76 (1%)
0/73 (0%)
0/67 (0%)
2/71 (3%)c
14/68 (21%)c
0/157 (0%)
0/73 (0%)
0/73 (0%)
0/70 (0%)
2/73 (3%)c
8/62 (13%)c
"Administered doses were averages calculated by the study authors based on animal B W and drinking water intake.
blncidences for F344 rats do not include animals from the 6- and 12-mo sacrifices and were further adjusted to
exclude (from the denominators) rats that died between 0 and 12 mos in the study.
Significantly different from controls (p < 0.01) as calculated by the study authors.
Table B-24. Summary of BMD modeling results based on incidence of
Zymbal gland tumors in F344 rats exposed to AN in drinking water for
2 years
Dose metric
Best-fit model3
X2 ^-valueb
AIC
BMD10C
BMDL10d
Males
Administered dose
CEO
AN
2°MS
2°MS
2°MS
0.78
0.77
0.78
116.51
116.53
116.52
5.73 mg/kg-d
2.73 x 10'3 mg/L
3.31 x 10"2mg/L
4.55 mg/kg-d
2.19 x 10-3mg/L
2.59 x 10"2 mg/L
Females
Administered dose
CEO
AN
2°MS
2°MS
1°MS
0.95
0.99
0.89
70.80
68.68
70.82
9. 16 mg/kg-d
3.78x 10"3mg/L
5.41 x 10"2mg/L
7.17 mg/kg-d
2.97 x 10"3 mg/L
3.35 x 10"2mg/L
"Dose-response models were fit using BMDS, version 1.4.1. "1°MS" indicates a one-stage multistage model and
"2°MS" indicates a two-stage multistage model.
bp value from the %2 goodness of fit test. Values <0.1 indicate a significant lack of fit.
CBMD10 = BMD at 10% extra risk.
dBMDL10 = 95% lower confidence limit on the BMD at 10% extra risk.
B-129
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for Zymbal gland tumors in F344 male rats employing
administered dose as a dose metric
Multistage Cancer IVbdd with 0.95 Confidence Level
0.35
0.3
0.25
CD
1 °'2
§ 0.15
•o
CO
it 0.1
0.05
0
F Multistage Cancer
; Linear extrapolation
r
F
F
F 4
X
> ?
0
^^^-'
1 2
' '
^
^/
BMDL
-_
-_
> F
F
BMD ;
345678
dose
14:41 01/262009
Multistage Cancer Model. (Version: 1.5; Date: 02/20/2007)
Input Data File: M:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\F344_MALE_ZYMBAL_DW.(d)
Gnuplot Plotting File: M:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_MALE_ZYMBAL_DW.plt
Mon Jan 26 14:41:15 2009
BMDS MODEL RUN
The form of the probability function is:
Dependent variable = Response
Independent variable = Dose
Total number of observations = 6
Total number of records with missing values
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
B-130
DRAFT- DO NOT CITE OR QUOTE
-------
Default Initial Parameter Values
Background = 0.00522088
Beta(l) = 0
Beta(2) = 0.00321913
and do not appear in the correlation matrix )
Background Beta(2)
Parameter Estimates
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
Model
Full model
Fitted model
Reduced model
AIC:
Log(likelihood) # Param's Deviance Test d.f.
-54.9964 6
-56.2568 2 2.52093 4
-77.5815 1 45.1702 5
116.514
Goodness of Fit
Est. Prob.
14
d.f. = 4
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
5.72676
4.54678
7.26717
Taken together, (4.54678, 7.26717) is a 90
interval for the BMD
% two-sided confidence
B-131
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for Zymbal gland tumors in F344 male rats employing CEO
in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
0.35
0.3
0.25
T3
CD
2 0.2
.1 °-15
CO
ii: 0.1
0.05
0
f Multistag
-
r
\7
\
^
? <
^
/
>— "^^
BMPL
BMD
^
\ '_
-_
0 0.0005 0.001 0.0015 0.002 0.0025 0.003 0.0035 0.004
dose
15:0209/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_MALE_ZYMBAL_BLOOD_CEO.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_MALE_ZYMBAL_BLOOD_CEO.plt
Thu Sep 27 15:02:53 2007
BMDS MODEL RUN
The form of the probability function is:
Dependent variable = Response
Independent variable = Dose
Total number of observations = 6
Total number of records with missing values = 0
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
B-132
DRAFT- DO NOT CITE OR QUOTE
-------
Default Initial Parameter Values
Background = 0.00480523
Beta(l) = 0
Beta(2) = 14327.9
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Beta(l)
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Background Beta(2)
1 -0.38
-0.38 1
Variable
Background
Beta(1)
Beta(2)
Indicates that this value is not calculated.
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
Model
Full model
Fitted model
Reduced model
AIC:
Log(likelihood) # Param's Deviance Test d.f.
-54.9964 6
-56.2667 2 2.54061 4
-77.5815 1 45.1702 5
P-value
Est. Prob.
14
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.00272678
0.00218772
0. 00345798
B-133
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for Zymbal gland tumors in F344 male rats employing AN in
blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
03
<
C
O
cc
0.35
0.3
0.25
0.2
0.15
0.1
0.05
0
Multistage
BMDL
BMD
0.01
0.02 0.03
dose
0.04
0.05
11:2709/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_MALE_ZYMBAL_BLOOD_AN.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELINGXCANCERXORAIA F344_MALE_ZYMBAL_BLOOD_AN.pit
Thu Sep 27 11:27:02 2007
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl-beta2*doseA2)]
Total number of observations = 6
Total number of records with missing values = 0
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
B-134
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Beta(l)
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Background
Background 1
Beta(2) -0.37
Variable Estimate Std. Err.
Background
Beta(1)
Beta(2)
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
Log(likelihood) # Param's Deviance Test d.f.
-54.9964 6
-56.2582 2 2.52365 4
-77.5815 1 45.1702 5
Goodness of Fit
Prob.
d.f. = 4
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.033067
0.0258907
0.0419928
% two-sided confidence
B-135
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for Zymbal gland tumors in F344 female rats employing
administered dose as a dose metric
Multistage Cancer IVbdd with 0.95 Confidence Level
0.25
0.2
CD
I °'15
0.1
0.05
Multistage Cancer
Linear extrapolation
BIVDL
BMD
8
10
dose
15:01 01/262009
Multistage Cancer Model. (Version: 1.5; Date: 02/20/2007)
Input Data File: M:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\F344_FEMALE_ZYMBAL_DW.(d)
Gnuplot Plotting File: M:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_FEMALE_ZYMBAL_DW.plt
Mon Jan 26 15:01:22 2009
BMDS MODEL RUN
The form of the probability function is:
Dependent variable = Response
Independent variable = Dose
Total number of observations = 6
Total number of records with missing values
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
B-136
DRAFT- DO NOT CITE OR QUOTE
-------
Default Initial Parameter Values
Background = 0
Beta(l) = 0.00476273
Beta(2) = 0.000742952
and do not appear in the correlation matrix )
Beta(l) Beta(2)
1 -0.95
-0.95 1
Parameter Estimates
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
Model
Full model
Fitted model
Reduced model
AIC:
Log(likelihood) # Param's Deviance Test d.f.
-33.0086 6
-33.4019 2 0.786523 4
-49.1799 1 32.3425 5
Goodness of Fit
Est. Prob.
d.f. = 4
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
9.16446
7.17181
13.5889
Taken together, (7.17181, 13.5889) is a 90
interval for the BMD
% two-sided confidence
B-137
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for Zymbal gland tumors in F344 female rats employing CEO
in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
CD
0.25
0.2
0.15
=§ 0.1
cc
0.05
0
Multistage
BMDL
0 0.0005 0.001 0.0015 0.002 0.0025 0.003 0.0035 0.004 0.0045
dose
15:1309/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_FEMALE_ZYMBAL_BLOOD_CEO.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_FEMALE_ZYMBAL_BLOOD_CEO.plt
Thu Sep 27 15:13:27 2007
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl-beta2*doseA2)]
Total number of observations = 6
Total number of records with missing values = 0
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
B-138
DRAFT- DO NOT CITE OR QUOTE
-------
Default Initial Parameter Values
Background = 0
Beta(l) = 9.4165
Beta(2) = 4928.95
and do not appear in the correlation matrix )
Beta (2)
1
Parameter Estimates
Variable
Background
Beta(1)
Beta(2)
Indicates that this value is not calculated.
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
Analysis of Deviance Table
Model Log(likelihood) # Param's Deviance Test d.f.
Full model -33.0086 6
Fitted model -33.3423 1 0.667255 5
Reduced model -49.1799 1 32.3425 5
of Fit
Est. Prob.
d.f. =
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0. 00377789
0.00297198
0.00541558
% two-sided confidence
B-139
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for Zymbal gland tumors in F344 female rats employing AN
in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
CD
0.25
0.2
0.15
=§ 0.1
cc
0.05
0
Multistage
BIVDL
BMD
0.01
0.02
0.03
dose
0.04
0.05
11:3709/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_FEMALE_ZYMBAL_BLOOD_AN.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELINGX CANCERX ORAIA F3 4 4_FEMALE_Z YMBAL_BLOOD_AN. pi t
Thu Sep 27 11:37:11 2007
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl)]
Total number of observations = 6
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-140
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Background
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Beta(1)
Beta(1)
Parameter Estimates
- Indicates that this value is not calculated.
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
Model
Full model
Fitted model
Reduced model
# Param's
6
1
1
Goodness of Fit
Est. Prob.
d.f. =
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.0541423
0.0335374
0.0957005
% two-sided confidence
B-141
DRAFT- DO NOT CITE OR QUOTE
-------
Tumor Site: Mammary Gland (Johannsen and Levinskas, 2002b; Biodynamics, 1980b)
Table B-25. Incidence of mammary gland tumors in F344 rats exposed to
AN in drinking water for 2 years
Sex
Female
Administered
animal dose
(ppm in drinking
water)
0
1
o
J
10
30
100
Equivalent
administered
animal dose"
(mg/kg-d)
0
0.12
0.36
1.25
3.65
10.90
Predicted internal dose metrics
AN-AUC in blood
(mg/L)
0
5.73 x 10"4
1.72 x 10"3
6.02 x 10"3
1.79 x 10"2
5.63 x 10"2
CEO-AUC in blood
(mg/L)
0
5.32 x 10"5
1.59 x 10"4
5.49 x 10"4
1.58 x 10"3
4.46 x 10"3
Incidence of
mammary gland
tumorsb
14/156 (9%)
8/80 (10%)
6/80 (8%)
9/80(11%)
12/80 (15%)
14/73 (19%)
"Administered doses were averages calculated by the study authors based on animal BW and drinking water intake.
Incidences for F344 rats do not include animals from the 6- and 12-mo sacrifices and were further adjusted to
exclude (from the denominators) rats that died between 0 and 12 mos in the study.
Table B-26. Summary of BMD modeling results based on incidence of
mammary gland tumors in F344 rats exposed to AN in drinking water for 2
years
Dose metric
Best-fit model3
X2/7-valueb
AIC
BMD10C
BMDL10d
Females
Administered dose
CEO
AN
1°MS
1°MS
1°MS
0.94
0.94
0.93
388.94
388.88
389.00
8.73 mg/kg-d
3.58x 10'3mg/L
4.51 x 10"2mg/L
4.77 mg/kg-d
1.97x 10'3mg/L
2.45 x 10"2 mg/L
aDose-response models were fit using BMDS, version 1.4.1. "1°MS" indicates a one-stage multistage model.
bp value from the %2 goodness of fit test. Values <0.1 indicate a significant lack of fit.
CBMD10 = BMD at 10% extra risk.
dBMDL10 = 95% lower confidence limit on the BMD at 10% extra risk.
B-142
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for mammary gland tumors in F344 female rats employing
administered dose as a dose metric
Multistage Cancer IVbdd with 0.95 Confidence Level
0.3
0.25
CD
I °'2
.S 0.15
CO
LL
0.1
0.05
Multistage Cancer
Linear extrapolation
BMDL
BMD
4 6
dose
8
10
15:0401/262009
Multistage Cancer Model. (Version: 1.5; Date: 02/20/2007)
Input Data File: M:\ACN DOSE-RESPONSE MODELING\CANCER\ORAL\F344_FEMALE_MAMMARY_DW.(d)
Gnuplot Plotting File: M:\ACN DOSE-RESPONSE
MODELING\CANCERXORAIA F344_FEMALE_MAMMARY_DW.p11
Mon Jan 26 15:04:09 2009
BMDS MODEL RUN
The form of the probability function is:
Dependent variable = Response
Independent variable = Dose
Total number of observations = 6
Total number of records with missing values
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-143
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.52
Beta(l) -0.52 1
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Test d.f.
P-value
Est. Prob.
14
8
d.f. = 4
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
8 . 72994
4.77219
27.9615
Taken together, (4.77219, 27.9615) is a 90
interval for the BMD
% two-sided confidence
B-144
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for mammary gland tumors in F344 female rats employing
CEO in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
0.3
0.25
CD
o 0.2
<
o 0.15
ro
0.1
0.05
Multistage
BMDL
BMD
0 0.0005 0.001 0.0015 0.002 0.0025 0.003 0.0035 0.004 0.0045
dose
15:1609/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_FEMALE_MAMMARY_BLOOD_CEO.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_FEMALE_MAMMARY_BLOOD_CEO.plt
Thu Sep 27 15:16:49 2007
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl)]
Total number of observations = 6
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-145
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.53
Beta(l) -0.53 1
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Test d.f.
P-value
Est. Prob.
14
8
d.f. = 4
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.00357812
0.00196521
0.011322
B-146
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for mammary gland tumors in F344 female rats employing
AN in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
CD
•G
0.3
0.25
0.2
o 0.15
•G
ro
u_
0.1
0.05
Multistage
BMD
0.01
0.02
0.03
dose
0.04
0.05
11:3909/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_FEMALE_MAMMARY_BLOOD_AN. (d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\ORAL\F344_FEMALE_MAMMARY_BLOOD_AN.pit
Thu Sep 27 11:39:30 2007
BMDS MODEL RUN
The form of the probability function is:
Total number of observations = 6
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
Default Initial Parameter Values
B-147
DRAFT- DO NOT CITE OR QUOTE
-------
Background =
Beta(l) =
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.52
Beta(l) -0.52 1
Parameter Estimates
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Log(likelihood) # Param's Deviance Test d.f.
-192.056 6
-192.5 2 0.887943 4
-195.631 1 7.14977 5
Goodness of Fit
P-value
Est. Prob.
12
14
d.f. = 4
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.0450726
0.0245301
0.146059
% two-sided confidence
B-148
DRAFT- DO NOT CITE OR QUOTE
-------
APPENDIX B-4. CANCER INHALATION DOSE-RESPONSE ASSESSMENT: BMD
MODELING RESULTS FOR TUMOR INCIDENCE DATA FROM RATS
CHRONICALLY EXPOSED TO AN VIA INHALATION
As summarized in Section 4.6.2, AN is a multisite carcinogen in chronic rodent
bioassays. Data from the only available chronic inhalation cancer bioassay with multiple
exposure levels to AN were used for dose-response assessment (Dow Chemical Co., 1992a;
Quast et al., 1980b). In this study, male and female Sprague-Dawley rats were exposed to AN in
air at concentrations of 0, 20, or 80 ppm 6 hours/day, 5 days/week for 2 years. At 80 ppm,
significantly increased incidences of CNS tumors (i.e., astrocytomas and glial cell proliferation)
and Zymbal's gland tumors were observed in males and females. Also at this concentration,
significantly increased incidences of malignant mammary gland tumors (i.e., adenocarcinomas)
in females, as well as intestinal and tongue tumors in males, were seen. At 20 ppm, male and
female rats exhibited increased incidences (although nonsignificant) of CNS tumors (i.e.,
astrocytomas and glial cell proliferation) and Zymbal's gland tumors.
In this appendix, detailed results of the dose-response modeling for each of the tumor
sites listed above are presented (Tables B-27 through B-36). For each tumor site, first a
summary of the dose-response data is presented, followed by a table summarizing the results of
the dose-response modeling. Finally, the standard output from EPA's BMDS, version 1.4.1, for
the selected dose-response model for each tumor site is presented.
In general, the multistage model was fit to all of the data sets with the BMR set at
0.1 (i.e., 10% extra risk). In fitting this model, successive stages of the multistage model,
starting with stage 1 and ending with the stage equal to the number of dose groups minus one,
were fit to the tumor incidence data at a particular site for each rat sex employing either
administered dose or the internal dose metric CEO in blood. Then, for each dose metric, all
stages of the multistage model that did not show a significant lack of fit (i.e.,/? > 0.1) were
compared using AIC. The stage of the multistage model with the lowest AIC was selected as the
best-fit model. For most tumor sites, the one-stage model exhibited the best fit. For data sets
that exhibited a significant lack of fit for all stages of the multistage model, dose groups were
dropped (starting with the highest dose group) until an adequate fit was achieved.
B-149 DRAFT- DO NOT CITE OR QUOTE
-------
Sprague-Dawlev Rats (Dow Chemical Co., 1992a; Quast et al., 1980b)
Tumor Site: Intestine
Table B-27. Incidence of intestinal tumors in Sprague-Dawley rats exposed
to AN in air for 2 years
Sex
Male
Administered AN concentration
(ppm in air)
0
20
80
Predicted CEO-AUC in blood
(mg/L)
0
2.17 x 10"3
8.20 x 10'3
Incidence of intestinal
tumors"
4/96 (4%)
3/93 (3%)
17/82 (21%)b
"Incidences for Sprague-Dawley rats do not include animals from the 6- and 12-mo sacrifices and were further
adjusted to exclude (from the denominators) rats that died between 0 and 12 mos in the study.
bSignificantly different from controls (p < 0.05) as calculated by the study authors.
Table B-28. Summary of BMD modeling results based on incidence of
intestinal tumors in Sprague-Dawley rats exposed to AN in air for 2 years
Dose metric
Best-fit model3
•£ p valueb
AIC
BMD10C
BMDL10d
Males
Administered dose
CEO
2°MS
2°MS
0.45
0.42
148.05
148.13
59.04 ppm
6.06 x 10"3 mg/L
42.68 ppm
4.47 x 10'3 mg/L
aDose-response models were fit using BMDS, version 1.4.1. "2°MS" indicates a two-stage multistage model
bp value from the %2 goodness of fit test. Values <0.1 indicate a significant lack of fit.
CBMD10 = BMD at 10% extra risk.
dBMDL10 = 95% lower confidence limit on the BMD at 10% extra risk.
B-150
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for intestinal tumors in Sprague-Dawley male rats employing
administered dose as a dose metric
Multistage IVbdel with 0.95 Confidence Level
0.3
0.25
S 0.2
o
t5
CO
0.15
0.1
0.05
Multistage
BIS/DL
0 10
20
30
40
dose
50
60
70
80
13:1004/132007
Multistage Model. $Revision: 2.1 $ $Date: 2000/08/21 03:38:21 $
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\INHALATION\SD_MALE_INTESTINE_INHALE_PPM.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\INHALATION\SD_MALE_INTESTINE_INHALE_PPM.plt
Fri Apr 13 13:10:33 200^
BMDS MODEL RUN
The form of the probability function is:
Dependent variable = Response
Independent variable = Dose
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
B-151
DRAFT- DO NOT CITE OR QUOTE
-------
Default Initial Parameter Values
Background = 0.0312892
Beta(l) = 0
Beta(2) = 3.12226e-005
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Beta(l)
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Background Beta(2)
1 -0.55
-0.55 1
Parameter Estimates
Variable
Background
Beta(1)
Beta(2)
Model
Full model
Fitted model
Reduced model
Log(likelihood) Deviance Test DF
-71.7319
-72.0247 0.585523 1
-81.082 18.7001 2
P-value
Chi-square =
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 59.036
BMDL = 42.6779
B-152
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for intestinal tumors in Sprague-Dawley male rats employing
CEO in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
03
<
C
O
cc
0.3
0.25
0.2
0.15
0.1
0.05
Multistage
BMDL
BMD
0.001 0.002 0.003 0.004 0.005 0.006 0.007 0.008
dose
15:2009/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\INHALATION\SD_MALE_INTESTINE_INHALE_BLOOD_CEO.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\INHALATION\SD_MALE_INTESTINE_INHALE_BLOOD_CEO.plt
Thu Sep 27 15:20:06 2007
BMDS MODEL RUN
The form of the probability function is:
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
Default Initial Parameter Values
B-153
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Beta(l)
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Background
Background 1
Beta(2) -0.55
Variable
Background
Beta(1)
Beta(2)
Indicates that this value is not calculated.
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
Model
Full model
Fitted model
Reduced model
Log(likelihood) # Param's Deviance Test d.f. P-value
-71.7319 3
-72.0636 2 0.663417 1 0.4154
-81.082 1 18.7001 2 <.0001
148.127
Goodness of Fit
Est._Prob. Expected Observed
d.f. = 1
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.00606044
0.004465
0.00809222
% two-sided confidence
B-154
DRAFT- DO NOT CITE OR QUOTE
-------
Tumor Site: CNS (Dow Chemical Co., 1992a; Quast et al., 1980b)
Table B-29. Incidence of CNS tumors in Sprague-Dawley rats exposed to
AN in air for 2 years
Sex
Male
Female
Administered AN concentration
(ppm in air)
0
20
80
0
20
80
Predicted CEO-AUC in blood
(mg/L)
0
2.11 x 10'3
8.20 x 10"3
0
2.18x 10"3
8.24 x 10'3
Incidence of CNS tumors"
0/96 (0%)
4/93 (4%)
22/82 (27%)b
0/93 (0%)
8/99 (8%)b
20/89 (22%)b
"Incidences for Sprague-Dawley rats do not include animals from the 6- and 12-mo sacrifices and were further
adjusted to exclude (from the denominators) rats that died between 0 and 12 mos in the study.
bSignificantly different from controls (p < 0.05) as calculated by the study authors.
Table B-30. Summary of BMD modeling results based on incidence of CNS
tumors in Sprague-Dawley rats exposed to AN in air for 2 years
Dose metric
Best-fit model3
X2/7-valueb
AIC
BMD10C
BMDL10d
Males
Administered dose
CEO
1°MS
1°MS
0.56
0.50
131.64
131.92
30.22 ppm
3.14x 10'3mg/L
22.23 ppm
2.3 Ix 10'3mg/L
Females
Administered dose
CEO
1°MS
1°MS
0.80
0.87
152.86
152.70
30.79 ppm
3.21 x 10"3mg/L
22.89 ppm
2.39 x 10"3 mg/L
"Dose-response models were fit using BMDS, version 1.4.1. "1°MS" indicates a one-stage multistage model.
bp value from the %2 goodness of fit test. Values <0.1 indicate a significant lack of fit.
CBMD10 = BMD at 10% extra risk.
dBMDL,n = 95% lower confidence limit on the BMD at 10% extra risk.
B-155
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for CNS tumors in Sprague-Dawley male rats employing
administered dose as a dose metric
Multistage IVbdel with 0.95 Confidence Level
0.35
0.3
•§ 0.25
0.2
• 0.15
0.1
0.05
0
Multistage
BMDL
BMD
0
13:1804/132007
10
20
30
40
dose
50
60
70
80
Multistage Model. $Revision: 2.1 $ $Date: 2000/08/21 03:38:21 $
Input Data File: G:\ACN DOSE-RESPONSE
MODE LING\CANCERXINHALATION\S D_MALE_CNS_INHALE_P PM. (d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODE LING\CANCERXINHALATION\S D_MALE_CNS_INHALE_P PM.p11
Fri Apr 13 13:18:45 200^
BMDS MODEL RUN
The form of the probability function is:
Dependent variable = Response
Independent variable = Dose
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-156
DRAFT- DO NOT CITE OR QUOTE
-------
Default Initial Parameter Values
Background = 0
Beta(1) = 0.00403595
and do not appear in the correlation matrix )
Beta(1)
Beta(l) 1
Parameter Estimates
Variable Estimate Std. Err.
Background 0 NA
Beta(l) 0.00348587 0.00147454
NA - Indicates that this parameter has hit a bound
implied by some inequality constraint and thus
has no standard error.
Model Log(likelihood) Deviance Test DF
Full model -64.1853
Fitted model -64.8194 1.26826 2
Reduced model -85.6554 42.9402 2
AIC: 131.639
Goodness of Fit
Est._Prob. Expected Observed
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 30.225
BMDL = 22.2323
B-157 DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for CNS tumors in Sprague-Dawley male rats employing
CEO in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
Affected
!_
o
t3
ro
u_
0.4
0.35
0.3
0.25
0.2
0.15
0.1
0.05
0
: ' . . ...
: Multistage
^
^
?
t T ' "
^^^-
>
i i-- x
; BIVPL
-^^^
J
'-.
\ -
^^^^ \
BIVD i
0.001 0.002 0.003 0.004 0.005 0.006 0.007 0.008
dose
15:2209/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\INHALATION\SD_MALE_CNS_INHALE_BLOOD_CEO.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\INHALATION\SD_MALE_CNS_INHALE_BLOOD_CEO.plt
Thu Sep 27
BMDS MODEL RUN
The form of the probability function is:
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
Default Initial Parameter Values
B-158
DRAFT- DO NOT CITE OR QUOTE
-------
Background =
Beta(l) =
Beta (1)
Beta(1)
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
Est. Prob.
0
4
22
d.f.
Benchmark Dose Computation
0.1
Extra risk
Confidence level = 0.95
BMD =
BMDL =
BMDU =
B-159
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for CNS tumors in Sprague-Dawley female rats
employing administered dose as a dose metric
Multistage IVbdel with 0.95 Confidence Level
0.3
0.25
0.2
0.15
0.1
0.05
Multistage
BMDL
BMD
10
20
30
40
dose
50
60
70
80
13:31 04/132007
Multistage Model. $Revision: 2.1 $ $Date: 2000/08/21 03:38:21 $
Input Data File: G:\ACN DOSE-RESPONSE
MODE LING\CANCERXINHALATION\S D_FEMALE_CNS_INHALE_P PM. (d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODE LING\CANCERXINHALATION\S D_FEMALE_CNS_INHALE_P PM.p11
Fri Apr 13 13:31:42 200^
BMDS MODEL RUN
The form of the probability function is:
Dependent variable = Response
Independent variable = Dose
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-160
DRAFT- DO NOT CITE OR QUOTE
-------
and do not appear in the correlation matrix )
Beta(1)
Beta(l) 1
Parameter Estimates
Variable
Background
Beta(1)
NA - Indicates that this parameter has hit a bound
implied by some inequality constraint and thus
has no standard error.
Model Log(likelihood) Deviance Test DF
Full model -75.2138
Fitted model -75.4293 0.431111 2
Reduced model -91.1284 31.8293 2
Goodness of Fit
Est._Prob. Expected Observed Size
0 93 0.000
8 99 0.237
20 89 -0.081
P-value = 0.7982
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 30.7901
BMDL = 22.8938
B-161 DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for CNS tumors in Sprague-Dawley female rats
employing CEO in blood as an internal dose metric
O)
cc
0.35
0.3
0.25
0.15
0.1
0.05
0
Multistage IVbdel with 0.95 Confidence Level
Multistage
B|VPL
BMD
0.001 0.002 0.003 0.004 0.005 0.006 0.007 0.008
dose
15:3809/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODE LING\CANCERXINHALATION\S D_FEMALE_CNS_INHALE_BLOOD_CE 0. (d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODE LING\CANCERXINHALATION\S D_FEMALE_CNS_INHALE_BLOOD_CE 0.p11
Thu Sep 27 15:38:
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl)]
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-162
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Background
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Beta(1)
Beta(1)
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
Analysis of Deviance Table
152.705
Goodness of Fit
Est._Prob. Expected Observed Size
P-value
0
d.f.
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.00321413
0.00238988
0.00446381
% two-sided confidence
B-163
DRAFT- DO NOT CITE OR QUOTE
-------
Tumor Site: Zymbal Gland (Dow Chemical Co., 1992a; Quast et al., 1980b)
Table B-31. Incidence of Zymbal gland tumors in Sprague-Dawley rats
exposed to AN in air for 2 years
Sex
Male
Female
Administered AN concentration
(ppm in air)
0
20
80
0
20
80
Predicted CEO-AUC in
blood
(mg/L)
0
2.11 x 10'3
8.20 x 10"3
0
2.18x 10"3
8.24 x 10'3
Incidence of Zymbal gland
tumors"
2/96 (2%)
4/93 (4%)
ll/82(13%)b
0/93 (0%)
1/98 (1%)
ll/89(12%)b
Incidences for Sprague-Dawley rats do not include animals from the 6- and 12-mo sacrifices and were further
adjusted to exclude (from the denominators) rats that died between 0 and 12 mos in the study.
bSignificantly different from controls (p < 0.05) as calculated by the study authors.
Table B-32. Summary of BMD modeling results based on incidence of
Zymbal's gland tumors in Sprague-Dawley rats exposed to AN in air for
2 years
Dose metric
Best-fit model3
^/J-value13
AIC
BMD10C
BMDL10d
Males
Administered dose
CEO
1°MS
1°MS
0.78
0.73
121.17
121.21
70.29 ppm
7.26 x 10"3 mg/L
42.53 ppm
4.40 x 10"3 mg/L
Females
Administered dose
CEO
1°MS
1°MS
0.50
0.46
81.47
81.67
75.70 ppm
7.90 x 10'3 mg/L
48.74 ppm
5.09 x 10'3 mg/L
aDose-response models were fit using BMDS, version 1.4.1. "1°MS" indicates a one-stage multistage model.
bp value from the %2 goodness of fit test. Values <0.1 indicate a significant lack of fit.
CBMD10 = BMD at 10% extra risk.
dBMDL10 = 95% lower confidence limit on the BMD at 10% extra risk.
B-164
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for Zymbal gland tumors in Sprague-Dawley male rats
employing administered dose as a dose metric
Multistage IVbdel with 0.95 Confidence Level
CD
0.2
0.15
-2 0.1
co
LL
0.05
0
Multistage
BMD
10
20
30
40
dose
50
60
70
80
13:2204/132007
Multistage Model. $Revision: 2.1 $ $Date: 2000/08/21 03:38:21 $
Input Data File: G:\ACN DOSE-RESPONSE
MODE LING\CANCERXINHALATION\S D_MALE_Z YMBAL_INHALE_P PM. (d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODE LING\CANCERXINHALATION\S D_MALE_Z YMBAL_INHALE_P PM.p11
Fri Apr 13 13:22:06 2001
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl)]
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-165
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.66
Beta(l) -0.66 1
Variable
Background
Beta(1)
Parameter Estimates
Model
Full model
Fitted model
Reduced model
AIC:
Goodness of Fit
Est._Prob. Expected Observed
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 70.2907
BMDL = 42.5347
2
4
11
P-value
B-166
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for Zymbal gland tumors in Sprague-Dawley male rats
employing CEO in blood as an internal dose metric
0.25
0.2
S 0.15
•B 0.1
o
CO
0.05
0
Multistage IVbdel with 0.95 Confidence Level
Multistage
BIN/PL
BMD
0 0.001
15:2509/272007
0.002 0.003 0.004 0.005 0.006 0.007 0.008
dose
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\INHALATION\SD_MALE_ZYMBAL_INHALE_BLOOD_CEO.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\INHALATION\SD_MALE_ZYMBAL_INHALE_BLOOD_CEO.plt
Thu Sep 27 1J
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl)]
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-167
DRAFT- DO NOT CITE OR QUOTE
-------
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.66
Beta(l) -0.66 1
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
# Param's Deviance Test d.f.
3
2 0.120008 1
1 9.9669 2
P-value
Est. Prob.
2
4
11
d.f. = 1
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.00726195
0.00440461
0.0159076
% two-sided confidence
B-168
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for Zymbal gland tumors in Sprague-Dawley female rats
employing administered dose as a dose metric
Multistage IVbdel with 0.95 Confidence Level
T3
CD
o
ts
co
0.2
0.15
0.1
0.05
Multistage
BMDL
0 10
20
30
40
dose
50
60
70
80
13:3504/132007
Multistage Model. $Revision: 2.1 $ $Date: 2000/08/21 03:38:21 $
Input Data File: G:\ACN DOSE-RESPONSE
MODE LING\CANCERXINHALATION\S D_FEMALE_Z YMBAL_INHALE_P PM. (d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODE LING\CANCERXINHALATION\S D_FEMALE_Z YMBAL_INHALE_P PM.p11
Fri Apr 13 13:35:30 2001
BMDS MODEL RUN
The form of the probability function is:
Dependent variable = Response
Independent variable = Dose
Total number of observations = 3
Total number of records with missing values
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
B-169
DRAFT- DO NOT CITE OR QUOTE
-------
Default Initial Parameter Values
Background = 0
Beta(1) = 0.0017365
and do not appear in the correlation matrix )
Beta(1)
Beta(l) 1
Parameter Estimates
Variable
Background
Beta(1)
NA - Indicates that this parameter has hit a bound
implied by some inequality constraint and thus
has no standard error.
Model Log(likelihood) Deviance Test DF
Full model -38.8683
Fitted model -39.7334 1.73007 2
Reduced model -49.5377 21.3387 2
Goodness of Fit
Est._Prob. Expected Observed
0
1
11
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 75. 6997
BMDL = 48.7384
B-170 DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for Zymbal gland tumors in Sprague-Dawley female rats
employing CEO in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
0.2
-o 0.15
<
c
o
CO
0.1
0.05
0
Multistage
BMDL
0.001 0.002 0.003 0.004 0.005 0.006 0.007 0.008
dose
15:41 09/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\INHALATION\SD_FEMALE_ZYMBAL_INHALE_BLOOD_CEO.(d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODELING\CANCER\INHALATION\SD_FEMALE_ZYMBAL_INHALE_BLOOD_CEO.plt
Thu Sep 27 15:41:40 2007
BMDS MODEL RUN
The form of the probability function is:
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
Default Initial Parameter Values
B-171
DRAFT- DO NOT CITE OR QUOTE
-------
Background =
Beta(l) =
0
Beta (1)
Beta(1)
Variable
Background
Beta(1)
Std. Err.
- Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
Est. Prob.
0
1
11
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.00789912
0.00508579
0.0132303
% two-sided confidence
B-172
DRAFT- DO NOT CITE OR QUOTE
-------
Tumor Site: Tongue (Dow Chemical Co., 1992a; Quast et al., 1980b)
Table B-33. Incidence of tongue tumors in Sprague-Dawley rats exposed to
AN in air for 2 years
Sex
Male
Administered AN concentration
(ppm in air)
0
20
80
Predicted CEO-AUC in blood
(mg/L)
0
2.11 x 10'3
8.20 x 10"3
Incidence of tongue tumors"
1/95 (1%)
0/14 (0%)
7/82 (9%)b
Incidences for Sprague-Dawley rats do not include animals from the 6- and 12-mo sacrifices and were further
adjusted to exclude (from the denominators) rats that died between 0 and 12 mos in the study.
bSignificantly different from controls (p < 0.05) as calculated by the study authors.
Table B-34. Summary of BMD modeling results based on incidence of
tongue tumors in Sprague-Dawley rats exposed to AN in air for 2 years
Dose metric
Best-fit model3
X2 /7-valueb
AIC
BMD10C
BMDL10d
Males
Administered
dose
CEO
1°MS
2°MS
0.51
0.63
63.76
63.37
11 1.06 ppm
9.48 x 10'3 mg/L
59.41 ppm
6.39 x 10'3 mg/L
aDose-response models were fit using BMDS, version 1.4.1. "1°MS" indicates a one-stage multistage model.
"2°MS" indicates a two-stage multistage model.
bp value from the %2 goodness of fit test. Values <0.1 indicate a significant lack of fit.
CBMD10 = BMD at 10% extra risk.
dBMDL10 = 95% lower confidence limit on the BMD at 10% extra risk.
B-173
DRAFT- DO NOT CITE OR QUOTE
-------
BMDS (version 1.4.1) output for tongue tumors in Sprague-Dawley male rats employing
administered dose as a dose metric
Multistage IVbdel with 0.95 Confidence Level
CD
o
t5
co
0.2
0.15
0.1
0.05
Multistage
BMDL
BMD
20
40
60
dose
80
100
120
13:2504/132007
Multistage Model. $Revision: 2.1 $ $Date: 2000/08/21 03:38:21 $
Input Data File: G:\ACN DOSE-RESPONSE
MODE LING\CANCERXINHALATION\S D_MALE_T ONGUE_INHALE_P PM. (d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODE LING\CANCERXINHALATION\S D_MALE_T ONGUE_INHALE_P PM.p11
Fri Apr 13 13:25:25 2007
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl)]
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
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Default Initial Parameter Values
Background = 0
Beta(l) = 0.00109944
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.65
Beta(l) -0.65 1
Variable
Background
Beta(1)
Parameter Estimates
Model
Full model
Fitted model
Reduced model
AIC:
Log(likelihood) Deviance Test DF
-29.4666
-29.8775 0.821799 1
-33.2127 7.4 9227 2
Goodness of Fit
Est._Prob. Expected Observed
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 111.061
BMDL = 59.4126
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BMDS (version 1.4.1) output for tongue tumors in Sprague-Dawley male rats employing
CEO in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
03
<
C
O
cc
0.25
0.2
0.15
0.1
0.05
0
Multistage
BMp
0 0.002
15:2809/272007
0.004 0.006
dose
0.008
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODE LING\CANCERXINHALATION\S D_MALE_T ONGUE_INHALE_BLOOD_CE 0. (d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODE LING\CANCERXINHALATION\S D_MALE_T ONGUE_INHALE_BLOOD_CE 0.p11
Thu Sep 27 15:28:11 2007
BMDS MODEL RUN
The form of the probability function is:
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
Default Initial Parameter Values
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Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Beta(l)
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Beta (2)
-0. 64
1
Parameter Estimates
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
Variable
Background
Beta(1)
Beta(2)
Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
Log(likelihood) # Param's Deviance Test d.f.
-29.4666 3
-29.6826 2 0.431994 1
-33.2127 1 7.49227 2
d.f. = 1
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.0094804
0.00638558
0.0279407
% two-sided confidence
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Tumor Site: Mammary Gland (Dow Chemical Co., 1992a; Quast et al., 1980b)
Table B-35. Incidence of mammary gland tumors in Sprague-Dawley rats
exposed to AN in air for 2 years
Sex
Female
Administered AN concentration
(ppm in air)
0
20
80
Predicted CEO-AUC in
blood
(mg/L)
0
2.18 x 10'3
8.24 x 10"3
Incidence of mammary gland
tumors"
9/93 (10%)
8/98 (8%)
20/99 (20%)b
"Incidences for Sprague-Dawley rats do not include animals from the 6- and 12-mo sacrifices and were further
adjusted to exclude (from the denominators) rats that died between 0 and 12 mos in the study.
bSignificantly different from controls (p < 0.05) as calculated by the study authors.
Table B-36. Summary of BMD modeling results based on incidence of
mammary gland tumors in Sprague-Dawley rats exposed to AN in air for
2 years
Dose metric
Best-fit model3
^/J-value13
AIC
BMD10C
BMDL10d
Females
Administered dose
CEO
1°MS
2°MS
0.28
0.57
219.42
218.52
66.48 ppm
7.3 Ix 10'3mg/L
37.82 ppm
4.33 x 10'3 mg/L
"Dose-response models were fit using BMDS, version 1.4.1. "1°MS" indicates a one-stage multistage model.
"2°MS" indicates a two-stage multistage model.
bp value from the %2 goodness of fit test. Values <0.1 indicate a significant lack of fit.
CBMD10 = BMD at 10% extra risk.
dBMDL10 = 95% lower confidence limit on the BMD at 10% extra risk.
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BMDS (version 1.4.1) output for mammary gland tumors in Sprague-Dawley female rats
employing administered dose as a dose metric
Multistage IVbdel with 0.95 Confidence Level
0.3
0.25
0.2
0.15
"G
CO
0.1
0.05
Multistage
BMD
10
20
30
40
dose
50
60
70
80
13:3804/132007
Multistage Model. $Revision: 2.1 $ $Date: 2000/08/21 03:38:21 $
Input Data File: G:\ACN DOSE-RESPONSE
MODELING\ CANCERX INHALATI ON\ SD_FEMALE_MAMMARY_INHALE_PPM. (d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODE LING\CANCERXINHALATION\S D_FEMALE_MAMMARY_INHALE_P PM.p11
Fri Apr 13 13:38:03 2001
BMDS MODEL RUN
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(
-betal*doseAl) ]
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
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Default Initial Parameter Values
Background = 0.0767126
Beta(l) = 0.00173168
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.68
Beta(l) -0.68 1
Variable
Background
Beta(1)
Model
Full model
Fitted model
Reduced model
AIC:
Log(likelihood) Deviance Test DF
-107.092
-107.71 1.23579 1
-110.714 7.24319 2
P-value
Goodness of Fit
Est._Prob. Expected Observed
Benchmark Dose Computation
Specified effect = 0.1
Risk Type = Extra risk
Confidence level = 0.95
BMD = 66.4843
BMDL = 37.8172
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BMDS (version 1.4.1) output for mammary gland tumors in Sprague-Dawley female rats
employing CEO in blood as an internal dose metric
Multistage IVbdel with 0.95 Confidence Level
0.3
0.25
CD
o 0.15
o
ro
0.1
0.05
Multistage
BIN/PL
BMD
0.001 0.002 0.003 0.004 0.005 0.006 0.007 0.008
dose
15:4409/272007
Multistage Model. (Version: 2.8; Date: 02/20/2007)
Input Data File: G:\ACN DOSE-RESPONSE
MODE LING\CANCERXINHALATION\S D_FEMALE_MAMMARY_INHALE_BLOOD_CE 0. (d)
Gnuplot Plotting File: G:\ACN DOSE-RESPONSE
MODE LING\CANCERXINHALATION\S D_FEMALE_MAMMARY_INHALE_BLOOD_CE 0.p11
Thu Sep 27 15:44:26 2007
BMDS MODEL RUN
The form of the probability function is:
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
Default Initial Parameter Values
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Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Beta(l)
have been estimated at a boundary point, or have been specified by the user,
and do not appear in the correlation matrix )
Beta (2)
-0. 6
1
Parameter Estimates
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
Variable
Background
Beta(1)
Beta(2)
Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
Log(likelihood) # Param's Deviance Test d.f.
-107.092 3
-107.258 2 0.332027 1
-110.714 1 7.24319 2
20
d.f. = 1
Specified effect =
Risk Type
Confidence level =
BMD =
BMDL =
BMDU =
0.1
Extra risk
0. 95
0.00730992
0.00432882
0.0152838
% two-sided confidence
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APPENDIX B-5. ANALYSIS TO ASSESS COMBINING TUMOR INCIDENCE DATA
FROM TWO CANCER BIOASSAYS EMPLOYING SPRAGUE-DAWLEY RATS
A statistical analysis was conducted to determine whether tumor dose-response data from
two 2-year drinking water bioassays in Sprague-Dawley rats were similar enough to be
combined. The first study employed AN drinking water concentrations of 0, 35, 100, and
300 ppm (Quast, 2002; Quast et al., 1980a), while the second study used AN drinking water
concentrations of 0, 1, and 100 ppm (Johannsen and Levinskas, 2002a). To conduct the analysis,
the multistage model in BMDS (version 1.3.2) was fit to the tumor incidence data from three
sites (forestomach, CNS, and Zymbal gland) in each sex across both studies, using administered
animal dose expressed in mg/kg-day. Using the best-fit model, a statistical test described by
r\
Stiteler et al. (1993), which employs a maximum likelihood ratio statistic distributed as a x , was
then used to test the null hypothesis that the corresponding data sets from the two studies are
compatible with a common dose-response model. If the null hypothesis is not rejected, this
analysis provides evidence that the results from the two studies may be pooled.
The results of this analysis showed that forestomach and Zymbal gland tumors in male
and female Sprague-Dawley rats were not compatible with a common dose-response model,
while CNS tumors in male and female Sprague-Dawley rats were compatible with a common
dose-response model. Because of these conflicting results, it was decided that the results from
the two Sprague-Dawley drinking water studies would not be pooled. Therefore, the final dose-
response analysis for deriving the oral slope factor for AN focused on the two rat drinking water
studies containing the most dose groups (i.e., the Sprague-Dawley rat bioassay reported by Quast
[2002] and the F344 rat bioassay reported by Johannsen and Levinskas [2002b]). For each tumor
site, summaries of the results of the statistical tests for compatibility are shown below.
Test 1: Forestomach tumors in male Sprague-Dawley rats
Tumor incidence data (Johannsen and Levinskas, 2002a):
• 0 ppm (0 mg/kg-day): 3/78
• 1 ppm (0.085 mg/kg-day): 3/78
• 100 ppm (8.53 mg/kg-day): 11/77
Best-fit model: 1 -stage multistage
Log (likelihood) = -57.0113
Tumor incidence data (Quast, 2002):
• 0 ppm (0 mg/kg-day): 0/80
• 35 ppm (3.42 mg/kg-day): 2/47
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• 100 ppm (8.53 mg/kg-day): 23/48
• 300 ppm (21.20 mg/kg-day): 39/48
Best-fit model: 2-stage multistage (highest dose group dropped)
Log (likelihood) = -42.4099
Tumor incidence data (combined):
• 0 ppm (0 mg/kg-day): 3/158
• 1 ppm (0.085 mg/kg-day): 3/78
• 35 ppm (3.42 mg/kg-day): 2/47
• 100 ppm (8.53 mg/kg-day): 34/125
• 300 ppm (21.20 mg/kg-day): 39/48
Best-fit model: 2-stage multistage
Log (likelihood) = -132.865
Conclusion: Based on the likelihood ratio test statistic, -2 In A = 2[132.865 - (42.4099 +
57.0113)] = 2 x 33.4438 = 66.89, the data sets are not compatible with a common dose-response
model at/?< 0.0001.
Test 2: Forestomach tumors in female Sprague-Dawley rats
Tumor incidence data (Johannsen and Levinskas, 2002a):
• 0 ppm (0 mg/kg-day): 1/80
• 1 ppm (0.11 mg/kg-day): 4/79
• 100 ppm (10.80 mg/kg-day): 7/79
Best-fit model: 1-stage multistage
Log (likelihood) = -45.8328
Tumor incidence data (Quast, 2002):
• 0 ppm (0 mg/kg-day): 1/80
• 35 ppm (4.36 mg/kg-day): 1/48
• 100 ppm (10.80 mg/kg-day): 12/48
• 300 ppm (25.00 mg/kg-day): 30/48
Best-fit model: 2-stage multistage
Log (likelihood) = -69.9834
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Tumor incidence data (combined):
• 0 ppm (0 mg/kg-day): 2/160
• 1 ppm (0.11 mg/kg-day): 4/79
• 35 ppm (4.36 mg/kg-day): 1/48
• 100 ppm (10.80 mg/kg-day): 19/127
• 300 ppm (25.00 mg/kg-day): 30/48
Best-fit model: 2-stage multistage
Log (likelihood) = -119.054
Conclusion: Based on the likelihood ratio test statistic, -2 In A = 2[119.054 - (69.9834 +
45.8328)] = 2 x 3.2378 = 6.48, the data sets are not compatible with a common dose-response
model atp = 0.011.
Test 3: CNS tumors in male Sprague-Dawley rats
Tumor incidence data (Johannsen and Levinskas, 2002a):
• 0 ppm (0 mg/kg-day): 2/78
• 1 ppm (0.085 mg/kg-day): 3/75
• 100 ppm (8.53 mg/kg-day): 23/77
Best-fit model: 1 -stage multistage
Log (likelihood) = -68.9254
Tumor incidence data (Quast, 2002):
• 0 ppm (0 mg/kg-day): 1/80
• 35 ppm (3.42 mg/kg-day): 12/47
• 100 ppm (8.53 mg/kg-day): 22/48
• 300 ppm (21.20 mg/kg-day): 30/48
Best-fit model: 1 -stage multistage
Log (likelihood) = -98.6909
Tumor incidence data (combined):
• 0 ppm (0 mg/kg-day): 3/158
• 1 ppm (0.085 mg/kg-day): 3/75
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• 35 ppm(3.42mg/kg-day): 12/47
• 100 ppm (8.53 mg/kg-day): 45/125
• 300 ppm (21.20 mg/kg-day): 30/48
Best-fit model: 1-stage multistage
Log (likelihood) = -168.873
Conclusion: Based on the likelihood ratio test statistic, -2 In A = 2[168.873 - (98.6909 +
68.9254)] = 2 x 1.2567 = 2.51, the data sets are compatible with a common dose-response model
tip = 0.113.
Test 4: CNS Tumors in female Sprague-Dawley rats
Tumor incidence data (Johannsen and Levinskas, 2002a):
• 0 ppm (0 mg/kg-day): 0/79
• 1 ppm (0.11 mg/kg-day): 1/80
• 100 ppm (10.80 mg/kg-day): 39/78
Best-fit model: 1 -stage multistage
Log (likelihood) = -59.5775
Tumor incidence data (Quast, 2002):
• 0 ppm (0 mg/kg-day): 1/80
• 35 ppm (4.36 mg/kg-day): 20/48
• 100 ppm (10.80 mg/kg-day): 25/48
• 300 ppm (25.00 mg/kg-day): 31/48
Best-fit model: Log-logistic
Log (likelihood) = -104.123
Tumor incidence data (combined):
• 0 ppm (0 mg/kg-day): 1/159
• 1 ppm (0.11 mg/kg-day): 1/80
• 35 ppm (4.36 mg/kg-day): 20/48
• 100 ppm (10.80 mg/kg-day): 64/126
• 300 ppm (25.00 mg/kg-day): 31/48
Best-fit model: Log-logistic
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Log (likelihood) = -164.485
Conclusion: Based on the likelihood ratio test statistic, -2 In A = 2[164.485 - (104.123 +
59.5775)] = 2 x 0.7845 = 1.57, the data sets are compatible with a common dose-response model
atp = 0.210.
Test 5: Zymbal gland tumors in male Sprague-Dawley rats
Tumor incidence data (Johannsen and Levinskas, 2002a):
• 0 ppm (0 mg/kg-day): 1/80
• 1 ppm (0.085 mg/kg-day): 0/71
• 100 ppm (8.53 mg/kg-day): 17/73
Best-fit model: 1-stage multistage
Log (likelihood) = -45.815
Tumor incidence data (Quast, 2002):
• 0 ppm (0 mg/kg-day): 3/80
• 35 ppm (3.42 mg/kg-day): 4/47
• 100 ppm (8.53 mg/kg-day): 3/48
• 300 ppm (21.20 mg/kg-day): 16/48
Best-fit model: 1 -stage multistage
Log (likelihood) = -70.1142
Tumor incidence data (combined):
• 0 ppm (0 mg/kg-day): 4/160
• 1 ppm (0.085 mg/kg-day): 0/71
• 35 ppm (3.42 mg/kg-day): 4/47
• 100 ppm (8.53 mg/kg-day): 20/121
• 300 ppm (21.20 mg/kg-day): 16/48
Best-fit model: 1 -stage multistage
Log (likelihood) = -118.806
Conclusion: Based on the likelihood ratio test statistic, -2 In A = 2[118.806 - (70.1142 +
45.815)] = 2 x 2.8768 = 5.75, the data sets are not compatible with a common dose-response
model atp = 0.016.
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Test 6: Zymbal gland tumors in female Sprague-Dawley rats
Tumor incidence data (Johannsen and Levinskas, 2002a):
• 0 ppm (0 mg/kg-day): 1/79
• 1 ppm (0.11 mg/kg-day): 0/75
• 100 ppm (10.80 mg/kg-day): 12/78
Best-fit model: 1-stage multistage
Log (likelihood) = -39.6428
Tumor incidence data (Quast, 2002):
• 0 ppm (0 mg/kg-day): 1/80
• 35 ppm (4.36 mg/kg-day): 5/48
• 100 ppm (10.80 mg/kg-day): 9/48
• 300 ppm (25.00 mg/kg-day): 18/48
Best-fit model: 1 -stage multistage
Log (likelihood) = -76.3975
Tumor incidence data (combined):
• 0 ppm (0 mg/kg-day): 1/159
• 1 ppm (0.11 mg/kg-day): 0/75
• 35 ppm (4.36 mg/kg-day): 5/48
• 100 ppm (10.80 mg/kg-day): 21/126
• 300 ppm (25.00 mg/kg-day): 18/48
Best-fit model: 1 -stage multistage
Log (likelihood) = -111.439
Conclusion: Based on the likelihood ratio test statistic, -2 In A = 2[111.439 - (76.3975 +
39.6428)] = 2 x (-4.6013) = -9.20 = 9.20, the data sets are not compatible with a common dose-
response model atp = 0.002.
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APPENDIX B-6. ESTIMATION OF COMPOSITE CANCER RISK FROM EXPOSURE
TO AN BY COMBINING RISK ESTIMATES ACROSS MULTIPLE TUMOR SITES
Increased tumor incidences were observed at multiple sites in male and female rats
following exposure to AN both orally and by inhalation. With this multiplicity of tumors, the
concern is that a potency or risk estimate based solely on one tumor site (e.g., the most sensitive
site) may underestimate the overall cancer risk associated with exposure to this chemical. The
most recent EPA cancer guidelines (U.S. EPA, 2005a) identified two ways to approach this
issue: 1) analyze the incidence of tumor-bearing animals, or 2) combine the potencies associated
with significantly elevated tumors at each site. The NRC (1994) concluded that an approach
based on counts of animals with one or more tumors would tend to underestimate overall risk
when tumor types occur independently, and thus an approach based on combining the risk
estimates from each separate tumor type should be used.
Because potencies are typically upper bound estimates, summing such upper bound
estimates across tumor sites is likely to overstate the overall risk. Therefore, following the
recommendations of the NRC (1994) and the Guidelines for Carcinogen Risk Assessment (U.S.
EPA, 2005a), a statistically valid upper bound on composite risk was derived in order to gain
some understanding of the overall risk resulting from tumors occurring at multiple sites. It is
important to note that this estimate of overall potency describes the risk of developing tumors at
any combination of the sites considered and is not the risk of developing tumors at all three sites
simultaneously. The combined risk estimate was derived assuming independence of tumors.
For modeling individual tumor data, the multistage model is specified as follows:
(1) P(d) = 1 - expf-(q0 + qid + q2d2 + ...
The model for the composite tumor risk is still multistage, with a functional form that has
the sum of stage-specific multistage coefficients as the corresponding multistage coefficient.
(2) Pc(d) = 1 - exp[-(Zq0l + dlqn + d2Iq2l + ... + fflqJJ, for i=l,. . ., k,
where k = total number of sites
The resulting equation for fixed extra risk (BMR) is polynomial in dose (when logarithms
of both sides are taken) and can be straightforwardly solved for the combined BMD. However,
confidence bounds for this BMD are not able to be estimated by the current version of BMDS.
Therefore, a Bayesian approach to finding confidence bounds on the combined BMD was
implemented using WinBugs (Spiegelhalter et al., 2003). WinBugs software is freely available
and implements Markov chain Monte Carlo (MCMC) computations. Use of WinBugs has been
demonstrated for derivation of a distribution of BMDs for a single multistage model (Kopylev et
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al., 2007) and can be straightforwardly generalized to derive the distribution of BMDs for the
combined tumor load, following the NRC (1994) methodology described above. The advantage
of a Bayesian approach is that it produces a distribution of BMDs that allows better
characterization of statistical uncertainty. For the current analysis, a diffuse (high variance or
low tolerance) Gaussian prior restricted to be nonnegative was used. The posterior distribution
was based on three chains with 50,000 burn-in (i.e., the initial 50,000 simulations were dropped)
and a thinning rate of 20, resulting in 150,000 simulations total. The median and 5 percentile of
the posterior distribution provided the BMDio (central estimate) and BMDLio (lower bound) for
combined tumor load, respectively.
The methodology above was applied to the dose-response data for the male and female
Sprague-Dawley (Quast, 2002) and F344 (Johannsen and Levinskas, 2002b) rat drinking water
studies, as well as to the data from male and female Sprague-Dawley rats in the Quast et al.
(1980b) inhalation study. As with the risk estimates generated for individual tumor sites, the
combined analysis used the internal dose metric CEO in blood (see Appendices B-3 and B-4).
The human equivalent PODs are presented in Tables B-37 (episodic oral exposure) and B-38
(continuous inhalation exposure). Estimates of composite risk were estimated by dividing the
BMR of 10% extra risk by the composite BMDLioS (0.1/BMDLio). Human equivalent
composite CSFs are presented in Table B-39, and composite unit risks are presented in
Table B-40. The slopes derived from the composite BMDios (O.I/ BMDio) are also included in
these tables for comparison.
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Table B-37. Summary of PODs for composite cancer risk associated with
episodic oral exposure to AN, using CEO-AUC levels in blood as dose metric
and multiple tumor incidence data in rats
Rat strain, sex
Tumor site
Sprague-Dawley, male
Forestomach, CNS, Zymbal gland,
tongue
Sprague-Dawley, female
Forestomach, CNS, Zymbal gland,
tongue, mammary gland
F344, male
Forestomach, CNS, Zymbal gland
F344, female
Forestomach, CNS, Zymbal gland,
mammary gland
PODs based on rat dose-response
using internal dose metric CEO-
AUC in blood (mg/L)
BMD10
4.2 x 10"4
2.6 x 10"4
4.6 x 10'4
6.7 x 10"4
BMDL10
3.6 x 10"4
2.0 x 10"4
3.3 x 10"4
5.2 x 10"4
Human equivalent PODsa
(mg/kg-d)
BMD10
0.042
0.026
0.046
0.066
BMDL10
0.036
0.020
0.033
0.051
aConverted using human PBPK model.
Table B-38. Summary of PODs for composite cancer risk associated with
inhalation exposure to AN, based on multiple tumor incidence data in rats
and CEO-AUC levels in blood
Rat strain, sex
Tumor site
Sprague-Dawley, male
Intestine, CNS, Zymbal gland, tongue
Sprague-Dawley, female
CNS, Zymbal gland, mammary gland
PODs for rat dose-response in
terms of CEO-AUC in blood
(mg/L)
BMC10
1.4 x 10'3
1.7 x 10'3
BMCL10
1.1 x 10'3
1.3 x 10'3
Human equivalent PODs:
AN in aira
(mg/m3)
BMC10
1.9
2.3
BMCL10
1.5
1.8
"Converted using human PBPK model.
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Table B-39. Estimated human oral CSFs for AN based on multiple tumor
incidence data in rats and CEO-AUC levels in blood
Rat strain, sex
Tumor site
Sprague-Dawley, male
Forestomach, CNS, Zymbal gland, tongue
Sprague-Dawley, female
Forestomach, CNS, Zymbal gland, tongue,
mammary gland
F344, male
Forestomach, CNS, Zymbal gland
F344, female
Forestomach, CNS, Zymbal gland,
mammary gland
Slope derived from composite
BMD10a
(mg/kg-d)1
2.4
3.8
2.2
1.5
Composite CSFb
(mg/kg-d)1
2.8
5.0
3.1
1.9
"Slope estimated by 0.1/BMD10.
bCSF = 0.1/BMDL10.
Table B-40. Estimated human lURs for AN based on multiple tumor
incidence data in rats and CEO-AUC levels in blood
Rat strain, sex
Tumor site
Sprague-Dawley, male
Intestine, CNS, Zymbal gland,
tongue
Sprague-Dawley, female
CNS, Zymbal gland, mammary gland
Slope to background from
overall BMC10a
(mg/m3)1
5.4 x 10"2
4.4 x 10"2
Overall IURb
(mg/m3)1
6.8 x 10"2
5.7 x 10"2
"Slope estimated by 0.1/BMD10.
bIUR = 0.1/BMDL10.
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APPENDIX B-7. STATISTICAL ANALYSIS OF BLAIR ET AL. (1998)
The Data
The data used in this analysis came from the best available epidemiological study (see
Section 5.4.1) by Blair et al. (1998). The raw data from the study were provided to EPA
courtesy of NCI. The process of data collection is described in Blair et al. (1998) and in more
detail in Stewart et al. (1998). The full raw data set contains 363 variables on 25,460 subjects.
The variables include demographic information, employment information, vital status, and
various measures of exposure per study year (1942-1983), including maximum and minimum
estimates, "best estimate," frequency of peaks, and several other ways of measuring exposure.
The raw data provided by NCI did not include the smoking data that were collected for
approximately 10% of the subjects. Therefore, smoking information was not part of the EPA's
statistical analysis. The uncertainty connected to lack of smoking information is discussed in
Section 6.4.1. The data on dates of exposure and corresponding exposure amounts, dates of
mortality due to lung cancer and other causes, a plant worked, and the birth year were retained
for the statistical analysis. Biological age was chosen as a time scale. Data handling and
statistical analysis were performed using S-Plus™ statistical software.
The Statistical Model
The semi-parametric Cox proportional hazards model (Cox, 1972) is a widely used model
in hazard regression. The Cox model with time-dependent covariates was chosen to model the
data for to two main reasons. Primarily, it allows taking into account individual covariate
history, allowing utilization of the extensive exposure data collected by Blair et al. (1998).
Additionally, the Cox model uses internal controls. Internal controls constitute an appropriate
comparison group, given the healthy worker effect observed by Blair et al. (1998) and further
demonstrated by Marsh et al. (2001).
In the Cox model, the conditional hazard function, given the covariate process Z(t), is
assumed to have the form:
?i(t\Z(t}} = /Lo(OexpG#TZ(0)
where P is the vector of regression coefficients and Xo(t) denotes the baseline hazard function.
No particular shape is assumed for the baseline hazard; it is estimated nonparametrically. The
contributions of covariates to the hazard are multiplicative. When pTZ(t) is small and Z(t)
represents exposure, the Cox proportional model is consistent with linearity of dose-response for
low doses. When no time-dependent covariates are present, the cumulative hazard function Ao(t)
is estimated using a Breslow (Breslow, 1974) estimator:
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_
~
where Zj is the (time-independent) covariate vector of the j* individual at age t;, /? is an
estimate of P from the Cox model, d; is the number of deaths at the age t;, and R(t;) is the risk set
at age t;. Risk set is defined as all the individuals who are under observation and at risk of death
at age t;. The Breslow estimator is piecewise constant with discrete jumps at times of death. The
corresponding estimate of the baseline hazard function is the size of the jump of the cumulative
hazard at ages of death and zero otherwise. The survival function S(t) = exp(-A(t)) is estimated
by plugging in the Breslow estimator:
S(t) = exp(- A(t))
The cumulative probability of death R(t) = l-S(t) is estimated by:
R(t) = l-
With time-dependent covariates, EPA followed the suggestion by Starr (2004) of a modified
Breslow estimator:
The survival function and cumulative probability of death are then estimated using the
modified cumulative hazard estimator. The modified estimator only approximates the true
cumulative hazard since a covariate path is not encountered in the estimator. Therneau and
Grambsch (2000) discuss clinical examples when this estimator could lead to inappropriate
results, but the occupational case is different from clinical examples, since at the baseline, not all
workers are exposed and exposure is nondecreasing and often ends well before death.
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Calculations of Excess Risk
Following Starr et al. (2004), EPA defined hazard functions for lung cancer mortality
Xlc(t) and other causes of death X°°(t), with the corresponding cause-specific cumulative hazard
functions, Alc (t) and A°°(t). The sum of Alc(t) and A°°(t) is the all-causes mortality hazard
Aac(t) with corresponding survival function Sac(t). Then, the cumulative risk of lung cancer
mortality by age t, given covariate path Z(t)=z(t) is:
Rk(t\z(ty) =
The cumulative risk of lung cancer mortality by age t in the absence of exposure:
Rlc(t\Z(t} = 0) = JX(z/|Z(0 = 0)SflC(z/|Z(0 = 0)
u
-------
males sex-race groups. Since the goal of modeling is extrapolation to the overall U.S.
population, sex and race were not included as covariates.
Mortality and Risk Estimates. The coxph function of S-Plus™ was used to obtain
estimates of p. The estimate of P ° was equal to 1.24 x 10~3with the standard error of 2.47 x 10~2
(p = 0.61). The estimate of P°° was equal to 6.7 x 10~4 with a standard error of 9.66 x 10~4
(p = 0.49). The estimates, covariates, and event times were then used to construct estimates of
cumulative hazard and risks, as described in the previous section. The exposures corresponding
to excess risks were divided by 2 to account for differences in volume of air inhaled during a
working day and during a whole day (10 vs. 20 m3/day). Resulting exposure estimates (by age
80): ECoi = 0.992 ppm and LECoi = 0.238 ppm. The corresponding unit risk was calculated to
equal 4.2 x lO^"1
Limitations of the Statistical Approach
The statistical approach followed the approach of Starr et al. (2004), and it shares the
limitations and uncertainties described there. The limitations and uncertainties include the
following:
• The Cox model fit the data adequately, but it was an empirical model fit rather than a
biologically based model.
• The estimator of the cumulative hazard does not account for the covariate path and hence,
is only an approximation.
• The estimate of risk is obtained using the first-order "linearized" approximation.
However, the results are consistent with assumptions of the first-order approximation
validity.
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APPENDIX C. PBPK MODEL DESCRIPTIONS AND SOURCE CODE
Model Description
The PBPK models used to calculate rat and human internal doses for the development of
candidate RfDs, oral CSFs, and lURs were those revised from Kedderis et al. (1996) (rat, with
EH activity towards CEO added and other parameters revised) and Sweeney et al. (2003)
(human). Both models are flow limited, lumped compartment models that predict amounts and
concentrations of AN and its metabolite of toxicological interest, CEO, in blood and seven
tissues: lung, brain, fat, stomach, liver, and rapidly and slowly perfused tissues. Both models
include portals of entry for oral, inhalation, and i.v. routes of exposure. Drinking water is
modeled as a continuous infusion to the GI lumen. (The model code allows one to define
episodic oral exposure as a series of boluses up to 6 times/day, but the simpler approach of
treating the exposure as continuous was chosen.) For rats, the infusion rate is set such that the
total amount consumed equals the average total amount ingested during the bioassays as
determined from water consumption and administered AN concentration. Inhalation duration
and frequency can be explicitly defined. Simulated metabolism of AN and CEO occur only in
the liver. Elimination of both compounds is accomplished via second-order GSH conjugation in
various tissues, saturable hepatic metabolism, and pulmonary excretion into exhaled air.
The rat and human models are based on the same structural framework but have two
primary differences: (1) physiological and metabolic parameters for each are species specific,
and (2) the human model simulates saturable, enzyme-mediated hydrolysis of CEO, a metabolic
process not observed in rats. In the absence of human in vivo data for metabolism, human
metabolic parameters were estimated from in vitro values extrapolated to in vivo values by using
in vitro/m vivo ratios in rats, which is described fully in Sweeney et al. (2003). In the Sweeney
et al. (2003) model, a conversion factor for in vitro to in vivo extrapolation of human metabolic
constants is implicitly defined in the calculation of Vmax values for AN oxidation and CEO
hydrolysis. Model parameter values used for both species are given in Table C-l.
The acslXtreme model code as used to calculate the various dose metrics, followed by the
corresponding Matlab model code used to perform the optimizations for fitting of the revised
parameters, are given at the end of this appendix. The Matlab code is divided into three .m files:
the top-level "optACN2.m" file, a secondary "RunACN.m" file that is called by the opt file and
returns the value of the objective function, and an "EqACN.m" file that defines the set of
differential equations.
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Table C-l. Rat and human PBPK model parameter values
Parameter
Total BW (kg)
Rat value"
0.25
Human value"
70
Percentage of BW
Liver
Brain
Stomach
Fat
Rapidly perfused
Slowly perfused
Venous blood
Arterial blood
4
0.6
0.63
7
3.77
75
3.9
2.1
2.57
2.00
0.21
21
42
5.22
58.58
4.35
2.74
Blood flows (L/h/kga74)
Cardiac output
Alveolar ventilation
14.0
14.0
13.4
12.9
Percentage of cardiac output
Liver
Brain
Stomach
Fat
Rapidly perfused
Slowly perfused
25
2.4
1.3
9
47.3
15
21.4
11.4
1
5
3
2
32.5
28.2
GSH content (mmol/L)
Liver
Brain
Stomach
Rapidly perfused
Slowly perfused
PCs
Blood:air
Liverblood
Brain:blood
Stomach:blood
Fatblood
Rapidly perfused:blood
Slowly perfused:blood
8.53
2.00
4.59
2.65
0.75
AN CEO
512 1,658
0.46 0.274
0.40 1.407
0.46 0.274
0.28 0.785
0.46 0.274
0.35 1.853
5.63
2.99
3.61
2.59
1.
AN
154
1.51
1.34
1.51
0.94
1.51
1.16
Blood binding (h"1)
Hb (kBC, kBC2)
Blood sulfhydryls (kFBC, kFBC2)
1.245(3.66) 1.134(3.33;
2.54 0.68
1.245 (3.66;
0.0008
AN oxidation
VmaxC (mg/h/kg07)
Km (mg/L)
5.0(7.7)
1.5(2.76)
13
CEO
1,658
0.274
1.407
0.274
0.785
0.274
1.853
1.134(3.33;
0.84
15.6 (22. 1)
0
8
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Table C-l. Rat and human PBPK model parameter values
Parameter
Rat value"
Human value"
CEO hydrolysis
VmaxcC2 (mg/h/kg07)
Km2 (mg/L)
kEHC (L/h/kg07)
—
—
— (3.92)
841 (-)
113(-)
-(2.20)
AN-GSH conjugation
Enzymatic: kFC (h/kg113)'1
Spontaneous: kso (mmol/h)"1
73 (50)
0.2584
113(77)
0.2584
CEO-GSH conjugation
Liver: kFC2 (h/kg03)'1
Brain/liver (kFBRC/kFC2)
Stomach/liver (kFSTC/kFC2)
Rapidly perfused/liver (kFRC/kFC2)
Slowly perfused/liver (kFSC/kFC2)
Oral absorption: Ka (h"1)
500
0.0234
0.0538
0.0311
0.00879
8.0 (4.2)
197
0.0531
0.0641
0.0460
0.0201
8.0 (4.2)
aRevised values used in this assessment are italicized in parentheses.
Sources: Sweeney et al. (2003); Kedderis et al. (1996).
Estimation of Internal Doses Corresponding to Bioassay Exposures
The model code from Sweeney et al. (2003) was modified to enable the explicit
definition of desired daily oral intake of AN either as a continuous infusion or in an episodic
pattern of up to six episodes (or both continuous and episodic). The continuous-infusion oral
exposure rate (STDOSE) is simply calculated as being equal to the total daily exposure
(STEADYODOSE*BW, by continuous oral dosing) divided by 24 hours. The episodic exposure
is calculated as the drinking water concentration (DRCONC) times the standard drinking water
volume (DRVOL, based on BW) times the ratio of actual water consumption to standard
consumption (FRACVOL) times the fraction of daily water consumed at episode "I" (DRP(I)),
with that quantity added as a bolus to the GI lumen compartment at episode time "I" (DRT(I)).
The first episodic exposure time is assumed to be TIME = 0, the same time at which a daily
gavage dose (ODOSE*BW) is also administered.
To simulate inhalation exposures as described in Quast (1980b), the model was run to
provide 6-hour continuous inhalation exposures, repeating every 24 hours (by setting model
variable TCHNG1 = 6.0) for 5 days/week (by setting model variable TCHNG2 = 120, the
number of hours in 5 days). Inhalation simulations were run for 800 continuous simulated hours
so that the effect of the 5 day/week exposures would be included in the calculation of average
lifetime daily AUCs. Study-specific rat BWs were used. In this way the PBPK model accounted
for the dynamics of concentration rise at the beginning of each exposure and clearance at the end
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of each exposure. The resulting internal dose metrics for inhalation exposure of the rat are listed
in Table 5-17.
Rat to Human Extrapolations
For inhalation and oral cancer and oral noncancer effects, similar approaches were used
to extrapolate PODs from rats to humans. Following derivation of PODs for noncancer and
cancer endpoints, as weekly average concentrations of AN and CEO in rat blood tissue
(C],tissue,rat(BMCLio) in mg/L), FtECs were then calculated using the human PBPK model and a
standard human BW of 70 kg, as the blood or brain concentration, assuming steady-state
exposure for a set of predefined exposure levels and that equal internal metrics in rats and
humans are associated with the same degree of response. The relationship between exposure
level and internal metric as predicted by the human PBPK model was then plotted in Excel and a
9
second-order polynomial of the form, FIEC = a x Ci + b x Ci, where Ci is the internal dose
metric, was fit to the relationship for each metric (blood and brain AN and CEO).
Inhalation Exposure
The relationship between inhalation exposure level and internal metric as predicted by the
human PBPK model is shown in Figures C-l and C-2 for AN and CEO, respectively. This
polynomial form gave an excellent fit and was fit over a range of concentrations that bracketed
the range of Ci!tiSSue,rat(BMCLio) for various cases. The resulting polynomials were then used to
convert the rat Q values to human equivalent exposures, reported in Chapter 5 for each toxicity
value.
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1000
Q. 100
D
W
O
g- 10
0)
c
O
"re
0.1
AN dosimetry
y = 38.6031 -C + [4305.55-C/(26.6520+
O ACN blood
n ACN brain
— AN blood fit
— AN brain fit
y = 68.3045-C + [1162.47-C/(26.6520 + C)]
0.0001 0.001 0.01 0.1 1 10
Internal average / steady state concentration (mg/L)
100
Points are steady-state internal AN concentrations predicted by the human PBPK
model for given inhalation exposure concentrations. Curves are polynomial
regressions.
Figure C-l. Human inhalation exposure level vs. internal AN concentration.
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Q.
a.
o
Q.
X:
a)
c
o
!is
(0
.c
c
100
10
0.1
+ CEO blood
X CEO brain
— Poly. (CEO brain)
— Poly. (CEO blood)
41.103x': + 618.54x
R2
y = 26.838X2 + 503.27X
R2 =
0.0001 0.001 0.01
Internal average / steady state concentration (mg/L)
0.1
Points are steady-state internal CEO concentrations predicted by the human PBPK
model for given inhalation exposure concentrations. Curves are polynomial
regressions.
Figure C-2. Human inhalation exposure level vs. internal CEO
concentrations.
Oral Cancer Risks
For extrapolation of oral cancer risks from exposure in drinking water, the exact pattern
of ingestion of drinking water by the rats in the bioassay was not measured and will be variable
among humans. Therefore, the rat PBPK model was used to estimate the internal doses of AN
and CEO at the exposure levels of the PODs as determined above (in mg/kg-day ingested), while
assuming two distinct exposure patterns that are expected to bound the truth: continuous
ingestion and ingestion in six boluses ("episodic").
The continuous pattern is simply described as continuous infusion to the stomach lumen
compartment (i.e., 24 hours/day) at a rate equal to the total daily ingestion (at the BMDLio).
This pattern leads to the lowest predicted peak concentration of any possible pattern that would
have the same total ingestion and, therefore, the least saturation of metabolism. The episodic
pattern assumes that rats consume their drinking water in six boluses, spaced at 4-hour intervals
during each day. The fraction of their total daily ingestion at each of these events is 23.3, 10, 10,
10, 23.3, and 23.4%, respectively. For continuous exposures, the steady-state blood and brain
concentrations of AN and CEO were calculated as internal dose metrics, while for the episodic
pattern, the daily average concentrations were determined.
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Strieker et al. (2003) measured water and food intake by Sprague-Dawley rats on diets
containing normal (1%) and high (8%) levels of NaCl. On days 5 and 10 of control exposure,
the rats were observed to consume water in 15.9 ± 2.0 and 19.3 ± 2.4 bouts, with amount
consumed per bout being 2.1 ± 0.3 and 1.8 ± 0.3 mL, respectively. Thus, actual consumption by
rats occurs in many more than six bouts, and the amount consumed per bout is more uniform
than the episodic pattern. The data also show that most of the consumption occurs during the 12-
hour dark/waking period (represented by the first and final two boluses in the EPA's pattern),
with less consumed during the light/sleeping period (represented by the second to fourth
boluses). Further, data on water consumption relative to controlled meal delivery showed that
drinking bouts can last as long as 80 seconds, rather than occurring in an instantaneous bolus.
Thus, actual consumption lies between an assumed continuous pattern and the idealized episodic
pattern. Because episodic consumption will lead to model prediction of the greatest metabolic
saturation, to obtain the desired prediction that represents an upper bound of that behavior
(including the possibility that water consumption in the bioassays, which included F344 as well
as Sprague-Dawley rats, may have differed from those observed by Strieker at al. [2003]), this
idealized episodic pattern rather than one designed to more closely match that reported by
Strieker et al. (2003) was chosen. The results of the PBPK simulations for continuous oral
infusion vs. an idealized episodic ingestion pattern for various oral BMDL values differed by no
more than 7%, indicating that for the rat internal dose metrics, the exact exposure pattern was not
significant.
To calculate the FIEDs, similarly two alternate assumptions of continuous and episodic
exposure to the corresponding rat metrics were applied. For the episodic pattern, it was again
assumed that this occurred in six bolus doses but occurring at 0, 3, 5, 8, 11, and 15 hours
computation time (i.e., from the time of first ingestion in the morning, assumed to occur at
breakfast), with the amount consumed in each bolus being 25, 10, 25, 10, 25, and 5% of the total
daily intake, respectively. As with the inhalation extrapolation, the human PBPK model was run
using a range of fixed input dose rates (using both continuous and episodic regimens), the
resulting internal doses calculated, and then an empirical regression performed in Excel to
interpolate between those doses. The resulting simulation values (points) and regressions (lines)
are shown in Figures C-3 to C-6.
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>.
C3
T3
OQ
D)
100
10
o
CL
"2 0.1
o
0.01
Continuous Exposure Model
AN blood fits
O ACN blood
AN brain fits
n ACN brain
y = 27.0624-C + [161.3074 C/(0.569868 + C)]
C + [138.101-C/(0.657976 + C)]
0.00001
0.0001 0.001 0.01 0.1
Internal steady state concentration (mg/L)
Points are PBPK model predictions; curves are polynomial regressions.
Figure C-3. Human oral exposure level vs. internal AN concentration for
continuous exposure.
100
1°
to
D)
A CEO blood
X CEO brain
— Poly. (CEO blood)
Poly. (CEO brain)
Continuous Exposure Model
y = 10.185X2 + 98.2x
R2 =
0.01
0.0001
0.001 0.01 0.1
Internal steady state concentration (mg/L)
Points are PBPK model predictions; curves are polynomial regressions.
Figure C-4. Human oral exposure level vs. internal CEO concentration for
continuous exposure.
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100
10
CO
TO
g.
X
01
15
O 0.1
0.01
Episodic Exposure Model —AN
AN blood fit
O AN blood
AN brain fit
n AN brain
0.00001
y = 51.3757-C + [1 5.2250-C/(0.0586390 + C)]
y = 40.6491 -C + [15.2251 -C/(0.074113 + C)]
0.0001 0.001 0.01
Internal average concentration (mg/L)
0.1
Points are PBPK model predictions; curves are nonlinear regressions.
Figure C-5. Human oral exposure level vs. internal AN concentration for
episodic exposure.
10
•8
0.01
Episodic Exposure Model — CEO
A CEO blood
A CEO brain
Poly. (CEO blood)
•Poly. (CEO brain)
y = 62.272X2 + 98.857x
R2 =
0.0001
0.001 0.01
Internal average concentration (mg/L)
Points are PBPK model predictions; curves are polynomial regressions.
Figure C-6. Human oral exposure level vs. internal CEO concentration for
episodic exposure.
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acslXtreme Model Source Code
PROGRAM
! ACRYLONITRILE and CYANOETHYLENE OXIDE MODEL w/ 1st order epoxide hydrolase
! Model of ACN administration and CEO production!
! Revised model of G.L. Kedderis!
! Original by M.L. Gargas 10/9/89, revised 5/6/92!
! Revised 2/21/94 by G.L. Kedderis!
! Refined 6/28-7/12/95, 8/8/95 by G.L. Kedderis!
Modified by SMH - 5/2/97 !
Modified for continuous oral dose by LMS 3/7/00!
!pst2, pr2, and pb2 revised to Kedderis et al 1996 values!
!periodic drinking water reinstated!
!modified for revised human scaling by LMS 7/26/01!
!built blood flow mass balance into equations LMS 7/27/01!
!built blood mass balance into equations, made blood binding!
!rates equal in venous and arterial blood LMS 7/31/01!
!modified to explicitly define DW cone, and daily ACN intake MHL 08/05!
!statement added to calculate daily ACN intake (mg/kg/day) ml 08/05!
! Modified 05/05/06 by Paul Schlosser and Allan Marcus to include!
! Linear-range approximation to epoxide hydrolase model of ACN to CEO!
!and to simplify the oral dosing calculations -- to avoid divide-by-zero!
!errors and allow for either or both continuous and/or periodic exposure!
! June '06 - Jan '07: Other modifications by P. Schlosser to allow both
!continuous and episodic drinking water dosing, and to update default
!parameters for rat
INTEGER I
REAL DRT(6)
DIMENSION
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
!counter for drinking water arrays!
DRP(6)
AO(16)
AO = 0,0,0,0
QPC = 14.
QCC = 14.
QLC =0.25
QFC =0.09
QBRC = 0.024
QSTC = 0.013
QSC = 0.15
!store drink water times, percents in array!
!Set of zeros to force values to be >/= 0!
0,0,0,0,0,0,0,0,0,0,0,0
Alveolar ventilation rate (L/hr)!
Cardiac output (L/hr)!
Fractional blood flow to liver!
Fractional blood flow to fat!
Fractional blood flow to brain!
Fractional blood flow to the stomach!
Fractional blood flow to slowly perfused tissue
BW = 0.523 ! Body weight (kg) male rat!
VLC =0.04 ! Fraction of liver tissue to total body!
VRC = 0.0377 ! Fraction of richly perfused tissue to total!
VSC =0.75 ! Fraction of slowly perfused tissue to total!
VBRC = 0.006 ! Fraction of brain tissue!
VFC =0.07 ! Fraction of fat tissue!
VSTC = 0.0063! Fraction of stomach tissue!
BVC =0.06 ! Fraction of blood volume!
WBC =0.65 ! Fraction of venous blood volume!
CONSTANT PER =0.40 ! AN brain/blood partition coefficient!
CONSTANT PBR2 = 1.407 ! CEO brain/blood partition coefficient!
CONSTANT PL = 0.46 ! AN liver/blood partition coefficient!
CONSTANT PL2 = 0.274 ! CEO liver/blood partition coefficient!
CONSTANT PST =0.46 ! AN stomach/blood partition coefficient!
CONSTANT PST2 = 0.274 ! CEO stomach/blood partition coefficient!
CONSTANT PF = 0.28 ! AN fat/blood partition coefficient!
CONSTANT PF2 = 0.785 ! CEO fat/blood partition coefficient!
CONSTANT PS = 0.35 ! AN slowly perfused tissue/blood partition!
CONSTANT PS2 = 1.853 ! CEO slowly perfused tissue/blood partition!
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CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
CONSTANT
PR = 0.46
PR2 = 0.274
PB = 512
PB2 = 1658
MW = 53.06
MW2 = 69.05
KEHC =3.92
VMAXC =7.1
VMAXC2 =0.0
KM = 2.76
KM2 = 113.
KFC =50
KFC2 = 500
KSO = 0.2584
GSHL =8.53
GSHBR = 2 .
GSHST =4.59
GSHS =0.75
GSHR = 2 . 65
KBC = 3. 66
KBC2 =3.33
KFBC =2.54
KFBC2 =0.68
KA = 4 .2
IVDOSE = 0.
CONC = 0.
CONC2 = 0 .
ODOSE= 0. !
! AN rapidly perfused tissue/blood partition!
! CEO rapidly perfused tissue/blood partition!
! AN blood/air partition coefficient!
! CEO blood/air partition coefficient !
! ACN molecular weight (g/mol) !
! CEO molecular weight (g/mol) !
! Effective Ist-order EH rate, liter/hr/kgA0 . 7
! Maximum velocity of metabolism (mg/hr-lkg) !
! Maximum velocity of metabolism, CEO (mg/hr-lkg) !
! Michaelis-Menten constant (mg/L) !
! Michaelis-Menten constant, CEO (mg/L) !
! ACN first order metabolism rate const (/hr-lkg) !
! CEO first order metabolism rate const (/hr-lkg) !
! ACN second order reaction with GSH (L/mMol/hr) !
! Liver GSH cone (mMol/L) !
! Brain GSH (inMol/L) !
! Stomach GSH (inMol/L) !
! Slowly perfused (muscle) GSH (mMol/L) !
! Rapidly perfused GSH (mMol/L) !
! ACN 1ST order binding to blood hb (/hr) !
! CEO 1ST order binding to blood hb (/hr) !
! ACN 1ST order binding to blood RSH (/hr) !
! CEO 1ST order binding to blood RSH (/hr) !
! Oral uptake rate (/hr) !
! IV dose (mg/kg) !
! Inhaled concentration (ppm) ACN!
! Inhaled concentration (ppm) CEO!
Oral dose in mg/kg-day, assumed given at t=0, 24, etc
! Timing commands!
CONSTANT TCHNG1 =24.0
CONSTANT TCHNG2 = 120.0
CONSTANT TINF = .003
! Length of exposure (hrs)!
! Allows for 5 day/week exposure!
! Length of IV infusion (hrs)!
CONSTANT STEADYODOSE = 0
!daily dose mg/kg-day, if assuming continuous oral infusion
!periodic drinking water section!
CONSTANT DRCONC=0.0 ! Cone, of ACN in drinking water (mg/L)
CONSTANT DRT=0,2,4,6,8,10
! Times for multiple oral drinks/day *after* 0
! Must be ascending, 0 <= times < 24 hr
! DRTIME(l) assumed = 0 and not used
CONSTANT DRP=1,0,0,0,0,0 !Percent consumed by drinking at those times
! Bolus of ODOSE*BW + DRPCT(1)*DRDOSE will be given at t=0,24,48, etc.
CONSTANT FRACVOL =1.0
! Actual daily water consumption / nominal or control volume
! used when calculating DRDOSE
INITIAL
DRVOL = 0.102*BW**0.7 ! Water consumption (L/d)
! 2L/d water consumption for a 70kg human
DRDOSE = DRCONC*DRVOL*FRACVOL
DAYDOSE = ODOSE*BW + DRP(1)*DRDOSE
! Once daily oral dose; given at t=0 via initial condition
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STDOSE = STEADYODOSE*BW/24 ! Steady oral dosing rate in mg/hr
IVR = IVDOSE*BW/TINF
NEWDAY =0; 1=2 ! First dose given as initial condition
SCHEDULE ORALDOSE.AT.DRT(2)
! Scaled parameters!
QC = QCC*BW**0.74
QP = QPC*BW**0.74
QL = QLC*QC
QF = QFC*QC
QBR = QBRC*QC
QST = QSTC*QC
QS = QSC*QC
QR = QC-QL-QF-QBR-QS-QST
VL = VLC*BW
VF = VFC*BW
VBR = VBRC*BW
VST = VSTC*BW
VS = VSC*BW
VR = VRC*BW
BV = BVC*BW
VAB = BV* (1-WBC)
WB = BV*WBC
GSHRB = GSHBR/GSHL
GSHRST = GSHST/GSHL
GSHRS = GSHS/GSHL
GSHRR = GSHR/GSHL
VMAX = VMAXC*BW**0.7
VMAX2 = VMAXC2*BW**0.7
KF = KFC/BW**0.3
KF2 = KFC2/BW**0.3
KEH = KEHC*BW**0.7
KFBRC = KFC2*.1*GSHRB
KFSTC = KFC2*.1*GSHRST
KFSC = KFC2*.1*GSHRS
KFRC = KFC2*.1*GSHRR
KFBR = KFBRC/BW**.3
KFST = KFSTC/BW**.3
KFS = KFSC/BW**.3
KFR = KFRC/BW**.3
KFB = KFBC
KFB2 = KFBC2
KB = KBC
KB2 = KBC2
!volume arterial blood!
! ratio of brain GSH to liver GSH!
!ratio of stomach GSH to liver GSH!
!ratio of slowly perfused GSH to liver GSH!
! ratio of richly perfused GSH to liver GSH!
! Liver P450 ACN to CEO!
Liver CEO Hydrolysis!
Liver ACN-GSH rate!
Liver CEO-GSH/RSH rate!
Approx. first-order rate in liver!
CEO brain reaction - 10% GSH ratio!
CEO stomach reaction - 10% GSH ratio!
CEO slowly perfused reaction - 10% GSH ratio!
CEO richly perfused reaction - 10% GSH ratio!
CEO brain RSH rate!
CEO stomach RSH rate!
CEO SPT RSH rate!
CEO RPT RSH rate!
DAILYDOSE = ODOSE+STEADYODOSE+(DRDOSE/BW)+(AI/BW/(TSTOP/24))
!daily dose mg/kg-day via oral and/or inhalation routes!
END
!of INITIAL
DYNAMIC
ALGORITHM IALG
MAXTERVAL MAXT
MINTERVAL MINT
CINTERVAL CINT
2 !Gear method for stiff systems!
l.Oe-3
l.Oe-9
0.5
DERIVATIVE
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ACRYLONITRILE
! CI = Concentration in inhaled air (mg/L)!
CIZONE = PULSE(0.0,24.0,TCHNG1) * PULSE(0.0,168.0,TCHNG2)
CI = CIZONE*CONC*MW/24450.
! AI = Amount inhaled (mg)!
RAI = QP*CI
AI = INTEG(RAI,0)
! MR = Amount in stomach lumen (mg)!
RMR = STDOSE - KA*MR
AMR = INTEG(RMR,DAYDOSE)
MR = MAX(AMR, AO(1)) !no negative values!
! CAL = Concentration in arterial lung blood (mg/L)!
CAL = (QC*CVB+QP*CI)/(QC+(QP/PB))
! CA = Cone, in systemic arterial blood (mg/1)!
RAB = QC*(CAL-CA)-(KB+KFB)*AB
AB = INTEG(RAB,AO(2))
CA = AB/VAB
! AX = Amount exhaled (mg)!
CX = CAL/PB
CXPPM = (0.7*CX+0.3*CI)*24450./MW
RAX = QP*CX
AX = INTEG(RAX,AO(3))
! AST = Amount in stomach tissue (mg)!
RAST = QST*(CA-CVST) - STGSH + KA*MR
AST = INTEG(RAST,AO(4))
CVST = AST/(VST*PST)
CST = AST/VST
STGSH = KSO*CVST*GSHST*VST
ASTG = INTEG(STGSH,AO(5))
! AS = Amount in slowly perfused tissues (mg)!
RAS = QS*(CA-CVS) - SGSH
AS = INTEG(RAS,AO(6))
CVS = AS/(VS*PS)
CS = AS/VS
SGSH = KSO*CVS*GSHS*VS
ASG = INTEG(SGSH,AO(7))
! AR = Amount in rapidly perfused tissues (mg)!
RAR = QR*(CA-CVR) - RGSH
AR = INTEG(RAR,AO(8))
CVR = AR/(VR*PR)
CR = AR/VR
RGSH = KSO*CVR*GSHR*VR
ARG = INTEG(RGSH,AO(9))
! AF = Amount in fat tissue (mg)!
RAF = QF*(CA-CVF)
AF = INTEG(RAF,AO(10))
CVF = AF/(VF*PF)
CF = AF/VF
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! ABR = Amount in brain tissue (mg)!
RABR = QBR*(CA-CVBR) - BRGSH
ABR = INTEG(RABR,AO(11))
CVBR = ABR/(VBR*PBR)
CBR = ABR/VBR
BRGSH = KSO*CVBR*GSHBR*VBR
ABRG = INTEG(BRGSH,AO(12))
! AL = Amount in liver tissue (mg)
RAL = QL*CA + QST*CVST - (QL+QST)*CVL - RAM1 - RAM2 - LGSH
AL = INTEG(RAL,AO(13))
CVL = AL/(VL*PL)
CL = AL/VL
AUCL = INTEG(CL,0.)
LGSH = KSO*CVL*GSHL*VL
ALG = INTEG ( (LGSH+RAM2+(KB*CVB*WB) + (KFB*CVB*WB) ) ,AO (14) )
! AMI = Amount metabolized, saturable (P450) and linear pathways (mg)!
RAM1 = VMAX*CVL/(KM+CVL)
RAM1M = RAM1*1000/MW
AMI = INTEG(RAM1,AO(15))
! AM2 = Amount metabolized, first-order pathway (GST) (mg)!
RAM2 = KF*CVL*VL
AM2 = INTEG(RAM2,0)
! CV = Mixed venous blood concentration (mg/L)!
IV = IVR*(T<=TINF) ! PULSE(0,24,TINF)
CV = (QF*CVF + (QL+QST)*CVL + QS*CVS + QR*CVR + QBR*CVBR + IV)/QC
! CVB = Mixed venous ACN cone, after binding (mg/L)!
RVB = QC*(CV-CVB) - (KB+KFB)*VB
VB = INTEG(RVB,AO(16))
CVB = VB/WB
! TMASS = mass balance (mg)!
TMASS = ABR+AF+AL+AS+AR+AST+AM1+AM2+AX+MR+ASTG+ASG+ARG+ABRG+ALG
CEO
! CI2 = Concentration in inhaled air (mg/L)!
CI2 = CIZONE*CONC2*MW2/24450.
! CAL2 = Concentration in arterial lung blood (mg/L)!
CAL2 = (QC*CVB2+QP*CI2)/(QC+(QP/PB2))
! CA2 = Cone, in systemic arterial blood (mg/1)!
RAB2 = QC*(CAL2-CA2)-(KB2+KFB2) *AB2
AB2 = INTEG(RAB2,0.)
CA2 = AB2/VAB
! AS2 = Amount in slowly perfused tissues (mg)!
RAS2 = QS*(CA2-CVS2) - KFS*CVS2*VS
AS2 = INTEG(RAS2,0.)
CVS2 = AS2/(VS*PS2)
CS2 = AS2/VS
C-14 DRAFT - DO NOT CITE OR QUOTE
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! AR2 = Amount in rapidly perfused tissues (mg)!
RAR2 = QR*(CA2-CVR2) - KFR*CVR2*VR
AR2 = INTEG(RAR2,0.)
CVR2 = AR2/(VR*PR2)
CR2 = AR2/VR
! AST2 = Amount in stomach (mg)!
RAST2 = QST*(CA2-CVST2) - KFST*CVST2*VST
AST2 = INTEG(RAST2,0.)
CVST2 = AST2/(VST*PST2)
CST2 = AST2/VST
! AF2 = Amount in fat tissue (mg)!
RAF2 = QF*(CA2-CVF2)
AF2 = INTEG(RAF2,0.)
CVF2 = AF2/(VF*PF2)
CF2 = AF2/VF
! ABR2 = Amount in brain tissue (mg)!
RABR2 = QBR*(CA2-CVBR2) - KFBR*CVBR2*VBR
ABR2 = INTEG(RABR2,0.)
CVBR2 = ABR2/(VBR*PBR2)
CBR2 = ABR2/VBR
! CEO Hydrolysis!
! VMAX2 = maximum velocity of metabolism (mg/hr-lkg)!
! KM2 = Michaelis-Menten constant (mg/L)!
! KEH is parameter for linear metabolism in liver to replace saturable
! AL2 = Amount CEO in liver tissue (mg)!
RAM = RAM1M*MW2/1000.
RAL2ADD = QL*CA2 + QST*CVST2 + RAM
RAL2M = VMAX2*CVL2/(KM2+CVL2) + KEH*CVL2
RAL2 = RAL2ADD - (QL+QST)*CVL2 - KF2*CVL2*VL - RAL2M
AL2 = INTEG(RAL2,0.)
CVL2 = AL2/(VL*PL2)
CL2 = AL2/VL
! CV2 = Mixed venous blood concentration (mg/L)!
CV2 = (QF*CVF2 + (QL+QST)*CVL2 + QS*CVS2 + QR*CVR2 + QBR*CVBR2)/QC
! CVB2 = Mixed venous CEO cone, after binding!
RVB2 = QC*(CV2-CVB2) - (KB2+KFB2)*VB2
VB2 = INTEG(RVB2,0.)
CVB2 = VB2/WB
! Calculation of the AUC for ACN and CEO in the brain and blood!
AUCBR = INTEG(CBR,0.) ! AUC for ACN brain cone.!
AUCBR2 = INTEG(CBR2,0.) ! AUC for CEO brain cone.!
AUCB = INTEG(CVB,0.) ! AUC for ACN blood cone.!
AUCB2 = INTEG(CVB2,0.) ! AUC for CEO blood cone.!
DAILYSTAUC = INTEG(CST, 0.)/(TSTOP/24) !Mean daily stomach AN
DAILYST2AUC = INTEG(CST2 , 0.)/(TSTOP/24) !Mean daily stomach CEO
END ! Derivative
! Code that is executed once at each communication interval goes here
CONSTANT TSTOP =24.0 ! Length of experiment (hrs)!
TERMT(T.GE.TSTOP, 'checked on communication interval: REACHED TSTOP1)
DISCRETE ORALDOSE
IF (I.EQ.l) THEN
C-15 DRAFT - DO NOT CITE OR QUOTE
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AMR = AMR + DAYDOSE ! Initial dose of day/daily dose
ELSE
AMR = AMR + DRP(I)*DRDOSE ! Drinking percent
END IF
I = MOD((1+1),6)
IF (I.EQ.l) THEN
SCHEDULE ORALDOSE.AT.NEWDAY ! Go to start of the next day
NEWDAY = NEWDAY + 24
ELSE
SCHEDULE ORALDOSE.AT.(NEWDAY+DRT(I)) ! Go to next drink time
END IF
END
! DISCRETE ORALDOSE
END ! Dynamic
TERMINAL ! Metrics calculated below!
DAILYBRAUC = AUCBR/(TSTOP/24) ! Mean daily stomach ACN
DAILYBR2AUC = AUCBR2/(TSTOP/24) ! Mean daily stomach CEO
DAILYBAUC = AUCB/(TSTOP/24) ! Mean daily blood ACN
DAILYB2AUC = AUCB2/(TSTOP/24) ! Mean daily blood CEO
PEAKPBR = MAX(0, CBR)
PEAKPB = MAX(0, CV)
PEAKMBR = MAX(0, CBR2)
PEAKMB = MAX(0, CV2)
Peak cone, parent compound in brain!
Peak cone, parent compound in blood!
Peak cone, metabolite in brain!
Peak cone, metabolite in blood!
! Calc. of the avg. values of ACN and CEO in the brain and blood!
AVEBR = AUCBR/TSTOP
AVEBR2 = AUCBR2/TSTOP
AVEB = AUCB/TSTOP
AVEB2 = AUCB2/TSTOP
Average of ACN brain cone.
Average of CEO brain cone.
Average of ACN blood cone.
Average of CEO blood cone.
END
TERMINAL
END
! PROGRAM
C-16
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Matlab source code - 3 programs (.m files)
% ACN / CEO PBPK model optimization file
%
% Main program, or SHELL for PBPK model.
%
% P = optACN2(pin, pvn, datfilename, switch)
%
% swtch = optimization choice (see below)
% pin = values of *all* parameters (input)
% pvn = cell-vector of parameter names TO BE VARIED
% datfilename = name of data file (.csv) to use
% P = values of the output parameters (optimized)
%
% where pin is the vector of paramters to input and swtch determines whether
% to run the simulation (swtch =0), or to optimize (swtch = 1 neld2
search).
%
% pvn = {'namel' ; 'name2' ; ... } Note: use curly brackets and semi-colons
function P = optACN(pin,pvn,dfn,sw)
format long
% globalize data sets-
global quan phys prs Ypt pv yname itr datfile pvar pbl j 1 jf ncv tsp pex
global ICs Ccafpt Ccafc nv Cost CO ps MW MW2 nv
% load dat set
quan.dat=load([dfn, ' .csv1] ) ;
swtch=sw;
%close all
% The sequence of columns in the data sets are as follows:
% tspan, CVB, CL, CBR, CVB2, CL2, CBR2, ODOSE(initial), IVDOSE(initial) ...
% % AN-GSH in urine, % CEO-GSH in urine, % Hb-bound material
yname={'CVB1;'CL';'CBR1;'CVB2';'CL2';'CBR2';'AGSHU';'CGSHU';'PBH';'MassBal'};
ncv=32;
% find indeces of parameters to be varied
pname={'VmaxC';'Km';'VmaxC2';'Km2';'kEHC';'kFC';'kFC2';'kA';'gammA';'gammC';'
kH'; 'kH2'} ;
%prs = [5.0; 1; l.e-18; 113.; 3.9; 73.; 500.; 8.; 1; 1];
%prs = [3.27;.000425; l.e-18; 113.; 3.9; 95.7; 500.; 8.; .0107; 1];
prs = [.102; 2.76; l.e-18; 113.; 0.98; 82.3; 14.1; 4.94; .0107; 0.941;
1.245; 1.134];
prs = [2.24; 2.76; l.e-18; 113.; 0.98; 82.3; 14.1; 4.94; .0107; 0.941;
1.245; 1.134];
%prs = [1.86; 1.5; l.e-18; 113.; 3.9; 15.7; 2240.; 8.; .425; 1.10];
%those above are default values
v)Kl— / ' I.I I.I I.I I.I I.I I.I I.I I.I I.I I.I I.I I 1 .
% KEHC = %Effective Ist-order EH rate, liter/hr/kgA0.7
% VMAXC = %Maximum velocity of metabolism (mg/hr-lkg)%
% VMAXC2 = %Max. vel. of metabolism, CEO (mg/hr-lkg)%
% KM = %Michaelis-Menten constant (mg/L)%
% KM2 = %Michaelis-Menten constant, CEO (mg/L)%
% KFC = %ACN first order metab rate const (/hr-lkg)%
% KFC2 = %CEO first order metab rate const (/hr-lkg)%
% KA = %0ral uptake rate (/hr)%
C-17 DRAFT - DO NOT CITE OR QUOTE
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fname='EqACN'; % file with set of model equations to use is fname.m
runfn='RunACN'; % Run file runfn.m
quan.ns = length(pname); pv = liquan.ns; pvar = pname;
if length(pin)>quan.ns
error('input parameter vector is too long')
end
if sum(pin) % if pin =[], then no optimization, use default params
popt=reshape(log(pin),length(pin), 1);
if strmatch('all',pvn, 'exact')
if length(pin)~=quan.ns
error('length of pin is wrong; must be ',num2str(quan.ns),' to
match pvn = all')
end
prs=pin;
else
pvar = pvn;
quan.ns = length(pvn);
pv = zeros(1,quan.ns);
if (length(pin)~=quan.ns)&(length(pin)~=length(pname))
error(['Length of pin is wrong; must be ',num2str(quan.ns),...
1 to match pvn or ',num2str(length(pname)),' to match
pname.'])
else
for i = 1:quan.ns
if strmatch(pvn(i),pname,'exact')
pv(i) = strmatch(pvn(i),pname,'exact');
else
error([char(pvn{i}),' is not in the named list of
parameters.'])
end
end
end
if length(pin)==length(pname)
prs=pin; popt=popt(pv);
end
end
else
popt=log(prs); swtch=0; 'No pin so running simulation only with default
params.'
end
save pv pv
% find indeces of start times
quan.t = quan.dat(:,1);
tci = (quan.t > 0);
datc=quan.dat(:, [2:7,12 : 14] ) ; % data columns used for fitting
nv = size(date,2); % number of variables to plot, etc
Ccafc=zeros(length(quan.t),(nv+1));
nplt=0;
for i=l:6
quan.ic{i} = find((date(:,i)>0) & tci);
quan.datc{i} = date(quan.ic{i}, i);
if quan.datc{i}
nplt=nplt+l;
end
quan.tc{i} = quan.t(quan.ic{i});
quan.nc(i) = sum ( (date(:,i)>0) & tci);
end
nplt2=0;
for i=7:nv
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quan.ic{i} = find((date(:,i)>0) & tci);
quan.datc{i} = date(quan.ic{i} , i) ;
if quan.datcfi}
nplt2=nplt2+l;
end
quan.tcfi} = quan.t(quan.ic{i});
quan.nc(i) = sum ( (date(:,i)>0) & tci);
end
C0=sum(quan.nc)*(log(2*pi)+l) ;
quan.mc=datc>0;
sz=length(quan.t) ;
quan.yc=0*[date,date(:,!)];
quan.jt = find(quan.t == 0);
% jt = vector of indeces of rows in data file starting with "0"
quan.idf= length(quan.j t) ;
quan.j t = [quan.jt; (sz + 1)] ;
% add a pseudo-index to mark the end of the data file
% constant parameters
ps.KSO = 0.2584; %ACN second order rxn with GSH (L/nnMol/hr)%
ps.GSHL = 8.53; %Liver GSH cone (inMol/L) %
ps.GSHBR =2.; %Brain GSH (inMol/L) %
ps.GSHST = 4.59; %Stomach GSH (inMol/L) %
ps.GSHS = 0.75; %Slowly perfused (muscle) GSH (inMol/L) %
ps.GSHR = 2.65; %Rapidly perfused GSH (inMol/L) %
%ps.KB = 1.245 + 2.54; %ACN 1ST order binding to blood hb + RSH (/hr)%
ps.KBO = 2.54; %ACN 1ST order binding to blood hb + RSH (/hr)%
ps.KBO
%ps.KH = 1.245;
ps.KFB = 2.54;
%ps.KB2 = 1.134 + 0.68 + 0.413;
ps.KB20 = 0.68 + 0.413;
%ps.KH2 = 1.134;
phys.KHR = 1.134/1.245;
ps.KFB2 = 0.68;
%CEO 1ST order binding to blood hb + RSH + chem. hydrol. (/hr)%
ps.tinf=.003; %blood infusion time (hr)
% ACN PCs
phys.PL = 0.46; % livenblood PC
phys.PST = 0.46; % stomach tissue:blood PC, also used for GI
phys.PER = 0.40; % brain tissue:blood PC
phys.PF = 0.28; % fat tissueiblood PC
phys.PS = 0.35; % slowly tissueiblood PC
phys.PR = 0.46; % richly tissueiblodd PC
phys.PB = 512; % blood tissueiblodd PC
%CEO PCs
phys.PL2 = 0.274; % liveriblood PC
phys.PST2 = 0.274; % stomach tissue:blood PC, also used for GI
phys.PBR2 = 1.407; % brain tissueiblood PC
phys.PF2 = 0.785; % fat tissueiblood PC
phys.PS2 = 1.853; % slowly tissueiblood PC
phys.PR2 = 0.274; % richly tissueiblodd PC
phys.PB2 = 1658; % blood tissueiblodd PC
QPC =14.; %Alveolar ventilation rate (L/hr)%
QCC =14.; %Cardiac output (L/hr)%
C-19 DRAFT - DO NOT CITE OR QUOTE
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QBRC = 0.024; % blood flow to brain
QSTC = 0.013; % Blood flow rate for stomach tissue (portal vein)
QFC =0.09; % blood flow to fat tissue
QLC = 0.25; % Blood flow rate for liver )
QSC = 0.15; % Blood flow rate for slowly
phys.BW=0.25; %body weight
VLC = 0.04; %Fraction of liver tissue to total body%
VRC = 0.0377; %Fraction of richly perf tiss to total%
VSC = 0.75; %Fraction of sloly perf tiss to total%
VBRC = 0.006; %Fraction of brain tissue%
VFC = 0.07; %Fraction of fat tissue%
VSTC = 0.0063; %Fraction of stomach tissue%
BV = 0.06*phys.BW; %Blood vol%
WBC = 0.65; %Fraction of venous blood vol%
MW = 53.06; %ACN molecular weight (g/mol)%
MW2 = 69.05; %CEO molecular weight (g/mol)%
% preset varying parameters
phys.QC = QCC*phys.BWA0.74;
phys.QP = QPC*phys.BWA0.74;
phys.QBR = QBRC*phys.QC; % blood flow to brain
phys.QST = QSTC*phys.QC; % Blood flow rate for stomach tissue (portal
vein)
phys.QS = QSC*phys.QC; % Blood flow rate for slowly
phys.QL = QLC*phys.QC; % Blood flow rate for liver (hepatic artery)
phys.QF = QFC*phys.QC; % blood flow to fat
phys.QR = phys.QC-phys.QL-phys.QF-phys.QBR-phys.QS-phys.QST;
phys.VBR = VBRC*phys.BW; % brain volume (L)
phys.VST = VSTC*phys.BW; % GI volume (use RP density)
phys.VL = VLC*phys.BW;
phys.VF = VFC*phys.BW;
phys.VBR = VBRC*phys.BW;
phys.VST = VSTC*phys.BW;
phys.VS = VSC*phys.BW;
phys.VR = VRC*phys.BW;
phys.VAB = BV*(1-WBC); %volume arterial blood%
phys.WB = BV*WBC;
phys.GSHRB = 0.l*ps.GSHBR/ps.GSHL; %ratio of brain GSH*GST to liver GSH%
phys.GSHRST = 0.l*ps.GSHST/ps.GSHL; %ratio of stomach GSH*GST to liver GSH%
phys.GSHRS = 0.l*ps.GSHS/ps.GSHL; %ratio of spt GSH*GST to liver GSH%
phys.GSHRR = 0.l*ps.GSHR/ps.GSHL %ratio of rpt GSH*GST to liver GSH%
ICs=[];
for i=l:quan.idf % idf is the number of simulations/initial conditions in
jl(i) = quan.jt(i); % index of initial data row for this
experiment/simulation
jf(i) = quan.jt(i+1)-1; % index of final data row for this
expt./simulation
tsp{i} = quan.t(jl(i) :jf(i)) ;
oralO=quan.dat(jl(i),8)*phys.BW;
ps.bloodO(i)=quan.dat(j1(i),9)*phys.BW/ps.tinf;
ps.cinh(i)=quan.dat(jl(i),10)*MW/24450; ps.tinh(i)=quan.dat(jl(i),11);
ICs=[ICs;[oralO,zeros(l,ncv-l)]]; % initial conditions
end
save ICs ICs
% call simulation or optimization procedure
C-20 DRAFT - DO NOT CITE OR QUOTE
-------
quan.pt=0; itr=0; P=prs;
if swtch==0
quan.pt=l;
y=eval([runfn,'(popt)']);
elseif swtch==l
pnew = exp(neld2(popt, runfn, le-6, 1000, 50000));
else
options = optimset('MaxFunEvals',8000,'Maxlter',8000);
pnew = exp(fminsearch(runfn, popt, options));
end
if swtch>0
P(pv) = pnew(1:quan.ns);
'Initial Parameter Values, Jinit ='
Cst=eval([runfn, ' (popt) '] ) ;
'Final Parameters, Jopt ='
quan.pt=l;
Cst=eval([runfn,'(log(P(pv)))']);
end
% final results/plots
co =
['rbkgmcrbkgmcrbkgmcrbkgmcrbkgmcrbkgmcrbkgmcrbkgmcrbkgmcrbkgmcrbkgmcrbkgmcrbk
gmcrbkgmcrbkgmcrbkgmc'];
%co = ['kkkkkkkkkkkkkkkkkkkkkkkkkkkkkkkkkkkkkkkkkkk'] ;
pc =
['+o*xsdAv<>ph+o*xsdAv<>ph+o*xsdAv<>ph+o*xsdAv<>ph+o*xsdAv<>ph+o*xsdAv<>ph+o*
xsdAv<>ph+o*xsdAv<>ph'];
linet = ['-
np=min(quan.idf, 96) ;
quan.j tp = [find(quan.tp == 0) ; (length(quan.tp) +1)] ;
% jt = vector of indeces of rows in data file starting with "0"; plus one
%close all
kfO=30;
kf=l+kfO;
figure(kf)
rows=ceil(nplt/2);
set(kf, 'Units', 'inches','Position', [0.1, 0.5, 6.3, (0.1+2.5*rows)]);
set(gcf,'DefaultTextFontSize',10,'DefaultAxesFontSize',12,'DefaultTextFontNam
e','Arial');
npt = -1;
for k = 1:6
if quan.datc{k}
npt=npt+2;
pos=mod(npt,(rows*2))+(npt>(rows*2));
h=subplot(rows,2,pos);
set(h,'Units','inches','Position',[(3.5-
2.95*mod(pos,2)),(0.45+2.35*(rows-ceil(pos/2))),2.7,2.1])
box on
jl=quan.jt (1) ;
J2=quan.jt(2)-l;
pl=quan.jtp (1) ;
p2=quan.j tp(2)-1;
C-21
DRAFT - DO NOT CITE OR QUOTE
-------
dt=quan . t ( j 1 : j 2 ) ;
dp = date ( j 1 : j 2 , k) ;
dj = find (dp > 0) ;
dk=find(datc( : ,k)>0) ;
xm=ceil (1. 01*max(quan.t ( dk ))*10)/10;
xl=floor (0.99*min(quan.t ( dk )*10)-1)/10;
if xl<0
xl = - round(2 . 5*xm) /100;
end
ym=max (loglO (1. 1* [Ccafpt (find(Ccafpt ( : , k) >0) , k) ;datc (dk, k) ] ) ) ;
if (10Aceil (ym) ) / (10Aym) > 2
ym = ( 10Aceil (ym) ) /2 ;
else
ym = 10Aceil (ym) ;
end
set (gca, 'YScale' , 'log' )
if k==l
axis ( [ xl , xm, .03, ym] )
end
if k==2
axis ( [ xl , xm, .01, ym] )
end
if k==3
axis ( [ xl , xm, .01, ym] )
end
if k==4
axis ( [ xl , xm, .001, ym] )
end
if k==5
axis ( [ xl , xm, .001, ym] )
end
if k==6
axis ( [ xl , xm, .001, ym] )
end
hold on
if dj
plot(dt(dj) ,dp(dj) , ['k',pc(l) ] ) ; %, [co(l) ,pc(l) ] )
plot (quan. tp (pl:p2) , Ccafpt (pi : p2 , k) , [ ' k ' , linet (1, : ) ] ) ;
end
if np > 1
for i = 2 : np
j l=quan. jt (i) ;
J2=quan. jt (i+1) -1;
dt=quan . t ( j 1 : j 2 ) ;
pl=quan. j tp (i) ;
p2=quan . j tp (i + 1) -1 ;
dp = date ( j 1 : j 2, k) ;
dj = find (dp > 0) ;
if dj
plot(dt(dj),dp(dj) , ['k',pc(i) ] ) ; %, [co(l) ,pc(l) ] )
plot (quan. tp (pl:p2) , Ccafpt (pi : p2 , k) , [ ' k ' , linet (i, : ) ] ) ;
end
end
end
yl=0. l*get (gca, 'YLim' ) ;
title (yname ( k) ,' Position ', [mean (get (gca, 'XLim' ) ) , yl (2) ] , ' Font Size ' , 14,
1 EdgeColor ' , ' k1 , ' BackgroundColor ' , ' w' , ' Fontname ' , 'Arial ' ) ;
C-22 DRAFT - DO NOT CITE OR QUOTE
-------
if k==2
legend ( ' show' )
end
hold off
end
end
kf=2+kfO;
if quan.datc{7}
figure ( kf ) ;
set (kf, 'Units' , 'inches' , 'Position' , [0.1, 0.5, 3.3, 7.6]);
set (gcf , ' DefaultTextFontSize ' , 14, ' Def aultAxesFontSize ' , 12, ' Def aultTextFontNam
e' , 'Arial' ) ;
for k = 7:9
h=subplot (3,l,k-6) ;
set (h, 'Units' , 'inches' , 'Position' , [ .55, (0.45+2.35* ( 9- k)), 2. 7, 2.1])
box on; hold on
dk=find(datc( : ,k)>0) ;
plot (quan.dat ( dk- 1,8) ,datc(dk,k) , [ ' k ' , pc ( k-6) ] ) ; %, [ co ( 1 ) , pc ( 1 ) ] )
plot (quan.dat ( dk- 1,8) ,Ccafc(dk,k) , [ 'k' , linet (k-6, : ) ] ) ;
yl=0.3*get (gca, 'YLim' ) ;
title (yname ( k) ,' Position ', [mean (get (gca, 'XLim' ) ) , yl (2) ] , ' Font Size ' , 14,
' EdgeColor ' , ' k' , ' BackgroundColor ' , ' w' , ' Fontname ' , 'Arial ' ) ;
hold off
end
end
k=nv+l; % show mass balance
figure(k)
set(k, 'Units', 'inches', 'Position', [1,1,4.1,4.1]);
hold on
title(yname(k));
pl=quan.jtp (1) ;
p2=quan.j tp(2)-1;
plot(quan.tp(pl:p2),Ccafpt(pl:p2,k), [co(1),linet (1, :)] ) ;
for i = 2:np
pl=quan.jtp(i) ;
p2=quan.j tp(i + 1)-1;
plot(quan.tp(pl:p2),Ccafpt(pi:p2 , k) , [co(i),linet(i, : ) ] ) ;
end
if swtch>0
'optimized params = '
[num2str([l:length(P)]'),char(pbl),char(pname),char(pbl),num2str(P)]
end
%RunACN file to return value of objective function, given input parameters
function Cost = RunACN(popt)
% globals
global phys prs yname quan itr pv pvar pbl jl jf ian ICs ncv tsp Ccafpt CO
global Cost Ccafc MW MW2 VMAX VMAX2 KF KF2 KEH KFBRC KFSTC KFSC KFRC KFBR
global KFST KFS KFR KA KM KM2 Ccafpt nv ps
tol=l.e-8;
options = odeset('RelTol',tol,'AbsTol',tol);
itr=itr+l;
C-23 DRAFT - DO NOT CITE OR QUOTE
-------
comp=0;
p = reshape(exp(popt),length(popt),1);
pu=prs;
pu(pv)=p;
save ptemp pu
[num2str(pv1),char(pbl(pv)),char(pvar),char(pbl(pv)),num2str(p)]
%quan.gam(1:5)=2./exp(exp(-popt( (quan.ns + 1) : (quan.ns+5) ) ) ) ;
%quan.gam(6:11)=quan.gam(5);
%quan.gam(1:5)
% display current values of parameters being valued (retransformed)
Ccafpt=[];Ccafw=[];
quan.tp=[];
comp=comp+l;
VMAXC = pu(l); %Max. vel. of metabolism, CEO (mg/hr-kgA0.7)%
KM = pu(2); %Michaelis-Menten constant (mg/L)%
VMAXC2=pu(3); %Max. vel of metabolism, CEO (mg/hr-jgA0.7)
KM2=pu(4); %CEO M-M constant (mg/L)
KEHC = pu(5); %Effective Ist-order EH rate, liter/hr/kgA0.7
KFC = pu(6); %ACN first order metab rate const (/hr-lkg)%
KFC2 = pu(7); %CEO first order metab rate const (/hr-lkg)%
KA = pu(8); %0ral uptake rate (/hr)%
gA = min(pu(9),2); % gammA
gC = min(pu(10),2); % gammC
ps.KH = pu(ll);
ps.KHR2 = pu(12);
ps.KB = ps.KH + ps.KBO; %ACN 1ST order binding to blood hb + RSH (/hr)%
ps.KH2 = ps.KH*phys.KHR*ps.KHR2;
ps.KB2 = ps.KH2 + ps.KB20
% VMAX = VMAXC*(phys.BWA0.7); %Liver P450 ACN to CEO%
VMAX2 = VMAXC2*(phys.BWA0.7); %Liver CEO Hydrolysis%
KF = KFC/(phys.BWA0.3); %Liver ACN-GSH rate%
KF2 = KFC2/(phys.BWA0.3); %Liver CEO-GSH/RSH rate%
KEH = KEHC*(phys.BWA0.7); % Approx. first-order rate in liver %
%CEO rxn rates%
KFBR = KF2*phys.GSHRB; %CEO brain RSH rate%
KFST = KF2*phys.GSHRST; %CEO stomach RSH rate%
KFS = KF2*phys.GSHRS; %CEO SPT RSH rate%
KFR = KF2*phys.GSHRR; %CEO RPT RSH rate%
% loop for running equations
% figure(15)
% set(15, 'Units', 'inches', 'Position',[1,1,4.1,4.1]);
% hold on
% title('RAM1');
for i=l:quan.idf % idf is the number of simulations/initial conditions in
% the data file this was computed in optACN.m
ian=i;
tsc = [0:1050] '*quan.t(jf (i) )/1000;
%ICs(i,[1,16])
[ts,Ys] = odelSs(@EqACN2,tsc,ICs(i,:),options); %
%[ts(1:4),Ys(1:4,1:16)]'
%[Ys(l:10,13),VMAX*Ys(1:10,13)./(KM+Ys(1:10,13)),(Ys(2:ll,15)-Ys(1:10,15))]
massb=sum(Ys(:,1:16) ,2) ;
% plot(ts,Ys(:,15)); %VMAX*Ys(:,13)./(KM+Ys(:,13)));
% The sequence of columns in the data sets are as follows:
C-24 DRAFT - DO NOT CITE OR QUOTE
-------
% tspan, CVB, CL, CBR, CVB2, CL2, CBR2, ODOSE(initial), IVDOSE(initial)
Yp=[Ys ( : ,16)/phys.WB, Ys ( : , 13)/phys . VL, Ys ( : , 11)/phys . VBR, ...
Ys ( : ,24)/phys.WB, Ys ( : , 23)/phys . VL, Ys ( : , 22 )/phys . VBR, ...
Ys(:,30:32)/phys.BW, massb];
Ccafc(jl(i) :jf(i) , :)=interpl(ts,Yp,tsp{i}) ;
quan.tp=[quan.tp;ts];
Ccafpt=[Ccafpt;Yp] ;
Ccafw=[Ccafw;[tsp{i},Ccafc(j1(i):jf(i),:)]];
% end loop
end
if quan.pt==l
yname = {'AN in blood'; 'AN in liver'; 'AN in brain'; ...
'CEO in blood'; 'CEO in liver'; 'CEO in brain'; ...
'AN-GSH in urine'; 'CEO-GSH in urine'; ' Hb binding'; 'mass
balance'};
fid=fopen(['ACN.txt'], 'w') ;
fprintf(fid, ['time', '\t', 'CVB', '\t' , 'CL' , '\t' , 'CBR', '\t', . . .
'CVB2','\t','CL2','\t','CBR2','\t','mball','\r']);
fprintf(fid, '%g\t%g\t%g\t%g\t%g\t%g\t%g\t%g\r',Ccafw') ;
%[quan.tp,Ccafpt]');
status=fclose(fid);
end
%Ccafpt=Ccafc;quan.tp=quan.t; % uncomment this line to plot only the fitted
%values
% cost function
Cost=CO+max(0,pu(9)-2)+max(0,pu(10)-2);gamm=[gA gA gA gC gC gC gC gC gC];
for i=l:nv
if quan.datcfi}
Cost = Cost + gamm(i)*sum(quan.datc{i}) + ...
quan.nc(i)*log( sum( ((quan.datc{ i} - Ccafc(quan.ic{i}, i)). A2)./...
(Ccafc(quan.ic{i}, i) .Agamm(i) ) )/quan.nc(i) );
end
end
['itr = ',num2str(itr),'; Cost = ',num2str(Cost)]
% ACN PBPK model equation file — returns derivatve dx = dx/dt, given
% state variable x and parameters passed through global statements
%
% function defined
function dx = eqs(t,x)
% global s
global prs phys ICs quan ps ian MW2 MW CINH TINH
global VMAX VMAX2 KF KF2 KEH KFBRC KFSTC KFSC KFRC KFBR KFST KFS KFR KA KM
KM2
% assign parameters to names
dx=zeros(size(x));
% "dx" = dx/dt, where x is the state vector
% i
% ACRYLONITRILE !
a I
C-25 DRAFT - DO NOT CITE OR QUOTE
-------
% MR = x(l); %MR = Amount in stomach lumen (mg)!
% AB = x(2); %Amount in systemic arterial blood (mg)!
CA = x(2)/phys.VAB; %concentration
%AX = x(3); % ACN exhaled (mg)!
%AST = x(4); % ACN in stomach tissue tissue (mg)!
CST = x(4)/phys.VST; CVST = CST/phys.PST; % tissue/venous concn's
%ASTG = x(5); % ACN GSH conjugated in stomach tissue (mg)
%AS = x(6); % ACN in slowly perfused (mg)
CVS = x(6)/(phys.VS*phys.PS); % venous concentration
%ASG = x(7); % ACN-GSH conjugated in slowly (mg)
%AR = x(8); % ACN in richly (mg)
CVR = x(8)/(phys.VR*phys.PR) ; % venous concentration
%ARG = x(9); % ACN-GSH conjugated in richly
%AF = x(10); % ACN in fat (mg)
CVF = x(10)/(phys.VF*phys.PF) ; % venous concentration
%ABR = x(ll); % ACN in brain (mg)
CBR = x(11)/phys.VBR; CVBR = CBR/phys.PER; % tissue/venous concn's
%ABRG = x(12); % ACN-GSH conjugated in brain
%AL = x(13); % ACN in liver (mg)
CL = x(13)/phys.VL; CVL = CL/phys.PL; % tissue/venous concn's
% ALG = x(14); ACN-GSH conjugated in liver (mg)
% AMI = x(15); ACN metabolized by P450 in liver (mg)
% AVB = x(16); ACN in mixed venous blood (mg)
CVB = x (16)/phys . WB; % concentration
RMR = KA*x(l); % rate of absorption from stomach
CI=ps.cinh(ian)*(t<=ps.tinh (ian) ) ; %Inhaled concentration
CAL = (phys.QC*CVB+phys.QP*CI)/(phys.QC+(phys.QP/phys.PB));
%CAL = Concentration in arterial lung blood (mg/L)!
%CX = CAL/phys.PB; % concentration in exiting pulmonary air
% CXPPM = (0.7*CX+0.3*CI)*24450./ps.MW
%RAB = ps.KB*x(2)CA*phys.VAB; % ACN binding in arterial blood
STGSH = ps.KSO*CVST*ps.GSHST*phys.VST; % stomach GSH conjugation of ACN
SGSH = ps.KSO*CVS*ps.GSHS*phys.VS; % slowly GSH conjugation of ACN
RGSH = ps.KSO*CVR*ps.GSHR*phys.VR; % richly GSH conjugation of ACN
BRGSH = ps.KSO*CVBR*ps.GSHBR*phys.VBR; % brain GSH conjugation of ACN
LGSH = ps.KSO*CVL*ps.GSHL*phys.VL; % liver GSH conjugation of ACN
% ACN metabolized,saturable (P450) and linear pathways (mg)!
RAM1 = VMAX*CVL/(KM+CVL) ; % (mg/hr)
%VMAX,KM,RAM1
%ACN metabolized,first-order pathway (GST) (mg)!
RAM2 = KF*CVL*phys.VL; % (mg/hr)
RTV =
phys.QF*CVF+(phys.QL+phys.QST)*CVL+phys.QS*CVS+phys.QR*CVR+phys.QBR*CVBR;
% mixed venous blood concentration
%RVB = ps.KB*CV*phys.WB; % ACN binding in venous blood
%[t,CVB/(RTV/phys.QC),CAL/CVB,CA/CAL,x(1),RMR]
dx(l) = - RMR; % ACN stomach lumen
dx(2) = phys.QC*(CAL - CA) - ps.KB*x(2); % ACN arterial blood after binding
dx(3) = phys.QP*CAL/phys.PB; % ACN exiting pulmonary air
dx(4) = phys.QST*(CA-CVST) - STGSH + RMR; % ACN stomach tissue
dx(5) = STGSH; % ACN-GSH conjugated in stomach tissue
dx(6) = phys.QS*(CA-CVS) - SGSH; % ACN slowly
dx(7) = SGSH; % ACN-GSH conjugated in slowly
dx(8) = phys.QR*(CA-CVR) - RGSH; % ACN in richly (mg)
dx(9) = RGSH; % ACN-GSH conjugated in richly
dx(10) = phys.QF*(CA-CVF); % ACN in fat (mg)!
dx(ll) = phys.QBR*(CA-CVBR) - BRGSH; % ACN in brain (mg)!
C-26 DRAFT - DO NOT CITE OR QUOTE
-------
dx(12) = BRGSH; % ACN-GSH conjugated in brain
dx(13) = phys.QL*CA + phys . QST*CVST - (phys . QL+phys . QST) *CVL -RAM1 -RAM2 -
LGSH;
dx(14) = LGSH + RAM2 + ps . KB* (x (2 ) +x ( 16) ) ;
% ACN-GSH conjugated in liver + bound in blood
dx(15) = RAM1; % ACN metabolized by P450 in liver
dx(16) = RTV - phys.QC*CVB - ps.KB*x(16) + ps . bloodO (ian) * ( t<=ps . tinf ) ;
%Mixed ven. ACN cone after binding (mg/L)
dx(30) = STGSH + SGSH + RGSH + BRGSH + LGSH + RAM2 + ps . KFB* (x (2 ) +x ( 16) ) ;
% amount of AN binding to GSH
CEO
%AB2 = x(17); % CEO in arterial blood (mg)
CA2 = x (17) /phys .VAB; % concentration
%AS2 = x(18); % CEO in slowly perfused (mg)
CVS2 = x(18) / (phys.VS*phys.PS2) ; % venous concn
SGSH2 = KFS*CVS2*phys.VS;
%AR2 = x(19); % CEO in richly (mg)
CVR2 = x(19) / (phys.VR*phys.PR2) ; % venous concn
RGSH2 = KFR*CVR2*phys .VR;
%AST2 = x(20); % CEO in stomach tissue tissue (mg) !
CST2 = x (20) /phys. VST; CVST2 = CST2/phys . PST2 ; % tissue/venous concn1 s
STGSH2 = KFST*CVST2*phys .VST;
%AF2 = x(21); % CEO in fat (mg)
CVF2 = x (21) / (phys .VF*phys. PF2) ; % venous concn
%ABR2 = x(22); % CEO in brain (mg)
CBR2 = x (22) /phys .VBR; CVBR2 = CBR2/phys . PBR2 ; % tissue/venous concn 's
BRGSH2 = KFBR*CVBR2*phys .VBR;
%AL2 = x(23); % CEO in liver (mg)
CL2 = x (23) /phys .VL; CVL2 = CL2/phys . PL2 ; % tissue/venous concn1 s
% AVB2 = x(24); % Mixed ven. CEO cone after binding!
CVB2 = x (24) /phys . WB; % concentration
RALG2 = (VMAX2*CVL2/ (KM2+CVL2) ) + KF2*CVL2*phys . VL;
% CEO hydrolysis, saturable + linear terms
% For fitting, either saturable or linear is set to zero
CV2 = (phys .QF*CVF2 + (phys . QL+phys . QST) *CVL2 + phys.QS*CVS2 + phys.QR*CVR2
+ phys .QBR*CVBR2) /phys .QC; % CEO mixed venous blood cone. (mg/L)
CAL2 = (phys .QC*CVB2) / (phys . QC+ (phys .QP/phys . PB2) ); % CEO in arterial lung
blood (mg/L) !
dx(17) = (phys.QC*CAL2)- (phys . QC*CA2 ) - (ps . KB2*x ( 17 ) ) ;
% CEO in systemic arterial after binding
dx(18) = phys .QS* (CA2-CVS2) - SGSH2; % CEO in slowly perfused tissues (mg) !
dx(19) = phys.QR* (CA2-CVR2) - RGSH2; % CEO in rapidly perfused tissues (mg) !
dx(20) = phys .QST* (CA2-CVST2) - STGSH2; % CEO in stomach (mg) !
dx(21) = phys.QF* (CA2-CVF2) ; % CEO in fat tissue (mg) !
dx(22) = phys . QBR* (CA2-CVBR2) - BRGSH2; % CEO in brain tissue (mg) !
dx(23) = phys .QL* (CA2-CVL2) + phys . QST* (CVST2-CVL2 ) - KEH*CVL2*phys . VL + ...
(RAM1*MW2/MW) - RALG2 ;
dx(24) = phys.QC*CV2 - phys.QC*CVB2 - ps . KB2*x (24 ) ;
% CEO mixed venous after binding
%Calculation of the AUC for ACN and CEO in liver, brain, and blood!
C-27 DRAFT - DO NOT CITE OR QUOTE
-------
dx(25) = CL; % AUC for ACN liver concentration
dx(26) = CBR; % AUC for ACN brain cone.!
dx(27) = CBR2; % AUC for CEO brain cone.!
dx(28) = CVB; % AUC for ACN blood cone.!
dx(29) = CVB2; % AUC for CEO blood cone.!
dx(31) = STGSH2 + SGSH2 + RGSH2 + BRGSH2 + RALG2 + ps.KFB2*(x(17)+x(24));
% amount of CEO binding to GSH
dx(32) = (ps.KH*(x(2)+x(16))) + (ps.KH2*(x(17)+x(24)));
% had problems with state variables going < 0 at one point (stiff system);
% the following is a fix for this.
%dx=(dx.*(x>=0) ) + (abs(dx) .*(x<0) ) ;
C-28 DRAFT - DO NOT CITE OR QUOTE
-------
APPENDIX D. UNCERTAINTIES ASSOCIATED WITH CHEMICAL-SPECIFIC
PARAMETERS EMPLOYED IN THE PBPK MODEL FOR AN DOSIMETRY
IN HUMANS
PBPK models are computational tools used to predict chemical/drug disposition. Models
are comprised of three distinct types of information: physiological, physicochemical, and
biochemical. The physiological data are chemically independent and describe such parameters
as organ volumes and blood flows. Physicochemical parameters are chemical specific and
specify parameters, such as PCs or permeability. Biochemical parameters define the rates of
chemical transformation or binding. In Appendix D, the EPA evaluates uncertainties in the
chemical-specific parameters employed in the PBPK model used to predict AN dosimetry in
humans, which is adapted from that of Sweeney et al. (2003). Only a small number of
(metabolic) parameters were changed in the EPA's adaptation, which are noted below.
Otherwise, it should be understood that any discussion of the model of Sweeney et al. (2003)
applies to the one used in this assessment.
PCs
PCs describe the extent of distribution of chemical into the body, including target tissues
such as the brain and lungs. PCs employed in the model of Sweeney et al. (2003) were derived
from rat tissue:air PCs and human blood:air PC for AN or CEO in the following way:
Ptissue:blood(human) = Ptissue:air(rat)/ Pblood:air(human)
This equation assumes that the relative difference in PCs between tissues is the same in
rats as in humans. The systematic difference between the two is accounted for by the blood:air
PC; for humans, the experimentally determined value for AN (154) differed from the value for
the rat (512) by threefold. Due to the absence of experimental data, Sweeney et al. (2003) set the
human blood:air PC for CEO equal to the rat value (1,658).
It may be asked whether the estimates produced by the above equation are reasonable
considering the chemical characteristics of AN and CEO and the distribution of AN- and
CEO-soluble components in human tissue. To address these issues, the method of Poulin and
Theil (2002) was used to estimate the PtiSSue:biood(human). The method is based on the additive
solubilities of a chemical in lipid and water, and the relative distribution of these substances in
tissues. Sensitivity analysis conducted by Sweeney et al. (2003) identified the Pstomach:biood and
Pbrain:biood for AN as influential on AN-related dose metrics and the Pbrain:biood for CEO as
influential on CEO-related dose metrics. Table D-l provides estimates of PCs for AN and CEO.
Note that the critical PCs, the Pstomach:biood for AN and the Pbrain:biood for AN and CEO, as
predicted by the method of Poulin and Theil were within a factor of about 1.5 of the values
employed by Sweeney et al. (2003).
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Table D-l. Tissuerblood PCs
PCs for AN
Adipose tissue
Brain
Stomach
Liver
Muscle
Richly perfused
Compa
0.30
1.12
1.00
1.02
0.98
1.00
Expb
0.94
1.34
1.51
1.51
1.16
1.51
Exp:comp ratio
3.11
.19
.52
.48
.19
.51
PCs for CEO
Adipose tissue
Brain
Stomach
Liver
Muscle
Richly perfused
Compa
0.22
0.99
0.89
0.94
0.83
0.97
Expb
0.79
1.40
0.27
0.27
1.84
0.27
Exp:comp ratio
3.52
1.41
0.31
0.31
1.97
0.28
"Computationally derived by Poulin and Theil (2002).
bExperimentally obtained/reported by Sweeney et al. (2003).
To evaluate the influence of the PC in the model, the revised human PBPK model was
implemented with both sets of PCs (Table D-l, with Pbiood:air as measured for AN using human
blood and for CEO using rat blood). For three different exposure scenarios, the peak AN and
CEO did not change by more than 30% of initial value (Table D-2). This finding may be
explained by consistency between both estimates of PC and a general lack of sensitivity to the
PC (see Sweeney et al., 2003). Indeed, the only PCs with normalized sensitivity coefficients that
exceeded 0.2 were the Pst0mach:biood (AN), Pbraimbiood (AN), and the Pbraimbiood (CEO). Consistent
with these results, Table D-2 shows a greater impact of changing the PC in the brain than the
blood (an averaged effect of other PCs). All of the predicted concentration in brain tended to
decrease when the model was implemented with PC as per the method of Poulin and Theil
(2002).
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Table D-2. Impact of method of estimation of PCs on the peak AN and CEO
model predictions
Exposure
Inhalation
2 ppm, 8 h
Inhalation
0.4 ppm, 1 wk
Oral
6 pulses in 24 h,
0.2 mg/kg total
PC method3
S
PT
S
PT
S
PT
PeakAN(jig/L)
Blood
4.87
4.87
0.0%
0.975
0.975
0%
2.50
2.68
7.3%
Brain
11.4
9.56
-16.4%
2.29
1.91
-16.4%
3.11
2.77
-10.9%
Peak CEO (jig/L)
Blood
0.998
0.998
0.0%
0.200
0.200
0%
8.36
8.64
3.4%
Brain
1.23
0.863
-29.6%
0.245
0.173
-29.6%
10.14
7.44
-26.7%
aS = PCs as estimated by Sweeney et al. (2003) (listed in Appendix C); PT = PCs estimated using the method of
Poulin and Theil (2002). AN and CEO are from model simulations using other human parameters as listed in
Appendix C.
Parameters of metabolic clearance
Sweeney et al. (2003) developed a PBPK model of AN and CEO disposition in humans
based on human in vitro data and the rat model of Kedderis et al. (1996). Human in vivo
pharmacokinetic data were not available for model development; therefore, the authors proposed
a rodent-human parallelogram approach to consider uncertainties in the scaling of in vitro
metabolism data. Critical metabolic pathways included the oxidation of AN to CEO, the
conjugation of AN and CEO, and the hydrolysis of CEO.
With the exception of the latter, the scaling of in vitro rate constants for these pathways
was first investigated by Kedderis et al. (1996), who observed that the scaled rate constants (rat
model) derived from rat liver microsomes, did not match those constants fit from in vivo rat data.
Sweeney et al. (2003) employed both in vitro and in vivo data to derive an empirical correction
factor (CF) to account for uncertainties in scaling (rat). The scaling constant was assumed to be
constant across species; thus, the preexisting rat model, the empirical CF, and human in vitro
data were employed to specify a human PBPK model.
Because EPA altered a number of the rat metabolic parameters to accommodate the rate
constant for AN hydrolysis extrapolated from in vitro data and to obtain parameters in a way that
was computationally reproducible, while following the same approach as Sweeney et al. (2003)
for metabolic parameter extrapolation, the EPA obtained different values for humans. First, the
derivation of the human Vmaxc for the oxidation of AN to CEO in the following equations was
described. The ratio of rates determined by human in vitro data and rat in vitro data is
considered (Kedderis et al. 1993c), as well as the allometrically scaled in vivo rate constant for
oxidation of AN in the rat.
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VmaxC(human) VmaxC(rat) , \03/ \ / \ f
(mg/h/kg07) = (mg/h/kga7)x BWhuman V J[mgMSP/gliver]human x VLChuman x Vmax,human
in vivo in vivo ^ BWrat J ^ [mgMSP/gliver]rat J [ VLCrat J [vmax,human
in vitro
22.1 = 7.1 x 5.422 x 1.4225 x 0.6425 x 0.626
Next, the value (and calculations) of the empirical CF = 5.785 is shown. This value represents
what is not accounted for in the in vitro to in vivo scaling in the rat.
0.7\
Fin vivo, in vitro = (mglhlkg"-')
in vivo
1 On
0-7\ ., rf^y*flyJOD / *~i Hi/s\r~\ ,, y -, \/i /^ -, O\A/ 0.3
(mglhlkg^')x [mgMSP I g liver]rat x -^- x \/LCraf
/n v/fro
5.785 = 7.17(0.001164 x 40 x 1000 x 0.04 x 0.659)
The allometrically scaled in vivo rate constant for human pseudo first-order GSH
conjugation of AN and CEO was calculated in the same way (not shown).
The parallelogram method was applied differently to estimate the rate constant for
enzymatic hydrolysis by EH in humans. Because the EPA extrapolated the EH rate constant in
rats from in vitro to in vivo without use of an adjustment factor, the same was done for humans.
Kedderis and Batra (1993) determined a Vmax and Km EH-mediated hydrolysis of CEO using
liver microsome samples from six individual humans. The lowest estimated Km in the group was
600 uM, so the EPA chose to describe the metabolism as first-order since in vivo concentrations
are expected to stay well below that value, using the ratio of Vmax/Km. The ratio of Vmax/Km was
first calculated for each individual since Vmax and Km tend to be statistically correlated due to the
way they are estimated, and an average value for the ratio was then determined to be 7.02 x 10"6
L/minute/mg MP. The value of 56.9 mg MP/g liver from Lipscomb et al. (2003), the liver
fraction of 25.7 g/kg BW, and the standard value of 70 kg BW for a human can be applied. The
rate constant for a standard human is then kEH = (7.02 x 10"6 L/minute/mg EH) x (56.9 mg MP/g
liver) x (25.7 g liver/kg BW) x (70 kg BW) x (60 minutes/hour) = 43.1 L/hour, and assuming it
also scales as BW0'7 (because it represents Vmax/Km, with Km assumed constant), kEnc =
kEH - (0.70 kg)0'7 = 2.20 L/hour-kg0'7.
The parallelogram approach described by Sweeney et al. (2003) is provocative,
considering that human in vivo pharmacokinetic data are sparse for many compounds. What
special assurances are required in using this approach?
First, what is the basis of the empirical CF? It is the basis to account for: (1) artifacts of
microsomes that may bias the estimation of rate constants (as suggested by Kedderis et al., 1996)
or (2) decay in activity due to handling and storage. An analysis by Lipscomb et al. (1998)
compared the in vivo Vmax that would be obtained by directly extrapolating the in vitro
metabolism of trichloroethylene to the value obtained by fitting in vivo data and found that the
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in-vitro-derived value came out to about one half of an in-vivo-fitted value. Thus, it appears that
a CF is appropriate to account for either or both of these factors. Sweeney et al. (2003) applied
the same approach to the enzyme-catalyzed GSH conjugation (cytosolic in origin). One concern
that had existed is that Sweeney et al. (2003) had been forced to assume that the same factor
holds for distinct enzymes (P450 and EH) that are membrane bound, because EH activity in the
rat was taken to be zero in the original model of Kedderis et al. (1996), but that assumption is no
longer required since the EPA was able to extrapolate the EH-mediated activity without
adjustment. The EPA must still assume that the factors for P450 and GST hold across species,
but, given that the in-vitro enzyme preparations and measurements of activity were made in the
same way for both animals and humans, this seems a reasonable assumption. Nevertheless, it
does represent a source of uncertainty.
The human PBPK model (with the EPA's revised parameters) can be tested by comparing
the data of Jakubowski et al. (1987) with mass-balance calculations from the PBPK model
(Table D-3). Unfortunately, there are some significant discrepancies between the data of
Jakubowski et al. (1987) and these simulations. The respiratory retention (nominally in the lung)
of AN from subjects exposed through a face mask ranged from 44 to 58%, with an average of
51.8%. The PBPK model predicts 99% uptake of AN inhaled to the pulmonary region.
Recognizing that -30% of inhaled air only enters the conducting airways (so-called "dead
space"), the PBPK model effectively predicts a retention of 69%, which is considerably higher
than measured. The respiration rates of the subjects in the Jakubowski et al. (1987) experiment
ranged from 366 to 625 L/hour, with an average of 508 L/hour, which corresponds to an alveolar
ventilation rate of 356 L/hour, while the PBPK model utilized a rate of 300 L/hour (for a 70-kg
person), but increasing the respiration rate in the model by this amount (while holding all other
parameters constant) only decreases the alveolar uptake fraction from 98.7 to 98.5%. A possible
explanation is that AN is subject to a considerable "wash-in/wash-out" effect, where some of the
material inhaled deposits temporarily in the conducting airways and then desorbs on expiration,
which is not accounted for in the model.
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Table D-3. PBPK mass balance predictions for an 8-hour human exposure
to 2 ppm AN
Parameter
AI
AX
ATI
AMt
AMo
AMgt
AMgn
ABD
CMo
CX
CTi
CEH
CC1
Ccti
CBD
Metric
AN mass inhaled
AN mass exhaled
AN final tissue mass (including blood)
AN metabolized by ...
AN oxidized
AN enzymatically conjugated (in liver)
AN non-enzymatically conjugated
Metabolic mass balance
(AMo + AMgt + AMgn)/ AMt
AN bound to Hb and sulfhydrils
Total AN mass balance
(AX + ATi + AMt + ABD)/AI
CEO formed from AN oxidation3
CEO mass exhaled
CEO final tissue mass
CEO hydrolyzed
CEO conjugated in liver
CEO conjugated in other tissues
CEO bound to Hb and sulfhydrils
CEO mass balance
(Cti + CEH + CC1 + Ccti + CBD)/CMo
Amount (mg)
10.389
0.133
0.725
5.102
3.904
0.280
0.918
Percent
1.3
7.0
49.1
37.6
2.7
8.8
100.00%
4.429
42.6
100.00%
5.080
0.001
0.097
1.128
2.596
0.283
0.975
0.0
1.9
22.2
51.1
5.6
19.2
100.00%
aCMo = AMo X MWcEo -
The subjects of Jakubowski et al. (1987) were then exposed in a chamber for 8 hours to
an average of 10.8 or 5.6 mg/m3 (5 or 2.6 ppm), and the total excretion of CEO in the urine was
found to be an average of 26.4 or 16.3% of the retained dose (portion not exhaled) at those
respective exposure levels. The PBPK model predicts that 38% of the retained AN is oxidized to
CEO and that 22% of that CEO is hydrolyzed, with these fractions being fairly constant at
<8 ppm. Assuming that all the hydrolysis product was excreted in the urine, that the acid-
extraction method of Jakubowski et al. (1987) would have reversed the hydrolysis, and that the
extract! on/HPLC technique would have separated that product from the GSH conjugates and
other metabolites, the predicted urinary excretion of "CEO" would then be 8.4% of the retained
dose. Alternately, if it is assumed that all CEO conjugated with GSH was excreted in the urine
and that the method of Jakubowski et al. (1987) had cleaved the GSH conjugates, then the total
CEO predicted in the urine would be 79% of the CEO formed or 30% of the retained AN.
Therefore, the measured CEO excretion as a fraction of the retained dose is bracketed by the
predicted levels, depending on what is assumed about the assay method. These results indicate
that the model predictions of AN and CEO metabolism in humans subsequent to absorption are
at least in the right range.
Kedderis et al. (1996) suggest that the overestimation of CEO by the rat model at the
early time points (which also occurs in the EPA's revision) may be due to an intrahepatic first
pass effects, as occurs with other epoxides formed in situ from their parent olefms (Filser and
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Bolt, 1984). The introduction of such a structure ("privileged access" as coined by Kohn and
Melnick, 2000) would contribute to the efficient clearance of CEO; however, whether the needed
differential reduction of CEO levels in the early phase would also be accomplished is not clear.
The issue is that the model predicts a very rapid rise for CEO in blood to peak/plateau levels
(Figure 3-5a, top panels), while the data show a gradual rise over the first -10 minutes. It is not
evident that a privileged access model would account for this dynamic difference since the
current model allows for reaction of AN with GSH in stomach tissue and liver before reaching
the general circulation. Further, the same lack-of-fit occurs vs. the CEO i.v.-exposure data
shown in Figure 3-3, which could not be explained by a first-pass effect from GI absorption.
An alternate explanation is that the model currently assumes constant GSH levels, while
GSH may be depleted during the early part of the exposure. A model that includes this depletion
could predict that at very short times when GSH is near control levels, relatively little AN would
remain unconjugated for conversion to CEO, but, as the depletion occurs, more and more AN
would be available for oxidative conversion to CEO, hence the gradual rise. To accurately
calibrate parameters that accomplish intrahepatic clearance would require a description of GSH
kinetics.
Thus, the inability to fit the rat blood CEO levels over the entire time course and the lack
of understanding of the biological basis for those kinetics indicate a level of uncertainty in the
model that could be addressed through further research. While the model's overprediction of
AN absorption in humans indicates that the description of gas uptake could be improved, this
discrepancy is only 10-30% (less than the apparent 3- to 10-fold overprediction of blood and
brain CEO levels in rats and [likely] in humans by the oral route). The model does, however,
predict AN and CEO blood and tissue levels from inhalation exposure in rats at the lowest
concentration measured (186 ppm) quite well (Figure 3-4a), indicating that the gas-uptake
description is adequate for prediction of rat dosimetry. For oral exposures, however, the greater
accuracy of the AN predictions vs. CEO (Figure 3-5b vs. 3-5a) suggests greater certainty in the
use of AN, although the mode of action suggests that CEO is the active metabolite.
Sensitivity and Uncertainty Analysis
Sensitivity and uncertainty analyses were performed to investigate the dependence of
PBPK model predictions for AN and CEO on specific model parameters and to estimate the
overall uncertainty in those predictions, given estimates of uncertainty in specific model
parameters.
The first step was to perform a sensitivity analysis on PBPK model predictions to identify
the degree to which model dose metric predictions depend on model parameters. For this
purpose, normalized sensitivity coefficients were calculated. For example, for the sensitivity of a
model prediction of concentration (e.g., AN concentration in blood) to parameter P (e.g.,
metabolic Vmaxc for AN oxidation), the normalized sensitivity coefficient is:
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S = (dC/dP) x (P/C)
The first term is the derivative of C with respect to P. (C is generally a function of time, so this
derivative will also depend on time.) This derivative is approximated by alternately increasing
and decreasing P by a small fraction, AP (taken to be 1% of P here) and calculating [C(P + AP) -
C(P - AP)] + (2 x AP). The second term normalizes the sensitivity with respect to the value of C
at the normal value of P and with respect to P, essentially yielding the percent change in C given
a 1% change in P. These sensitivities typically vary between ±1, where values near 1 show high
sensitivity and values near 0 show low sensitivity.
Sensitivities were determined for four dose metrics, the blood and brain concentrations of
AN and CEO, under an inhalation and an oral exposure scenario for the human PBPK model.
For simplicity, the EPA considered continuous exposures to either 0.4 ppm AN by inhalation or
an oral absorption rate equal to 1 mg/kg-day and analyzed the values of these metrics at
24 hours, by which time, the human body is predicted to reach steady state. Under such
conditions, the AUC for each metric is just 24 hours times the steady-state concentration
(subsequent to reaching steady state), and hence, the sensitivity of the AUC is identical to the
sensitivity of the concentration, given that the coefficients are normalized as described above.
Since these concentrations are in the linear range of the model, they apply across the low-dose,
linear range of concentrations. The sensitivity coefficients for each parameter for which at least
one coefficient had absolute value greater than 0.1 are listed in Tables D-4 and D-5 for the
inhalation and oral scenarios, respectively.
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Table D-4. Sensitivity of AN and CEO metrics for continuous inhalation
exposure
Parameter
Qcc
QPC
VLC
VBRC
Vsc
VABC
VVBC
QLC
QBRC
Qsc
GSHL
GSHBR
GSHS
Vmaxc
Km
kFC2
ksHC
kBC
kBC2
PER
?BR2
Dose metrics" (concentrations)
AN in blood
-0.49
0.99
-0.01
-0.01
-0.14
-0.07
-0.12
-0.76
0.00
-0.02
0.00
-0.01
-0.13
-0.09
0.09
0.00
0.00
-0.19
0.00
0.00
0.00
AN in brain
-0.57
0.99
-0.01
-0.04
-0.10
-0.07
-0.08
-0.53
0.03
-0.01
0.00
-0.04
-0.09
-0.06
0.06
0.00
0.00
-0.16
0.00
0.97
0.00
CEO in blood
0.51
0.99
-0.64
-0.04
-0.28
-0.12
-0.20
0.65
0.00
-0.10
0.27
-0.04
-0.27
0.11
-0.11
-0.86
-0.25
-0.16
-0.13
0.00
0.00
CEO in brain
0.64
0.98
-0.64
-0.14
-0.28
-0.15
-0.20
0.65
0.10
-0.10
0.38
-0.14
-0.27
0.11
-0.11
-0.96
-0.25
-0.16
-0.15
0.00
1.00
aValues are normalized sensitivity coefficients (i.e., (dC/dP) x (P/C), for sensitivity of concentration C to parameter
P) for steady state blood and brain concentrations predicted for humans during an exposure to 0.4 ppm AN.
Sensitivities are approximated by increasing and decreasing each parameter by 1% of its default value. Values of
0.2 or greater are indicated in bold and values between 0.1 and 0.2 in italics. Parameters for which all coefficients
are less than 0.1 are omitted.
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Table D-5. Sensitivity of AN and CEO metrics for continuous oral exposure
Parameter
BW
Qcc
VLC
VBRC
Vsc
VVBC
QLC
GSHL
GSHBR
GSHS
Vmaxc
Km
kFC2
kEHC
kBc
kBC2
PER
PBR2
Dose metrics" (concentrations)
AN in blood
0.21
0.30
-0.07
-0.01
-0.10
-0.12
0.29
0.00
-0.01
-0.09
-0.89
0.89
0.00
0.00
-0.17
0.00
0.00
0.00
AN in brain
0.20
0.36
-0.07
-0.04
-0.10
-0.12
0.29
0.00
-0.04
-0.09
-0.89
0.89
0.00
0.00
-0.19
0.00
0.97
0.00
CEO in blood
0.24
0.27
-0.64
-0.03
-0.19
-0.13
0.35
0.27
-0.03
-0.18
0.11
-0.11
-0.86
-0.25
-0.01
-0.13
0.00
0.00
CEO in brain
0.24
0.39
-0.64
-0.13
-0.19
-0.13
0.35
0.38
-0.13
-0.18
0.11
-0.11
-0.96
-0.25
-0.01
-0.15
0.00
1.00
aValues are normalized sensitivity coefficients (i.e., (dC/dP) x (P/C), for sensitivity of concentration C to parameter
P), for steady state blood and brain concentrations predicted for humans during a continuous oral infusion equal to
1 mg/kg-d. Sensitivities are approximated by increasing and decreasing each parameter by 1% of its default value.
Values of 0.2 or greater are indicated in bold, and values between 0.1 and 0.2 in italics. Parameters for which all
coefficients are less than 0.1 are omitted.
The EPA's results can be compared to those of Sweeney et al. (2003), derived for the
original model, with previous parameter values. For the inhalation scenario (Table D-4), the
EPA's results for AN are essentially identical to those of Sweeney et al. (2003), and those for
CEO show the same qualitative pattern and only have a few notable differences. The sensitivity
to the rate constant for CEO hydrolysis (kEnc in the EPA's model; Vmaxc2 in Sweeney et al.,
2003) is -0.25 in the revised model and -0.5 in their previous model, indicating somewhat less
importance for this parameter now. On the other hand, the sensitivity to CEO-GSH conjugation
in the liver parameterized by kpc2 increased from -0.6-0.7 for Sweeney et al. (2003) to -0.9 with
the revised model, showing more significance now. Thus, there is a shift in the relative
importance of these two parameters and the rate constant for GSH conjugation in humans is an
especially important factor in the EPA's predictions, while the exact rate of CEO hydrolysis is
less influential, though still somewhat important.
Important factors in comparing the EPA's results for an oral exposure to those of
Sweeney et al. (2003) are that the EPA held the dose in mg/kg-day constant, so that it varied in
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proportion to BW when that parameter was varied, while Sweeney et al. (2003) held the dose in
mg/day constant, and they calculated sensitivities of the AUC and peak concentrations for only
the AN metrics, while the EPA used steady-state concentrations of AN and CEO. Thus, it is not
surprising that Sweeney et al. (2003) had a negative sensitivity to BW, since the same total dose
given to a larger body would result in a lower concentration, while the EPA's is positive,
reflecting that a proportionately higher dose is expected to yield higher-than-proportional
concentrations because the rates of metabolic elimination scale as BW07.
It is interesting to note that the liver volume (fraction), which effects the rates of first-
and second-order reactions in the model, and the rate constants for CEO removal but not CEO
production (Vmaxc) significantly affect the CEO metrics (absolute values from Sweeney et al.
(2003) are about 0.17-0.18 for sensitivity of CEO metrics to Vmax and Km). The EPA's lack of
significant dependence on blood flow rates to stomach (Qstc) and slowly-perfused tissues (Qsc)
is due to the fact that the EPA is analyzing a continuous infusion, as these do significantly affect
the peak concentration after a bolus exposure. Thus, these are only important for specific dosing
scenarios.
Likewise the rate of absorption (kA) from the stomach significantly affects the peak
concentration (verified but not shown for the revised model) but has a smaller affect on the
AUC. These results for absorption are not surprising, since changes in kA do not alter the fact
that the entirety of an oral bolus is predicted to be absorbed. Whether or not there is an impact
on the steady-state concentrations depends on what one believes about oral absorption. In
particular, if one assumes 100% absorption, where the rate of absorption increases with the
amount in the GI tract, then at steady state, the rate of absorption must equal the rate of ingestion
and the exact rate of absorption is unimportant. It is only if orally ingested material may be
eliminated in the feces or otherwise transformed in the gut without absorption that the exact rate
constant for absorption would be important.
Finally, uncertainty in these steady-state model predictions was estimated using the
equation of Sweeney et al. (2003) for the approximate coefficient of variation (CV):
CF_ =
where CVm is the estimated CV for metric "m," Sm; is the sensitivity coefficient of metric "m" to
parameter p;, as tabulated above, and CV; is the CV of p;. For this calculation, the EPA decided
to use the individual parameter CV values as estimated and used by Sweeney et al. (2003, see
Table 3-5 in that paper). The CVs for physiological parameters (e.g., blood flow rates, tissue
fractions) would indeed be identical, and, while the EPA reestimated some metabolic parameters
and kA, the underlying in vitro data on which the human extrapolation is based are the same, and
hence, they are expected to be approximately the same.
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The resulting CV values for the EPA's four metrics under the two exposure scenarios are
listed in Table D-6.
Table D-6. Estimated CVs for AN and CEO metrics
Exposure scenario
0.4 ppm inhalation
1 mg/kg-d oral
Dose metrics (concentrations)
AN in blood
0.620
0.790
AN in brain
0.656
0.873
CEO in blood
0.893
0.711
CEO in brain
1.15
1.01
The overall CVs in Table D-6 are interesting in themselves, in that they indicate the
overall level of confidence in the model's prediction of those dose metrics. But even more
interesting is the contribution of the individual parameters to each of these metrics, as listed in
Tables D-7 and D-8. What those contributions show is that the vast majority of the uncertainty
arises from a small number of parameters and that most of those parameters are physiological
values—alveolar ventilation (Qcc), cardiac output (Qpc), and blood flow to the liver (VLC)—and
the brain:blood PC in the case of the brain metrics.
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Table D-7. Parameter contributions to overall CVs for inhalation exposure
Parameter
Qcc
QPC
VLC
VBRC
Vsc
VABC
VVBC
QLC
QBRC
Qsc
GSHL
GSHBR
GSHS
Vmaxc
Km
kFC2
kEHC
kBc
kBC2
PER
?BR2
Dose metrics" (concentrations)
AN in blood
9.3
37.9
0.0
0.0
0.7
0.4
1.1
44.4
0.0
0.0
0.0
0.0
0.9
1.1
0.4
0.0
0.0
2.9
0.0
0.0
0.0
AN in brain
11.5
33.9
0.0
0.1
0.3
0.4
0.5
19.8
0.1
0.0
0.0
0.2
0.4
0.5
0.2
0.0
0.0
1.8
0.0
30.0
0.0
CEO in blood
5.0
18.2
7.7
0.1
1.4
0.6
1.5
15.9
0.0
0.4
1.1
0.1
1.8
0.7
0.3
39.2
3.4
1.0
0.7
0.0
0.0
CEO in brain
4.6
11.0
4.6
0.5
0.9
0.5
0.9
9.6
0.2
0.2
1.3
0.7
1.1
0.4
0.2
29.8
2.1
0.6
0.5
0.0
29.9
"Values are 100 x Smk x CVk/E(Smi x CV,). Values >5 (i.e., 5% contribution) are in bold.
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Table D-8. Parameter contributions to overall CVs for oral exposure
Parameter
BW
Qcc
VLC
VBRC
Vsc
VVBC
QLC
GSHL
GSHBR
GSHS
Vmaxc
Km
kFC2
kEHC
kBc
kBC2
PER
PBR2
Dose metrics" (concentrations)
AN in blood
1.1
2.2
0.1
0.0
0.2
0.7
4.1
0.0
0.0
0.3
66.1
23.1
0.0
0.0
1.4
0.0
0.0
0.0
AN in brain
0.8
2.5
0.1
0.1
0.2
0.5
3.4
0.0
0.1
0.2
54.1
18.9
0.0
0.0
1.5
0.0
16.9
0.0
CEO in blood
1.7
2.1
12.1
0.1
1.0
0.9
7.1
1.8
0.1
1.3
1.2
0.4
61.9
5.4
0.0
1.0
0.0
0.0
CEO in brain
0.8
2.3
6.0
0.5
0.5
0.5
3.5
1.7
0.8
0.6
0.6
0.2
38.6
2.7
0.0
0.7
0.0
38.7
"Values are 100 x Smk2 x CVk/E(Smi2 x CV;). Values >5 (i.e., 5% contribution) are in bold.
Since the PCs are expected to be similar in rats and humans, and the extrapolation really
depends on the rathuman ratio of those, the true resulting uncertainty is likely to be very small.
The nominal uncertainty in cardiac ventilation and blood flows is expected to be a composite of
uncertainty and inter-individual variability. And the only metabolic parameter to contribute
significantly to the uncertainty for inhalation exposure is the CEO-GSH conjugation rate. Thus,
a significant improvement in the overall uncertainty can be gained by further investigation of
only a small number of parameters, half of which are applicable to every PBPK model that might
be considered for gases.
The distribution of parameter influence for the oral simulation scenario, as shown in
Table D-8, is similar to that for inhalation in that most of the influence is distributed among a
few parameters, with the brain:blood PCs strongly influencing brain concentrations. Also similar
is that the CEO-GSH conjugation rate (kpc2) is a significant factor for the CEO metrics. Unlike
the inhalation case, alveolar ventilation (Qpc) has negligible influence, which is to be expected,
but also cardiac output (Qcc) has only small influence. This later case occurs because of the
strong first-pass effect for AN as it is absorbed through the liver, which also explains the very
high influence of the oxidation parameters, Vmaxc and Km, on the AN metrics.
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These results for oral exposure are quite similar to the results obtained by Sweeney et
al. (2003), except that their results indicate somewhat higher influence by blood flow to the liver
(i.e., 8-12% on AN metrics), liver GSH (6%), and a higher influence of CEO hydrolysis (ksHC
here, parameterized by a Vmax and Km with influence of 12-30%) on CEO metrics. The shift of
influence from hydrolysis to GSH conjugation for CEO metrics between Sweeney et al. (2003)
and the EPA's results arises from the shift in the relative rates through those two pathways, but in
both, the rate of CEO removal is a strong determinant of the CEO metric.
D-15 DRAFT - DO NOT CITE OR QUOTE
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