QRAFT EPA/600/8-90/057E
DO NOT CITE OR QUOTE July 20.00
SAB Review Draft
Health Assessment Document
for Diesel Exhaust
NOTICE
THIS DOCUMENT IS A DRAFT. It has not been formally released by the
Environmental Protection Agency and should not at this stage be construed to
represent Agency policy. It is being circulated for comment on its technical
accuracy and policy implications.
National Center for Environmental Assessment-Washington Office
Office of Research and Development
U.S. Environmental Protection Agency
Washington, DC
-------
DISCLAIMER
This document is a draft for review purposes only and does not constitute U.S.
Environmental Protection Agency policy. Mention of trade names or commercial products does
not constitute endorsement or recommendation for use.
07/25/00 ii DRAFT—DO NOT CITE OR QUOTE
-------
CONTENTS
LIST OF TABLES xii
LIST OF FIGURES xv
PREFACE xviii
AUTHORS AND CONTRIBUTORS xix
ACKNOWLEDGMENTS xxiii
1. EXECUTIVE SUMMARY 1-1
1.1. INTRODUCTION 1-1
1.2. COMPOSITION OF DIESEL EXHAUST 1-1
1.3. DIESEL EMISSIONS 1-2
1.4. ATMOSPHERIC TRANSFORMATION OF DIESEL EXHAUST 1-2
1.5. EXPOSURE TO DIESEL EXHAUST 1-3
1.6. HEALTH EFFECTS 1-3
1.6.1. Acute Effects 1-3
1.6.2. Chronic Noncancer Respiratory Effects 1-4
1.6.3. Carcinogenic Effects 1-4
1.7. SOURCES OF UNCERTAINTIES 1-5
2. DIESEL EMISSIONS CHARACTERIZATION, ATMOSPHERIC
TRANSFORMATION, AND EXPOSURES 2-1
2.1. INTRODUCTION 2-1
2.2. PRIMARY DIESEL EMISSIONS 2-4
2.2.1. History of Dieselization 2-4
2.2.1.1. Dieselization of the On-Road Fleet 2-4
2.2.1.2. Dieselization of Railroad Locomotive Engines 2-5
2.2.2. Diesel Combustion and Formation of Primary Emissions 2-6
2.2.3. Diesel Emission Standards and Emission Trends Inventory 2-10
2.2.4. Historical Trends in Diesel Fuel Use and Impact of Fuel
Properties on Emissions 2-13
2.2.5. Chronological Assessment of Emission Factors 2-16
2.2.5.1. On-Road Vehicles 2-16
2.2.5.2. Locomotives 2-22
2.2.6. Engine Technology Description and Chronology 2-22
2.2.6.1. Indirect and Direct Injection High-Speed
Diesel Engines 2-23
2.2.6.2. Injection Rate 2-25
2.2.6.3. Turbocharging, Charge-Air Cooling, and
Electronic Controls 2-26
2.2.6.4. Two-Stroke and Four-Stroke High-Speed
Diesel Engines 2-28
2.2.7. Air Toxic Emissions 2-29
2.2.7.1. Aldehyde Emissions 2-29
2.2.7.2. Dioxin and Furans 2-30
07/25/00 iii DRAFT—DO NOT CITE OR QUOTE
-------
CONTENTS (Continued)
2.2.8. Physical and Chemical Composition of Diesel
Exhaust Particles 2-32
2.2.8.1. Organic and Elemental Carbon Content of Particles 2-35
2.2.8.2. PAHs and Nitro-PAH Emissions 2-39
2.2.8.3. Particle Size 2-43
2.3. ATMOSPHERIC TRANSFORMATION OF DIESEL EXHAUST 2-45
2.3.1. Gas-Phase Diesel Exhaust 2-46
2.3.1.1. Organic Compounds 2-46
2.3.1.2. Inorganic Compounds 2-47
2.3.1.3. Atmospheric Transport of Gas-Phase Diesel Exhaust 2-48
2.3.2. Particle-Phase Diesel Exhaust 2-48
2.3.2.1. Particle-Associated PAH Photooxidation 2-49
2.3.2.2. Particle-Associated PAH Nitration 2-50
2.3.2.3. Particle-Associated PAH Ozonolysis 2-50
2.3.2.4. Atmospheric Transport of Diesel Exhaust
Paniculate Matter '. 2-51
2.3.3. Diesel Exhaust Aging • 2-51
2.4. AMBIENT DIESEL EXHAUST CONCENTRATIONS AND
EXPOSURES 2-52
2.4.1. Diesel Exhaust Gases in the Ambient Atmosphere 2-52
2.4.2. Ambient Concentrations of Diesel Particulate Matter 2-53
2.4.2.1. Source Apportionment Studies 2-54
2.4.2.2. EC Surrogate for Diesel Particulate Matter2- 2-57
2.4.2.3. Dispersion Modeling Results 2-60
2.4.3. Exposures to Diesel Exhaust 2-61
2.4.3.1. Occupational Exposure to Diesel Exhaust 2-61
2.4.3.2. Ambient Exposure to Diesel Exhaust 2-64
2.5. SUMMARY 2-69
2.5.1. History of Diesel Engine Use, Standards, and Technology 2-69
2.5.2. Physical and Chemical Composition of Diesel Exhaust 2-71
2.5.3. Atmospheric Transformation of Diesel Exhaust 2-74
2.5.4. Ambient Concentrations and Exposure to Diesel Exhaust 2-75
2.6. REFERENCES 2-139
3. DOSIMETRY OF DIESEL PARTICULATE MATTER 3-1
3.1. INTRODUCTION 3-1
3.2. CHARACTERISTICS OF INHALED DPM 3-2
3.3. REGIONAL DEPOSITION OF INHALED DPM 3-2
3.3.1. Deposition Mechanisms 3-3
3.3.1.1. Biological Factors Modifying Deposition . 3-4
3.3.2. Particle Clearance and Translocation Mechanisms ../::. 3-6
3.3.2.1. Extrathoracic Region ....-;. 3-6
07/25/00 iv DRAFT—DO NOT CITE OR QUOTE
-------
CONTENTS (Continued)
3.3.2.2. Tracheobronchial Region 3-7
3.3.2.3. A Region 3-9
3.3.3. Translocations of Particles to Extra-Alveolar Macrophage
Compartment Sites 3-15
3.3.3.1. Clearance Kinetics 3-17
3.3.3.2. Interspecies Patterns of Clearance 3-17
3.3.3.3. Clearance Modifying Factors and Susceptible
Populations 3-18
3.3.3.4. Respiratory Tract Disease 3-18
3.4. PARTICLE OVERLOAD 3-18
3.4.1. Introduction 3-18
3.4.2. Relevance to Humans 3-20
3.4.3. Potential Mechanisms for an AM Sequestration Compartment
for Particles During Particle Overload 3-21
3.5. BIOAVAILABILITY OF ORGANIC CONSTITUENTS
PRESENT ON DIESEL EXHAUST PARTICLES 3-23
3.5.1. In Vivo Studies 3-23
3.5.1.1. Laboratory Investigations 3-23
3.5.1.2. Studies in Occupationally Exposed Humans 3-24
3.5.2. In Vitro Studies 3-25
3.5.2.1. Extraction of Diesel Particle-Associated Organics
by Biological Fluids 3-25
3.5.2.2. Extraction of DPM-Associated Organics by Lung Cells
and Cellular Components 3-26
3.5.3. Modeling Studies 3-27
3.5.4. Bioavailability/Deposition of Organics 3-29
3.6. MODELING THE DEPOSITION AND CLEARANCE OF
PARTICLES IN THE RESPIRATORY TRACT 3-29
3.6.1. Introduction 3-29
3.6.2. Dosimetry Models for DPM 3-30
3.6.2.1. Introduction 3-30
3.6.2.2. Human Models 3-31
3.6.2.3. Animal Models 3-32
3.6.2.4. Combined Models (for Interspecies Extrapolation) 3-34
3.6.2.5. Use of the Yu et al. (1991) Model for Interspecies
Extrapolation 3-39
3.7. SUMMARY 3-40
3.8. REFERENCES 3-50
4. MUTAGENICITY 4-1
. - 4.1. GENE MUTATIONS 4-1
- 4.2. CHROMOSOME EFFECTS 4-4
4.3. OTHER GENOTOXIC EFFECTS 4-6
07/25/00 v DRAFT—DO NOT CITE OR QUOTE
-------
CONTENTS (Continued)
4.4. SUMMARY 4-7
4.5. REFERENCES 4-8
5. NONCANCER HEALTH EFFECTS OF DIESEL EXHAUST 5-1
5.1. HEALTH EFFECTS OF WHOLE DIESEL EXHAUST 5-1
5.1.1. Human Studies 5-1
5.1.1.1. Short-Term Exposures 5-1
5.1.1.2. Long-Term Exposures 5-12
5.1.2. Laboratory Animal Studies 5-16
5.1.2.1. Acute Exposures 5-17
5.1.2.2. Short-Term and Subchronic Exposures 5-18
5.1.2.3. Chronic Exposures 5-21
5.2. MODE OF ACTION OF DIESEL EMISSIONS-INDUCED
NONCANCER EFFECTS 5-53
5.2.1. Comparison of Health Effects of Filtered and
Unfiltered Diesel Exhaust 5-53
5.2.2. Mode of Action for the Noncarcinogenic Effects of DPM 5-56
5.3. INTERACTIVE EFFECTS OF DIESEL EXHAUST 5-57
5.4. COMPARATIVE RESPONSIVENESS AMONG SPECIES TO THE
HISTOPATHOLOGIC EFFECTS OF DIESEL EXHAUST 5-59
5.5. DOSE-RATE AND PARTICULATE CAUSATIVE ISSUES 5-60
5.6. SUMMARY AND DISCUSSION 5-64
5.6.1. Effects of Diesel Exhaust on Humans 5-64
5.6.2. Effects of Diesel Exhaust on Laboratory Animals 5-66
5.6.2.1. Effects on Survival and Growth 5-67
5.6.2.2. Effects on Pulmonary Function 5-67
5.6.2.3. Histopathological and Histochemical Effects 5-68
5.6.2.4. Effects on Airway Clearance 5-68
5.6.2.5. Neurological and Behavioral Effects 5-69
5.6.2.6. Effects on Immunity and Allergenicity 5-69
5.6.2.7. Other Noncancer Effects 5-69
5.6.3. Comparison of Filtered and Unfiltered Diesel Exhaust 5-69
5.6.4. Interactive Effects of Diesel Exhaust 5 70
565. Ccnclus'r>r|s 5-70
5.7. REFERENCES 5-100
6. QUANTITATIVE APPROACHES TO ESTIMATING HUMAN
NONCANCER HEALTH RISKS OF DIESEL EXHAUST 6-1
6.1. INTRODUCTION 6-1
6.2. DEVELOPMENT OF THE PM25 NAAQS 6-2
6.3. DPM AND THE PM25 NAAQS 6-6
6.4. THE INHALA11UN REFERENCE CONCENTRATION APPROACH 6-7
07/25/00 vi DRAFT—DO NOT CITE OR QUOTE
-------
CONTENTS (Continued)
6.5. CHRONIC REFERENCE CONCENTRATION FOR
DIESEL EXHAUST 6-9
6.5.1. Principal Studies for Dose-Response Analysis: Chronic,
• Multiple-Dose Level Rat Studies 6-10
6.5.2. HEC Derivation 6-11
6.5.3. Consideration of Uncertainty Factors for the RfC 6-14
6.5.4. Derivation of the RfC for DE 6-15
6.6. CHARACTERIZATION OF THE NONCANCER
ASSESSMENT FOR DE 6-15
6.7. SUMMARY 6-17
6.8. REFERENCES 6-22
7. CARCINOGENICITY OF DIESEL EXHAUST 7-1
7.1. INTRODUCTION 7-1
7.1.1. Overview 7-1
7.1.2. Ambient PM-Lung Cancer Relationships 7-2
7.2. EPIDEMIOLOGIC STUDIES OF THE CARCINOGENICITY
OF EXPOSURE TO DIESEL EXHAUST 7-4
7.2.1. Cohort Studies 7-5
7.2.1.1. Waller (1981): Trends in Lung Cancer in London
in Relation to Exposure to Diesel Fumes 7-5
7.2.1.2. Howe et al. (1983): Cancer Mortality (1965 to 1977)
in Relation to Diesel Fumes and Coal Exposure
in a Cohort of Retired Railroad Workers 7-7
7.2.1.3. Rushtonetal. (1983): Epidemiological Survey
of Maintenance Workers in the London Transport
Executive Bus Garages and Chiswick Works 7-9
7.2.1.4. Wong et al. (1985): Mortality Among Members of a
Heavy Construction Equipment Operators Union
With Potential Exposure to Diesel Exhaust Emissions 7-10
7.2.1.5. Edling et al. (1987): Mortality Among Personnel
Exposed to Diesel Exhaust 7-13
7.2.1.6. Boffetta and Stellman (1988): Diesel Exhaust
Exposure and Mortality Among Males in the
American Cancer Society Prospective Study 7-14
7.2.1.7. Garshick et al. (1988): A Retrospective
Cohort Study of Lung Cancer and Diesel
Exhaust Exposure in Railroad Workers 7-16
7.2.1.8. Gustavsson et al. (1990): Lung Cancer and
Exposure to Diesel Exhaust Among Bus
Garage Workers 7-20
07/25/00 vii DRAFT—DO NOT CITE OR QUOTE
-------
CONTENTS (Continued)
7.2.1.9. Hansen(1993): A Followup Study on the
Mortality of Truck Drivers 7-21
7.2.1.10. Saverinetal. (1999): Diesel Exhaust and Lung
Cancer Mortality in Potash Mining 7-23
7.2.2. Case-Control Studies of Lung Cancer 7-24
7.2.2.1. Hall and Wynder (1984): A Case-Control Study
of Diesel Exhaust Exposure and Lung Cancer 7-24
7.2.2.2. Damber and Larsson( 1987): Occupation and
Male Lung Cancer, a Case-Control Study in
Northern Sweden 7-26
7.2.2.3. Lerchen et al. (1987): Lung Cancer and
Occupation in New Mexico 7-27
7.2.2.4. Garshick et al. (1987): A Case-Control Study
of Lung Cancer and Diesel Exhaust Exposure
in Railroad Workers 7-29
7.2.2.5. Benhamou et al. (1988): Occupational Risk
Factors of Lung Cancer in a French
Case-Control Study 7-32
7.2.2.6. Hayes etal. (1989): Lung Cancer in Motor
Exhaust-Related Occupations 7-33
7.2.2.7. Steenland etal. (1990): A Case-Control Study
of Lung Cancer and Truck Driving in the
Teamsters Union 7-35
7.2.2.8. Steenland etal. (1998): Diesel Exhaust
and Lung Cancer in the Trucking Industry:
Exposure-Response Analyses and Risk Assessment 7-37
7.2.2.9. Boffetta etal. (1990): Case-Control Study
on Occupational Exposure to Diesel Exhaust
and Lung Cancer Risk 7-39
7.2.2.10. Emmelin et al. (1993): Diesel Exhaust Exposure
and Smoking: A Case-Referent Study of Lung
Cancer Among Swedish Dock Workers 7-40
7.2.2.11. Swanson etal. (1993): Diversity in the
Accnriatinn Retween Occupation gnrl T.nno
.. — — — ~ . ^.... _ ._^
Cancer Among Black and White Men 7-42
7.2.2.12. Hansen etal. (1998): Increased Risk of
Lung-Cancer Among.Different Types.of
Professional Drivers in Denmark 7-43
7.2.2.13. Briiske-Hohlfeld etal. (1999): Lung Cancer Risk
in Male Workers Occupationally Exposed to
Diesel Motor Emissions in Germany 7-44
07/25/00 viii DRAFT—DO NOT CITE OR QUOTE
-------
CONTENTS (Continued)
7.2.3. Summaries of Studies and Meta-Analyses of Lung Cancer 7-47
7.2.3.1. Cohen and Higgins (1995): Health Effects of
Diesel Exhaust: Epidemiology 7-47
7.2.3.2. Bhatiaetal. (1998): Diesel Exhaust
Exposure and Lung Cancer 7-48
7.2.3.3. Lipsett and Campleman( 1999): Occupational
Exposure to Diesel Exhaust and Lung Cancer:
A Meta-Analysis 7-50
7.2.4. Summary and Discussion 7-51
7.2.4.1. Summary of the Cohort Mortality Studies 7-53
7.2.4.2. Summary of the Case-Control Studies of Lung Cancer 7-56
7.2.4.3. Summary of the Reviews and Meta-Analyses of
Lung Cancer 7-61
7.2.4.4. Discussion of Relevant Methodologic Issues 7-62
7.2.4.5. Evaluation of Causal Association 7-65
7.3. CARCINOGENICITY OF DIESEL EMISSIONS IN
LABORATORY ANIMALS 7-68
7.3.1. Inhalation Studies (Whole Diesel Exhaust) 7-69
7.3.1.1. Rat Studies 7-69
7.3.1.2. Mouse Studies 7-77
7.3.1.3. Hamster Studies 7-80
7.3.1.4. Monkey Studies 7-81
7.3.2. Inhalation Studies (Filtered Diesel Exhaust) 7-82
7.3.3. Inhalation Studies (Diesel Exhaust Plus Cocarcinogens) 7-82
7.3.4. Lung Implantation or Intratracheal Instillation Studies 7-84
7.3.4.1. Rat Studies 7-84
7.3.4.2. Syrian Hamster Studies 7-87
7.3.4.3. Mouse Studies 7-88
7.3.5. Subcutaneous and Intraperitoneal Injection Studies 7-88
7.3.5.1. Mouse Studies 7-88
7.3.6. Dermal Studies 7-90
7.3.6.1. Mouse Studies 7-90
7.3.7. Summary and Conclusions of Laboratory Animal
Carcinogenicity Studies 7-93
7.4. MODE OF ACTION OF DIESEL EMISSION-INDUCED
CARCINOGENESIS 7-97
7.4.1. Potential Role of Organic Exhaust Components
in Lung Cancer Induction 7-98
7.4.2. Role of Inflammatory Cytokines and Proteolytic Enzymes
in the Induction of Lung Cancer in Rats by Diesel Exhaust 7-101
7.4.3. Role of Reactive Oxygen Species in Lung Cancer
Induction by Diesel Exhaust 7-102
07/25/00 ix DRAFT—DO NOT CITE OR QUOTE
-------
CONTENTS (Continued)
7.4.4. Relationship of Physical Characteristics of Particles to
Cancer Induction 7-105
7.4.5. Integrative Hypothesis for Diesel-Induced Lung Cancer 7-106
7.4.6. Summary 7-107
7.5. WEIGHT-OF-EVIDENCE EVALUATION FOR POTENTIAL
HUMAN CARCINOGENICITY 7-108
7.5.1. Human Evidence 7-109
7.5.2. Animal Evidence 7-110
7.5.3. Other Key Data 7-111
7.5.4. Mode of Action 7-112
7.5.5. Characterization of Overall Weight of Evidence:
EPA's 1986 Carcinogen Risk Assessment Guidelines 7-112
7.5.6. Weight-of-Evidence Hazard Narrative: EPA's
Proposed Revised Carcinogen Risk Assessment
Guidelines (1996, 1999) 7-113
7.6. EVALUATIONS BY OTHER ORGANIZATIONS 7-114
7.7. CONCLUSION 7-114
7.8. REFERENCES 7-143
8. DOSE-RESPONSE ASSESSMENT: CARCINOGENIC EFFECTS 8-1
8.1. INTRODUCTION 8-1
8.2. MODE OF ACTION AND DOSE-RESPONSE APPROACH 8-1
8.3. USE OF EPIDEMIOLOGIC STUDIES FOR QUANTITATIVE
RISK ASSESSMENT 8-3
8.3.1. Sources of Uncertainty 8-3
8.3.2. Evaluation of Key Epidemiologic Studies for
Potential Use in Quantitative Risk Estimates 8-5
8.3.2.1. Railroad Worker Studies 8-5
8.3.2.2. Teamsters Union Trucking Industry Studies 8-8
8.3.3. Conclusion 8-11
8.4. PERSPECTIVES ON CANCER RISK 8-11
8.5. SUMMARY 8-14
8.6. REFERENCES 8-17
9. CHARACTERIZATION OF POTENTIAL HUMAN HEALTH EFFECTS
OF DIESEL EXHAUST: HAZARD AND DOSE-RESPONSE ASSESSMENTS 9-1
9.1. INTRODUCTION .. ... 9-1
9.2. PHYSICAL AND CHEMICAL COMPOSITION OF DIESEL EXHAUST 9-1
9.2.1. Diesel Exhaust Components of Possible Health Concern 9-2
9 2 i .2 Organic Ccnvounds 9-3 •
9.2.2. "Fresh" Versus "Aged" Diesel Exhaust 9-4-
9.2.3. Changes of DE Emissions and Composition Over Time 9-4
07/25/00 x DRAFT—DO NOT CITE OR QUOTE
-------
CONTENTS (Continued)
9.3. AMBIENT CONCENTRATIONS AND EXPOSURE TO DIESEL EXHAUST .... 9-5
9.4. HAZARD CHARACTERIZATION 9-7
9.4.1. Acute and Short-Term Exposures 9-8
9.4.1.1. Acute Irritation 9-8
9.4.1.2. Respiratory Effects 9-8
9.4.1.3. Immunological Effects 9-8
9.4.2. Chronic Exposure 9-9
9.4.2.1. Noncancer Effects 9-9
9.4.2.2. Carcinogenic Effects 9-10
9.5. DOSE-RESPONSE ASSESSMENT 9-15
9.5.1. Evaluation of Risk for Noncancer Health Effects 9-16
9.5.1.1. Chronic Reference Concentrations for Diesel Exhaust 9-16
9.5.1.2. Risks Based on Ambient PM25 9-18
9.5.1.3. Apportionment Method Based on Ambient PM25 9-18
9.5.1.4. Conclusions 9-18
9.5.2. Evaluation of Cancer Risks 9-19
9.6. SUMMARY AND CONCLUSIONS 9-22
9.7. REFERENCES 9-24
APPENDIX A: CALCULATION OF HUMAN EQUIVALENT CONTINUOUS
EXPOSURE CONCENTRATIONS (HECs) A-l
APPENDIX B: BENCHMARK CONCENTRATION ANALYSIS OF
DIESEL DATA B-l
APPENDIX C: KEY PARTICULATE MATTER (PM) EPIDEMIOLOGIC
FINDINGS RELATED TO PM NAAQS DECISIONS C-l
APPENDIX D: A SUMMARY REVIEW OF CANCER DOSE-RESPONSE
ANALYSES ON DIESEL EXHAUST D-l
07/25/00
XI
DRAFT—DO NOT CITE OR QUOTE
-------
LIST OF TABLES
2-1. Vehicle classification and weights for on-road trucks 2-77
2-2. Total (gas and diesel) and diesel trucks in the fleet in 1992 2-77
2-3. Typical chemical composition of fine particulate matter 2-78
2-4. U.S. emission standards: HD highway diesel engines 2-78
2-5. U.S. emission standards: locomotives (g/bhp-hr) 2-79
2-6. U.S. emission standards for nonroad diesel equipment (g/bhp-hr) 2-80
2-7. Comparison of in-use truck fleet with truck fleet tested on
chassis dynamometer, percent of total vehicles 2-81
2-8. Diesel engine emissions data from engine dynamometer tests 2-82
2-9. HD diesel emissions results from tunnel tests 2-85
2-10. Remote sensing results for HD vehicles 2-86
2-11. Summary of CDD/CDF emissions from diesel-fueled vehicles 2-87
2-12. Baltimore Harbor Tunnel Study: estimated CDD/CDF emission
factors for HD vehicles 2-88
2-13. Organic and elemental carbon fractions of diesel and
gasoline engine PM exhaust 2-89
2-14. Emission rates of PAH (mg/mi) from LD and HD diesel vehicles 2-90
2-15. Polycyclic aromatic hydrocarbons identified in extracts of diesel
particles from LD diesel engine exhaust 2-91
2-16. Emission rates of particle-bound PAH (ug/mi) from diesel and
gasoline engines 2-92
2-17. Concentrations of nitro-PAHs identified in LD diesel particulate
extracts 2-93
2-18. Average emission rates for polycyclic aromatic hydrocarbons for
different fuel types 2-94
2-19. Classes of compounds in diesel exhaust 2-95
2-20. Calculated atmospheric lifetimes for gas-phase reactions of
selected compounds present in automotive emissions with
important reactive species 2-96
2-21. Major components of gas-phase diesel engine emissions, their
known atmospheric transformation products, and the biological
impact of the reactants and products 2-97
2-22. Major components of particle-phase diesel engine emissions,
their known atmospheric transformation products, and the
biological impact of the reactants and products 2-98
2-23. Ambient DPM concentrations reported from chemical mass
balance modeling .. 2-99
2-24. Comparison of DPM concentrations reported by CMB
and EC surrogate calculation 2-100
2-25. Ambient diesel particulate matter concentrations from elemental
carbon measurements in urban locations 2-101
2-26. Ambient diesel particulate matter concentrations from
dispersion modeling 2-102
07/25/00 xii DRAFT—DO NOT CITE OR QUOTE
-------
LIST OF TABLES (Continued)
2-27. Occupational exposure to DPM 2-103
2-28. Ranges of occupational exposure to diesel paniculate matter by
job category with estimates of equivalent environmental exposures 2-104
2-29. Annual average nationwide DPM exposure estimates (ng/m3) from
on-road sources for rural, urban, and urban demographic groups in
1990, 1996, 2007, and 2020 using HAPEM-MS3 2-105
2-30. Annual average DPM exposures for 1990 and 1996 in the general
population and among the highest exposed demographic groups in
nine urban areas and nationwide from (on-road sources only) using
HAPEM-MS3 2-106
2-31. Modeled and estimated concentrations of DPM in microenvironments
for California for all sources 2-107
2-32. Estimated indoor air and total air exposures to DPM in California
in 1990 2-108
3-1. Predicted doses of inhaled DPM per minute based on total lung
volume (M), total airway surface area (M,), or surface area in
alveolar region (M2) 3-43
3-2. Alveolar clearance in laboratory animals exposed to DPM 3-44
5-1. Human studies of exposure to diesel exhaust 5-72
5-2. Short-term effects of diesel exhaust on laboratory animals 5-77
5-3. Effects of chronic exposures to diesel exhaust on survival and
growth of laboratory animals 5-79
5-4. Effects of chronic exposures to diesel exhaust on organ weights
and organ-to-body-weight ratios 5-81
5-5. Effects of diesel exhaust on pulmonary function of laboratory animals 5-83
5-6. Histopathological effects of diesel exhaust in the lungs of
laboratory animals 5-84
5-7. Effects of exposure to diesel exhaust on the pulmonary defense
mechanisms of laboratory animals 5-87
5-8. Effects of inhalation of diesel exhaust on the immune system
of laboratory animals 5-90
5-9. Effects of diesel paniculate matter on the immune response of
laboratory animals 5-92
5-10. Effects of exposure to diesel exhaust on the liver of laboratory animals 5-93
5-11. Effects of exposure to diesel exhaust on the hematological and
cardiovascular systems of laboratory animals 5-94
5-12. Effects of chronic exposures to diesel exhaust on serum chemistry
of laboratory animals 5-95
5-13. Effects of chronic exposures to diesel exhaust on microsomal
enzymes of laboratory animals 5-96
07/25/00 • xiii DRAFT—DO NOT CITE OR QUOTE
-------
LIST OF TABLES (Continued)
5-14. Effects of chronic exposures to diesel exhaust on behavior
and neurophysiology 5-97
5-15. Effects of chronic exposures to diesel exhaust on reproduction
and development in laboratory animals 5-98
5-16. Composition of exposure atmospheres in studies
comparing unfiltered and filtered diesel exhaust 5-99
6-1. Histopathological effects of diesel exhaust in the lungs of
laboratory animals 6-18
6-2. Human equivalent continuous concentrations (HECs) calculated with
the model of Yu et al. (1991) from long-term repeated exposure
rat studies of DPM exposure 6-20
6-3. Decision summary for the quantitative noncancer RfC assessment for
continuous exposure to diesel particulate matter (DPM) 6-21
7-1. Epidemiologic studies of the health effects of exposure to diesel
exhaust: cohort mortality studies 7-116
7-2. Epidemiologic studies of the health effects of exposure to diesel
exhaust: case-control studies of lung cancer 7-121
7-3. Summary of animal inhalation carcinogenicity studies 7-127
7-4. Tumor incidences in rats following intratracheal instillation of
diesel exhaust particles (DPM), extracted DPM, carbon black (CB),
benzo[a]pyrene (BaP), or particles plus BaP 7-133
7-5. Tumorigenic effects of dermal application of acetone extracts of DPM 7-134
7-6. Tumor incidence and survival time of rats treated by surgical lung
implantation with fractions from diesel exhaust condensate (35 rats/group) 7-135
7-7. Dermal tumorigenic and carcinogenic effects of various emission extracts 7-136
7-8. Cumulative (concentration * time) exposure data for rats exposed
to whole diesel exhaust 7-137
7-9. Evaluations of diesel exhaust as to human carcinogenic potential 7-139
8-1. DPM exposure margins for occupational vs. environmental exposures 8-16
07/25/00 xiv DRAFT—DO NOT CITE OR QUOTE
-------
LIST OF FIGURES
2-1. Diesel truck sales (domestic) for the years 1939-1997 2-109
2-2. Diesel truck sales as a percentage of total truck sales for the
years 1939-1997 2-110
2-3. Percentage of truck miles attributable to diesel trucks 2-111
2-4. Model year distribution of in-use heavy HD truck fleet in 1997 2-112
2-5. Model year distribution of vehicle miles traveled by the in-use
heavy HD truck fleet in 1997 2-113
2-6. A comparison of IDI (A) and DI (B) combustion systems of
high-speed HD diesel truck engines 2-114
2-7. Schematic diagram of diesel engine exhaust particles 2-115
2-8. Typical chemical composition for diesel particulate matter
(PM2.5) from new (post-1990) HD diesel vehicle exhaust 2-116
2-9. Trends in PM10 emissions from on-road and nonroad diesel engines
from 1970 to 1998 and projections of emissions to 2007 and 2030 2-117
2-10. Trends in NOX emissions from on-road and nonroad diesel engines
from 1970 to 1998 2-118
2-11. Trends in SO, emissions from on-road diesel engines from 1970 to
1998 and nonroad diesel engines from 1990 to 1998 . 2-119
2-12. Trends in VOC emissions from on-road and nonroad diesel engines
from 1970 to 1998 2-120
2-13. Trends in CO emissions from on-road and nonroad diesel engines
from 1970 to 1998 2-121
2-14. Percentage of total motor fuel use that is on-road diesel fuel since 1949 2-122
2-15. On-highway diesel fuel consumption since 1943, values in
thousands of gallons 2-123
2-16. Model year trends in PM, NOX, HC, and CO emissions from HD
diesel vehicles (g/mile) 2-124
2-17. Diesel engine certification data for NOx emissions as a function
of model year 2-125
2-18. Diesel engine certification data for PM emissions as a function
of model year 2-125
2-19. Emission factors from HD diesel vehicles from tunnel studies 2-126
2-20. Line-haul and switch emissions data 2-127
2-21. Effect of turbocharging and aftercooling on NOX and PM 2-128
2-22. Comparison of diesel engine dynamometer PM emissions for
4-stroke, naturally aspirated and turbocharged engines 2-128
2-23. An example of uniflow scavenging of a two-stroke diesel engine with a
positive displacement blower 2-129
2-24. Comparison of two- and four-stroke vehicle diesel PM emissions
from chassis dynamometer studies 2-129
2-25. Comparison of two- and four-stroke engine diesel PM emissions
from engine dynamometer studies 2-130
07/25/00 xv DRAFT—DO NOT CITE OR QUOTE
-------
LIST OF FIGURES (Continued)
2-26. Diesel engine dynamometer SOF emissions from two- and four-stroke
engines. SOF obtained by dichloromethane extraction in most studies 2-130
2-27. Diesel engine aldehyde emissions measured in chassis
dynamometer studies 2-131
2-28. Diesel engine aldehyde emissions from engine dynamometer studies 2-131
2-29. Trend in SOF emissions based on chassis dynamometer testing of
heavy-duty diesel vehicles 2-132
2-30. Trend in SOF emissions for transient engine dynamometer testing
of HD diesel engines 2-132
2-31. Trend in SOF emissions as a percent of total PM based on chassis
dynamometer testing of HD diesel vehicles 2-133
2-32. Trend in SOF emissions as a percentage of total PM from engine
dynamometer testing 2-134
2-33. EC emission rates for diesel vehicles 2-135
2-34. EC content as percent of total carbon content for DPM samples
obtained in chassis dynamometer studies 2-135
2-35. Diesel engine emissions of benzo[a]pyrene and 1-nitropyrene
measured in chassis dynamometer studies 2-136
2-36. Diesel engine dynamometer measurements of benzo[a]pyrene
and 1-nitropyrene emissions from HD diesel engines 2-137
2-37. Particle size distribution in diesel exhaust 2-138
3-1. Modeled deposition distribution patterns of inhaled diesel exhaust
particles in the airways of different species 3-45
3-2. Modeled clearance of poorly soluble 4-um particles deposited in
tracheobronchial and alveolar regions in humans 3-46
3-3. Short-term thoracic clearance of inhaled particles as determined by
model prediction and experimental measurement 3-47
3-4. Clearance from lungs of rats of 134Cs-FAP fused aluminosilicate tracer
particles inhaled after 24 months of diesel exhaust exposure at concentrations
of 0 (control), 0.35 (low), 3.5 (medium), and 7.1 (high) mg DPM/m3 3-48
3-5. Lung burdens of DPM within rats exposed to 0.35 (low), 3.5 (medium),
and 7.1 (high) mg ppm/m3 3-49
7-1. Pooled relative risk estimates and heterogeneity-adjusted 95% confidence
intervals for all studies and subgroups of studies included in the meta-analysis ... 7-140
7-2. Pooled estimates of relative risk of lung cancer in epidemiological
studies involving occupational exposure to diesel exhaust
(random-effects models) 7-141
7-3. Pathegenesis of lung disease in rats with chronic, high-level
exposures to particles 7-142
07/25/00 xvi DRAFT—DO NOT CITE OR QUOTE
-------
PREFACE
This draft health risk assessment document was prepared by the National Center for
Environmental Assessment (NCEA), which is the health risk assessment program in EPA's
Office of Research and Development. The assessment has been prepared for EPA's Office of
Transportation and Air Quality, which requested advice regarding the potential health hazards
associated with diesel exhaust exposure. As diesel exhaust emissions also affect air toxics and
ambient particulate matter, other EPA air programs also have an interest in this assessment. The
previous draft of this assessment was released for public comment in November 1999, and the
Agency's Clean Air Scientific Advisory Committee (CASAC) met in public session in
December 1999 to review the draft. This July 2000 draft is a revision of that 1999 draft,
prepared in response to CASAC advice.
The scientific literature search for this assessment is generally current through January
1999, although a number of more recent publications on key topics have been included.
This July 2000 draft assessment will be reviewed by CASAC in the fall of 2000, and
concurrently, public comments will be accepted for a limited time. Following the receipt of
comments from CASAC and the public, NCEA plans to finalize the assessment.
07/25/00 xvii DRAFT—DO NOT CITE OR QUOTE
-------
AUTHORS AND CONTRIBUTORS
The National Center for Environmental Assessment (NCEA), within EPA's Office of
Research and Development (ORD), was responsible for the preparation of this document.
Authors and chapter managers for this draft health assessment document are listed below.
CHAPTER 1. EXECUTIVE SUMMARY
Authors
NCEA Diesel Team
CHAPTER 2. DIESEL EMISSIONS CHARACTERIZATION, ATMOSPHERIC
TRANSFORMATION, AND EXPOSURES
Chapter Manager/Author
Marion Hoyer, Office of Transportation and Air Quality (OTAQ), U.S. Environmental Protection
Agency, Ann Arbor, MI.
Contributors
Chad Bailey, OTAQ, U.S. Environmental Protection Agency, Ann Arbor, MI.
Tom Baines, OTAQ, U.S. Environmental Protection Agency, Ann Arbor, MI.
David Cleverly, National Center for Environmental Assessment, U.S. Environmental Protection
Agency, Washington, DC.
William Ewald, National Center for Environmental Assessment, U.S. Environmental Protection
Agency, Research Triangle Park, NC.
Robert McCormick, Colorado School of Mines, Golden, CO
Joseph McDonald, OTAQ, U.S. Environmental Protection Agency, Ann Arbor, MI.
Joseph Somers. OTAQ. U.S. Environmental Protection Agency, Ann Arbor. MI.
07/25/00 xviii DRAFT—DO NOT CITE OR QUOTE
-------
Janet Yanowitz, Colorado School of Mines, Golden, CO.
Barbara Zielinska, Desert Research Institute, Reno NV.
CHAPTER 3. DOSIMETRY OF DIESEL PARTICULATE MATTER
Authors
James McGrath, National Center for Environmental Assessment, U.S. Environmental Protection
Agency, Research Triangle Park, NC.
William Pepelko, National Center for Environmental Assessment, U.S. Environmental
Protection Agency, Washington, DC.
Contributor
Gary Foureman, National Center for Environmental Assessment, U.S. Environmental Protection
Agency, Research Triangle Park, NC.
CHAPTER 4. MUTAGENICITY
Author
Lawrence Valcovic, National Center for Environmental Assessment, U.S. Environmental
Protection Agency, Washington, DC.
CHAPTER 5. NONCANCER HEALTH EFFECTS OF DIESEL EXHAUST
Authors
James McGrath, National Center for Environmental Assessment, U.S. Environmental Protection
Agency, Research Triangle Park, NC.
Contributor
Gary Foureman, National Center for Environmental Assessment, U.S. Environmental Protection
Agency, Research Triangle Park, NC.
07/25/00 xix DRAFT—DO NOT CITE OR QUOTE
-------
CHAPTER 6. QUANTITATIVE APPROACHES TO ESTIMATING
HUMAN NONCANCER HEALTH RISKS OF DIESEL EXHAUST
Authors
Gary Foureman, National Center for Environmental Assessment, U.S. Environmental Protection
Agency, Research Triangle Park, NC.
Contributor
James McGrath, National Center for Environmental Assessment, U.S. Environmental Protection
Agency, Research Triangle Park, NC.
CHAPTER 7. CARCINOGENICITY OF DIESEL EXHAUST
Authors
Aparna Koppikar, National Center for Environmental Assessment, U.S. Environmental
Protection Agency, Washington, DC.
William Pepelko, National Center for Environmental Assessment, U.S. Environmental Protection
Agency, Washington, DC.
Contributors
Drew Levy, University of Washington, Seattle, WA
Robert Young, Oak Ridge National Laboratory, Oak Ridge, TN
CHAPTER 8. DOSE-RESPONSE ASSESSMENT: CARCINOGENIC EFFECTS
Authors
Chao Chen, National Center for Environmental Assessment, U.S. Environmental Protection
Agency, Washington, DC.
William Pepelko, National Center for Environmental Assessment, U.S. Environmental Protection
Aaencv. Washington. DC.
07/25/00 xx DRAFT—DO NOT CITE OR QUOTE
-------
Contributor
Charles Ris, National Center for Environmental Assessment, U.S. Environmental Protection
Agency, Washington, DC.
CHAPTER 9. CHARACTERIZATION OF POTENTIAL HUMAN HEALTH EFFECTS
OF DIESEL EXHAUST: HAZARD AND DOSE-RESPONSE ASSESSMENTS
Author
Charles Ris, National Center for Environmental Assessment, U.S. Environmental Protection
Agency, Washington, DC.
Contributors
NCEA Diesel Team
This document was preceded by four earlier drafts: a Workshop Review Draft
(EPA/600/8-90/057A, July 1990), an External Review Draft (EPA/600/8-90/057B, December
1994), an SAB Review Draft (EPA/600/8-90/057C, February 1998), and an SAB Review Draft
(EPA/600/8-90/057D, November 1999). The Science Advisory Board's Clean Air Scientific
Advisory Committee (CASAC) reviewed the 1994 draft in public sessions in May 1995, the
1998 draft in May 1998, and the 1999 draft in December 1999. Public comment periods also
were conducted concurrently with the CASAC reviews. In addition, many reviewers both within
and outside the Agency provided assistance at various review stages.
07/25/00 xxi DRAFT—DO NOT CITE OR QUOTE
-------
ACKNOWLEDGMENTS
Document Review
Vanessa Vu
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Washington, DC
Document Production
Terri Konoza
Judy Theisen
National Center for Environmental Assessment
U.S. Environmental Protection Agency
Washington, DC
Kay Marshall
Eric Sorensen
Clara Laucho
Harold Rogers
The CDM Group, Inc.
Chevy Chase, MD
07/25/00 xxii DRAFT—DO NOT CITE OR QUOTE
-------
1. EXECUTIVE SUMMARY
1 1.1. INTRODUCTION
2 The Health Assessment Document for Diesel Exhaust (DE) represents the Agency's first
3 comprehensive review of the potential health effects from ambient exposure to exhaust from
4 diesel engines. This assessment identifies and characterizes the potential human health hazards of
5 DE (i.e, hazard assessment) and characterizes the related dose-response associated with the key
6 health effects (i.e., dose-response assessment). This is part of the information needed for a
7 complete risk assessment of DE in support of EPA's Clean Air Act regulatory programs. A full
8 exposure assessment and risk characterization, the other two components of a complete risk
9 assessment, are beyond the scope of this document.
10 The report has nine chapters (including this chapter) and four appendices. Chapter 2
11 provides a characterization of diesel emissions, atmospheric transformation, and human
12 exposures to DE to provide a context for the hazard evaluation of DE. Chapters 3, 4, 5, and 7
13 provide a review of relevant information for the evaluation of potential health hazards of DE,
14 including dosimetry (Chapter 3), mutagenicity (Chapter 4), other noncancer health effects
15 (Chapter 5), and carcinogenicity (Chapter 7). Chapters 6 and 8 contain dose-response analyses to
6 provide insight about the significance of the potential noncancer and cancer hazards,
7 respectively. Chapter 9 characterizes the overall nature of the potential health hazard and risk
18 from environmental exposure to DE and discusses the overall confidence and uncertainties of the
19 assessment. Major conclusions of the health assessment for DE are provided below.
20
21 1.2. COMPOSITION OF DIESEL EXHAUST
22 DE is a complex mixture of hundreds of constituents in either a gas or particle phase.
23 Gaseous components of DE include carbon dioxide, oxygen, nitrogen, water vapor, carbon
24 monoxide, nitrogen compounds, sulfur compounds, and low-molecular-weight hydrocarbons.
25 Among the gaseous components of DE that are of toxicologic relevance are the aldehydes (e.g.,
26 formaldehyde, acetaldehyde, acrolein), benzene, 1,3-butadiene, and polycyclic aromatic
27 hydrocarbons (PAHs) and nitro-PAHs.
28 The particles present in DE (i.e., diesel particulate matter or DPM) are composed of
29 elemental carbon, adsorbed organic compounds, and small amounts of sulfate, nitrate, metals,
30 and other trace elements. DPM consists of fine and ultrafine particles. These particles are highly
31 respirable and have a very large surface area, which make them an excellent carrier for adsorbed
32 inorganic and organic compounds. The most lexicologically relevant organic compounds that are
adsorbed onto the particles include PAHs, nitro-PAHs, and oxidized PAH derivatives. PAHs
34 and their derivatives comprise about 1% or less of the DPM mass. Many of the organic
7/25/00 1 -1 DRAFT—DO NOT CITE OR QUOTE
-------
1 compounds present on the particle and in the gases are known to have mutagenic and
2 carcinogenic properties.
3
4 1.3. DIESEL EMISSIONS
5 DE is emitted from "on-road" diesel engines (vehicle engines) or "nonroad" diesel
6 engines (e.g., locomotives, marine vessels, heavy-duty equipment, etc). Nationwide, data in
7 1998 indicate that DE as measured by DPM made up about 6% of the total ambient PM2 5
8 inventory (i.e., particles with aerodynamic diameter of 2.5 microns or less) and about 23% of the
9 inventory excluding natural and miscellaneous sources. Estimates of the DPM percentage of the
10 total inventory in urban centers can be higher. For example, estimates range from 10% to 36% in
11 some areas in California, Colorado, and Arizona. Available data indicate that over the years,
12 there have been significant reductions in DPM emissions from the exhaust of on-road diesel
13 engines, whereas very limited data suggest that exhaust emissions from nonroad engines have
14 increased.
15 DE emissions vary significantly in chemical composition and particle sizes with different
16 engine types (heavy-duty, light-duty), engine operating conditions (idle, accelerate, decelerate),
17 and fuel formulations. The mass of particles emitted and the organics on the particles from on-
18 road diesel engines have been reduced over the years. Available data indicate that lexicologically
19 relevant organic components of DE (e.g., PAHs, nitro-PAHs) were present in DPM and DE
20 emitted from older vehicle engines and are still present in emissions from newer engines. There
21 is insufficient information, however, to characterize the changes in the composition of DPM
22 from nonroad diesel engines over time.
23
24 1.4. ATMOSPHERIC TRANSFORMATION OF DIESEL EXHAUST
25 After emission from the tailpipe, DE undergoes dilution and chemical and physical
26 transformations in the atmosphere, as well as dispersion and transport in the atmosphere. The
27 atmospheric lifetime for some compounds present in DE ranges from hours to days. DPM is
28 either directly emitted from diesel-powered engines (primary particulate matter) or is formed
29 from the gaseous compounds emitted by diesel engines (secondary particulate matter). Limited
30 information is available about the physical and chemical transformation of DE in the atmosphere.
31 It is not clear what the overall toxicological consequence of DE^aging is, because some
32 compounds in the DE mixture are altered during aging to more toxic forms while others are made
33 less toxic.
7/25/00 1 -2 DRAFT—DO NOT CITE OR QUOTE
-------
1 1.5. EXPOSURE TO DIESEL EXHAUST
2 DPM mass (expressed as ng/m3 of DPM ) has historically been measured as a surrogate
3 for whole DE. Although considerable uncertainty exists as to whether DPM is the most
4 appropriate dosimeter for human health effects, it is considered a reasonable choice until more
5 definitive information about the mechanisms or mode(s) of action of DE becomes available. In
6 the ambient environment, exposure to DE comes from both on-road and nonroad engine exhaust.
7 A large percentage of the U.S. population is exposed to ambient PM25, of which DE is a part.
8 Estimates suggest that nonroad sources of DE contribute as much to the nationwide PM
9 inventory as do on-road DE sources. With limited information from actual measurements of DE,
10 various types of models and assumptions are used to estimate human exposure to on-road
11 generated DE as measured by DPM. Exposure information is useful to provide a context for the
12 health effects information, and estimates for the early to mid-1990s suggest that annual average
13 DE exposure from on-road engines alone was in the range of about 0.5 to close to 1.0 u.g
14 DPM/m3 of inhaled air in many rural and urban areas, respectively. For urban areas where
15 people spend a large portion of their time outdoors, the exposures may range up to 4.0 p,g
16 DPM/m3 of inhaled air. Exposure estimates are adjusted to account for time spent outdoors.
17 Exposures could be higher still, if there is a nonroad DE source that adds to the on-road-
18 generated exposure.
\B
20 1.6. HEALTH EFFECTS
21 Available evidence indicates that adverse human health effects may result from current-
22 day environmental inhalation exposure to DE. DE exposure may cause acute and chronic
23 noncancer respiratory effects and has the potential to cause lung cancer in humans.
24
25 1.6.1. Acute Effects
26 Available information for characterizing potential health effects associated with acute or
27 short-term exposure is limited. On the basis of available human and animal evidence, it is
28 concluded that DE can cause acute irritation (e.g., eye, throat, bronchial irritation),
29 neurophysiological symptoms (e.g., lightheadedness, nausea), and respiratory symptoms (cough
30 and phlegm). There is also evidence for possible immunologic effects and/or exacerbation of
31 allergenic responses to known allergens. The lack of exposure-response information precludes
32 the development of recommendations about levels of exposure that would be protective for these
33 effects.
34
7/25/00 1-3 DRAFT—DO NOT CITE OR QUOTE
-------
1 1.6.2. Chronic Noncancer Respiratory Effects
2 The information in available human studies is inadequate for a definitive evaluation of
3 possible noncancer health effects from chronic exposure to DE. However, on the basis of
4 extensive animal evidence, DE may pose a chronic respiratory hazard to humans. Chronic
5 animal inhalation studies show a spectrum of dose-dependent chronic inflammation and
6 histopathological changes in the lung in several animal species including rats, mice, hamsters,
7 and monkeys.
8 This assessment provides an estimate of an air-level exposure of DE (as measured by
9 DPM) to which humans may be exposed throughout their lifetime without experiencing any
10 untoward or adverse noncancer health effects. This exposure level, known as the reference
11 concentration (RfC), for DE of 14 ng/m3 of DPM was derived on the basis of dose-response data
12 from four chronic rat inhalation studies. This value is almost the same as the long-term PM2 5
13 NAAQS (National Ambient Air Quality Standard) of 15 p.g/m3.
14
15 1.6.3. Carcinogenic Effects
16 This assessment concludes that DE is likely to be carcinogenic to humans by inhalation at
17 any exposure condition. This characterization is based on the totality of evidence from human,
18 animal, and other supporting studies. There is considerable evidence demonstrating an
19 association between DE exposure and increased lung cancer risk among workers in different
20 occupations. The human evidence is considered strong but less than sufficient to definitively
21 conclude that DE exposure is causally associated with lung cancer, because of the possible
22 confounding effects of smoking and the lack of actual DE exposure data for the workers. In
23 addition to the human evidence, there is extensive evidence for the induction of lung cancer in
24 the rat from chronic inhalation exposure to high concentrations of DE, and supporting evidence
25 of carcinogenicity of DPM and associated organic compounds in rats and mice by noninhalation
26 routes of exposure. Other supporting evidence includes the demonstrated mutagenic and
27 chromosomal effects of DE and its organic constituents. There is also suggestive evidence for
28 the bioavailability of the organics from DE in humans and animals. The precise role of DPM
29 v/ith its organic component in DE-induced carcinogenicity is unclear, although in high-exposure
30 animal test systems, DPM and its elemental carbon core are shown to be the most important
31 fraction of DE. - -- - - --------- - —
32 Although the available human evidence shows the hazard to be present at exposures
33 generally higher than ambient levels, it is reasonable to presume that the hazard extends to
34 ambient environmental exposure levels. Because of an incomplete understanding of the mode of
35 action for DE-induced lung cancer in humans, and some evidence for a mutagenic mode of
36 action, it is a prudent public health policy to presume a cancer hazard for DE at any exposure
7/25/00 1-4 DRAFT—DO NOT CITE OR QUOTE
-------
1 condition. This presumption pertains only to the carcinogenic hazard and does not inform about
2 the magnitude of the risk at ambient levels. Overall, the evidence for a potential cancer hazard to
3 humans resulting from chronic inhalation exposure to DE is persuasive, even though assumptions
4 and thus uncertainties are involved.
5 Given a carcinogenicity hazard, EPA typically performs a dose-response assessment of
6 human or animal data to develop a cancer unit risk estimate that can be used with exposure
7 information to characterize the potential cancer disease impact on an exposed population. For
8 DE, the exposure-response data in rat studies are not deemed appropriate for the estimation of
9 human risk. Exposure-response data in available human studies are considered too uncertain to
10 derive a confident quantitative estimate of cancer unit risk. Therefore, EPA has chosen not to
11 derive a quantitative estimate of cancer unit risk.
12 In the absence of a unit risk to assess environmental cancer risk, simple analyses are
13 performed to provide a perspective of the range of the possible lung cancer risk from
14 environmental exposure to DE. The analyses make use of epidemiologic findings of lung cancer
15 risks from occupational exposures to DE, and consider the exposure margins between
16 occupational and environmental exposures to DE. The magnitude of the possible lifetime cancer
17 risk, based on the simple analyses, indicates the significance of the potential lung cancer hazard
J 8 from ambient exposure to DE. These analyses, however, are subject to considerable
19 uncertainties, and should not be viewed as a definitive quantitative characterization of risk.
20
21 1.7. SOURCES OF UNCERTAINTIES
22 Even though the overall evidence for potential human health effects of DE is persuasive,
23 many uncertainties exist because of the use of assumptions to bridge data and knowledge gaps
24 about human exposures to DE, and the underlying mechanisms by which DE causes observed
25 toxicities in humans and animals. A major uncertainty of this assessment is how the physical and
26 chemical nature of the past exposures to DE compares with present-day exposures, and how
27 representative the DE exposure-response data are from occupational and toxicological studies for
28 the characterization of possible hazard and risk from present-day environmental exposures.
29 Available data are not sufficient to provide definitive answers to these questions, as changes in
30 DE composition over time cannot be confidently quantified and the modes of action for DE
31 toxicity and carcinogenicity are unknown in humans. Despite these uncertainties, this assessment
32 assumes that prior-year toxicologic and epidemiologic findings can be applied to more current
33 exposures, both of which use ug/m3 of DPM mass as the dosimeter.
34 Other uncertainties include the assumptions that health effects observed at high dose may
15 be applicable to low dose, and that toxicologic findings in laboratory animals are predictive of
36 human responses. In the absence of more complete understanding of how DE may cause adverse
7/25/00 1 -5 DRAFT—DO NOT CITE OR QUOTE
-------
1 health effects in humans and laboratory animals, the assumptions used in this assessment (i.e., a
2 biological threshold for chronic respiratory effects) and absence of a threshold for lung cancer are
3 considered prudent and reasonable.
4 The assessment addresses the potential DE health hazards for average healthy adults.
5 There is no DE-specific information that provides direct insight to the question of variable
6 susceptibility within the general human population and vulnerable subgroups, including infants
7 and children, and people with preexisting health conditions, particularly respiratory conditions.
8 Despite these uncertainties, the default approach of using an uncertainty factor of 10 to account
9 for possible interindividual variation to DE in the derivation of the RfC is appropriate and
10 reasonable given the lack of DE-specific data.
11 In providing a perspective on the significance of the environmental cancer hazard of DE,
12 this assessment considers the differences in the magnitude of DE exposures between the
13 occupational and environmental settings. Variation in DE exposure is a source of uncertainty.
14 Because of variation in activity patterns, different population subgroups could potentially receive
15 higher or lower exposure to DE depending on their proximity to DE sources. Accordingly, DE
16 exposure estimates used in this assessment have included possible high-end exposures.
17 Lastly, this assessment considers only potential heath effects from exposures to DE alone.
18 DE exposure could be additive or synergistic to concurrent exposures to many other air
19 pollutants. However, in the absence of more definitive data demonstrating interactive effects
20 (e.g., potentiation of allergenicity effects, potentiation of DPM toxicity by ambient ozone and
21 oxides of nitrogen) from combined exposures to DE and other pollutants, it is not possible to
22 address this issue at this time. Further research is needed to improve the knowledge and database
23 on DE exposures and potential human health effects, and thereby reduce uncertainties of future
24 risk assessments of DE.
7/25/00 1-6 DRAFT—DO NOT CITE OR QUOTE
-------
2. DIESEL EMISSIONS CHARACTERIZATION, ATMOSPHERIC
TRANSFORMATION, AND EXPOSURES
1 2.1. INTRODUCTION
2 This chapter provides background information relating to the diesel engine, the pollutants
3 it emits, the history of its use in highway vehicles and railroad locomotives, diesel exhaust
4 composition and emissions trends, and ah" pollution regulatory standards for diesel engines in the
5 United States. The chapter also provides specific information about physical and chemical
6 composition of diesel exhaust, descriptions of its atmospheric transformations, observations of
7 measured and modeled ambient concentrations (considered alone and as a component of
8 atmospheric particles in general), and some estimates of population exposures. In addition, this
9 chapter gives background information that is used in conjunction with the toxicology and
10 epidemiology data to formulate the conclusions about human health hazards that are discussed in
11 later chapters of this document. The exposure information does not represent a formal or
12 rigorous exposure assessment; it is intended only to provide a context for the health effects data
13 and health hazard findings.
14 The diesel engine was patented in 1892 by Rudolf Diesel, who conceived it as a prime
15 mover that would provide much improved fuel efficiency compared with spark-ignition engines.
B To the present day, the diesel engine's excellent fuel economy remains one of its strongest
17 selling points. In the United States, the diesel engine is used mainly in trucks, buses, agricultural
18 and other off-road equipment, locomotives, and ships.
19 The chief advantages of the diesel engine over the gasoline engine are its fuel economy
20 and durability. Diesel engines, however, emit more paniculate matter per mile driven compared
21 with gasoline engines of a similar weight class. Over the past decade, modifications of diesel
22 engine components have substantially reduced particle emissions from both diesel and gasoline
23 engines (Hammerle et al., 1994; Sawyer and Johnson, 1995).
24 The diesel engine compresses air to high pressure and temperature. Fuel, when injected
25 into this compressed air, autoignites, releasing its chemical energy. The expanding combustion
26 gases do work on the piston before being exhausted to the atmosphere. Power output is
27 controlled by the amount of injected fuel rather than by throttling the air intake. Compared to its
28 spark-ignition counterpart, the diesel engine's superior efficiency derives from a higher
29 compression ratio and no part-load throttling. To ensure structural integrity for prolonged
30 reliable operation at the higher peak pressures brought about by a higher compression ratio and
31 autoignition, the structure of a diesel engine generally is more massive than its SI counterpart.
Diesel engines (also called compression-ignition) may be broadly identified as being
either two- or four-stroke cycle, injected directly or indirectly, and naturally aspirated or
7/25/00 2-1 DRAFT—DO NOT CITE OR QUOTE
-------
1 supercharged. They also are classified according to service requirements such as light-duty (LD)
2 or heavy-duty (HD) automotive/truck, small or large industrial, and rail or marine.
3 All diesel engines use hydraulic fuel injection in one form or another. The fuel system
4 must meet four main objectives if a diesel engine is to function properly over its entire operating
5 range. It must: (1) meter the correct quantity of fuel, (2) distribute the metered fuel to the
6 correct cylinder, (3) inject the metered fuel at the correct time, and (4) inject the fuel so that it is
7 atomized and mixes well with the in-cylinder air. The first two objectives are functions of a
8 well-designed injection pump, and the last two are mostly functions of the injection nozzle. Fuel
9 injection systems are moving toward the use of electronic components for more flexible control
10 than is available with purely mechanical systems to obtain lower exhaust emissions without
11 diminishing fuel efficiency.
12 Both the fuel and the lubricants that are used to service diesel engines are highly finished
13 petroleum-based products combined with chemical additives. Diesel fuel is a mixture of many
14 different hydrocarbon molecules from about C7 to about C35, with a boiling range from roughly
15 350 °F to 650 °F. Many of the fuel and oil properties, such as its specific energy content (which
16 is higher than gasoline), ignition quality, and specific gravity, are related to its hydrocarbon
17 composition. Therefore, fuel and lubricant composition affects many aspects of engine
18 performance, including fuel economy and exhaust emissions.
19 Complete and incomplete combustion of fuel in the diesel engine results in the formation
20 of a complex mixture of gaseous and particulate exhaust. Because of concerns over health
21 effects associated with diesel particulate emissions, EPA began regulating emissions from diesel
22 engines in 1970 (for smoke) and then added regulations for gaseous emissions. EPA first
23 regulated particulate emissions from HD diesels in 1988.
24 For the purposes of this document, carbonaceous matter, diesel exhaust, diesel particulate
25 matter (DPM), elemental carbon (EC), organic carbon (OC), soluble organic fraction (SOF), and
26 soot are defined as listed below.
27 Carbonaceous matter: Carbon-containing compounds, includes organic carbon and
28 elemental carbon that are associated with particulate matter in diesel exhaust. In this
29 document, th.e term carbonaceous matter includes all organic and elemental carbcn-
30 containing compounds that are found in the particle phase. In other documents, this term
31 is sometimes used interchangeably to refer to the insoluble fraction of diesel particulate
32 matter or the soot fraction.
33
34 Diesel exhaust (DE): Gaseous and panicle phase emissions resulting from the
35 combustion of diesei fuel in an internal combustion engine (e.g., compression ignition
36 engine). Diesel exhaust includes emissions from a diesel engine or diesel vehicle
7/25/00 2-2 DRAFT—DO NOT CITE OR QUOTE
-------
(inclusive of aftertreatment devices) but does not include emissions from brake and tire
wear.
3
4 Diesel particulate matter: The particle-phase compounds emitted in diesel exhaust.
5 DPM can refer to both primary emissions as well as secondary particles that are formed
6 by atmospheric processes. In this document, DPM refers to primary particles. Primary
7 diesel particles are considered fresh after being emitted and undergo aging (oxidation,
8 nitration, or other chemical and physical changes as discussed in Section 2.3) in the
9 atmosphere. As used in this document, DPM refers to both fresh and aged diesel
10 particulate matter unless a distinction is made.
11
12 Elemental carbon: Carbon that has undergone pyrolysis (e.g., has been stripped of
13 hydrogen). In pure form, elemental carbon contains only carbon atoms although
14 elemental carbon as it exists in combustion particulate matter is likely to contain some
15 hydrogen atoms. In this document, the terms elemental carbon and organic carbon are
16 used to refer to the carbon-containing components of DPM, and collectively they are
17 referred to as the carbonaceous fraction of a diesel particle.
I
(9 Organic carbon: Carbon- and hydrogen-containing molecules emitted in diesel exhaust
20 largely as the result of unburned diesel fuel and to a lesser extent resulting from engine
21 lubrication oil. Organic carbon compounds can also contain oxygen, nitrogen, and sulfur
22 as well as other elements in small quantities. In this document, the terms elemental
23 carbon and organic carbon are used to refer to the carbon-containing components of
24 DPM, and collectively they are referred to as the carbonaceous fraction of a diesel
25 particle.
26
27 Soluble organic fraction: The portion of DPM that can be extracted from the particle
28 matrix into solution. Extraction solutions and procedures vary and are described in
29 Section 2.2.8.1.
30
31 Soot: Agglomerations of elemental carbon and organic carbon particles. Often
32 characterized as the insoluble portion of DPM.
33
7/25/00 2-3 DRAFT—DO NOT CITE OR QUOTE
-------
1 This chapter begins with the history of dieselization for on-road vehicles and
2 locomotives, followed by an introductory discussion of the formation of primary diesel
3 emissions to assist the reader in understanding the complex factors that influence the formation
4 of particulate matter (PM) and other diesel exhaust emissions. The next section is a summary of
5 EPA emission standards for on-road and locomotive diesel engines and a description of the
6 national trends in emissions from on-road and nonroad diesel engine sources based on inventory
7 modeling. The chapter continues with a discussion of diesel fuel use and the impact of fuel
8 properties on emissions. The chronological assessment of emissions factors is presented in
9 summaries of chassis and engine dynamometer testing and tunnel tests. This is followed by a
10 description of engine technologies and their effect on emissions and a description of the
11 chemical and physical nature of emissions. The data describing the important atmospheric
12 transformations of diesel exhaust are summarized. The chapter concludes with a summary of the
13 available literature regarding atmospheric concentrations of diesel particulate matter and
14 exposures to diesel exhaust. EPA has assessed national and urban-area annual average exposure
15 to diesel particulate matter using the Hazardous Air Pollutant Exposure Model - Mobile Sources
16 version 3 model, and this assessment is presented in Section 2.4.3. A full exposure assessment
17 including the distribution of ambient diesel exhaust exposures hi different geographic regions
18 and among different demographic groups; the most highly exposed (90* percentile); exposures
,19) in microenvironments for short and long durations; and the maximum exposure range (98th
20 percentile) and number of maximum-exposed individuals is not currently available. EPA is
21 developing tools to provide a more complete exposure assessment.
22
23 2.2. PRIMARY DIESEL EMISSIONS
24 2.2.1. History of Dieselization
25 Information regarding diesel engine use including market share, vehicle miles traveled
26 (VMT) and fleet turnover are important aspects in understanding potential health effects of past
27 and present diesel exhaust exposure. In this section, the dieselization of the on-road truck fleet
28 and locomotives are discussed.
23
30 2.2.1.1. Dieselization of the On-Road Fleet
31 Because of their durability and fuel economy, the use of diesel engines, particularly in
32 long-distance applications, has increased over the years. The Census of Transportation, Truck
33 Inventory and Use Survey (TIUS) indicates that among Class 3-8 trucks, diesel engine use has
34 increased more rapidly than gasoline engine use in the past 20 years. Truck classes are defined
3§ by gross vehicle weight as described in Table 2-1. Dieselization first occurred among Class 7
^S and 8 (HHD) trucks The TTUS indicates that 81.5% of diesel trucks on the road in 1963 were
37 Class 7 or 8 (HHD) trucks (Table 2-2). Class 7 sales became predominantly (e.g., >50%) diesel
7/25/00 2-4 DRAFT—DO NOT CITE OR QUOTE
-------
1 in the 1970s and Class 8 sales became predominantly diesel in the 1960s. Diesels did not
)2 comprise a majority of class 5 and 6 sales until the 1990s (Figures 2-1 and 2-2). Heavy-duty
3 trucks (LHD and HMD) have historically constituted the majority of diesel sales and mileage.
4 However, an increasing number of light-duty (LD) diesel trucks have been sold domestically in
5 recent years. In the 1990s, approximately one in three diesel trucks sold were Class 1 and Class
6 2 vehicles. Diesel trucks have historically been driven more miles per truck than gasoline
7 trucks. For example, the TIUS indicates that 59% of diesel trucks were driven more than 50,000
8 miles in 1963, compared to 3% of gasoline trucks. Among combination trucks, consisting of
9 tractor-trailers and single-unit trucks with trailers, diesel vehicles have driven a majority of the
10 miles since at least 1963, the earliest year in which TIUS was conducted (Figure 2-3).
11 The longevity of diesel trucks is an important factor to understand past, current, and
12 projected exposures to diesel exhaust because older vehicles are subject to less stringent
13 regulations and may remain in use for several decades after their manufacture. American
14 Automobile Manufacturers Association publications (AAMA, 1927-1997) indicate that 53% of
15 trucks from model years 1947-1956 were still on the road after 14 years. The proportion of
16 trucks in use after 14 years was 63% for model years 1974-1983, suggesting that the lifespan of
17 trucks built in later years is longer. According to the 1997 TIUS, vehicles older than 10 years
18 make up 40% of Class 7 and 8 (HHD) trucks and drove 16% of Class 7-8 VMT (Figures 2-4 and
^ 2-5). Almost all HDD trucks were diesel vehicles in the period 1982-1997 (93% in 1982 and
20 99% in 1997).
21
•22 2o2.L2. DlessMzsaHom ®f Rmiln&d Loc®m®tive Engines
23 Early in the 20th century the political and economic pressure on the railroads to replace
24 steam locomotives was substantial. Railroads were losing business to other forms of transport.
25 The diesel-electric locomotive provided 90% in-service time, compared with only 50% for steam
26 locomotives, and had three times the thermal efficiency (Klein, 1991; Kirkland, 1983).
27 Additionally, several cities had passed laws barring steam locomotives within the city limits
28 because the large quantities of smoke obscured visibility, creating a safety hazard. The first
29 prototype diesel locomotive was completed in 1917. By 1924 General Electric (GE) was
30 producing a standard line of switching locomotives on a production basis. Electro-Motive
31 Corporation was founded the same year to produce diesel locomotives in competition with GE.
32 This company was purchased in 1929 by General Motors (GM) and became the Electro-Motive
33 Division. After this acquisition, GM began to develop the two-stroke engine for this application.
34 Up to this time, all locomotive diesel engines were four-stroke. Two-strokes offered a much
35 higher power-to-weight ratio, and GM's strategy was to get a large increase in power by moving
to the two-stroke cycle. The first true high-speed two-stroke diesel-electric locomotives were
37 produced by GM in 1935. However, because of the economic climate of the Great Depression,
7/25/00 2-5 DRAFT—DO NOT CITE OR QUOTE
-------
J few of these were so Id until after the Second World War. At the end of the war most
/ *^
. 2r locomotives were still steam-driven but were more than 15 years old, and the railroads were
3 ready to replace the entire locomotive fleet. Few if any steam locomotives were sold after 1945
4 because the entire fleet was converted to diesel (Coifhian, 1994).
5 The locomotive fleet has included significant percentages of both two- and four-stroke
6 engines. The four-stroke diesel engines were naturally aspirated in the 1940s and 1950s. It is
7 unlikely that any of the two-stroke engines used in locomotive applications were strictly
8 naturally aspirated. Nearly all two-stroke diesel locomotive engines are uniflow scavenged, with
9 a positive-displacement blower for scavenging assistance. In 1975, it was estimated that 75% of
10 the locomotives in service were two-stroke, of which about one-half used one or more
11 turbochargers in addition to the existing positive-displacement blower for additional intake boost
12 pressure.
13 Almost all of the four-stroke locomotive engines were naturally aspirated in 1975.
14 Electronic fuel injection for locomotive engines was first offered in the 1994 model year (U.S.
15 EPA, 1998b). All locomotive engines manufactured in recent years are turbocharged,
16 aftercooled or interceded four-stroke engines. In part, this is because of the somewhat greater
17 durability of four-strokes, although impending emissions regulations may have also been a factor
18 in this shift. The typical lifespan of a locomotive has been estimated to be more than 40 years
1^ (U.S. EPA, 1998b). Many of the smaller railroads are still using engines built in the 1940s,
20 although the engines may have been rebuilt several times since their original manufacture.
21
22 2.2.2. Diesel Combustion and Formation of Primary Emissions
23 A basic understanding of diesel combustion processes can assist in understanding the
24 complex factors that influence the formation of DPM and other diesel exhaust emissions. Unlike
25 spark-ignition combustion, diesel combustion is a fairly nonhomogenous process. Fuel is
26 sprayed at high pressure into the compressed cylinder contents (primarily air with some residual
27 combustion products) as the piston nears the top of the compression stroke. The turbulent
28 mixing of fuel and air that takes place is enhanced by injection pressure, the orientation of the
23 intake ports (e.g., inducement of intake-swirl tangential to the cylinder wall), piston motion, and
30 piston bowl shape. In some cases, fuel and air mixing is induced via injection of the fuel into a
31 turbulence-generating pre-chamber or swirl chamber located adjacent to the main chamber
32 (primarily in older, higher speed engines and some LD diesels). Examples of typical direct
33 injection and indirect injection combustion systems are compared in Figure 2-6. Diesel
34 combustion can be considered to consist of the following phases (Heywood, 1988; Watson and
Janota, 1982):
i
7/25/00 2-6 DRAFT—DO NOT CITE OR QUOTE
-------
1. An ignition delay period, which starts after the initial injection of fuel and continues
until the initiation of combustion. The delay period is governed by the rate of fuel
3 and air mixing, diffusion, turbulence, heat transfer, chemical kinetics, fuel
4 vaporization, and fuel composition. Fuel cetane rating is an indication of ignition
5 delay.
6 2. Rapid, premixed burning of the fuel and air mixture from the ignition delay period.
7 3. Diffusion-controlled burning, where the fuel burns as it is injected and diffuses into
8 the cylinder.
9 4. A very small amount of rate-controlled burning during the expansion stroke, after
10 the end of injection.
11
12 Engine speed and load are controlled by the quantity of fuel injected. Thus, the overall
13 fuel-to-air ratio varies greatly as engine speed and load vary. On a macro scale, the cylinder
14 contents are always fuel-lean. Depending on the time available for combustion and the
15 proximity of oxygen, the fuel droplets are either completely or partially oxidized. At
16 temperatures above 1,300 K, much of the unburned fuel that is not oxidized is pyrolized
17 (stripped of hydrogen) to form elemental carbon (Dec and Espey, 1995). In addition to
18 elemental carbon, other carbonaceous matter is present, largely from unburned fuel. The
^fc agglomeration of elemental and organic carbon (OC) forms particles that are frequently referred
20 to as "soot" particles. In this document, the terms "elemental carbon" and "organic carbon" are
21 used to refer to the carbon-containing components of diesel paniculate matter, and collectively,
22 they are referred to as the carbonaceous fraction of a diesel particle.
23 Carbonaceous particle formation occurs primarily during the diffusion-burn phase of
24 combustion, and it is highest during high load and other conditions consistent with high fuel-air
25 equivalence ratios. Most of the carbonaceous matter formed (80% to 98%) is oxidized during
26 later stages of combustion, most likely by hydroxyl radicals formed during combustion
27 (Kittelson et al., 1986; Foster and Tree, 1994). The remainder of the carbonaceous particulate
28 leaves as a component of DPM emissions from the engine.
29 DPM is defined by the measurement procedures summarized in the Code of Federal
30 Regulations Title 40 CFR, Part 86, Subpart N (CFR 40:86.N). These procedures define DPM
31 emissions as the mass of material collected on a filter at a temperature of 52 °C or less after
32 dilution of the exhaust with air. As the exhaust is diluted and cooled, nucleation, condensation,
33 and adsorption transform volatile material to solid and liquid DPM. Diesel exhaust particles are
34 aggregates of primary spherical particles consisting of solid carbonaceous material and ash and
35 containing adsorbed organic and sulfur compounds (sulfate) combined with other condensed
^P material (Figure 2-7). These particles have a very large surface area per gram of mass, which
37 make them an excellent carrier for adsorbed inorganic and organic compounds that can
7/25/00 2-7 DRAFT—DO NOT CITE OR QUOTE
-------
effectively reach the lowest airways of the lung. The elemental carbon core has a high specific
surface area of approximately 30-50 mVg (Frey and Corn, 1967). Because of this high surface
3 area, the elemental carbon core is able to adsorb large quantities of ash, organic compounds, and
4 sulfate. Pierson and Brachaczek (1976) report that after the extraction of adsorbed organic
5 material, the surface area of the diesel particle core is approximately 90 m2/g.
6 The organic material associated with diesel particles originates from unburned fuel,
7 engine lubrication oil, and small quantities of partial combustion and pyrolysis products. This is
8 frequently quantified as the soluble organic fraction, which is discussed in much more detail hi
9 Section 2.2.7. The formation of sulfate hi diesel exhaust depends primarily on fuel sulfur
10 content. During combustion, sulfur compounds present in the fuel are oxidized to sulfur dioxide
11 (SO2). Approximately 1% to 4% of fuel sulfur is oxidized to SO3, which combines with water
12 vapor in the exhaust to form sulfuric acid (H2SO4) (Wall et al., 1987; Khatri et al., 1978;
13 Baranescu, 1988; Barry et al., 1985). Upon cooling, sulfuric acid and water condense into an
14 aerosol that is nonvolatile under ambient conditions. The mass of sulfuric acid DPM is more
15 than doubled by the mass of water associated with the sulfuric acid under typical DPM
16 measurement conditions (50% relative humidity, 20-25 °C) (Wall et al., 1987).
17 Oxide of nitrogen (NOX) emissions from combustion engines, primarily (at least initially)
.18 in the form of NO, are generally thought to be formed via the Zeldovich mechanism, which is
1^ highly temperature dependent. High combustion temperatures cause reactions between oxygen
20 and nitrogen to form NO and some NO2. The majority of NO2 formed during combustion is
21 rapidly decomposed. NO can also decompose to N2 and O2 but the rate of decomposition is very
22 slow because of the rapidly decreasing temperatures from the expansion of combustion gases
23 during the expansion stroke (Heywood, 1988; Watson and Janota, 1982). Thus, almost all of the
24 NOX emitted is NO.
25 Some organic compounds from unburned fuel and from lubricating oil consumed by the
26 engine can be trapped hi crevices or cool spots within the cylinder and thus are not sufficiently
27 available to conditions that would lead to their oxidation or pyrolysis. These compounds are
28 emitted from the engine and either contribute to gas-phase organic emissions or to DPM
29 emissions, depending on their volatility. Within the exhaust system, temperatures gre sufficiently
30 High that these compounds are entirely present within the gas phase (Johnson and Kittelson,
31 1996). Upon cooling and mixing with ambient air in the exhaust plume, some of the less volatile
32 organic compounds can adsorb to the surfaces of the elemental carbon agglomerate particles.
33 Lacking sufficient elemental carbon adsorption sites, the organic compounds may condense on
34 sulfuric acid nuclei to form a heterogeneously nucleated organic aerosol (AbJul-Khalek et al.,
•« ^^ ^*- ^*\
iy?9).
wmie not unique to diesel exhaust, the high content of elemental carbon associated with
37 typical DPM emissions has long been used by some investigators to distinguish diesel engine
7/25/00 2-8 DRAFT—DO NOT CITE OR QUOTE
-------
1 sources of this particle from other combustion aerosols. Diesel particles from newer HD engines
^ are typically composed of-75% elemental carbon (EC can range from 33-90%), -20% organic
3 carbon (OC can range from 7-49%), and small amounts of sulfate, nitrate, trace elements, water,
4 and unidentified components (Figure 2-8). Metallic compounds from engine component wear,
5 and from compounds in the fuel and lubricant, contribute to DPM mass (1% to 5% of mass is
6 attributed to metals as discussed in more detail below). Ash from oil combustion also
7 contributes trace amounts to DPM mass.
8 In contrast to the composition of diesel PM2.5, ambient PM2.5 measured in the eastern
9 United States is dominated by sulfate (34%), while ambient PM2.5 in the western United States
10 is dominated by organic carbon (39%) (Table 2-3) (U.S. EPA, 1996a). The organic carbon
11 fraction of DPM is increasingly being used to assist investigators in identifying the contribution
12 of diesel engine emissions to ambient PM2.5. In particular, hopane and sterane compounds
13 (aromatic compounds, >C30) have been used in addition to other polycyclic aromatic HCs
14 (P AHs) and long chain alkanes to distinguish DPM from other mobile source PM and from
15 ambient PM (Schauer et al., 1996; Fujita et al., 1998). While PAH compounds comprise 1% or
16 less of DPM mass, diesel emissions have been observed to have elevated concentrations of
17 methylated naphthalenes and methylated phenanthrene isomers compared to other combustion
18 aerosols (Benner et al., 1989; Lowenthal et al., 1994; Rogge et al., 1993). Enrichment of
|) benzo[a]anthracene and benzofajpyrene in DPM has also been observed under some conditions
20 and has been used to assess the relative contribution of diesel exhaust to ambient PM.
21 While specific organic carbon species are being identified to help distinguish DPM
22 aerosols from other combustion aerosols, up to 90% of the organic fraction associated with DPM
23 is currently classified as unresolvable complex material. While accounting for the majority
24 (50% to 90%) of the number of particles, ultrafine DPM (5-50 nm) accounts for only 1% to 20%
25 of the mass of DPM. A study conducted by Gertler( 1999) in the Tuscarora Mountain tunnel
26 demonstrated an increase in 20 nm diameter particles as the fraction of diesel vehicles in the
27 tunnel increased from 13% to 78%. The contribution of nuclei mode particles from a freeway on
28 an ambient aerosol size distribution was reported by Whitby and Sverdrup (1980).
29 In general, the major factors that distinguish DPM from ambient PM are: (1) the high
30 portion of elemental carbon, (2) the large surface area associated with the carbonaceous particles
31 in the 0.2 Jim size range, (3) enrichment of certain polycyclic organic compounds, and (4)
32 50%-90% of the number of DPM particles in diesel engine exhaust are in the nuclei mode size
33 range, with a mode of 20 nm. The physical and chemical composition of DPM and changes in
34 DPM chemical composition are discussed in more detail in Section 2.2.8.
35
7/25/00 2-9 DRAFT—DO NOT CITE OR QUOTE
-------
2.2.3. Diesel Emission Standards and Emission Trends Inventory
EPA set a smoke standard for on-road HD diesel engines beginning with the 1970 model
3 year and then added a carbon monoxide (CO) standard and a combined hydrocarbon (HC) and
4 NOX standard for the 1974 model year, as detailed in Table 2-4. Beginning in the 1979 model
5 year, EPA added a HC standard while retaining the combined HC and NOX standard. All of the
6 testing for HC, CO, and NOX was completed using a steady-state test procedure. Beginning in
7 the 1985 model year, EPA added a NOX standard (10.7 g/bhp-hr), dropped the combined HC and
8 NOX standard, and converted from steady-state to transient testing for HC, CO, and NOX
9 emissions. EPA introduced a particulate standard for 1988 model year diesel engines using the
10 transient test (0.6 g/bhp-hr). Transient testing involves running an engine on a dynamometer
11 over a range of load and speed set points.
12 Since the 1985 model year, only the NOX and particulate standards have been tightened
13 for on-road diesel engines. For truck and bus engines, the particulate standard was reduced to
14 0.25 g/bhp-hr in 1991, and it was reduced again in 1994 for truck engines to 0.1 g/bhp-hr. For
15 urban bus engines, the particulate standard was reduced in 1994 to 0.07 g/bhp-hr and again in
16 1996 to 0.05 g/bhp-hr. The NOX standard was reduced to 4.0 g/bhp-hr in 1998 for all on-road
17 diesel engines (bus and truck engines). For 2004, the standards were further lowered in a 1997
18 rulemaking, with limits on nonmethane hydrocarbon (NMHC) and NOX combined, but no further
j)b reductions in CO, PM, or smoke. These lower NMHC-plus-NOx levels are under examination as
20 part of EPA's 1996 technology review. EPA has proposed further reductions in NO^NMHC,
21 and PM for the post-2004 timeframe.
• 22 In December 1997, EPA adopted emission standards for NOX, HC, CO, PM, and smoke
23 for newly manufactured and remanufactured railroad locomotives and locomotive engines. The
24 rulemaking, which takes effect in the year 2000, applies to locomotives originally manufactured
25 in 1973 or after, and any time they are manufactured or remanufactured (locomotives originally
26 manufactured before 1973 are not regulated). Three sets of emission standards have been
27 adopted (Tier 0, 1, and 2); they apply to locomotives and locomotive engines originally
28 manufactured from 1973 through 2001 (Tier 0), from 2002 through 2004 (Tier 1), and in 2005
29 and later (Tier 2) (Table 2-5; see EPA web page at http://www.epa.gov/omswww/ or
30 http://www.diesehiet.com/standards/ for current information on mobile source emission
31 standards). The emissions are measured over two steady-state resi cycles that icpresent two
32 different types of service, including the line-haul (long-distance transport) and switch (involved
33 in all transfer and switching operations in switch yards) locomotives.
34 Emission standards for nonroad equipment are not as stringent as current standards for
on-road equipment and are being phased in within the next decade. Currently, Federal PM
standards exist for nonroad equipment of several horsepower ratings. For equipment between
37 175 and 750 horsepower, the PM standard was set at 0.4 g/bhp-hr in 1996 and will decrease to
7/25/00 2-10 DRAFT—DO NOT CITE OR QUOTE
-------
0.15 g/bhp-hr between 2001 and 2003 depending on the power rating (Table 2-6). This
equipment includes construction, agricultural, and industrial equipment such as bulldozers,
3 graders, cranes, and tractors. The current PM standard for this equipment is only slightly lower
4 than the 0.6 g/bhp-hr PM standard in place for on-road HD diesel engines in the late 1980s.
5 The EPA emission trends report (U.S. EPA, 2000a) provides emission inventories for
Q criteria pollutants (PM10, PM2.5, SO2, NOX, volatile organic compounds [VOC], CO, Pb, and
7 NH3) from point, area, and mobile sources, which indicate how emissions have changed from
8 1970 to 1997. The emission trends are based on the EPA mobile source inventory models
9 MOBILE, PARTS, and the draft NONROAD model. PARTS derives paniculate emission rates
10 for HD diesel vehicles using data generated for new engine certification purposes. PARTS is
11 currently being modified to account for deterioration, in-use emissions, poor maintenance, and
12 tampering effects, all of which would increase emission factors. PM, SO2, NOX, and VOC
13 emissions trends from the report are discussed below. Ambient urban/suburban PM samples
14 rarely reflect the large fraction of natural and miscellaneous sources suggested by the national
15 inventory due to removal of a large portion of these emissions close to their source as well as
16 dispersion from these sources to urban/suburban sites. The removal of natural and
17 miscellaneous PM10 (largely fugitive dust) near their source is a result of the lack of inherent
18 thermal buoyancy, low release height, and interaction with their surroundings (impaction and
(•§) filtration by vegetation). For the summaries presented here, natural and miscellaneous sources
20 are excluded from the national PM and NOX inventories.
21 Mobile sources of PM include both gasoline- and diesel-powered on-road vehicles and a
22 variety of nonroad equipment. Nonroad diesel engine sources include construction equipment,
23 agricultural equipment, marine vessels, locomotives, and other sources. The EPA emission
24 trends report (U.S. EPA, 2000a) indicates that, excluding natural and miscellaneous (mainly
25 fugitive emissions) sources, mobile sources were responsible for 25% of PM10 emissions in
26 1998. Diesel engines (on-road and nonroad combined) were estimated to contribute 72% of
27 mobile source PM 10 emissions and 18% of total PM 10 in 1998 (excluding natural and
28 miscellaneous emissions). Due to the high concentration of fine particles in engine emissions,
29 diesel engines (on-road and nonroad combined) were estimated to contribute 77% of mobile
30 source PM2.5 emissions and 23% of total PM2.5 in 1998 (excluding natural and miscellaneous
31 emissions). If natural and miscellaneous PM2.5 sources are included in the inventory, diesel
32 PM2.5 contributes 6% to the national inventory.
33 Gram per mile particulate emissions from diesel vehicles are much greater than those
34 from gasoline-fueled vehicles, accounting for the large contribution of diesel engine emissions to
35 the national inventory in spite of the smaller number of diesel engines in use. Particulate
^_} emissions (PM10) from gasoline-fueled engines decreased dramatically in 1975 with the
37 widespread introduction of unleaded gasoline. Particulate emissions from diesel highway
7/25/00 2-11 DRAFT—DO NOT CITE OR QUOTE
-------
1 vehicles have decreased recently because of EPA emission standards for new model year HD
2* diesel trucks that were first implemented in 1988 and became increasingly stringent in 1991 and
3 1994, as presented in Table 2-4. A decrease in on-road HD DPM emissions since the mid-1980s
4 is confirmed by in-use vehicle testing, as described in Section 2.2.5. Due to the implementation
5 of existing regulations, DPM emissions from on-road sources are expected to decrease 37% from
6 1998 to 2007; however, nonroad DPM emissions are expected to increase 15% in the same time
7 period (Figure 2-9).
8 The EPA emission trends report (U.S. EPA, 2000a) indicates that annual on-road vehicle
9 PM10 emissions decreased from 397,200 tons to 257,080 tons from 1980 to 1998.1 Passenger
10 car particulate emissions decreased 53% (from 119,000 to 56,000 tons) in this timeframe, while
11 on-road diesel vehicle PM10 emissions decreased 27% (from 208,000 to 152,000 tons) (Figure
12 2-9). Nonroad diesel engine PM10 emissions increased 17% (from 314,000 tons in 1980 to
13 369,000 tons in 1998). Emissions data for PM2.5 are available only for the period from 1990 to
14 1998. Between 1990 and 1998, PM2.5 emissions from mobile sources decreased by 14%,
15 largely as the result of decreased on-road emissions. Diesel engines also contribute to secondary
16 PM formation from NOX and SO2 emissions that are converted to nitrate and sulfate. VOCs from
17 diesel engines also contribute to secondary particle formation.
18 In 1998, 53% of total emitted NOX came from mobile sources, with diesels responsible
M y for 57% of the mobile source contribution. Overall, NOX emissions from mobile sources have
20 remained relatively constant over time, increasing an estimated 7% from 1980 to 1998. While
21 NOX from LD gasoline vehicles decreased from 1980 to 1998, resulting in an overall decrease in
22 on-road NOX emissions of 9%, NOX from diesel trucks and buses increased 7% (from 2,463,390
23 tons in 1980 to 2,630,120 tons in 1998) owing to the use of NOX control defeat devices as
24 discussed in Section 2.2.5. NOX emissions from nonroad diesel engines (including commercial
25 marine and locomotives) have increased 46% (from 3,251,600 tons in 1980 to 4,752,800 tons in
26 1998) (Figure 2-10).
27 About 7% of SO2 came from mobile sources in 1998, with diesels responsible for 74% of
28 that total. EPA regulations on fuel sulfur have significantly reduced SO2 emissions from
29 highway diesels, SO2 emissions from highway diesel engines have decreased 72% (from
30 303,000 tons in 1980 to 85,000 tons in 1998) (Figure 2-11). Similar trends are not apparent for
'Exhaust emissions constitute the majority of PM emissions from mobile sources, with tire and brake wear
contributing the remainder. To compare trends estimates from past years with future projections (which are
provided for exhaust emissions only), the fraction of brake and tire wear would need to be omitted from these
estimates as reported in the emission trends report (U.S. EPA, 2GGOa). On average in the late 1990s 39% and 64%
ot gasoline vehicle particulate emissions originated from exhaust and 95% and 98% of on-road diesel emissions
originated from exhaust for PM10 and PM2.5, respectively.
7/25/00 2-12 DRAFT—DO NOT CITE OR QUOTE
-------
1 nonroad diesels, although in 1998, nonroad diesel engines, excluding commercial marine
/2 vessels, emitted 785,000 tons of SO2, comprising 56% of mobile source SO2 emissions in 1998.
3 Diesel engines are not a large source of VOC emissions compared to gasoline engines.
4 VOC emissions from diesel engines in 1998 were estimated at 2% of the total emissions from all
5 sources. VOC emissions from diesel mobile sources decreased 9% (from 779,000 tons in 1980
6 to 721,000 tons in 1998) (Figure 2-12).
7 Diesel engines are also not a large source of CO emissions compared to gasoline engines.
8 In 1998, mobile sources emitted 79% of all CO and diesel engines accounted for 4% of the
9 mobile source CO. CO emissions from on-road diesel vehicles increased 34% between 1980 and
10 1998, during which time nonroad diesel emissions of CO increased 45% (Figure 2-13).
11
12 2,2 A, Historical Tremdls nm Ofiesd Funs! Us® and Empactf of FTC! Propiginlnes nn Emissioms
13 Use of diesel fuel has increased steadily in the second half of this century. According to
14 statistics from the Federal Highway Administration (1995, 1997), in 1949 diesel fuel was
15 approximately 1% of the total motor fuel used, and in 1995 it was about 18%. Over the same
16 time, diesel fuel consumption increased from about 400 million gallons to 26 billion gallons per
17 year in the United States, an increase by a factor of more than 60 (Figures 2-14 and 2-15).
18 The chemistry and properties of diesel fuel have a direct effect on emissions of regulated
jp) pollutants from diesel engines. Researchers have studied the NOX and DPM effect of sulfur
20 content, total aromatic content, polyaromatic content, fuel density, oxygenate content, cetane
21 number, and T90 on emissions of regulated pollutants. T90 is the 90% distillation point
22 temperature. An increase in T90 has been observed to cause an increase in DPM emissions
23 (Cunningham et al., 1990; Sienicki et al., 1990). Cetane number is a measure of the ignition
24 quality, or ignition delay time, of a diesel fuel. The percent of cetane (less commonly referred to
25 as hexadecane, C16H34) by volume in a blend with alpha-methylnaphthalene (C10H7CH3) defines
26 the cetane number that provides the same ignition delay time as the fuel in use.
27 The chemical makeup of diesel fuel has changed over time, in part because of new
28 regulations and in part because of technological developments in refinery processes. EPA
29 currently regulates on-road diesel fuel and requires sulfur content to be less than 500 ppm for on-
30 road applications and cetane index (a surrogate for actual measurements of cetane number) to be
31 greater than or equal to 40, or the maximum aromatic content to be 35% or less (CFR 40:80.29).
32 California has placed additional restrictions on the aromatic content of diesel fuel (California
33 Code of Regulations, Title 13, Sections 2281-2282) and requires a minimum cetane number of
34 50 and an aromatics cap of 10% by volume, with some exceptions for small refiners and
alternative formulations as long as equivalent emissions are demonstrated. Diesel fuel from
larger refiners is limited to 10% aromatic content, and for three small refiners (a small fraction
37 of diesel sales) to 20% aromatic content. The refiners can also certify a fuel with higher
7/25/00 2-13 DRAFT—DO NOT CITE OR QUOTE
-------
J aromatic content as being emissions-equivalent to the 10% (or 20%) aromatic content fuels by
• 2r performing a 7-day engine dynamometer emissions test. This method is chosen by most, if not
3 all, California refiners, and so a typical California diesel fuel has an aromatic content above
4 20%. Emissions equivalence has been obtained through use of cetane enhancers, oxygenates,
5 and other proprietary additives. Nonroad diesel fuel is not regulated, and consequently, cetane
6 index, aroriiatic content, and sulfur content vary widely with nominal values for cetane number
7 around 43, 31% aromatics, and sulfur approximately 3,000 ppm.
8 Before 1993, diesel fuel sulfur levels were not federally regulated in the United States,
9 although the State of California had such regulations. Industry practices that were in place (e.g.,
10 the ASTMD 975 specification for No. 2 oils) limited sulfur to 0.5%. During the years 1960 to
11 1986, fuel sulfur content showed no chronological increasing or decreasing trends and ranged
12 from 0.23 to 0.28 wt% (NIPER, 1986). A maximum allowable diesel fuel sulfur content in the
13 United States for on-road diesel fuel was established at 0.05 mass % in 1993, in advance of the
14 1994 0.10 g/bhp-hr PM standard for HD on-highway trucks. Nationally, on-road fuels averaged
15 0.032% sulfur in 1994 while nonroad fuels averaged 10-fold the sulfur level of on-road fuel, or
16 0.32% (Dickson and Sturm, 1994). The reduction in diesel fuel sulfur reduced total DPM mass
17 emissions through reduction of sulfate PM (primarily present as sulfuric acid). Approximately
18 1 % to 4% of fuel sulfur is oxidized to SO3, which rapidly forms sulfuric acid in the presence of
(1 9) water vapor in the exhaust (Wall et al., 1987; Khatri et al., 1978; Baranescu, 1988).
20 Considerably higher sulfuric acid PM emissions are possible with diesel exhaust aftertreatment
21 systems containing precious metals (oxidation catalysts, lean NOX catalysts, catalyzed DPM
22 traps). At temperatures over 350 °C to 500 °C (depending on device), SO2 in the exhaust can be
23 oxidized to SO3 and increase sulfuric acid PM emissions (McClure et al., 1992; McDonald et al.,
24 1995; Wall, 1998). Sulfur content remains at unregulated levels for off-highway diesel fuels and
25 fuels used in railroad locomotives.
26 The average cetane number of U.S. diesel fuel declined steadily from 50.0 to 45.1, or
27 about 0.2 per year, from 1960 to 1986 (NIPER, 1986). The decline in cetane number was likely
28 accompanied by an increase in aromatic content and density (Lee et al., 1998). A number of
29 EPA-sponsored studies refer to fuels with nominally 22% aromatics content as "national average
30 fuel" during the 1970s (Hare, 1977; Springer, 1979), while by the 1980s a so-called national
31 average fuel contained 30% aromatics (Martin, 1981a,b). Shelton (1979, 1977) has reported a
32 trend of increasing T90 from 1960 through the late 1970s, which is consistent with increasing
33 density, aromatic content, and polyaromatic content. Unfortunately, aromatic content was not
34 commonly measured before the 1980s. The reason for the decline in cetane number and increase
«in aromatics is that as diesel demand grew, straight-run diesel became a smaller part of the pool
and light-cycle oil from catalytic cracking became available iu make up for the increased
37 demand. Light-cycle oil is high in aromatics and PAHs.
7/25/00 2-14 DRAFT—DO NOT CITE OR QUOTE
-------
Studies measuring the emissions impact of changes in cetane number and aromatic
content for roughly 1990 model year engine technology find that increasing the aromatic content
3 from 20% to 30%, with an accompanying decrease in the cetane number from 50 to 44, results in
4 a 2% to 5% increase in NOX and a 5% to 10% increase in total DPM (McCarthy et al., 1992;
5 Ullman et al., 1990; Sienicki et al., 1990; Graboski and McCormick, 1996). These ranges may
6 be reasonable upper bounds for the effect of changes in fuel quality on NOX and DPM emissions
7 during the years 1960-1990.
8 In the northern United States during wintertime, on-road No. 2 diesel may contain some
9 percentage of No. 1 diesel to improve cold-flow properties. Discussions with refiners indicate
10 that a typical wintertime No. 1 diesel blending level is 15 volume percent; however, this number
11 must be taken as a rough estimate. Blending of No. 1 may lower the aromatic content, resulting
12 in improved emissions performance. Nationally, on-highway No. 1 fuels averaged 17%
13 aromatic content in 1994 (Dickson and Sturm, 1994). Thus, there may also be some small but
14 perceptible seasonal changes in emissions from diesel engines.
15 Railroad-grade diesel fuel is significantly different from on-road fuel and is not required
16 to meet the ASTM specification for No. 2 oils. Diesel fuel oil surveys (Shelton, 1979) show that
17 railroad grade diesels have lower cetane number, higher density, and higher T90. Also, the
18 cetane index for these fuels can be as much as 9 cetane units higher than the cetane number, an
m indication of a high aromatic content in railroad grade diesels.
20 Fuel chemistry is also important for emission of particle-associated PAHs. In studies
21 performed over more than a decade, Williams and Andrews of the University of Leeds have
.22 shown that the solvent extractable PAHs from diesel particulate originate almost entirely in the
23 fuel (Williams et al., 1987; Andrews et al., 1998). The PAH molecules are relatively refractory
24 so that a significant fraction survive the combustion process and condense onto the DPM. These
25 studies have been confirmed by other research groups (Crebelli et al., 1995; Tancell et al., 1995).
26 There is a consensus among these researchers that pyrosynthesis of PAHs occurs only at the
27 highest temperature operating conditions in a diesel engine. Under these conditions, most of the
28 DPM and other pyrolysis products are ultimately burned before exiting the cylinder. These
29 results indicate that emissions of PAHs are more a function of the PAH content of the fuel than
30 of engine technology. For a given refinery and crude oil, diesel fuel PAH correlates with total
31 aromatic content and T90. Representative data on aromatic content for diesel fuels in the United
32 States do not appear to be available before the mid-1980s. However, the decreasing trend in
33 cetane number, increasing trend in T90, and the increasing use of light cycle oil from catalytic
34 cracking beginning in the late 1950s suggest that diesel PAH content has increased over the past
35 40 years. Historical trends in PAH-measured emissions are discussed in Section 2.2.8.2.
7/25/00 2-15 DRAFT—DO NOT CITE OR QUOTE
-------
1 2.2.5. Chronological Assessment of Emission Factors
2/ 2.2.5.1. On-Road Vehicles
3 Numerous studies have been conducted on emissions from in-use on-road HD diesel
4 vehicles. HD vehicles are defined as having a rated gross vehicle weight (GVWR) of greater
5 than 8,500 Ib, and most over-the-road trucks have a GVWR of 80,000 Ib. Emissions of regulated
6 pollutants from these studies have been reviewed (Yanowitz et al., 2000); the review findings,
7 which encompass vehicles from model years 1976 to 1998, are summarized below. In addition,
8 a large amount of engine dynamometer data on HD diesel engines have been published since the
9 mid-1970s. These data are used below to confirm and expand upon the findings from in-use
10 vehicle testing.
11 Figure 2-16 shows chassis dynamometer data for more than 200 different vehicles
12 (approximately one-half of which are transit buses), reported in 20 different published studies, as
13 well as a large amount of additional data collected by West Virginia University (Yanowitz et al.,
14 1999; Warner-Selph and Dietzmann, 1984; Dietzmann et al., 1980; Graboski et al, 1998a;
15 McCormick et al., 1999; Clark et al., 1997; Bata et al., 1992; Brown and Rideout, 1996, Brown
16 et al., 1997; Clark et al., 1995; Dunlap et al., 1993; Ferguson et al., 1992; Gautam et al., 1992;
17 Katragadda et al., 1993; Rideout et al., 1994; Wang et al., 1993,1994; Williams et al., 1989;
18 Whitfield and Harris, 1998; Graboski et al., 1998b; West Virginia University data available on
V19f the World Wide Web at www.afdc.nrel.gov). The results from vehicles tested more than once
20 using the same test cycle, and without any additional mileage accumulated between tests, are
21 averaged and reported as one data point. Buses were tested using the Central Business District
22 (CBD) cycle, while most trucks were tested using the Urban Dynamometer Driving Schedule
23 (UDDS), also known as the Schedule Id cycle. Some of the trucks were tested using the West
24 Virginia 5-peak cycle, which generates considerably lower g/mi emissions than the CBD or
25 UDDS (Yanowitz et al., 1999). Emissions results from vehicles tested under different test cycles
26 or at different points in the engine's life cycle have been reported as separate data points. Note
27 that all NOX mass emissions data are reported as equivalent NO2.
28 Table 2-7 compares the make-up of the fleet of trucks that was tested with the in-use
29 truck fleet according to the 1997 Vehicle Inventory and Use Survey (U.S. Bureau of the Census,
30 1999a). The tested fleet is mostly vehicles in the 33,000-60,000 ib range. Analysis of the tested
31 fleet also shows that the model year distribution is skewed toward ne\ver vehicles. The 1997
32 Vehicle Inventory and Use Survey indicates a flat distribution with roughly the same number of
33 in-use vehicles for each of the model years in the decade preceding 1997. The 1992 Truck
34 .Inventory and Use Survey (U.S. Bureau of the Census, 1995) shows the same trend, as shown in
Figure 2-1. Analysis of odometer mileage for the tested fleet shows that 45% of the vehicles had
less than 50,000 miles at the time of testing. Only 10% of the vehicles had more than 250,000
37 miles. While the mileage distribution ot the in-use fled is unkncv/n, it seer"1? unlikely to be as
7/25/00 2-16 DRAFT—DO NOT CITE OR QUOTE
-------
heavily weighted to low mileage vehicles. Because of the relatively low-mileage of most of the
vehicles tested, deterioration of emissions may not be reflected in the results. Yanowitz and co-
3 workers (2000) report that average emissions of regulated pollutants for vehicles of the different
4 classes listed in Table 2-7 are approximately the same. This is clearly a reflection of the small
5 number of vehicles in the lighter weight classes for this dataset, but it also indicates no real
6 difference in emissions for vehicles in Classes 6-8. The data are mainly for vehicles of 19,500
7 Ib and greater GVWR (Classes 6 and 7 and heavier), and predominantly for vehicles of 33,000 Ib
8 and greater GVWR (Class 8 trucks and buses).
9 Figure 2-16 shows emissions trends ing/mi. Least-squares linear regressions and 95%
10 confidence intervals are plotted on each graph and yield the following equations for predicting
11 emissions trends (applicable to the years 1976-98):
12
13 Log NOX (g/mile) = (Model year * -0.008) + 16.519 R2 = 0.024 (2-1)
14 Log PM (g/mile) = (Model year*-0.044)+ 88.183 R2=0.28 (2-2)
15 Log HC (g/mile) = (Model year * -0.055) + 109.39 R2 = 0.27 (2-3)
16 Log CO (g/mile) = (Model Year*-0.041)+ 82.876 R2 = 0.22 (2-4)
17
18 As shown in Figure 2-16, changes in NOX emissions have been relatively small, with an
^B emission rate averaging about 26 g/mi. The data reported in Figure 2-16 are real-world, in-use
20 emissions measurements and therefore more accurately reflect emission factors than engine test
21 data during this time period. There are two potential causes for the relative constancy of NOX
22 emissions as described by Figure 2-16. The first is emissions deterioration due to engine wear.
23 Weaver and Klausmeier (1988) have shown that diesel engine deterioration results in lower NOX
24 emissions and higher DPM emissions, and this finding has recently been confirmed by
25 McCormick and co-workers (2000). Wear of mechanical devices that limit smoke, fuel pumps,
26 and fuel injectors alters the effective injection timing to decrease NOX. Since deterioration is
27 more a function of maintenance than vehicle age or mileage, deterioration introduces a wide
28 range in NOX emission factors measured in the chassis dynamometer studies. The lack of a
29 decreasing trend in NOX emissions can also be attributed to the use of illegal emissions control
30 defeat devices, an issue addressed by EPA in its recent settlement with the diesel engine
31 manufacturers. The defeat devices produced low NOX emissions on the transient test (HD FTP)
32 but operated in a high NOx/high fuel economy mode in-use under highway cruise type
33 conditions.
34 Figure 2-17 shows engine certification data for NOX emissions reported in the many
§ studies that have employed the transient test over the past 25 years. The engine testing data are
also listed in Table 2-8. The data compiled in Figure 2-17 show a significant decline in NOX
37 emissions, and all engines would appear to meet the regulatory standards for their year of
7/25/00 2-17 DRAFT—DO NOT CITE OR QUOTE
-------
1 manufacture due to the defeat devices. During the period from 1980 to 1997, the EPA emissions
2 trends report (U.S. EPA, 1998a) predicted a decline in NOX emissions from HD diesel vehicles
3 since these data are based on engine test data. The emissions trend includes the growth in
4 vehicle miles traveled over time as well as changes in emission factors. The more recent trends
5 inventory (U.S. EPA, 2000a, discussed earlier), accounts for the defeat devices and accordingly
6 demonstrates a slight increase in NOx emissions from on-road HD diesel vehicles in the period
7 from 1990 to 1998.
8 DPM, CO, and THC emissions, although widely variable within any model year, have
9 shown a pronounced declining trend (Figure 2-16). DPM emissions from chassis dynamometer
10 tests decreased from an average of 3-4 g/mi in 1977 to an average of about 0.5 g/mi in 1997,
11 suggesting a decrease in DPM emissions of a factor of about 6. Note that these data are for
12 vehicles or engines tested on in-use or industry average fuel at the time they were tested.
13 Indications are that the observed decline in DPM is caused primarily by changes in engine
14 technology that often result from emission standards, as well as by the lowering of on-road
15 diesel fuel sulfur content in 1993.
16 While a substantial decreasing trend in DPM emissions from in-use chassis dynomometer
17 testing and engine testing (Figure 2-18) is evident, these data reflect a wide range in emission
18 factors within any given model year. For example, emission factors for model year 1996 range
19 from less than 0.1 g/mi to over 1 g/mi (Yanowitz et al., 2000; Graboski et al., 1998b). The high
20 variability in DPM emissions measured in the chassis dynamometer tests is observed due to
21 several factors, including differences in measurement methods and test conditions at the various
22 testing facilities, deterioration, and engine-to-engine variation. While there can be excellent
23 agreement between chassis dynamometer testing facilities (Graboski et al., 1998a), there is no
24 standard HD chassis dynamometer Federal test procedure, and no detailed procedures for such
25 testing are described in any authoritative source such as the Code of Federal Regulations, which
26 does contain such procedures for engine dynamometer testing used for EPA emission
27 regulations. Therefore, each facility has developed its own approach to HD testing. Clark et al.
28 (1999) report that the test cycle can have a substantial effect on DPM emissions, with higher
29 DPM emissions reported from test cycles that incorporate full power accelerations. Test cycles
30 incorporating full power accelerations reflect urban HD vehicle driving for several types of
31 vehicles (garbage trucks, buses) operaling in urban areas Clark et al. (1999) also report that
32 aggressive acceleration produces higher DPM emission rates compared to conservative
33 acceleration, and Clark and co-workers suggest that real in-use driving is more likely to mimic
34 aggressive acceleration. While currently unquantified, it is generally believed that the majority
35 of PPM is generated under transient conditions such as heavy acceleration.
35 Weaver and Klausmeier (1988) have examined potential causes and frequency of DPM
37 emissions deterioration for in-use HD diesel vehicles. Potential causes include manufacturing
7/25/00 2-18 DRAFT—DO NOT CITE OR QUOTE
-------
defects and malfunctions such as retarded timing, fuel injector malfunction, smoke limiting
mechanism problems, clogged air filter, wrong or worn turbocharger, clogged intercooler, engine
3 mechanical failure, excess oil consumption, and electronics that have been tampered with or
4 have failed. The recent report by McCormick and co-workers (2000) indicates that many of
5 these malfunctions can have very large effects on DPM emissions, resulting in DPM increases of
6 typically 50% to 100%. While Yanowitz and co-workers (1999) found that DPM emissions
7 were positively correlated with odometer mileage for a fleet of 21 vehicles, it is more likely that
8 the vehicle state of maintenance will be more important for determining the degree of emissions
9 deterioration than mileage. In fact, in a similar analysis performed on the chassis dynamometer
10 results included in the review of Yanowitz et al. (2000), DPM emissions could not be correlated
11 with odometer mileage. Differences in testing methods between various facilities, as well as
12 varying states of maintenance for vehicles of the same mileage and model year probably account
13 for this lack of correlation.
14 It is difficult, given current information, to quantitatively assess the contribution of high-
15 emitting or smoking diesel vehicles to ambient DPM. Emission models used to prepare diesel
16 paniculate emission inventories presently do not account for deterioration. The relative
17 contribution of high-emitting diesel vehicles to the total mass and overall chemical composition
18 of diesel particulates is presently being quantified. Some studies report numerous smoking
jjn diesel trucks. A study of the smoke opacity based inspection and maintenance program in
20 California found failure rates of 20% and higher, suggesting that high-emitting vehicles are not
21 uncommon (CARB/EEAI, 1997). In the Northeast, smoke opacity testing conducted on 781 HD
22 trucks found that 15% of the vehicles failed the smoke standard (40% opacity for 1991 and
23 newer HD diesel vehicles and 50% opacity for pre-1991 HD diesel vehicles) (Cooper, 1999).
24 While the correlation between smoke and paniculate emissions tends to be qualitative or
25 semiquantitative (discussed below), there is a good correlation between opacity and elemental
26 carbon concentrations, and it is expected that high-emitting diesel vehicles may be an important
27 part of the DPM emission inventory.
28 Others have attempted to determine if the effects of deterioration could be detected for
29 in-use vehicles. In a study of 21 vehicles (Yanowitz et al., 1999), a linear multivariate
30 regression analysis found that DPM emissions were positively correlated with odometer mileage
31 (several other correlation factors were also identified, including model year). A similar analysis
32 performed on the chassis dynamometer results included in the review of Yanowitz et al. (2000)
33 found that DPM emissions could not be correlated with odometer mileage, probably because of
34 differences in testing methods between the various facilities.
35 Other approaches for measuring emissions from in-use on-road diesel vehicles include
/jlSfe \
\™J tunnel tests and remote sensing, the latter of which measures gaseous, but not DPM, emissions.
37 The literature reports of those studies are summarized in Tables 2-9 and 2-10. Several tunnel
7/25/00 2-19 DRAFT—DO NOT CITE OR QUOTE
-------
1 test studies have reported DPM emission factors (Pierson and Brachaczek 1976; Szkarlat and
2 Japar 1983; Pierson et al., 1983; Kirchstetter et al., 1999; Gertler et al., 1995, 1996).
3 The method for determining emission rates for vehicles traveling through a tunnel are
4 explained in detail by Pierson et al. (1996). Briefly, the emissions of a species are determined by
5 measuring the concentration of a pollutant entering and leaving a tunnel along with knowledge
6 of the cross section of the tunnel and measurements of the wind flux at the inlet and outlet of the
7 tunnel. The emission rate is calculated by dividing the mass of the pollutant by the number of
8 vehicles that passed through the tunnel and the length of the tunnel. The diesel and gasoline
9 vehicle contributions to the total emission of the pollutant are separated by a simple regression
10 analysis where the intercepts (100% HD and 100% LD) are the diesel and gasoline emission
11 rates, respectively.
12 Emission factors from tunnel studies provide a snapshot of real-world emissions under
13 driving conditions experienced in the tunnel and reflect emission factors representative of the
14 mix of in-use vehicles and the atmospheric dilution and short-term transformation processes of
15 diesel exhaust. Emission factors derived from tunnel studies are often used as one source of
16 information to study the impact of improved technology and fleet turnover on emissions because
17 they allow random sampling of large numbers of vehicles, including a range of ages and
18 maintenance conditions. However, tunnel studies are limited in that they represent driving
19 conditions on a single roadway passing through a tunnel and represent mostly steady-state
20 driving conditions, whereas most DPM is generated during transient modes of operation; also
21 tunnel studies do not include cold-start operations. Both of these factors need to be assessed to
•22 understand emission rates for DPM to which people are exposed (U.S. EPA 1992, 1995). DPM
23 emission factors from in-use fleets derived from tunnel studies in the 1970s and 1980s compared
24 with the 1990s suggest approximately a fivefold decrease in DPM mass emission factors, with
25 the most recent data from 1999 reporting an emission factor of 0.29 g/mi for the on-highway HD
26 diesel fleet (Figure 2-19).
27 Emission factors vary substantially for the various tunnels, with NOX emissions ranging
28 from 9.7 to 23.8 g/mi in the 1990s, CO emissions ranging from 6 to 14 g/mi, and THC emissions
29 ranging from 0.16 to 2.55 g/mi.
30 Remote sensing reports emission factors in terms of pollutant emissions per unit of fuel,
31 not on a per-mile basis. Agreement between remote sensing and tunnel studies for NOX
32 emissions is reasonably good for the fleet as a whole, suggesting an average level for the fleet of
33 about 130 g/'gal, comparable to the average emissions factor measured in chassis dynamometer
34 studies (remote sensing can measure emissions from an individual vehicle, while tunnel studies
35 . measure emissions from the fleet as a whole). Generally, chassis dynamometer tests and engine
35 dynamometer test results are corrected for ambient humidity, in accordance "with the Federal
37 Test Procedure (CFR 40, Subpart N). Tunnel tests and remoie sensing icsts have typically not
7/25/00 2-20 DRAFT—DO NOT CITE OR QUOTE
-------
1 included corrections for humidity. Appropriate humidity corrections for NOX and DPM can be
2 greater than 20% and 10%, respectively (or a total difference of more than 45% and 20%,
3 respectively, between low- and high-humidity areas), under normally occurring climatic
4 conditions. Additionally, the remote sensing literature has not addressed how to determine the
5 correct value for the NO/NOX ratio, and there is reason to believe that this value may differ
6 systematically from site to site, although almost all of the NOX is NO as it leaves the vehicle.
7 In addition to the humidity correction discussed above, several factors must be taken into
8 account when comparing DPM measurements from tunnel tests to chassis dynamometer
9 measurements (Yanowitz et al., 2000): (1) Chassis testing measures only tailpipe emissions;
10 tunnel tests can include emissions from other sources (tire wear, etc.), and (2) tunnel tests
11 typically measure emissions under steady-speed freeway conditions, whereas most chassis
12 dynamometer tests are measured on cycles that are more representative of stop-and-go urban
13 driving conditions. This latter limitation also applies to remote sensing readings, which measure
14 instantaneous emissions versus emissions over a representative driving cycle.
15 Because THC emissions for diesel vehicles are very low in total mass in comparison to
16 gasoline vehicles, tunnel test results for THC have a high degree of uncertainty. A regression
17 analysis to determine the contribution of the limited number of HD vehicles to THC emissions is
18 unstable; small errors in the total measurements can change estimates substantially. Similarly,
)9 CO emissions are comparable to automobile emissions on a per-vehicle-mile basis, but since
20 there are generally many more automobiles than HD diesels in tunnel tests, CO measurements
21 from diesels may also have a high degree of uncertainty.
22 As the discussion above indicates, there is a reasonable amount of data upon which to
23 base emission factor estimates for late 1970s and later HD vehicles. However, almost no
24 transient test data are available on engines earlier than the mid-1970s. Nevertheless, there are
25 engine design factors that allow reasonable assumptions to be made about emissions from
26 engines in the mid-1950 to 1970 timeframe. For example, the same means of controlling the
27 engine's air-fuel ratio was used over the period from 1950 to 1970. That is, air-fuel ratios richer
28 than the "smoke limit" of about 22:1 were avoided by fuel system control. This air-to-fuel ratio
29 control, in essence, formed an "upper limit" on DPM emissions and was implemented before the
30 advent of EPA smoke standards for customer satisfaction reasons. There is only a qualitative
31 correlation between smoke (measured as smoke opacity) and particle emissions over the
32 transient driving cycle, but there is semiquantitative correlation between smoke and particle
33 emissions over steady-state operating modes (McGuckin and Rykowski, 1981). Additionally,
34 there is good correlation between smoke and elemental carbon emissions (Alkidas, 1984). The
35 fact that engines, turbocharged or not, were controlled to avoid smoky operation makes it
5>> reasonable to assume that they had emissions roughly at the mid-1970s level for many years
37 before this timeframe. Other than the increased use of turbochargers, HD diesel engine
7/25/00 2-21 DRAFT—DO NOT CITE OR QUOTE
-------
1 technology was reasonably stable since the 1950s, with most engines having direct injection and
2 a nearly constant fraction of two-stroke versus four-stroke engines. Thus, it is likely that the
3 emission factors reported above for mid-1970s engines adequately represent the engines in use in
4 the 1950—70 timeframe. Note that the impact of engine technology developments on emissions
5 is specifically discussed in Section 2.2.6.
6
7 2.2.5.2. Locomotives
8 Locomotive engines generally range from 1,000 horsepower up to 6,000 horsepower.
9 Similar to the much smaller truck diesel engines, the primary pollutants of concern are NOX,
10 DPM, CO, and HC. Unlike truck engines, most locomotive engines are not mechanically
11 coupled to the drive wheels. Because of this decoupling, locomotive engines operate in specific
12 steady-state modes rather than the continuous transient operation normal for trucks. Because the
13 locomotive engines operate only at certain speeds and torques, the measurement of emissions is
14 considerably more straightforward for locomotive engines than for truck engines. Emissions
15 measurements made during the relatively brief transition periods from one throttle position to
16 another indicate that transient effects are very short and thus could be neglected for the purposes
17 of overall emissions estimates.
18 Emissions measurements are made at the various possible operating modes with the
19 engine in the locomotive, and then weighting factors for typical time of operation at each throttle
20 position are applied to estimate total emissions under one or more reasonable operating
21 scenarios. In the studies included in this analysis, two scenarios were considered: line-haul
22 (movement between cities or other widely separated points) and switching (the process of
23 assembling and disassembling trains in a switchyard).
24 The Southwest Research Institute made emissions measurements for three different
25 engines in locomotives in 1972 (Hare and Springer, 1972) and five more engines in locomotives
26 using both low- and high-sulfur fuel in 1995 (Fritz, 1995). Two engine manufacturers (the
27 Electro-Motive Division of GM, and GE Transportation Systems) tested eight different engine
28 models and reported the results to EPA (U.S. EPA, 1998b). All available data on locomotives
29 are summarized in the regulatory impact assessment and shown in Figure 2-20.
30
31 2.2.6. Engine Technology Description and Chronology
32 NOX emissions, DPM emissions, and brake-specific fuel consumption (BSFC) are among
33 the parameters that are typically considered during the development of a diesel engine. Many
34 engine variables that decrease NOX can also increase DPM and BSFC. One manifestation of the
35 interplay among NOX, DPM, and BSFC is that an increase in combustion temperatures will tend
35 to increase NO formation via the Zeldovich mechanism. Higher temperatures will also often
37 improve thermal efficiency, can improve BSFC, and can increase the rate of DPM oxidation.
7/25/00 2-22 DRAFT—DO NOT CITE OR QUOTE
-------
1 thus lowering DPM emissions. One example of this is the tradeoff of DPM emissions and BSFC
2 versus NOX emissions with fuel injection timing. Many recent advances in reducing the
3 emissions of diesel engines without aftertreatment are combinations of technologies that provide
4 incremental improvements hi the tradeoffs among these emissions and fuel consumption. The
5 sum total, however, can be considerable reductions in regulated emissions within acceptable
6 levels of fuel consumption.
7 The majority of current HD diesel truck engines certified for use in the United States
8 utilize:
9 • A four-stroke cycle
10 • Direct-injection, high-pressure (1,200 bar to >2,000 bar) fuel injection systems with
11 electronic control of injection timing and, in some cases, injection rate
12 • Centrally located multihole injection nozzles
13 • Three or four valves per cylinder
14 • Turbochargers
15 • In many cases, air-to-air aftercooling
16 • In some cases, the use of an oxidation catalyst.
17
18 These features have phased into use with HD truck engines because they offer a
(19 relatively good combination of fuel consumption, torque-rise, emissions, durability, and the
20 ability to better "tune" the engines for specific types of applications. Fuel consumption, torque-
21 rise, and drivability have been maintained or improved while emissions regulations have become
22 more stringent. Many Class 8a and 8b diesel truck engines are now capable of 700,000 to
23 1,000,000 miles of driving before their first rebuild and can be rebuilt several times because of
24 their heavy construction and the use of removable cylinder liners. These engines are expected to
25 last longer and therefore have a useful life longer than the regulatory estimate of full useful life
26 for HD engines (-1,000,000 miles) previously used by EPA (for 1980 engines that were driven
27 less than 300,000 miles between rebuilds and were rebuilt up to three times). Current four-
28 stroke locomotive engines use engine technology similar to on-highway diesel engines, except
29 that electronic controls have only recently been introduced.
30 It is difficult to separate the components of current high-speed diesel engines for
31 discussion of their individual effects on emissions. Most of the components interact in numerous
32 ways that affect emissions, performance, and fuel consumption.
33
34 2.2.6.1. Indirect and Direct Injection High-Speed Diesel Engines
35 Prior to the 1930s, diesel engine design was limited to relatively low-speed applications
because sufficiently high-pressure fuel injection equipment was not available. With the advent
37 of high-speed and higher pressure pump-line-nozzle systems, introduced by Robert Bosch in the
7/25/00 2-23 DRAFT—DO NOT CITE OR QUOTE
-------
1 1930s, it became possible to inject the fuel directly into the cylinder for the first tune, although
2 indirect injection (IDI) diesel engines continued in use for many years. As diesels were
3 introduced into the heavy truck fleet in the 1930s through the 1950s, both IDI and direct
4 injection (DI) naturally aspirated variants were evident. A very low-cost rotary injection pump
5 technology was introduced by Roosa-Master in the 1950s, reducing the cost of DI systems and
6 allowing their introduction on smaller displacement, higher speed truck engines. After this time,
7 only a small fraction of truck engines used an IDI system.
8 DI diesel engines have now all but replaced IDI diesel engines for HD on-highway
9 applications.2 IDI engines typically required much more complicated cylinder head designs but
10 generally were capable of using less sophisticated, lower pressure injection systems with less
11 expensive single-hole injection nozzles. IDI combustion systems are also more tolerant of lower
12 grades of diesel fuel. Fuel injection systems are likely the single most expensive component of
13 many diesel engines. Caterpillar continued producing both turbocharged and naturally aspirated
14 IDI diesel engines for some on-highway applications into the 1980s. Caterpillar and Deutz still
15 produce engines of this type, primarily for use in underground mining applications. IDI
16 combustion systems are still used in many small-displacement (<0.5 L/cylinder), very high-
17 speed (>3,000 rpm rated speed) diesel engines for small nonroad equipment (small imported
18 tractors, skid-steer loaders), auxiliary engines, and small generator sets, and they were prevalent
19 in diesel automotive engines in the 1980s; IDI designs continue to be used in automotive diesel
20 engines.
21 IDI engines have practically no premixed burn combustion and thus are often quieter and
22 have somewhat lower NOX emissions than DI engines. Electronic controls, high-pressure
23 injection (e.g., GM 6.5), and four-valve/cylinder designs (e.g., the six-cylinder Daimler LD
24 engine) can be equally applied to IDI diesel engines as in DI, but they negate advantages in cost
25 over DI engines. DI diesel engines of the same power output consume 15% to 20% less fuel
26 than IDI engines (Heywood, 1988). Considering the sensitivity of the HD truck market to fuel
27 costs, this factor alone accounts for the demise of IDI diesel engines in these types of
28 applications. Throttling and convective heat transfer through the chamber-connecting orifice,
29 and heat rejection from the increased surface area of IDI combustion systems, decrease their
30 efficiency and can cause cold-start difficulties when compared to DI designs. Most IDI diesel
31 engine designs require considerably higher than optimum compiession ratios (from an efficiency
32 standpoint) to aid in cold-starting (19:1 to 21:1 for IDI engines vs. -15:1 to 17:1 for DI engines).
33 Because of the early introduction of DI technology into truck fleets, it is likely that by the
34 end of the 1960s, only a small fraction of the HD diesel engines sold for on-highway use were
11 he GM rovvcrtrzir-'AM funeral 6.5L electronically controlled, turbocharged IDI-swirl chamber engine, certified
as a light HD diesel truck engine, is the last remaining HD on-hignway IDI engine am
:5/00 2-24 DRAFT—DO NOT CITE OR QUOTE
-------
1 IDI engines. It is unlikely that the shift from IDI to DI engine designs through the 1950s and
!2 1960s occurred rapidly and likely that this shift had little significant impact on emissions.
3 Springer (1979) reports a comparison of nearly identical Caterpillar 3406 engines (turbocharged
4 and aftercooled) hi DI and IDI configurations tested on an engine dynamometer under steady-
5 state conditions, which limits the usefulness of these data. There is no significant difference in
6 emissions of DPM, soluble organic fraction (SOF), aldehydes, or DPM-associated
7 benzo[a]pyrene (Table 2-8). Note that IDI designs continue to be used hi automotive diesel
8 engines.
9
10 2.2.6.2. Injection Rate
11 Decreasing the duration of diffusion combustion and promoting elemental carbon
12 oxidation during the expansion stroke can reduce formation of elemental carbon agglomerates
13 (Stone, 1995) and reduce the particulate carbon fraction at high load (Needham et al., 1989).
14 Both of these effects are enhanced by increasing the fuel injection rate. The primary means of
15 accomplishing this is by increasing fuel injection pressure. In 1977 Robert Bosch introduced a
16 new type of high-pressure pump capable of producing injection pressures of 1,700 bar at the
17 nozzle (Voss et al., 1977). This increased fuel injection pressure by roughly a factor of 10. Unit
18 injection, which combines each fuel injection nozzle with individual cam-driven fuel pumps, can
9 achieve very high injection pressures (>2,000 bar). The first combination of unit injectors with
20 electronically controlled solenoids for timing control was offered in the United States by Detroit
21 Diesel Corporation in the 1988 model year (Hames et al., 1985). Replacement of the injection
22 cam with hydraulic pressure, allowing a degree of injection rate control, was made possible with
23 the hydraulic-electronic unit injection jointly developed by Caterpillar and Navistar, introduced
24 on the Navistar T444E engine (and variants) in 1993.
25 It is widely known that high fuel injection pressures have been used to obtain compliance
26 with the PM standards that went into effect in 1988 (Zelenka et al., 1990). Thus, it is likely that
27 a transition to this technology began hi the 1980s, with the vast majority of new engine sales
28 employing this technology by 1991, when the 0.25 g/bhp-hr Federal PM standard went into
29 effect.
30 The use of electronic control of injection rate is rapidly increasing on medium HD diesel
31 engines. Engines are currently under development, perhaps for 2002—2004 introduction, that use
32 common-rail fuel injection systems with even more flexible control over injection pressure and
33 timing than previous systems.
34 Increased injection rate and pressure can significantly reduce elemental carbon
35 emissions, but it can also increase combustion temperatures and cause an increase in NOX
6 emissions (Springer, 1979; Watson and Janota, 1982; Stone, 1995). Low NOX, low DPM, and
7/25/00 2-25 DRAFT—DO NOT CITE OR QUOTE
-------
1 relatively good BSFC and brake mean engine pressure (BMEP) are possible when combined
2 with turbocharging, aftercooling, and injection timing retard.
3
4 2.2.6.3. Turbocharging, Charge-Air Cooling, and Electronic Controls
5 Use of exhaust-driven turbochargers to increase intake manifold pressure has been
6 applied to both IDI and DI diesel engines for more than 40 years. Turbocharging can decrease
7 fuel consumption compared to a naturally aspirated engine of the same power output.
8 Turbocharging utilizes otherwise wasted exhaust heat and pressure to generate intake boost. The
9 boosted intake pressure effectively increases air displacement and increases the amount of fuel
10 that can be injected to achieve a given fuel-air equivalence ratio. Turbocharging increases the
11 power density of an engine. Boosting intake pressure via turbocharging and reducing fuel-to-air
12 ratio at a constant power can significantly increase both intake temperatures and NOX emissions.
13 Increased boost pressure can significantly reduce ignition delay, which reduces VOC and DPM
14 SOF emissions (Stone, 1995) and increases the flexibility in selection of injection timing.
15 Injection timing on turbocharged engines can be retarded further for NOX emission control with
16 less of an effect on DPM emissions and fuel consumption. This allows a rough parity in NOX
17 emissions between turbocharged (non-aftercooled) and naturally aspirated diesel engines
18 (Watson and Janota, 1982).
19 Turbocharging permits the use of higher initial injection rates (higher injection pressure),
20 which can reduce particulate emissions. Although this may offer advantages for steady-state
21 operation, hard accelerations can temporarily cause overly fuel-rich conditions because the
22 turbocharger speed lags behind a rapid change in engine speed (turbo-lag). This can cause
23 significant increases in DPM emissions during accelerations. Before the advent of electronic
24 controls, the effect of acceleration on DPM emissions could be limited by mechanically delaying
25 demand for maximum fuel rate with a "smoke-puff eliminator." Since this device also limited
26 engine response, there was considerable incentive for the end-users to remove or otherwise
27 render the device inactive. Charge-air cooling, for example, using an air-to-air aftercooler (air-
28 cooled heat exchanger) between the turbocharger compressor and the intake manifold, can
29 greatly reduce intake air and peak combustion temperatures. When combined with injection
30 timing retard, charge-air cooling allows a significant reduction in NOX emissions with acceptable
31 BSFC and DPM emissions when compared to either non-aftercooied or naturally aspirated diesel
32 engines (Hardenberg and Fraenkle, 1978; Pischinger and Cartellieri, 1972; Stone, 1995). The
33 use of charge-air cooling effectively shifts the NOX-DPM tradeoff curve, as shown in Figure 2-
34 21.
35 Electronic control of fuel injection timing allowed engine manufacturers to carefully
36 tailor the start and length of the fuel injection events much more precisely than through
37 mechanical means. Because of this, newer on-highway turbocharged truck engines have
7/25/00 2-26 DRAFT—DO NOT CITE OR QUOTE
ii
-------
virtually no visible smoke on acceleration (although emissions of DPM are substantial during
this driving mode). Electronic controls also allowed fuel injection retard under desirable
3 conditions for NOX reduction, while still allowing timing optimization for reduced VOC
4 emissions on start-up, acceptable cold-weather performance, and acceptable performance and
5 durability at high altitudes. Previous mechanical unit injected engines (e.g., the 1980s Cummins
6 LI 0, the non-DDEC DDC 6V92) were capable of reasonably high injection pressures, but they
7 had fixed injection timing that only varied based on the hydraulic parameters of the fuel system.
8 Many other engines with mechanical in-line or rotary injection pumps had only coarse injection
9 timing control or fixed injection timing.
10 Precise electronic control of injection timing over differing operating conditions also
11 allowed HD engine manufacturers to retard injection timing to obtain low NOX emissions during
12 highly transient urban operation, similar to that found during emissions certification. HD engine
13 manufacturers also advanced injection timing during less transient operation (such as freeway
14 driving) for fuel consumption improvements (~3% to 5%) at the expense of greatly increased
15 NOX emissions (approximately three to four times regulated levels). This particular situation
16 resulted in the recent consent decree settlements between the Federal Government and most of
17 the HD engine manufacturers to assure effective NOX control in all driving conditions, including
18 on-highway high speed steady-state driving.
(ll) Turbocharged engines entered the market very slowly beginning in the 1960s. Data for
20 DPM emissions from naturally aspirated engines of model years 1976 to 1983 are compared
21 with DPM emissions from turbocharged engines in Figure 2-22. It is apparent that there is no
22 consistent difference in DPM emissions between turbocharged and naturally aspirated engines.
23 While not plotted, the data also show no difference in emissions of NOX, DPM SOF, or DPM-
24 associated benzo[a]pyrene and 1-nitropyrene.
25 Charge-air cooling was introduced during the 1960s and was initially performed in a heat
26 exchanger using engine coolant. Cooling of the charge-air using ambient air as the coolant was
27 introduced into heavy trucks by Mack in 1977 with production of the ETAY(B)673A engine
28 (Heywood, 1988). Use of ambient air allowed cooling of the charge-air to much lower
29 temperatures. Most HD diesel engines sold today employ some form of charge-air cooling, with
30 air-to-air aftercooling being the most common. Johnson and co-workers (1994) have presented a
31 comparison of similar engines that differ in that the charge-air is cooled by engine coolant (1988
32 engine) and by ambient air, with a higher boost pressure for the second (1991 engine). The 1991
33 engine also used higher pressure fuel injectors. The 1991 engine exhibited both lower DPM
34 emissions (50% lower than the 1988 engine) and lower NOX emissions. Higher injection
5 pressure is likely to have enabled the reduced DPM emissions, while the lower charge-air
temperature and the ability to electronically retard the injection timing under some conditions
37 likely enabled the lower NOX emissions.
7/25/00 2-27 DRAFT—DO NOT CITE OR QUOTE
-------
1 It is apparent on the basis of both the literature and certification data that turbochargers
2 with aftercoolers can be used in HD engines in conjunction with other changes to produce a
3 decrease in emissions. Upon the advent of a NOX standard in 1 985, NOX was probably reduced
4 on the order of 10% to 30% in turbocharged aftercooled engines with retarded injection timing.
5 This decrease in emissions is not evident in the in-use chassis testing data because of
6 deterioration and the use of illegal defeat devices as described above. Overall, it is expected that
7 engines in the 1950 to mid-1970s timeframe would have similar DPM emission rates, while
8 post- 1970 engines would have somewhat lower DPM emission rates.
9
1 0 2.2.6.4. Two-Stroke and Four-Stroke High-Speed Diesel Engines
1 1 A detailed discussion of the two- and four-stroke engine cycles can be found in the
1 2 literature (Heywood, 1988; Taylor, 1990; Stone, 1995). Nearly all high-speed two-stroke diesel
1 3 engines utilize uniflow scavenging assisted by a positive-displacement blower (Figure 2-23).
1 4 Uniflow-scavenged two-stroke diesels use poppet exhaust valves similar to those found hi four-
1 5 stroke engines. The intake air enters the cylinder through a pressurized port in the cylinder wall.
16 A crankshaft-driven, positive-displacement blower (usually a roots-type) pressurizes the intake
1 7 port to ensure proper scavenging. A turbocharger may be added to the system to provide
1 8 additional boost upstream of the blower at higher speeds and to reduce the size and parasitic
1 9 losses associated with the positive-displacement blower.
20 Two-stroke diesel engines can achieve efficiency comparable to four-stroke counterparts
2 1 and have higher BMEP (torque per unit displacement) (Heywood, 1988). It is useful to note that
i
22 the two-stroke cycle fires each cylinder once every revolution, while the four-stroke cycle fires
23 every other revolution. Thus, for a given engine size and weight, two-strokes can produce more
24 power. However, two-stroke diesel engines are less durable than their four-stroke counterparts.
25 Lubricating oil is transferred from the piston rings to the intake port, which causes relatively
26 high oil consumption relative to four-stroke designs. Durability and low oil consumption are
27 desirable for on-highway truck applications. This may be why four-stroke engines have been
28 favored for these applications since the beginning of dieselization in the trucking industry, with
29 the notable exception of urban bus applications. Although it is no longer in production, the
30 Detroit Diesel 6V92 series of two-stroke diesel engines is still the most popular for urban bus
3 1 applications, where the high power density allows the engine to be more easily packaged within
32 limited spaces. The primary reason that two-stroke engines like the 6V92 are no longer offered
33 for urban bus applications is excessive DPM emissions. The lubricating oil control with two-
34 strokes tends to be lower for two-stroke engines, and therefore, emissions have higher VOC and
35 organic DPM emissions relative to four-stroke designs. This was particularly problematic for
35 nrhan hns applications because urban bus engines must meet tighter Federal and California PM
37 emissions standards. The current urban bus PM standard (0.05 g/bhp-hr) is one-half of the
FJP A F
-------
current on-highway HD diesel engine PM standard, although EPA is in the process of proposing
more strict standards for HD diesel truck engines along with further reductions in diesel fuel
3 sulfur levels. No two-stroke diesel engine designs have been certified to meet the most recent
4 urban bus PM emissions standards, and Detroit Diesel Corporation has not certified a two-stroke
5 diesel engine for on-highway truck use since 1995.
6 A comprehensive review of emissions from hundreds of vehicles (1976-98 model years)
7 that had been tested on chassis dynamometers found that DPM emissions vary substantially
8 within a given model year and that within that variation there are no discernable differences in
9 DPM emissions between two- and four-stroke vehicles (Figure 2-24) (Yanowitz et al., 2000).
10 DPM emission factors reported for engine tests also indicate that two- and four-stroke engines
11 have comparable emission factors, since these engines all had to meet the same regulatory
12 standard (Figure 2-25). In contrast to DPM emissions, evidence suggests that mid-1970s two-
13 stroke engines exhibited very high SOF levels compared to four-stroke engines during that
14 timeframe with later model years showing similar SOF emissions for two- and four-stroke
15 engines (Figure 2-26). For aldehydes, benzo[a]pyrene, and 1-nitropyrene, data are available for
16 only one two-stroke engine, but they indicate no significant difference in emissions from
17 comparable model year four-stroke engines. Overall, regulated emissions changes due to
18 changing proportions of two- and four-stroke engines in the in-use fleet do not appear to have
^ influenced DPM emission levels, but the transition to four-stroke engines in the 1970s would
20 have decreased the fraction of SOF associated with the DPM. It appears that the proportion of
21 two-stroke engines in the in-use fleet was relatively constant until the late 1980s when it began
22 to decline.
23
24 2.2.7. Air Toxic Emissions
25 Heavy-duty diesel vehicle exhaust contains several substances that are known, likely, or
26 possible human or animal carcinogens, or that have serious noncancer health effects. These
27 substances include, but are not limited to, benzene, formaldehyde, acetaldehyde, 1,3-butadiene,
28 acrolein, dioxin, PAH and nitro-PAH (the complete list of chemically characterized compounds
29 present in diesel exhaust is provided in Section 2.3.1). Very few historical data are available to
30 examine changes in emission rates over time. In this section, trends in aldehyde emissions over
31 time and a summary of dioxin emission factors are presented. PAH and nitro-PAH emission
32 factors are discussed in Section 2.2.8.2.
33
34 2.2.7.1. Aldehyde Emissions
35 Among the gaseous components emitted by diesel engines, the aldehydes are a
Pb particularly important component because they comprise an important fraction of the gaseous
37 emissions and they are probable carcinogens that also produce noncancer health effects.
7/25/00 2-29 DRAFT—DO NOT CITE OR QUOTE
-------
1 Formaldehyde comprises the majority of the aldehyde emissions (65% to 80%), with
2 acetaldehyde being the second most abundant aldehyde in HD diesel emissions. Total aldehyde
3 emissions reported from chassis dynamometer testing suggest that aldehyde emissions have
4 declined since 1980; however, only two tests reported aldehydes from engines made after 1985
5 (Figure 2-27). Engine dynamometer studies also suggest a downward trend in the emissions of
6 aldehydes in the time period from 1976 to 1994 (Figure 2-28). Engine dynamometer studies
7 report aldehyde emission levels of 150-300 mg/bhp-hr for late 1970s engines with no significant
8 effect of turbocharging, or IDI versus DI. High-pressure fuel injection may have resulted in a
9 marginal increase in aldehyde emissions (Springer, 1979). By comparison, 1991 model year
10 engines (DI, turbocharged) exhibited aldehyde emissions in the 30-50 mg/bhp-hr range
11 (Mitchell etal., 1994).
12
13 2.2.7.2. Dioxin and Furans
14 Ballschmiter et al. (1986) reported detecting polychlorinated dibenzo-p-dioxins (CDDs)
15 and polychlorinated dibenzofurans (CDFs) in used motor oil and thus provided some of the first
16 evidence that CDDs and CDFs might be emitted by the combustion process in diesel-fueled
17 engines. Incomplete combustion and the presence of a chlorine source in the form of additives
18 in the oil or the fuel were speculated to lead to the formation of CDDs and CDFs. Since 1986,
19 several studies have been conducted to measure or estimate CDD/CDF concentrations in
20 emissions from diesel-fueled vehicles. These studies can be characterized as direct
21 measurements from the engine exhaust and indirect measurements as indicated by the sampling
22 of air within transportation tunnels.
23 Table 2-11 is a summary of various CDD/CDF emission characterization studies reported
24 in the United States and Europe for diesel-fueled cars and trucks. Hagenmaier et al. (1990)
25 reported an emission factor for LD diesel vehicles of 24 pg TEQ per liter of diesel fuel
26 consumed. TEQ, or the toxic equivalency factor, rates each dioxin and furan relative to that of
27 2,3,7,8-TCDD, which is arbitrarily assigned a TEQ of 1.0. Schwind et al. (1991) and Hutzinger
28 et al. (1992) studied emissions of CDDs/CDFs from German internal combustion engines
29 running on commercial diesel fuels and reported a range of CDD/CDF emission rates across the
30 test conditions (in units of pg TEQ per liter of diesel fuel consumed) of 10-130 pg TEQ/L for
31 diesel car exhaust and 70-81 pg TEQ/L for diesel truck exhaust.
32 In 1994, Hagenmaier reported CDD/CDF emissions from a diesel-fueled bus and found
33 no detectable levels in the exhaust (at a detection limit of 1 pg /L of fuel consumed) for
34 individual congeners. In 1987, the California Air Resources Board (CARB) produced a draft
35 report of a HD engine tested under steady-state conditions indicating a TF.O emission factor of
36 7,290 pg'L of fuel burned (or 1 ,^00 pg/krn driven) if nondetected values are treated as one-half
37 the detection limit. Treating nondetected values as zeros yields a TEQ concentration equivalent
7/25/00 2-30 DRAFT—DO NOT CITE OR QUOTE
-------
to 3,720 pg/L of fuel burned (or 663 pg/km driven) (Lew, 1996). Norbeck et al. (1998c)
reported emission factors for dioxin and furans from a Cummins L10 HD diesel engine running
3 on pre-1993 fuel of 0.61 pg/L and 0.41 pg/L for the same engine running on reformulated fuel.
4 The low emission factors reported by Norbeck et al. (1998c) were attributed to losses of dioxin
5 and furan compounds to the dilution tunnel walls.
6 EPA has directly sampled the exhaust from a HD diesel truck for the presence and
7 occurrence of CDDs/CDFs (Gullett and Ryan, 1997). The average of five tests (on highway and
8 city street driving conditions) was 29.0 pg TEQ/km with a standard deviation of 38.3 pg
9 TEQ/km; this standard deviation reflects the 30-fold variation in the two city driving route tests.
10 Tunnel studies are an indirect means of measuring contaminants that may be associated
11 with emissions from cars and trucks. In these studies, scrapings of carbonaceous matter from the
12 interior walls of the transportation tunnel or the tunnel air are sampled and analyzed for the
13 target contaminants. Several European studies and one recent U.S. study evaluated CDD/CDF
14 emissions from vehicles by measuring the presence of CDDs/CDFs in tunnel air. This approach
15 has the advantage that it allows random sampling of large numbers of vehicles passing through
16 the tunnel, including a range of ages and maintenance levels. The disadvantage of this approach
17 is that it relies on indirect measurements (rather than tailpipe measurements), which may
§ introduce unknown uncertainties into the interpretation of results.
Oehme et al. (1991) reported the emission rates associated with HD diesel trucks as
20 follows: uphill = 9,500 pg TEQ/km; downhill = 720 pg TEQ/km; mean = 5,100 pg TEQ/km.
21 The mean values are the averages of the emission rates corresponding to the two operating
22 modes: vehicles moving uphill on a 3.5% incline at an average speed of 37 mi/hr and vehicles
23 moving downhill on a 3.5% decline at an average speed of 42 mi/hr.
24 Wevers et al. (1992) measured the CDD/CDF content of air samples taken during the
25 winter of 1991 inside a tunnel in Antwerp, Belgium. The results obtained indicated that the
26 tunnel air had a dioxin TEQ concentration about twice as high as the outside air (80.3 fg TEQ/m3
27 for tunnel air vs. 35 fg TEQ/m3 for outside air for one set of measurements and 100 fg TEQ/m3
28 for tunnel air vs. 58 fg TEQ/m3 for outside air for a second set of measurements).
29 During October/November 1995, Gertler et al. (1996, 1998) measured CDDs/CDFs in the
30 Fort McHenry Tunnel in Baltimore, Maryland. The emission factors calculated, assuming that
31 all CDDs/CDFs emitted in the tunnel were from HD vehicles, are presented in Table 2-12. The
32 average TEQ emission factor was reported to be 172 pg TEQ/km. The major uncertainties in the
33 study were tunnel air volume measurement, sampler flow volume control, and analytical
34 measurement of CDDs/CDFs (Gertler et al., 1996, 1998).
The relative strengths of the Gertler et al. (1996; 1998) study include: (1) The study is a
recent study conducted in the United States and thus reflects current U.S. fuels and technology;
37 (2) virtually no vehicle using the tunnel used leaded gasoline, which is associated with past
7/25/00 2-31 DRAFT—DO NOT CITE OR QUOTE
-------
1 emissions of CDDs and CDFs from gasoline-powered vehicles; (3) the tunnel walls and streets
2 were cleaned 1 week before the start of sampling, and in addition, the study analyzed road dust
3 and determined that resuspended road dust contributed only about 4% of the estimated emission
4 factors; (4) HD vehicles made up, on average, a relatively large percentage (25.7%) of vehicles
5 using the tunnel; and (5) a large number of HD vehicles, approximately 33,000, passed through
6 the tunnel during the sampling period, which generates confidence that the emission factor is
7 representative of interstate trucks.
8 The EPA Office of Research and Development's dioxin source emission inventory
9 estimates that 33.5 g of dioxin TEQ (total 2,3,7,8-TCDD equivalents) were emitted from HD
10 U.S. trucks in 1995. This is a very small contribution(1.2 %) compared with the national annual
11 emission of 2,800 gCDDs/CDFs.
12
13 2.2.8. Physical and Chemical Composition of Diesel Exhaust Particles
14 DPM is defined by the measurement procedures summarized in Title 40 CFR, Part 86,
15 subpart N. These procedures define DPM emissions as the mass of material collected on a filter
16 at a temperature of 52 °C or less after dilution of the exhaust. As the exhaust is diluted and
17 cooled, nucleation, condensation, and adsorption transform volatile material to solid and liquid
18 DPM. As a consequence of this process, diesel exhaust particles are aggregates of primary
19 spherical particles that consist of solid carbonaceous material and ash and that contain adsorbed
20 organic and sulfur compounds (sulfate) combined with other condensed material. The organic
21 material includes unburned fuel, engine lubrication oil, and low levels of partial combustion and
22 pyrolysis products.
23 The organic material is absorbed to the elemental carbon core and is also found in a
24 heterogeneously nucleated aerosol. This fraction of the DPM is frequently quantified.as the
25 SOF, (i.e., the fraction that can be extracted by an organic solvent). Because of the toxicological
26 significance of the organic components associated with DPM, it is important to understand to the
27 extent possible, the historical changes in the composition of SOF and potential changes in the
28 fraction of SOF associated with DPM.
23 Various researchers have attempted to apportion the SOF to unburned oil and fuel
30 sources by thermogravimetric analysis and have found that the results vary with test cycle and
31 engine (Abbass et al., 1991; Wachter, 1990). Kittelson (1998) estimates that a typical
32 composition of SOF is about one-fourth unburned fuel and three-fourths unburned engine
33 lubrication oil. Partial combustion and pyrolysis products represented a very small fraction of
34 the SOF on a mass basis (Kittelson, 1998), which is confirmed in numerous other studies.
*?R A mimlv»r nf inv^ctionlwrc Viai7f» trif»H tr> c^rtarntp tVip nrcranir fVartinn intn varirmc Har»mr«"»M«Hc f?<->mi=»t-7l<=» flOR"^ s»nalv7(=rl thp Hirhlrvrr\mptVian<» f^Ytrnrt nfDPK/f frnm a T T)
—' •"- — — —• ••*— — — f— — —• •— • — — — — TT-l \ - "" •"" / —————_/ — — — — •— ——— — — — —— — —•• •• -— - — — • — —— — — —.
37 diesel engine and found that approximately 57% of the extracted organic mass is contained in the
7/25/00 2-32 DRAFT—DO NOT CITE OR QUOTE
-------
nonpolar fraction. About 90% of this fraction consists of aliphatic HCs from approximately C,4
to about C40 (Black and High, 1979; Pierson et al., 1983). PAHs and alkyl-substituted PAHs
3 account for the remainder of the nonpolar mass. The moderately polar fraction (~9% w/w of
4 extract) consists mainly of oxygenated PAH species, substituted benzaldehydes, and nitrated
5 PAH). The polar fraction (~32% w/w of extract) is composed mainly of n-alkanoic acids,
6 carboxylic and dicarboxylic acids of PAH, hydroxy-PAH, hydroxynitro-PAH, and nitrated N-
7 containing heterocyclic compounds (Schuetzle, 1983; Schuetzle et al., 1985).
8 Rogge et al. (1993) reported the composition of the extractable portion of fine DPM
9 emitted from two HDdiesel trucks (1987 model year). The DPM filters were extracted twice
10 with hexane then three times with a benzene/2-propanol mixture. The extract was analyzed by
11 capillary gas chromatography/mass spectrometry (GC/MS) before and after derivatization to
12 convert organic acids and other compounds having an active H atom to their methoxylated
13 analogues. Unidentified organic compounds made up 90% of the eluted organic mass and were
14 shown to be mainly branched and cyclic HCs. From the mass fraction that was resolved as
15 discrete peaks by GC/MS, -42% were identified as specific organic compounds. Most of the
16 identified resolved organic mass (~60%) consisted of n-alkanes, followed by n-alkanoic acids
17 (-20%). PAH accounted for -3.5% and oxy-PAH (ketones and quinones) for another -3.3%.
18 The distribution of the emissions between the gaseous and particulate phases is
£| determined by the vapor pressure of the individual species, by the amount and type of the DPM
20 present (adsorption surface available), and by the temperature (Ligocki and Pankow, 1989).
21 Two-ring and smaller compounds exist primarily in the gas phase (e.g., naphthalene), while
22 five-ring and larger compounds (e.g., benzo[a]pyrene) are almost completely adsorbed on the
23 particles. Three- and four- ring compounds are distributed between the two phases. The vapor
24 pressures of these intermediate PAHs can be significantly reduced by their adsorption on various
25 types of surfaces. Because of this phenomenon, the amount and type of DPM present play an
26 important role, together with temperature, in the vapor-particle partitioning of semivolatile
27 organic compounds (SOCs).
28 The measurements of gas/particulate phase distribution are often accomplished by using
29 a high-volume filter followed by an adsorbent such as polyurethane foam (PUF), Tenax, or
30 XAD-2 (Cautreels and Van Cauwenberghe, 1978; Thrane and Mikalsen, 1981; Yamasaki et al.,
31 1982). The pressure drop behind a high-volume filter or cascade impactor can contribute to
32 volatilization of the three- to five-ring PAHs from the PM (e.g., "blow-off') proportional to their
33 vapor pressures. The magnitude of this blow-off artifact depends on a number of factors,
34 including sampling temperature and the volume of air sampled (Van Vaeck et al., 1984; Coutant
|5 etal., 1988). Despite these problems from volatilization, measurements with the high-volume
filters followed by a solid adsorbent have provided most estimates of vapor-particle partitioning
37 of SOCs in ambient air, as well as insights into the factors influencing SOC adsorption onto
7/25/00 2-33 DRAFT—DO NOT CITE OR QUOTE
-------
1 aerosols. Significant fractions of phenanthrene, anthracene, and their alkylated derivatives,
2 along with fluoranthene and pyrene, exist in the gas phase. PAHs with molecular weight greater
3 than that of pyrene are typically not observed on PUF samples. During the collection of
4 particulate organic compounds, adsorption of semivolatile PAHs can also occur as well as
5 chemical transformation of the semivolatile compounds (Schauer et al., 1999; Cantrell et al.,
6 1988; Feilberg et al., 1999; Cautreels and Van Cauwenberghe, 1978).
7 Most of the sulfur in the fuel is oxidized to SO2, but a small amount (1% to 4%) is
8 oxidized to SO3 and converted to sulfate and sulfuric acid in the exhaust. Sulfate emissions are
9 roughly proportional to sulfur in the fuel. Since the reduction of the allowable sulfur content in
10 diesel fuel in 1993, sulfate emissions have declined from roughly 10% of the DPM mass to
11 around 1%. Particulate emissions from numerous vehicles tested using low sulfur fuel were
12 found to have a sulfate content of only about 1% (Yanowitz et al., 1999). Water content is on
13 the order of 1.3 times the amount of sulfate (Wall et al., 1987).
14 Metal compounds and other elements in the fuel and engine lubrication oil are exhausted
15 as ash. Hare (1977) examined 1976 Caterpillar 3208 and DDC 6V-71 engines and found the
16 most abundant elements emitted from the 6V-71 engine were silicon, copper, calcium, zinc, and
17 phosphorus. From the Caterpillar engine the most abundant elements were lead, chlorine,
18 manganese, chromium, zinc, and calcium. Calcium, phosphorus, and zinc were present in the
19 engine lubrication oil. The two-stroke 6V-71 engine had higher engine lubrication oil emissions
20 and therefore emitted higher levels of zinc, calcium, and phosphorus than the Caterpillar 3208
21 engine. Other elements may have been products of engine wear or contaminants from the
22 exhaust system. Springer (1979), in his study of 1977 Mack ETAY(B)673A and Caterpillar
23 3208 (EGR) engines, found that calcium was the most abundant metallic element in DPM
24 samples, with levels ranging from 0.01 to 0.29 wt% of the DPM. Phosphorus and silica were the
25 next most abundant elements reported, and sodium, iron, nickel, barium, chromium, and copper
26 were either present at very low levels or were below detection limits. Roughly 1 wt% of the
27 total DPM was represented by the analyzed metals. There was no consistent difference in metal
28 emissions between the engines tested by Springer or between modes. Springer tested both
23 engines on a 13-mode steady-state test. Dietzmann and co-workers (1980) examined metal
30 emission rates from four HD vehicles tested using the UDDS chassis cycle. For the single two-
31 stroke engine tested (1977 DDC 8V-71), calcium, phosphorus, and zinc emission rates were
32 more than 10 times higher than metal levels observed for three 1979 model year four-stroke
33 engines because of higher engine lubrication oil emissions. Metals accounted for 0.5% to 5% of
34 total DPM, depending on engine model. In addition to these studies, other source profiles for
35 HD diesel engine emissions report levels of chromium, manganese, mercury compounds, and
3fi nickel at levels above the detection limit (Cooper et al., 1987).
7/25/00 2-34 DRAFT—DO NOT CITE OR Q
-------
In more recent studies, Hildemann and co-workers (1991) examined metals in DPM from
the same two 1987 trucks (four-stroke engines) studied by Rogge and co-workers (1993).
3 Aluminum, silicon, potassium, and titanium were the only metals observed at statistically
4 significant levels. Taken together these made up less than 0.75 wt% of total DPM mass.
5 Lowenthal and co-workers (1994) also report metals emission rates for a composite sample of
6 several diesel vehicles. The most abundant metals were zinc, iron, calcium, phosphorus, barium,
7 and lanthanum. Together these represented less than 0.3% of total DPM mass, with an
8 emissions rate of 3.3 mg/mi. Norbeck and co-workers (1998b) report engine transient test
9 emissions of metals for a 1991 Cummin L10 engine. Silicon, iron, zinc, calcium, and
10 phosphorus were observed and together made up about 0.5% of total DPM, with an emissions
11 rate of 1.2 mg/bhp-hr.
12
13 2.2.8.1. Organic and Elemental Carbon Content of Particles
14 2.2.8.1.1. Measurement of the organic fraction. Various methods have been used to quantify
15 the organic fraction of DPM. The most common method has been Soxhlet extraction with an
16 organic solvent. Following extraction, the solvent can be evaporated and the mass of extracted
17 material (the SOF) determined, or alternatively the PM filter is weighed before and after
18 extraction and the extracted material can be further analyzed to determine concentrations of
^P individual organic compounds. Vacuum oven sublimation is used to measure a comparable
20 quantity, the volatile organic fraction (VOF), which can be further speciated by GC with a flame
21 ionization detector. Other methods have also been employed, including thermal methods,
•22 microwave extraction, sonication with an organic solvent, supercritical fluid extraction,
23 thermogravimetric analysis, and thermal desorption GC. Abbass et al. (1991) compared various
24 methods, including vacuum oven sublimation and 8 hours of Soxhlet extraction, with 4:1
25 benzene/methanol solvent for determination of SOF and found reasonably good agreement
26 between the two methods. The VOF value was typically 10% higher; however, this variation
27 was less than the coefficient of variation between measurements using the same method.
28 Thermal methods of organic carbon analysis include thermal optical reflectance (TOR)
29 and thermal optical transmittance (TOT). The extractable portion of total carbon, although
30 commonly used as a measure of organic compound content, is not equivalent to the organic
31 carbon fraction as measured by TOR or TOT. Japar et al. (1984) reported an average ratio of
32 organic compound measured by TOR to extractable mass of 0.70 ± 0.05, when the extraction
33 solvent was a toluene/propanol-1 mixture. This ratio may have resulted from the presence of
34 both oxygenated organic compounds and inorganic sulfate in the extracted mass.
*Levson (1988) reviewed literature regarding the extraction efficiency of various solvents
and found contradictory results in many cases. He concluded that there is strong evidence that
37 the most commonly used solvent, dichloromethane, leads to poor recoveries of higher molecular
7/25/00 2-35 DRAFT—DO NOT CITE OR QUOTE
-------
1 weight PAH. More recently, Lucas et al. (1999) reported the effect of varying
2 dichloromethane/benzene ratios in the solvent (from 25% to 100% dichloromethane) and
3 changing extraction tunes and found that the most effective extraction (i.e., the largest extracted
4 mass) utilized a 70% dichloromethane/30% benzene mixture and extraction times several times
5 longer than the commonly used 8-hour extraction period. Extractions of 70 hours using pure
6 dichloromethane were found to result in about twice as much SOF as extractions of only 12
7 hours. Between 6 and 24 hours of extraction time (the typical range of extraction times used),
8 the SOF recovered increased by about one-third. Using the most effective extraction conditions
9 (Soxhlet, 70 hours, 70:30 dichloromethanerbenzene ratio), Lucas et al. (1999) were able to
10 extract over 90% of the total paniculate mass.
11 Other researchers have investigated the relative quantities of mass removed by sequential
12 extraction by polar, moderately polar, and nonpolar solvents. The extracted nonpolar fraction
13 (cyclohexane) ranged from 56% to 90% of the SOF, the moderately polar (dichloromethane)
14 from 6% to 22%, and the polar fraction (acetonitrile) from 4% to 29% (Dietzmann et al., 1980).
15 Water and sulfate are not soluble in cyclohexane or dichloromethane but are soluble in
16 acetonitrile.
17 Although the reports on the extraction efficiencies for PAHs are in part contradictory, it
18 appears that Soxhlet extraction and the binary solvent system composed of aromatic solvent and
19 alcohol yield the best recovery of PAHs, as determined by C-B[a]PM (benzo[a]pyrene) spiking
20 experiments (Schuetzle and Perez, 1983). Limited recovery studies have shown that there is
21 little degradation or loss of diesel POM on the HPLC column. More than 90% of the mass and
22 70% to 100% of the Ames S. typhimuriwn-active material injected onto the column have been
23 recovered (Schuetzle et al., 1985).
24 The solvent polarity and extraction time used for measuring the SOF can have a
25 significant effect on the quantity reported. In the discussion that follows, every effort has been
26 made to compare only studies using comparable methods and to state what extraction methods
27 were employed.
28
29 2.2.8.1.2. Trends in SOF emissions. SOF emission values are highly dependent on the test
30 cycle used. Various studies have shown that SOF generally increases at light engine loads and
31 high engine speeds because these conditions lead to low exhaust temperatures, where fuel and oil
32 are not as effectively oxidized (Scholl et al., 1982; Kittleson, 1998; Springer, 1979; Schuetzle
33 and Perez, 1983; Martin, 1981b; Shi et ai., 2000). These conditions are more typically observed
34 in LD diesel vehicle applications, and thus DPM from these vehicles typically emits DPM with a
35 higher SOF component than HD Hiesel vehicles (Norbeck et al., 1998c). Acceleration modes
36 normally cause increased emission nf elemental carbon and an increase in total DPM emissions,
37 while organic components are more dominant when motoring (Wachter, 1990). Additionally,
7/25/00 2-36 DRAFT—DO NOT CITE OP. QUOTE
-------
cold-start test emissions of SOF have been shown to be approximately 25% higher than hot-start
emissions (Wachter, 1990).
3 The quantity of sulfur in diesel fuel has been suggested to have a role in the quantity of
4 SOF emitted (Sienicki et al., 1990; Tanaka et al., 1998). Sienicki et al. (1990) reported an
5 approximate 25% increase in SOF when sulfur concentrations are increased from 0.08% to
6 0.33%. The cause is unclear but several explanations have been put forth, including increased
7 absorption of organic compounds from the vapor phase onto the DPM by sulfates or sorbed
8 sulfuric acid. Alternately, it has been proposed that the measured SOF may include some
9 sulfate, so that the apparent increase in organic material is due instead to sulfate. Other fuel
10 effects include an increase in SOF emissions with a higher T90 (or T95) and with an increase in
11 aromatic content (Barry et al., 1985; Sienicki et al., 1990; Tanaka et al., 1998; Rantanen et al.,
12 1993).
13 Figures 2-29 and 2-30 show SOF emissions as a function of year for transient emissions
14 tests on chassis and engines, respectively. Both figures suggest a significant decline in SOF
15 emissions, on the order of a factor of 5 since about 1980. The highest SOF emissions are for
16 two-stroke engines built in the 1970s. These data indicate that SOF emission factors for newer
17 model year vehicles are lower than SOF emission factors for pre-1990 model year vehicles and
18 that this decrease is similar to that observed for emissions of total DPM by model year.
^ Steady-state testing conducted on late 1970s engines reported SOF at levels between 0.1
20 and 0.9 g/bhp-hr, while engines from the late 1980s and 1990s all emit 0.03 g/bhp-hr or less
21 (Table 2-8). Hori and Narusawa (1998) measured emissions from engines produced two decades
22 apart, using identical analytical procedures, and found that SOF emission factors and the percent
23 contribution of SOF to DPM were lower in the new engine compared with the old engine, under
24 all tested engine load and speed conditions, and with different fuels. The authors reported that
25 the decrease in SOF was due to lower emissions of both lubricating oil and unbumed fuel. It is
26 evident that to meet the 1991 and 1994 U.S. emission standards, SOF emission rates would need
27 to be reduced from the levels of the previous decade, although one may expect differences in
28 SOF fractions of DPM with transient cycles used to determine compliance with emission
29 standards verus steady-state conditions used in earlier test programs (Kawatani et al., 1993;
30 Wachter, 1990). Finally, in the past three decades for economic reasons, engine manufacturers
31 have made efforts to reduce oil consumption and increase the fuel efficiency of diesel engines,
32 both of which would be expected to reduce SOF emissions. Problems in achieving SOF
33 reductions from two-stroke engines have been one factor leading to the phase-out of these
34 engines for on-road use during the 1990s. No data are available prior to 1976 on SOF emissions
from HD diesel vehicles. The engine technology changes that occurred between the mid-1950s
and mid-1970s (high pressure direct injection and turbocharging, primarily) might be expected to
37 increase the efficiency of combustion and thereby reduce fuel-related SOF. SOF emissions
7/25/00 2-37 DRAFT—DO NOT CITE OR QUOTE
-------
1 levels in the mid- to late 1970s may be used as a conservative (low) estimate of SOF emissions
2 during the preceding two decades.
3 The fraction of DPM attributed to SOF from chassis dynamometer studies also shows a
4 decreasing trend over time from SOFs that ranged up to approximately 50% in the 1980s to 20%
5 SOF or less in the 1990s (Figure 2-31). The wide range in SOF as a percent of DPM displayed
6 in Figure 2-31 is suspected to result from factors such as engine deterioration and test cycle. The
7 vehicle emissions data reported in Figure 2-31 do not overrepresent buses that are likely to emit
8 DPM with a greater fraction of SOF than other vehicles. Figure 2-32 presents SOF as a fraction
9 of DPM from the same engine dynamometer studies reported in Figure 2-30. These data do not
10 reflect a downward trend hi SOF as a fraction of DPM. Since similar extraction methods were
11 used in reports of the SOF in both the chassis and engine dynamometer studies, this does not
12 appear to be a source of the wide variability observed in the fraction of SOF reported. In some
13 of the engine studies, improved ainfuel ratio control was tested hi an attempt to lower
14 carbonaceous DPM formation. Therefore, substantial differences in SOF as a percent of total
15 DPM could be the result of different engine technology or test conditions. The engine
16 dynamometer results presented in Figures 2-30 and 2-32 are from new, or relatively new,
17 engines, that is, engines with no deterioration, whereas the older engines tested on a chassis
18 dynamometer may have experienced significant deterioration that would increase SOF emissions
19 as a percent of DPM. One of the main differences suspected for the lack of a decreasing trend in
20 the percent of SOF in the engine dynamometer studies is the test cycle used. The engine
21 dynamometer tests typically include test modes, such as high speed and low load, or low speed
22 lugging modes, which produce much higher SOF relative to DPM than the driving cycles used
23 on the chassis tests.
24 It appears that as a fraction of total DPM, SOF from new model year HD diesel vehicles
25 is lower than that from older (pre-1990) HD diesel vehicles. However, as with total DPM
26 emissions, a wide range in the fraction of SOF can be observed under different driving
27 conditions and from vehicles with extensive engine wear. In general, DPM emissions have a
28 lower fraction of organic matter compared to gasoline PM (Table 2-13). Recent testing of HD
29 engines at the Desert Research Institute suggests that the organic carbon fraction of DPM is
30 approximately 19%, while earlier studies reported in the U.S. EPA SPECIATE database suggest
31 a slightly higher organic fraction of DPM from HD diesel vehicles ranging from 21% to 36%.
32 The SPECIATE database represents older vehicles that, as discussed above, tend to have higher
33 SOF emissions. The organic carbon emissions from LD diesel vehicles recently reported by
34 Norbeck et al. (1998c) and those reported by the U.S. EPA SPECIATE suggest that LD diesel
35 vehicles emit DPM with a slightly higher organic content than HD diesel vehicles, ranging from
36 22% to 43%. Gasoline engine PM emissions have recently been analyzed at the Desert Research
37 Institute by Fujita et al. (1998) and Watson et al. (1998) for hot stabilized, visibly smoking
n^/rsrs ?_IQ r>PAFT r\r\ XTOT OTTP r>P or TOTP
—DO NOT CITE OR Q
-------
vehicles, and cold-starts. These data all indicate that LD gas vehicles emit PM with a higher
fraction of organic matter than diesel vehicles, with the highest organic content measured from
3 smoking and high-emitting gasoline vehicles (averaging 76% organic carbon). One new finding
4 from the data reported by Fujita et al. (1998) is the roughly equivalent emission of organic and
5 elemental carbon from cold-start emissions of gasoline vehicles. Additional information is
6 needed to characterize a range of organic carbon for DPM from smoking and high-emitting
7 diesel vehicles as well as cold-start HD diesel vehicles.
8
9 2.2.8.O. Trends in EC comteml The EC content of ambient PM samples is one component of
10 the source profile commonly used to determine the contribution of diesel vehicles to ambient PM
11 samples (i.e., in source apportionment via chemical mass balance modeling). EC is not, strictly
12 speaking, a regulated pollutant, and so EC emissions are not routinely measured in tests of diesel
13 vehicles and engines. The scant data available on measured EC emissions from HD diesel
14 vehicles are plotted in Figure 2-33. Approximately one-half of these data, from the study of
15 Norbeck et al. (1998b), are for pickup trucks equipped with light and medium HD diesel engines
16 which represent a small portion of the diesel vehicle fleet. All of the 1994 and later model year
17 pickup trucks are equipped with diesel oxidation catalysts that oxidize the SOF component of
18 DPM. EC emissions from these vehicles are uniformly low. Results for the other three studies,
(w) all performed on HD trucks, suggest a decline in EC emission rates by model year since the early
20 1980s. In a study conducted in 1992, four HD vehicles of unknown vintage were tested and a
21 combined EC emission rate of 0.81 g/mi was reported, which consistent with the 1990 timeframe
22 in Figure 2-33 (Lowenthal et al., 1994). EC as a percentage of total DPM in these studies ranged
23 from 30% to 90%, most likely as a result of different testing cycles and different engines.
24 Figure 2-34 presents these data as a fraction that EC comprises of total carbonaceous
25 matter in the emitted DPM. The carbonaceous matter attributable to EC varied widely in the
26 1980s from approximately 20% to 90%, while in more recent years, the data suggest a smaller
27 range in the EC fraction from approximately 50% to 90% (with one data point at 30%). Recent
28 emission profiles for HD diesel vehicles suggest that 75% ± 10% of the DPM is attributable to
29 EC, whereas approximately 25% of gasoline PM is composed of EC, except for PM emissions
30 during gasoline vehicle cold-starts, which were found to have an EC content of approximately
31 42% (Table 2-13). These data also provide evidence that newer model year HD engines
32 generally emit DPM that is more rich in EC than older HD engines.
33 2.2.8.2. PAHs andNi&ro-PAHEmissions
34 PAHs, nitro-PAHs, and oxidized derivatives of these compounds have attracted
35 considerable attention because of their known mutagenic and, in some cases, carcinogenic
\^J character (National Research Council, 1982). In this section, PAH and nitro-PAH
37 concentrations and emission rates and trends in emissions over time are presented.
7/25/00 2-39 DRAFT—DO NOT CITE OR QUOTE
-------
1 2.2.8.2.1. PAHs identified in diesel exhaust At least 32 PAHs have been identified in the
2 exhaust of LD diesel vehicles and HD diesel vehicles (Table 2-14) (Watson et al., 1998;
3 Zielinska et al., 1998). Table 2-15 lists the PAHs and thioarenes identified in three LD diesel
4 vehicles' DPM extracts, reported as ng/g of DPM (Tong et al., 1984). SOF fractions accounted
5 for 11% to 15% of the total DPM mass for the LD diesel vehicles reported by Tong et al. (1984),
6 which is lower than the LD diesel vehicles organic fraction reported by Norbeck et al. (1998c) in
7 Table 2-13. Among the PAHs reported by Watson et al. (1998) and Zielinska et al. (1998), the
8 higher molecular weight compounds (pyrene through coronene) that are expected to partition to
9 the particle phase have emission rates from HD diesel vehicles ranging from below detection
10 limits up to 0.071 mg/mi. HD diesel vehicle emission rates for the lower molecular weight
11 PAHs ranged up to 2.96 mg/mi for dimethylnaphthalenes. In general, among the vehicles tested,
12 PAH emission rates were higher for LD diesel vehicles compared to HD diesel vehicles. Table
13 2-16 presents emission rates of four representative particle-phase PAHs from HD diesel vehicles,
14 LD diesel vehicles, and gasoline (with and without catalytic converter) engines. Emission rates
15 for benzo[a]pyrene were higher in diesel emissions compared with gasoline emissions, except
16 for the report by Rogge et al. (1993), who used extraction methods different from those in other
17 studies (discussed above).
18
19 2.2.8.2.2. Nitro-PAHs identified in diesel exhaust. Positive isomer identification for 16 nitro-
20 PAHs has been made utilizing the GC retention times of authentic standards and low- and high-
21 resolution mass spectra as identification criteria. These include 1-nitropyrene; 2-methyl-l-
22 nitronaphthalene; 4-nitrobiphenyl; 2-nitrofluorene; 9-nitroanthracene; 9-methyl-10-
23 nitroanthracene; 2-nitroanthracene; 2-nitrophenanthrene; l-methyl-9-nitroanthracene; 1-methyl-
24 3-nitropyrene; 1-methy 1-6-nitropyrene; l-methyl-8-nitropyrene; 1,3-, 1,6-, and 1,8-dinitropyrene;
25 and 6-nitrobenzo[a]pyrene. In addition, two nitrated heterocyclic compounds were identified, 5-
26 and 8-nitroquinoline. Forty-five additional nitro-PAHs were tentatively identified in this diesel
27 particulate extract (Paputa-Peck et al., 1983). The concentration of nitro-PAHs adsorbed on
28 diesel particles varies substantially from sample to sample. Usually 1-nitropyrene is the
29 predominant component, and concentrations ranging from 7 to 165 |ig/g of particles are reported
30 (Levsen, 1988).
31 Table 2-17 gives the approximate concentrations of several of the abundant nitro-PAHs
32 quantified in the early 1980s LD diesel particulate extracts (with the exception of 3-
33 nitrobenzanthrone, reported recently) in |ig/g of particles. Concentrations for some of the nitro-
34 PAHs identified range from 0.3 |ig/g for 1,3-dinitropyrene to 8.6 ^ig/g for 2,7-dinitro-9-
35 fluorenone and 75 jig/g for 1-nitrcpyrcnc. More recent nitro-FAK and PAH data for HD diesei
38 engines are reported in units cf g/bhp hr or mass/volume of cxliauat, making it impossible to
37 directly compare them to the older data (Norbeck et al., 1998; Bagley et al., 1996, 1998;
7/25/00 2-40 DRAFT—DO NOT CITE OR QUOTE
-------
Baumgard and Johnson, 1992; Opris et al., 1993; Hansen et al., 1994; Harvey et al., 1994;
Kantola et al., 1992; Kreso et al., 1998b; McClure et al., 1992; Pataky et al., 1994).
3
4 2.2.8.2.3. PAH and nitro-PAH emission changes over time. It is difficult to compare PAH
5 emissions from different studies because not all investigators analyze for total PAH or the same
6 suite of PAH compounds. Most studies have reported emissions of benzo[a]pyrene (B[a]P), or
7 1-nitropyrene (1-NP) due to their lexicological activity. The results of chassis dynamometer
8 studies in which B[a]P or 1-NP were measured are displayed in Figure 2-35. Dietzmann and co-
9 workers (1980) examined four vehicles equipped with late 1970s turbocharged DI engines.
10 Emissions of B[a]P from particle extracts ranged from 1.5 to 9 Hg/mi. No correlation with
11 engine technology (one of the engines was two-stroke) was observed. Rogge and co-workers
12 (1993) reported total particle-associated PAH and B[a]P emissions from two 1987 model year
13 trucks (averaged together, four-stroke and turbocharged engines). The total particle-phase PAH
14 emission rate was 0.43 mg/mi and the B[a]P emission rate was 2.7 fig/mi. Particle-phase PAH
15 in the Rogge et al. (1993) study accounted for approximately 0.5% of total DPM mass. Schauer
16 and co-workers (1999) recently reported a particle-phase PAH emission rate of 1.9 mg/mi
17 (accounting for about 0.7% of total DPM mass) for a 1995 MD turbocharged and intercooled
18 truck. B[a]P emissions were not reported, but emissions of individual species of similar
^P molecular weight were approximately 10 [ig/mi. Schauer et al. (1999) also reported gas-phase
20 PAH emission rate of 6.9 mg/mi for the same truck. Measurements of particle- and gas-phase
21 PAHs conducted for the Northern Front Range Air Quality Study in Colorado (Zielinska et al.,
22 1998) showed an average B[a]P emission rate of 13 fig/mi for 15 vehicles ranging from 1983-
23 1993 model years. The combined gas- and particle-phase PAH emission rate reported for the
24 NFRAQS study was 13.5 mg/mi. B[a]P emissions from chassis studies are summarized in
25 Figure 2-35. Zielinska (1999) reports a decreasing trend in particle-associated diesel exhaust
26 PAH from 11 measurements made on vehicles from model year 1984 to 1993 with a low
27 correlation coefficient of 0.29.
28 B[a]P emissions reported from diesel engine dynamometer studies are summarized in
29 Figure 2-36. Springer (1979) compared B[a]P emissions from naturally aspirated and
30 turbocharged engines and found that naturally aspirated engines emitted about 1 fig B[a]P/bhp-
31 hr, and DI and IDI engines emitted about 0.15 fig B[a]P/bhp-hr (Table 2-8). The difference
32 between 1 and 0.15 |ig/bhp-hr could not be attributed to specific technology changes. The
33 majority of engine test data indicate that B[a]P emissions have generally ranged from
34 approximately 1 to 4 fig/bhp-hr over the past 25 years.
35 Emissions reported for 1 -NP from diesel engines tested by chassis dynamometer ranges
VI from 0.1 to 12 fig/mi (Figure 2-35), and diesel engines dynamometer studies report 1-NP
7/25/00 2-41 DRAFT—DO NOT CITE OR QUOTE
-------
1 emission factors ranging from 1 to 4 |ig/bhp-hr (Figure 2-36). Too few measurements are
2 available to discern trends in the emission rates of these compounds.
3 As discussed in Section 2.2.4, Williams et al. (1987) and Andrews et al. (1998) of the
4 University of Leeds have demonstrated that the solvent-extractable PAH from diesel paniculate
5 originates primarily in the fuel. PAH molecules are relatively refractory so that a significant
6 fraction survives the combustion process and is exhausted as DPM. These studies have been
7 confirmed by other research groups (Crebelli et al., 1995; Tancell et al., 1995) that included the
8 use of isotopic labeling of fuel PAH. Additionally, engine oil was found to be a reservoir for
9 PAH that originates in the fuel. Pyrosynthesis of PAH occurs during very high temperature
10 conditions in a diesel engine, and under these conditions much of the DPM and other pyrolysis
11 products are ultimately oxidized before exiting the cylinder. Thus, pyrogenic formation of PAH
12 is thought to contribute a small fraction of the total PAH in diesel engine exhaust. As discussed
13 above, fuel PAH content is expected to have slowly increased over a 30-year period until 1993,
14 after which PAH content of diesel fuel is expected to have remained constant. Increasing use of
15 catalytic cracking over time may lead to increasing proportions of PAH in distillates; however,
16 fuel standards limit the aromaticity of fuel to 35%.
17 Recently, Norbeck et al. (1998a) reported on the effect of fuel aromatic content on PAH
18 emissions. Three diesel fuels were used in a Cummins L10 engine: pre-1993 fuel containing
19 33% aromatic HC and 8% PAH; low aromatic fuel containing a maximum content of 10%
20 aromatic HC and maximum of 1.4% PAH; and a reformulated fuel containing 20% to 25%
21 aromatic HC and 2% to 5% PAH. The investigators found that emission rates for the low
22 molecular weight PAHs (PAHs < 3 rings) were significantly lower when the engine was tested
23 using the low aromatic fuel compared to when the engine was run on the pre-1993 or
24 reformulated fuel (Table 2-18). While emission rates reported for several higher molecular
25 weight (particle-associated) PAHs were lower (ranging from four to28% lower) for the low
26 aromatic fuel compared with the other two fuels, the differences were not statistically significant
27 except for coronene.
28 On the basis of these limited data it is difficult to draw a precise, quantitative conclusion
29. regarding how PAH, B[a]P or 1-NP emissions have changed over time and in response to fuel
30 and engine changes. A decrease in the range cf PAH emission factors from pest 1990 model
31 year vehicles and engines compared to pre-1990 vehicles and engines is suggested by the data;
32 however, the data also suggest that differences in a vehicles' engine type and make, general
33 engine condition, fuel composition, and test conditions can influence the emissions levels of
34 PAH.
7/25/00 2-42 DRAFT—DO NOT CITE OR QUOTE
-------
2.2.83. Particle Size
Figure 2-37 shows a generic size distribution for diesel participate based on mass and
3 particle number. Approximately 50% to 90% of the number of particles in diesel exhaust are in
4 the ultrafine size range (nuclei mode), with the majority of diesel particles ranging in size from
5 0.005 - 0.05 fim and the mode at about 0.02 p,m. These are believed to be aerosol particles
6 formed from exhaust constituents during cooling and to consist of sulfuric acid droplets, ash
7 particles, condensed organic material, and perhaps primary carbon spherules (Abdul-Khalek et
8 al., 1998; Baumgard and Johnson, 1996). While accounting for the majority of the number of
9 particles, ultrafine DPM accounts for only 1% to 20% of the mass of DPM.
10 Approximately 80% to 95% of diesel particle mass is in the size range from 0.05 to 1.0
11 [am, with a mean particle diameter of about 0.2 \im. These particles have a very large surface
12 area per gram of mass, which make them an excellent carrier for adsorbed inorganic and organic
13 compounds that can effectively reach the lowest airways of the lung. The elemental carbon core
14 has a high specific surface area of approximately 30 to 50 m2/g (Frey and Corn, 1967), and
15 Pierson and Brachaczek (1976) report that after the extraction of adsorbed organic material, the
16 surface area of the diesel particle core is approximately 90 m2/g.
17 Considerable caution is required when reporting particle size measurements from diesel
18 engine exhaust because dilution conditions during the measurement process significantly affect
(jfj size distributions (e.g., the size distribution is largely a function of how it was measured), and
20 DPM size distributions obtained in dilution tunnel systems may not be relevant to size
21 distributions resulting from the physical transformation of engine exhaust in the atmosphere.
22 Measurements made on diluted diesel exhaust typically show higher numbers of nuclei mode
23 particles than do measurements made on raw exhaust because of condensation to form nuclei
24 mode aerosol upon cooling of the exhaust. Dilution ratio, sampling temperature, humidity,
25 relative concentrations of carbon and volatile matter, and other sampling factors can therefore
26 have a large impact on the number and makeup of nuclei mode particles (Abdul-Khalek et al.,
27 1999; Shi and Harrison, 1999; Luders et al., 1998). Dilution air temperature and humidity can
28 have a large effect on particle number and size distribution especially in the size range below
29 0.05 (o.m (also referred to as nanoparticles). Shi and Harrison (1999) report that a high dilution
30 ratio and high relative humidity favor the production of ultrafine particles in diesel engine
31 exhaust. Abdul-Khalek et al. (1998) report that an increase in the residence time of the exhaust
32 during dilution resulted in an increase in the number of particles in exhaust. Khatri et al. (1978)
33 report increased gas-phase HC condensation to DPM with a decrease in dilution air temperature.
34 Some studies report no peak in diesel particles in the ultrafine size range (Kleeman et al., 2000).
Kittelson (2000) reports that nanoparticle formation can be prevented by an oxidizing catalyst,
which burns organic components of the exhaust, making them unavailable for nucleation or
37 condensation to form an aerosol.
7/25/00 2-43 DRAFT—DO NOT CITE OR QUOTE
-------
1 While dilution tunnel conditions clearly do not simulate what occurs when a diesel
2 exhaust plume enters the atmosphere, Gertler (1999) demonstrated an increase in 0.02 \im
3 particles as the fraction of diesel vehicles in the Tuscarora Mountain tunnel increased from 13%
4 to 78%. These data suggest that the mode at 0.02 \im for ultrafine DPM from diesel exhaust is
5 evidenced under real-world conditions.
6 Several groups have shown that decreasing sulfur content decreases the number of nuclei
7 mode particles measured in the exhaust, assuming temperature is low enough and residence time
8 is long enough for nucleation and condensation of sulfate aerosol and water in the dilution tunnel
9 (Baumgard and Johnson, 1992, 1996; Opris et al., 1993; Abdul-Khalek et al., 1999). The
10 application of this finding to real-world conditions is difficult to predict, as the number of nuclei
11 mode particles formed from sulfate and water in the atmosphere will be determined by
12 atmospheric conditions, not by dilution tunnel conditions. With all other factors held constant, it
13 appears that reducing fuel sulfur content reduces the number of sulfate nuclei mode particles.
14 Thus the reduction in on-road fuel sulfur content that occurred in 1993 is expected to have
15 reduced the number of sulfate particles emitted by diesel vehicles, or formed secondarily in the
16 atmosphere from diesel exhaust in many urban areas. As discussed above, the contribution of
17 sulfate to total DPM mass ranges from 1% to 5% and is therefore not a substantial portion of
18 DPM mass.
19 More controversial is the suggestion that the DPM emission size distribution from newer
20 technology engines (1991 and later) may be shifted to have a much higher number concentration
21 of nuclei mode particles, independent of fuel sulfur content (Kreso et al., 1998b; Abdul-Khalek
22 et al., 1998; Baumgard and Johnson, 1996; Bagley et al., 1996). For example, Kreso and co-
23 workers (1998b) compared emissions from a 1995 model year engine with measurements made
24 on 1991 and 1988 model year engines in earlier studies (Bagley et al., 1993, 1996). Nuclei mode
25 particles made up 40% to 60% of the number fraction of DPM emissions for the 1988 engine and
26 97%+ of the DPM from the 1991 and 1995 engines. Number concentrations were roughly two
27 orders of magnitude higher for the newer engines. SOF made up 25% to 30% of DPM in the
28 1988 engine and 40% to 80% of DPM for the newer engines. Total DPM mass was significantly
29 reduced for the newer engines. It was suggested that increased fuel injection pressure leads to
30 improved fuel atomization and evaporation, leading to smaller primary carbonaceous particles.
31 Dilution conditions (relatively low temperature, low primary dilution ratio, long residence time
32 of more than 3 seconds) strongly favor the formation of nucleation products. The 1991 and 1988
33 engines were tested with 100 ppm sulfur fuel while the 1995 engine was tested with 310 ppm
34 sulfur fuel, which may confound the results to some extent.
3 5 The results of Kreso and co-workers (199Sb) and cf Bagley and co-workers (1993, 1996)
3fi have been called into question because the high level cf SGF emitted by the 1991 engine,
37 particularly at high-load test modes, was inconsistent with SOF values measured for other
7/25/00 2-44 DRAFT—DO NOT CITE OR QUOTE
-------
engines using similar types of technology (Last et al., 1995; Ullman et al., 1995). Kittelson
(1998) notes that there is far less carbonaceous DPM formed in newer engines. Accumulation
3 mode particles may have provided a high surface area for adsorption of sulfate and unburned
4 organic compounds. In the absence of this surface area for adsorption, higher number
5 concentrations of small particles are formed from nucleation of HCs and sulfuric acid.
6 A study performed at EPA by Pagan (1999) suggested that increased injection pressure
7 can lead to the formation of more nuclei mode particles hi the exhaust. Particle size distributions
8 were measured for diluted exhaust from an engine in which injection pressure could be varied
9 from roughly 35 to 110 MPa (about 5,000-16,000 psi), comparable to pressures obtained with
10 injection technology introduced in the 1980s. The dilution system and particle size measurement
11 set-up were identical in all experiments, removing some of the uncertainty in earlier studies that
12 compared engine tests performed years apart. The results showed a clear increase in the number
13 of nuclei mode particles and a decrease in the number of accumulation mode particles as
14 injection pressure was increased. This shift did not occur, however, at high engine speeds and
15 loads but only at low to intermediate speeds and loads. The increase in number concentration of
16 nuclei mode particles was much lower than the two orders of magnitude increase reported by
17 Kreso et al. (1998b) or Bagley et al. (1996). One must use caution in applying the results of
18 Pagan to modem high-injection pressure diesel engines with turbocharging/charge-air cooling
because the engine used by Pagan was a naturally aspirated engine to which high-pressure
20 common rail injection was applied. This would likely preclude this particular engine from
21 meeting current on-highway PM or NOX standards. While some studies have suggested that
22 increased injection pressure can lead to elevated ultrafine DPM number counts, Kittelson et al.
23 (1999) cite a German study that reported a decrease in ultrafine DPM number and mass with
24 increasing injection pressure.
25 While the majority of particles in diesel exhaust from modern on-road diesel engines are
26 in the ultrafine size range, evidence regarding a change in the size distribution over time is
27 unclear. To understand the size distribution of DPM to which people are exposed will require
28 measurements under conditions that more closely resemble ambient conditions.
29
30 2.3. ATMOSPHERIC TRANSFORMATION OF DIESEL EXHAUST
31 Primary diesel emissions are a complex mixture containing hundreds of organic and
32 inorganic constituents in the gas and particle phases, the most abundant of which are listed in
33 Table 2-19. The more reactive compounds with short atmospheric lifetimes will undergo rapid
34 transformation in the presence of the appropriate reactants, whereas more stable pollutants can
be transported over greater distances. A knowledge of the atmospheric transformations of
gaseous and paniculate components of diesel emissions and their fate is important in assessing
37 environmental exposures and risks. This section describes some of the major atmospheric
7/25/00 2-45 DRAFT—DO NOT CITE OR QUOTE
-------
1 transformation processes for gas-phase and particle-phase diesel exhaust, focusing on the
2 primary and secondary organic compounds that are of significance for human health. For a more
3 comprehensive summary of the atmospheric transport and transformation of diesel emissions,
4 see Winer and Busby ( 1 995).
5
6 2.3.1. Gas-Phase Diesel Exhaust
7 Gas-phase diesel exhaust contains organic and inorganic compounds that undergo various
8 chemical and physical transformations hi the atmosphere depending on the abundance of
9 reactants and meteorological factors such as wind speed and direction, solar radiation, humidity,
1 0 temperature, and precipitation. Gaseous diesel exhaust will primarily react with the following
1 1 species (Atkinson, 1988):
12
13 • Sunlight, during daylight hours
14 • Hydroxyl (OH) radical, during daylight hours
15 • Ozone (O3), during daytime and nighttime
16 • Hydroperoxyl (HO2) radical, typically during afternoon/evening hours
17 • Gaseous nitrate (NO3) radicals or dinitrogen pentoxide (N2O5), during nighttime
1 8 hours
19 • Gaseous nitric acid (HNO3) and other species such as nitrous acid (MONO) and
20 sulfuric acid (H2SO4).
21
22 The major loss process for most of the diesel exhaust emission constituents is oxidation,
23 which occurs primarily by daytime reaction with OH radical (Table 2-20). For some pollutants,
24 photolysis, reaction with O3, and reactions with NO3 radicals during nighttime hours are also
25 important removal processes. The atmospheric lifetimes do not take into consideration the
26 potential chemical or biological importance of the products of these various reactions. For
27 example, the reaction of gas-phase PAHs with NO3 appears to be of minor significance as a PAH
28 loss process, but it is more important as a route of formation of mutagenic nitro-PAHs. The
29 reaction products for some of the major gaseous diesei exhaust compounds are listed in Table 2-
30 21 and arc discussed briefly below.
31
32 2.3.1.1. Organic Compounds
33 The organic fraction of diesel is a complex mixture of compounds, very few of which
34 have been characterized. The atm exhaust ospheric chemistry of several organic constituents of
••»«- J-- --- 1 --- 1 ----- * /.-.u:_U __„ ~!-~ ~~~A,,~<*A U*r ^*U«- ~,-,,~,U,,o«-;^.-, .-.~,,-~,,,.\ Unr. U^«« r.4-.,^;»/4 A •£»„,
VV 1111/11 CUl* aiOW ^IVJUUWWU \JJ VJUll*!. ^wwl V**AVSW. J. V/A t*. **W*A&£>AWI.W *>b*A.LJLtJLAM* J
37 atmospheric chemistry of organic combustion products, see Seinfeld and Pandis (1998).
7/25/00 2-46 DRAFT— DO NOT CITE OR QUOTE
-------
Acetaldehyde forms peroxyacetyl nitrate, which has been shown to be a direct-acting
mutagen toward S. typhimurium strain TA100 (Kleindienst et al., 1985) and is phytotoxic.
3 Benzaldehyde, the simplest aromatic aldehyde, forms peroxybenzoyl nitrate or nitrophenols
4 following reaction with oxides of nitrogen (Table 2-21).
5 For those PAHs present in the gas phase, reaction with the OH radical is the major
6 removal route, leading to atmospheric lifetimes of a few hours in daylight. The gas-phase
7 reaction of PAHs containing a cyclopenta-fused ring such as acenaphthene, acenaphthylene, and
8 acephenanthrylene with the nitrate radical may be an important loss process during nighttime
9 hours. Relatively few data are available concerning the products of these gas-phase reactions. It
10 has been shown that, in the presence of NOX, the OH radical reactions with naphthalene, 1- and
11 2-methylnaphthalene, acenaphthylene, biphenyl, fluoranthene, pyrene, and acephenanthrylene
12 lead to the formation of nitroarenes (Arey et al., 1986; Atkinson, 1986,1990; Zielinska et al.,
13 1988, 1989a; Arey, 1998). In addition, in a two-step process involving OH radical reaction and
14 NO2 addition, 2-nitrofluoranthene and 2-nitropyrene can be formed and eventually partition to
15 the particle phase, as will other nitro-PAHs.
16 The addition of the NO3 radical to the PAH aromatic ring leads to nitroarene formation
17 (Sweetman et al., 1986; Atkinson et al., 1987, 1990; Zielinska et al., 1989a). The gas-phase
reactions of NO3 radical with naphthalene, 1- and 2-methylnaphthalene, acenaphthene,
phenanthrene, anthracene, fluoranthene, and pyrene produce, in general, the same nitro-PAH
20 isomers as the OH radical reaction, but with different yields (Arey et al., 1989; Sweetman et al.,
21 1986; Atkinson et al., 1987, 1990; Zielinska et al., 1986, 1989a). For example, the same 2-
22 nitrofluoranthene is produced from both OH radical and NO3 gas-phase reactions, but the
23 reaction with NO3 produces a much higher yield. The production of several nitroarene
24 compounds has been studied in environmental chambers (Arey et al., 1989; Zielinska et al.,
25 1990; Atkinson and Arey, 1994; Arey, 1998; Feilberg et al., 1999), and generally the same nitro-
26 PAH isomers formed from reaction with OH and NO3 radicals are observed in ambient air
27 samples. Secondary formation of nitroarenes through the gas-phase reactions of the 2-, 3-, and
28 4-ring PAHs is the major source for many of the nitroarenes observed in ambient air (Pitts et al.,
29 1985a, b, c; Arey et al., 1986; Zielinska et al., 1988). Photolysis is the major removal pathway
30 for nitroarenes with lifetimes of approximately 2 hours (Feilberg et al., 1999; Nielsen and
31 Ramdahl, 1986).
32
33 2.3.1.2. Inorganic Compounds
34 SO2 and oxides of nitrogen (primarily NO) are emitted from diesel engines. SO, is
« readily oxidized by the OH radical in the atmosphere, followed by formation of the HO2 radical
and HSO3, which rapidly reacts with water to form H,SO4 aerosols. Because SO2 is soluble in
37 water, it is scavenged by fog, cloud water, and raindrops. In aqueous systems, SO2 is readily
7/25/00 2-47 DRAFT—DO NOT CITE OR QUOTE
-------
1 oxidized to sulfate by reaction with hydrogen peroxide (H2O2), O3, or O2 in the presence of a
2 metal catalyst (Calvert and Stockwell, 1983). Sulfur emitted from diesel engines is
3 predominantly (~98%) in the form of SO2, a portion of which will form sulfate aerosols by the
4 reaction described above. Nonroad equipment, which typically uses fuel containing 3,300 ppm
5 sulfate, emits more SO2 than on-road diesel engines, which use fuels currently containing an
6 average of 340 ppm sulfur because of EPA regulations effective in 1993 decreasing diesel fuel
7 sulfur levels. EPA estimates that mobile sources are responsible for about 7% of nationwide SO2
8 emissions, with diesel engines contributing 74% of the mobile source total (the majority of the
9 diesel SO2 emissions originate from nonroad engines) (U.S. EPA, 1998b).
10 NO is also oxidized in the atmosphere to form NO2 and particulate nitrate. The fraction
11 of motor vehicle NOX exhaust converted to particulate nitrate in a 24-hour period has been
12 calculated using a box model to be approximately 3.5% nationwide, a portion of which can be
13 attributed to diesel exhaust (Gray and Kuklin, 1996). EPA estimates that in 1997, mobile
14 sources were responsible for about 50% of nationwide NOX emissions, with diesel engines being
15 responsible for approximately one-half of the mobile source total (U.S. EPA, 1998b).
16
17 2.3.1.3. Atmospheric Transport of Gas-Phase Diesel Exhaust
18 Gas-phase diesel exhaust can be dry deposited, depending on the deposition surface,
19 atmospheric stability, and the solubility and other chemical properties of the compound. Dry
20 deposition of organic species is typically on the order of weeks to months, with dry deposition
21 velocities of approximately 10"4 cm/sec (Winer and Busby, 1995). In contrast, inorganic species
22 such as SO2 and nitric acid have relatively fast deposition rates (0.1-2.5 cm/sec) and will remain
23 in the atmosphere for shorter time periods compared with the organic exhaust components.
24 Some gas-phase species will also be scavenged by aqueous aerosols and potentially deposited
25 via precipitation. These processes can greatly reduce the atmospheric concentration of some
26 vapor-phase species. Atmospheric lifetimes for several gas-phase components of diesel exhaust
27 are on the order of hours or days, during which tune atmospheric turbulence and advection can
28 disperse these pollutants widely.
29
30 2.3.2. Particle-Phase Diesel Exliausi
31 Particle-associated diesel exhaust is composed of primarily carbonaceous material
32 (organic and elemental carbon) with a very small fraction composed of inorganic compounds
33 and metals. The organic carbon fraction adsorbed on DPM is composed of high-molecular-
34 weight compounds, such as PAHs, which are generally more resistant to atmospheric reactions
35 than PAIIs in the gas phase. The elemental carbon component of diesei exhaust is inert to
3S atmospheric degradation, while the PAII cumpouuus aic degraded by reaction with the following
37 species:
7/25/00 2-48 DRAFT—DO NOT CITE OR QUOTE
-------
Sunlight, during daylight hours
• O3, during daytime and nighttime
3 • NO3 and N2O5 during nighttime hours
4 OH and HO2
5 • NO2, during nighttime and daytime hours
6 • H2O2
7 • HNO3 and other species such as HONO and H2SO4.
8
9 Since many of the PAH derivatives formed by reaction with some of the reactants listed
10 above have been found to be highly mutagenic, a brief discussion of PAH photolysis, nitration,
11 and oxidation follows. Some of the major degradation products from particulate diesel exhaust
12 and their biological impact are listed in Table 2-22.
13
14 2.3.2.1. Particle-Associated PAH Photooxidation
15 Laboratory studies of photolysis of PAHs adsorbed on 18 different fly ashes, carbon
16 black, silica gel, and alumina (Behymer and Kites, 1985,1988) and several coal stack ashes
17 (Yokely et al., 1986; Dunstan et al., 1989) have shown that the extent of photodegradation of
,8 PAHs depends very much on the nature of the substrate to which they are adsorbed. The
dominant factor in the stabilization of PAHs adsorbed on fly ash was the color of the fly ash,
20 which is related to the amount of black carbon present. It appears that PAHs were stabilized if
21 the carbon black content of the fly ash was greater than approximately 5%. On black substrates,
22 half-lives of PAHs studied were on the order of several days (Behymer and Kites, 1988). The
23 environmental chamber studies of Kamens et al. (1988) on the daytime decay of PAHs present
24 on residential wood smoke particles and on gasoline internal combustion emission
25 particles showed PAH half-lives of approximately 1 hour at moderate humidities and
26 temperatures. At very low angle sunlight, very low water vapor concentration, or very low
27 temperatures, PAH daytime half-lives increased to a period of days. The presence and
28 composition of an organic layer on the aerosol seems to influence the rate of PAH photolysis
29 (Jang and McDow, 1995; McDow et al., 1994; Odum et al., 1994).
30 Because of limited understanding of the mechanisms of these complex heterogeneous
31 reactions, it is currently impossible to draw any firm conclusion concerning the photostability of
32 particle-bound PAHs in the atmosphere. Because DPM contains a relatively high quantity of
33 elemental carbon, it is reasonable to speculate that PAHs adsorbed onto these particles might be
34 relatively stable under standard atmospheric conditions, leading to an anticipated half-life of 1 or
more days.
7/25/00 2-49 DRAFT—DO NOT CITE OR QUOTE
-------
1 23.2.2. Particle-Associated PAH Nitration
2 Since 1 978, when Pitts et al. ( 1 978) first demonstrated that B[a]P deposited on glass-fiber
3 filters exposed to air containing 0.25 ppm NO2 with traces of HNO3 formed nitro-B[a]P,
4 numerous studies of the heterogeneous nitration reactions of PAHs adsorbed on a variety of
5 substrates in different simulated atmospheres have been carried out (Finlayson-Pitts and Pitts,
6 1986). PAHs deposited on glass-fiber and Teflon-impregnated glass-fiber filters react with
7 gaseous N2O5, yielding their nitro derivatives (Pitts et al., 1985b,c). The most abundant isomers
8 formed were 1-NP from pyrene, 6-nitro-B[a]P from B[a]P, and 3-nitroperylene from perylene.
9 The formation of nitro-PAHs during sampling may be an important consideration for
1 0 DPM collection because of the presence of NO2 and HNO3 (Feilberg et al., 1 999). However,
1 1 Schuetzle (1 983) concluded that the artifact formation of 1 -NP was less than 1 0% to 20% of the
1 2 1-NP present in the diesel particles if the sampling time was less than 23 min (one FTP cycle)
1 3 and if the sampling temperature was not higher than 43 °C. The formation of nitroarenes during
1 4 ambient high-volume sampling conditions has been reported to be minimal, at least for the most
1 5 abundant nitropyrene and nitrofluoranthene isomers (Arey et al., 1988).
1 6 DPM contains a variety of nitroarenes, with 1-NP being the most abundant among
1 7 identified nitro-PAHs. The concentration of 1 -NP was measured in the extract of particulate
1 8 samples collected at the Allegheny Mountain Tunnel on the Pennsylvania Turnpike as 2. 1 ppm
1 9 and ~5 ppm by mass of the extractable material from diesel and SI vehicle PM, respectively.
20 These values are much lower than would be predicted on the basis of laboratory measurements
21 for either diesel or SI engines (Gorse et al., 1983). Several nitroarene measurements have been
'22 conducted in airsheds heavily affected by motor vehicle emissions (Arey et al., 1987; Atkinson
23 et al., 1988; Zielinska et al., 1989a,b; Ciccioli et al., 1989, 1993). Ambient PM samples were
24 collected at three sites in the Los Angeles Basin during two summertime periods and one
25 wintertime period. Concentrations of 1-NP ranged from 3 pg/m3 to 60 pg/m3 and 3-
26 nitrofluoranthene was also present in DPM at concentrations ranging from not detectable to 70
27 pg/m3.
28
29 2.3.2.3. Particle-Associated PAH Ozonoiysis
30 Numerous laboratory studies have shewn that PAHs deposited on combi^tiGn-generated
3 1 fine particles and on model substrates undergo reaction with O3 (Katz et al., 1 979; Pitts et al.,
32 1980, 1986; Van Vaeck and Van Cauwenberghe, 1984; Finlayson-Pitts and Pitts, 1986). The
33 dark reaction toward O3 of several PAHs deposited on model substrates has been shown to be
34 relatively fast under simulated atmospheric conditions (Katz et al., 1979; Pitts et al., 1980,
"3C 1 QQ^ ITal-P. l«Tnao f\n «-li«» f\r-Aav r\f 1 <•/-> o«»i r/M-ol U^i.^o -11 jo-c. ^a»»^»*a^ f^— +U« ~* ~—~ -~.,,,,-»:, ,~ T> A Uc
^ w LS*J\JJ- JL J.IAA.L 14 » WkJ Wii bAAW WAMWA \J i i Iw >3 W » WA MA llV/lUk? VvwAW IWLJW1LWIA i.V/1 U.1W iiL\Ji W 1 ^fclW i& W- J. /"1-L.L^,
n^Tnln^l-U^n^o^n fV nt~, ~+ «1
•* A AA» ^M J «AAA •.* AA Wh^'^'i A W yJL^.b*V^, Wb WLA . , LSI ^ I.
7/25/00 2-50 DRAFT— DO NOT CITE OR QUOTE
-------
The reaction of PAHs deposited on diesel particles with 1.5 ppm O3 under high-volume
sampling conditions has been shown to be relatively fast, and half-lives on the order of 0.5 to 1
3 hour have been reported for most PAHs studied (Van Vaeck and Van Cauwenberghe, 1984).
4 The most reactive PAHs include B[a]P, perylene, benz[a]anthracene, cyclopenta[cd]pyrene, and
5 benzo[ghi]perylene. The benzofluoranthene isomers are the least reactive of the PAHs studied,
6 and benzo[e]perylene is less reactive than its isomer B[a]P. The implications of this study for
7 the high-volume sampling ambient POM are important: reaction of PAHs with O3 could
8 possibly occur under high-volume sampling conditions during severe photochemical smog
9 episodes, when the ambient level of 03 is high. However, the magnitude of this artifact is
10 difficult to assess from available data.
11
12 2.3.2.4. Atmospheric Transport of Diesel Exhaust Paniculate Matter
13 Ultrafine particles emitted by diesel engines undergo nucleation, coagulation, and
14 condensation to form fine particles. DPM can be removed from the atmosphere by dry and wet
15 deposition. Particles of small diameter (<1 \im), such as DPM, are removed less efficiently than
16 larger particles by wet and dry deposition and thus have longer atmospheric residence times.
17 Dry deposition rates vary depending on the particle size. Because of their small size, diesel
IS exhaust particles have residence times of several days (dry deposition velocities of
^P approximately 0.01 cm/sec) (Winer and Busby, 1995). Diesel particulates may be removed by
20 wet deposition if they serve as condensation nuclei for water vapor deposition or are scavenged
21 by precipitation in- or below-cloud.
22 In a study designed to assess the atmospheric concentrations and transport of diesel
23 exhaust particles, Horvath et al. (1988) doped the sole source of diesel fuel in Vienna with an
24 organometallic compound of the heavy earth element dysprosium. The authors found that in
25 some of the more remote sampling areas, DPM composed more than 30% of the paniculate
26 mass, indicating that DPM can be dispersed widely.
27
28 2.3.3. Diesel Exhaust Aging
29 After emission from the tailpipe, diesel exhaust undergoes dilution, reaction, and
30 transport in the atmosphere. The primary emission is considered "fresh," while "aged" diesel
31 exhaust is considered to have undergone chemical and physical transformation and dispersion
32 over a period of a day or two. Laboratory dilution tunnel measurements represent a
33 homogeneous environment compared to the complex and dynamic system into which real-world
34 diesel exhaust is emitted. The physical and chemical transformation of diesel exhaust will vary
depending on the environment into which it is emitted. In an urban or industrial environment,
diesel exhaust may enter an atmosphere with high concentrations of oxidizing and nitrating
7/25/00 2-51 DRAFT—DO NOT CITE OR QUOTE
-------
1 radicals as well as nondiesel organic and inorganic compounds that may influence the toxicity,
2 chemical stability, and atmospheric residence tune.
3 In general, secondary pollutants formed in an aged aerosol mass are more oxidized, and
4 therefore have increased polarity and water solubility (Finlayson-Pitts and Pitts, 1986). Kamens
5 et al. (1988) reported that photooxidation of particle-bound PAH is enhanced as relative
6 humidity is increased. Weingartner et al. (1997a) evaluated the hygroscopic growth of diesel
7 particles and found that freshly emitted diesel particles demonstrated minimal hygroscopic
8 growth (2.5%) while aged particles subjected to UV radiation and ozonolysis exhibited greater
9 hygroscopic growth. In addition, increasing sulfur content in the fuel has been observed to result
1 0 in greater water condensation onto diesel particles. Dua et al. (1999) reported that unlike many
1 1 other types of particles, diesel particles do not appear to undergo hygroscopic growth once
1 2 emitted to the atmosphere. To the extent that diesel exhaust components are oxidized or nitrated
1 3 in the atmosphere, they may be removed at rates different from their precursor compounds and
1 4 may exhibit different biological reactivities.
15 In a recent experiment, the biological activity of DPM exposed to 0. 1 ppm ozone for 48
1 6 hours was compared with that of DPM not exposed to ozone (Ohio et al., 2000). Instillations of
1 7 the ozonated DPM in rat lung resulted in an increase in biological activity (neutrophil influx,
1 8 increased protein, and lactate dehydrogenase activity) compared with DPM that had not been
1 9 treated with ozone. These data suggest that ambient levels of ozone can alter DPM constituents
20 causing an increase in toxicity compared with nonozonated DPM.
2 1 In addition to changes in particle composition with aging, particle size distributions may
22 vary depending on aggregation and coagulation phenomena in the aging process. People in
23 vehicles, near roadways (e.g., cyclists, pedestrians, people in nearby buildings), and on
24 motorcycles will be exposed to more fresh exhaust than the general population. In some settings
25 where emissions are entrained for long periods through meteorological or other factors,
26 exposures would be expected to include both fresh and aged diesel exhaust. The complexities of
27 transport and dispersion of emission arising from motor vehicles have been the subject of
28 extensive modeling and experimental studies over the past decades and have been summarized
29 by Sampson (1988); exposures to DPM are discussed in the next section of this chapter.
30 The major organic constituents of diesel exhaust and their potential degradation
31 pathways described above provide evidence for (1) direct emission of PAHs, (2) secondary
32 formation of nitroarenes, and (3) secondary sulfate and nitrate formation. Since nitro-PAH
33 products are often more mutagenic than their precursors, the formation, transport, and
34 concentrations of these compounds in an aged aerosol mass are of significant interest.
35
OR *) A A 1\/1T>T1?'M'T1 T»¥T7O1?T T'Vljr A TTOT f-_. M. »_»_•• i-n^iiin i »-\_ri i jn_nio /V11U fWVJTVSOUItIL>O
37 2.4.1. Diesel Exhaust Gases in the Ambient Atmosphere
7/25/00 2-52 DRAFT— DO NOT CITE OR QUOTE
-------
Although emissions of several diesel exhaust components have been measured, few
studies have attempted to elucidate the contribution of diesel-powered engines to atmospheric
3 concentrations of these components. The emission profile of gaseous organic compounds is
4 different for diesel and SI vehicles; the low-molecular-weight aromatic HCs and alkanes (C10) and aromatic
6 HCs (such as naphthalene, methyl- and dimethyl- naphthalenes, methyl- and dimethyl-indans)
7 are more characteristic of diesel engine emissions. These differences were the basis for
8 apportionment of gasoline- and diesel-powered vehicle emissions to ambient nonmethane
9 hydrocarbon (NMHC) concentrations in the Boston and Los Angeles (South Coast Air Basin)
10 urban areas.
11 The chemical mass balance receptor model (described below) was applied to ambient
12 samples collected in these areas, along with appropriate fuel, stationary, and area source profiles
13 (Fujita et al., 1997). The average of the sum of NMHC attributed to diesel exhaust, gasoline-
14 vehicle exhaust, liquid gasoline, and gasoline vapor was 73% and 76% for Boston and the South
15 Coast Air Basin (SoCAB), respectively. The average source contributions of diesel exhaust to
16 NMHC concentrations were 22% and 13% for Boston and the SoCAB, respectively. Diesel
17 vehicles emit lower levels of NMHC in the exhaust compared to gasoline vehicles. The relative
18 contribution of diesel exhaust clearly depends on several factors, including fleet composition,
^ sampling location (e.g., near a bus station vs. near a highway or other sources), and the
20 contribution from point and area sources. The contribution of diesel exhaust to ambient NMHC
21 showed large variations among sampling sites in the Boston area. The source apportionment in
•22 the Fujita et al. (1997) study indicates that mobile vehicle-related emissions account for the
23 majority of ambient NMHC in the two urban areas studied and the results can likely be
24 extrapolated to other urban areas with similar source compositions. Other source apportionment
25 methods such as those used by Henry et al. (1994) have been applied to speciated HC data to
26 separate the mobile source direct emission from gasoline evaporative emissions. This method
27 uses a combination of graphical analysis (Graphical Ratio Analysis for Composition Estimates,
28 GRACE) and multivariate receptor modeling methods (Source Apportionment by Factors with
29 Explicit Restrictions, SAFER) and was not used to identify the diesel engine contribution to the
30 HCs measured.
31
32 2.4.2. Ambient Concentrations of Diesel Particulate Matter
33 Since DPM is chemically complex, an assessment of ambient DPM concentrations relies
34 primarily on (1) studies that collect ambient samples and adequately characterize their chemical
composition or (2) modeling studies that attempt to recreate emissions and atmospheric
conditions. Ambient concentrations of DPM also have been reported from studies using
37 surrogate species. The results of these studies are summarized below. Studies conducted in
7/25/00 2-53 DRAFT—DO NOT CITE OR QUOTE
-------
1 Europe and Japan were reviewed but for the most part were not included because of questions
2 surrounding the applicability of measurements in locations that use different diesel technology
3 and control measures from those in the United States.
4
5 2.4.2.1. Source Apportionment Studies
6 Receptor models are used to infer the types and relative contributions of sources to
7 pollutant measurements made at a receptor site. Receptor models assume that the mass is
8 conserved between the source and receptor site and that the measured mass of each pollutant is a
9 sum of the contributions from each source. Receptor models are referred to as "top-down" in
10 contrast to "bottom-up" methods, which use emission inventory data, activity patterns, and
11 dispersion modeling from the source to predict concentrations at a receptor site.
12 The most commonly used receptor model for quantifying concentrations of DPM at a
13 receptor site is the chemical mass balance (CMB) model. Input to the CMB model includes
14 measurements of PM mass and chemistry made at the receptor site as well as measurements
15 made of each of the source types suspected to impact the site. Because of problems involving
16 the elemental similarity between diesel and gasoline emission profiles and their co-emission in
17 time and space, chemical molecular species that provide markers for separation of these sources
18 have been identified (Lowenthal et al., 1992). Recent advances in chemical analytical
19 techniques have facilitated the development of sophisticated molecular source profiles, including
20 detailed speciation of PM-associated organic compounds that allow the apportionment of PM to
21 gasoline and diesel sources with increased confidence. CMB analysis that uses speciation of
22 organic compounds in the source profiles is typically referred to as extended species CMB.
23 Older studies that made use of only elemental carbon, total organic carbon, trace elements, and
24 major ions in the source profiles (conventional CMB) have been published and are summarized
25 here, but they are subject to more uncertainty. It should be noted that since receptor modeling is
26 based on the application of source profiles to ambient measurements, estimates of DPM
27 concentration generated by this method include the contribution from on-road and nonroad
28 sources to the extent the source profiles are similar (which would include military sources
29 depending on the sampling locations and fleet composition). In addition, this method identifies
30 sources of primary emissions of DPM only, and the contribution of secondary aerosols is not
31 attributed to sources.
32 The CMB model has been used to assess concentrations of DPM in areas of California,
33 Phoenix, Denver, and Manhattan (Table 2-23). DPM concentrations reported by Schauer et al.
34 (1996) for samples collected in California in 1982 ranged from 4.4 |ig/m3 in west Los Angeles to
35 11 6 jig/m3 in downtown Los Angeles. The average contribution of PM to total PM mass ranged
3R from 11% in RnKiHrmv tn 36% in downtov.Ti Los Angelss. As msnticiisci dbcvc, this mode!
37 accounts for primary emissions of DPM only; the contribution of secondary aerosol formation
7/25/00 2-54 DRAFT—DO NOT CITE OR QUOTE
-------
(both acid and organic aerosols) is not included. In sites downwind from urban areas, such as
Rubidoux in this study, secondary nitrate formation can account for a substantial fraction of the
3 mass (25% of the fine mass measured in Rubidoux was attributed to secondary nitrate), a portion
4 of which comes from diesel exhaust (Gray and Kuklin, 1996).
5 The California Environmental Protection Agency (Cal EPA) reported ambient DPM
6 concentrations for 15 air basins in California based on ambient measurements taken statewide
7 from 1988 to 1992 across California (Cal EPA, 1998a). Cal EPA used CMS analysis of ambient
8 measurements from the San Joaquin Valley (1988-89), South Coast (1986), and San Jose
9 (winters for 1991-92 and 1992-93) to determine mobile source contributions and then applied
10 the California 1990 PM10 emissions inventory to determine the fraction of mobile source PM10
11 attributable to diesel emissions. The results of this analysis indicate that annual average basin-
12 wide levels of direct DPM may be as low as 0.2 |J,g/m3 and may range up to 2.6 [J,g/m3 for basins
13 that are largely nonurban but may have one or more densely populated areas (such as Palm
14 Springs in the Salton Sea basin). DPM concentrations for air basins that are moderately or
15 largely urbanized ranged from 1.8 [J,g/m3 to 3.6 y,g/m3.
16 Two studies using CMB analysis have been conducted in the Phoenix area that report
17 DPM concentrations. A wintertime study conducted in the Phoenix area in 1989—90 reported
DPM concentrations for nonurban areas ranging from 2 fig/m3 to 5 ^g/m3 and DPM
concentrations for central and south Phoenix urban areas ranging from 10 p,g/m3 to 13 (J,g/m3
20 (Chow et al., 1991). Chow et al. (1991) reported that DPM levels on single days can range up to
21 22 y,g/m3 at the central Phoenix site. A more recent study conducted from November 1994
22 through March 1995 reported DPM concentrations for Phoenix averaging 2.4 |J,g/m3 and
23 reaching 5.3 fJ.g/m3 (Maricopa Association of Governments, 1999). The extended species CMB
24 was used for this study, providing a more confident identification of DPM separate from
25 gasoline PM emissions than the earlier Phoenix study. DPM accounted for an average 15% of
26 ambient PM2.5, and gasoline PM accounted for an average of 52% of ambient PM2.5 in the
27 1994-95 Phoenix study.
28 During the winter of 1997, a study was conducted that assessed DPM concentrations at
29 two urban sites in the Denver area (Fujita et al., 1998). The Northern Front Range Air Quality
30 Study (NFRAQS), initiated to assess the sources of the "brown cloud" observed along
31 Colorado's Front Range, conducted air quality sampling during the winter of 1996, summer of
32 1996, and winter of 1997. For a 60-day period from December 1996 through January 1997,
33 ambient samples collected at two urban Denver sites were analyzed for organic carbon species
34 for use in the extended species CMB. The average DPM concentrations reported for the urban
site at Welby, CO, and the suburban site at Brighton, CO, were 1.7 (J,g/m3 and 1.2 [ig/m3,
respectively. During the study period, DPM concentrations exceeded 5 p.g/m3 on two occasions
37 in Welby, with reported DPM concentrations of 5.7 (ig/m3 and 7.3 |lg/m3. DPM accounted for
7/25/00 2-55 DRAFT—DO NOT CITE OR QUOTE
-------
1 an average of 10% of ambient PM2.5, and gasoline PM accounted for an average of 27% of
2 ambient PM2.5.
3 One of the major claims from the NFRAQS was a substantial contribution of EC from
4 gasoline-powered vehicles, mainly from cold-start and high-emitting gasoline vehicles. At the
5 Welby site, the contribution of diesel and gasoline emissions to EC measurements was 52% and
6 42%, respectively. At the Brighton site, the contribution of diesel and gasoline emissions to EC
7 measurements was 71% and 26%, respectively. The findings from the NFRAQS are compelling
8 and suggest the need for further investigations to quantify the contribution from cold-start and
9 high-emitting vehicle emissions for both gasoline and diesel vehicles. Geographical, temporal,
10 and other site-specific parameters that influence PM concentrations, such as altitude, must be
11 considered when extrapolating the NFRAQS findings to other locations.
12 In addition to the need for urban and rural average DPM concentrations, an assessment of
13 potential health effects resulting from DPM exposure includes an assessment of people in
14 environments with potentially elevated levels of DPM. Limited data are available to allow a
15 characterization of DPM concentrations in "hotspots" such as near heavily traveled roadways,
16 bus stations, train stations, and marinas. Only one CMB study has attempted to apportion PM
17 measured hi an urban hotspot. Wittorff et al. (1994) reported results of conventional CMB
18 performed on PM samples collected in the spring of 1993 over a 3-day period at a site adjacent
19 to a major bus stop on Madison Avenue in midtown Manhattan. Buses hi this area idle for as
20 long as 10 minutes, and PM emissions are augmented by the elevated levels of DPM emitted
21 during acceleration away from the bus stop (discussed in Section 2.2.5). DPM concentrations
22 reported from this study ranged from 13.0 flg/m3 to 46.7 |J.g/m3. This study attributed, on
23 average, 53% of the PM10 to diesel exhaust. The DPM concentrations resulting from the source
24 apportionment method used in this study require some caution since the CMB model
25 overpredicted PM10 concentrations by an average 30%, which suggests that additional sources
26 of the mass were not accounted for in the model. The relevance of the Manhattan bus stop
27 concentrations and potential exposure for large urban populations provides strong motivation for
28 further studies in the vicinity of such hotspots.
29 In summary, source apportionment studies of ambient samples collected before 1990
30 suggest that seasonal and annual average diesel PM concentrations for noiiurban areas ranged
31 from 2 fig/m3 to 5 (ig/m3. DPM concentrations reported from CMB studies for urban areas in the
32 pre-1990 timeframe ranged from 4.4 |J.g/m3 to 13 Jig/m3 with concentrations on individual days
33 ranging up to 22 u.g/mj. Source apportionment applied to ambient measurements taken in 1990
34 or later suggest that seasonal or annual average DPM levels in suburban/nonurban locations can
35 range from 0.2 [ag/m3 to 2.6 }ig/ir.3 with maximum reported values ranging up to 3.4 p.g/m3.
36 DPM rrmr?n_tr?tions reported from CMB studies in urban areas duiiug 1990 or later range from
37 1.7 |J.g/m3 to 3.6 (ig/m3 with maximum concentrations up to 7.3 (J-g/m3. The highest DPM
7/25/00 2-56 DRAFT—DO NOT CITE OR QUOTE
-------
concentrations reported from CMB analysis of ambient measurements were those in the vicinity
of a bus stop in midtown Manhattan, which ranged from 13.2 ng/m3 to 46.7 Jlg/m3.
3
4 2.4.2.2. EC Surrogate for Diesel Paniculate Matter
5 EC is a major component of diesel exhaust, contributing approximately 50% to 85% of
6 diesel particulate mass, depending on engine technology, fuel type, duty cycle, engine
7 lubrication oil consumption, and state of engine maintenance (Graboski et al., 1998b; Zaebst et
8 al., 1991; Pierson and Brachaczek, 1983; Warner-Selph et al., 1984). In urban ambient
9 environments, diesel exhaust is one of the major contributors to EC, with other potential sources
10 including spark-engine exhaust; combustion of coal, oil, or wood; charbroiling; cigarette smoke;
11 and road dust. While coldstart emissions from gasoline combustion vehicles were reported to be
12 an important source of EC in wintertime samples collected in two cities in the Denver area
13 (Fujita et al., 1998), it is currently unclear to what extent these results are transferable to other
14 locations. It is noteworthy that the EC content of the cold-start emissions from gasoline
15 combustion vehicles was lower than that from diesel combustion engines in the same study by
16 almost a factor of 2.
17 Fowler (1985) evaluated several components of diesel exhaust and concluded that EC is
18 the most reliable overall measure of ambient diesel exhaust exposure. Because of the large
portion of EC in DPM, and the fact that diesel exhaust is one of the major contributors to EC in
20 many ambient environments, DPM concentrations can be bounded using EC measurements.
21 Surrogate calculations of DPM have been based on the fraction of ambient EC measured in a
22 sample that is attributable to diesel engine exhaust and the fraction of the diesel particle mass
23 accounted for by EC. In the recent Multiple Air Toxics Exposure Study in the South Coast Air
24 Basin (MATES-II, SCAQMD, 2000), EC measurements were used to estimate DPM
25 concentrations by the following relationship: approximately 67% of fine EC in the ambient air
26 in the Los Angeles area originates from diesel engine exhaust (Gray, 1986), and the average EC
27 fraction of diesel particles measured was 64%. Therefore, in the MATES-II study, the South
28 Coast Air Quality Management District calculated DPM concentrations from EC measurements
29 by multiplying a measured EC concentration by 67% and dividing by the fraction of DPM mass
30 accounted for by EC of 64%, for example, DPM concentration = (EC * 0.67)70.64, or DPM =
31 HC * 1.04. This calculation relies on data collected in the 1982 timeframe and may not
32 accurately represent the current day contributions of diesel engines to the ambient EC inventory.
33 An alternative calculation can be derived using more recent studies. The fraction of EC
34 attributable to diesel exhaust can be estimated from detailed source profiles applied to a CMB
model as discussed above. The contribution of diesel engines to EC averaged 68% ± 20% for
Brighton, CO, and 49% ± 26% at Welby, CO, as part of the winter 1996-1997 NFRAQS. In
37 Phoenix, diesel engine exhaust was estimated to account for approximately 46% ± 22% of the
7/25/00 2-57 DRAFT—DO NOT CITE OR QUOTE
-------
1 ambient EC. For some environments, such as certain occupational settings in which diesel
2 engines are in proximity to workers, all the EC may realistically be attributed to diesel exhaust as
3 a reasonable upper bound estimate of DPM concentrations.
4 As discussed in Section 2.2, the EC content of DPM can vary widely depending on
5 engine type, load conditions, and the test cycle. However, typical profiles for HD and LD diesel
6 engines have been determined and the typical EC fraction of DPM ranges from approximately
7 52% to 75%.
8 Ambient EC attributed to diesel exhaust hi the studies described above ranges from 46%
9 to 68%. A lower bound estimate of DPM from ambient EC measurements in areas with similar
10 source contributions to those in the Phoenix and Denver areas can be estimated using the
11 equation:
12 DPM = (EC * 0.46)/0.75 or DPM = HC * 0.62
13
14 An upper bound estimate uses the equation:
15
16 DPM = (EC * 0.68)/0.52 or DPM = HC * 1.31
17
18 Using the average of the ranges provides the equation:
19
20 DPM = HC * 0.89.
21
22 Clearly the choice of a point estimate can provide a surrogate calculation of DPM that
23 can vary by at least a factor of 2. To assess the usefulness and applicability of the surrogate
24 calculation, the average DPM concentration predicted by extended CMB analysis can be
25 compared with DPM concentrations predicted by the EC surrogate calculation. The average
26 DPM concentrations reported by CMB for the Colorado and Phoenix sites are within the range
27 of DPM concentrations estimated by the EC surrogate method (Table 2-24).
28 While a recommended surrogate DPM calculation method is not provided here, it is
29 evident that on an annually averaged basis, a surrogate calculation such as the one used above
30 may provide a reasonable estimate of DPM that bounds the actual concentration in those ambient
31 environments with similar source contributions to those in the Denver and Phoenix areas. The
32 surrogate DPM calculation is used here to illustrate the usefulness of this approach for
33 estimating DPM in the absence of a more sophisticated receptor modeling analysis for locations
34 where fine PM EC concentrations are available.
35 One source of variability in EC concentrations reported for ambient studies is the
36 measurement method used to quantify EC. As discussed in Section 2.2.8.1, EC and OC are
37 operationally defined. Ambient samples are typically analyzed for EC using thermal optical
7/25/00 2-58 DRAFT—DO NOT CITE OR QUOTE
-------
reflectance or thermal optical transmittance. The measurement technique for the studies
2 discussed here is noted since TOR methods often report higher EC levels compared to TOT
3 analyses (Birch, 1998; Morris et al., 2000).
4 Table 2-25 provides a lower and upper bound DPM estimate from annual average EC
5 concentrations for three urban areas in addition to DPM concentrations reported from EC
6 measurements for the MATES- II (SCAQMD, 2000). Under an EPA research grant with the
7 Northeastern States for Coordinated Air Use Management (NESCAUM), PM2.5 samples were
8 collected every 6 days for 1 year (1995) in Boston (Kenmore Square), MA, and Rochester, NY,
9 and were analyzed for EC using TOT (Salmon et al., 1997). DPM concentrations are estimated
10 to be in the range from 0.8 |ig/m3 to 1.7 ng/m3 in Boston, and from 0.4 |ig/m3 to 0.8 p-g/m3 in
11 Rochester (Table 2-25).
12 The Interagency Monitoring of Protected Visual Environments (IMPROVE) project
13 being conducted by the National Park Service includes an extensive aerosol monitoring network
14 mainly in rural or remote areas of the country (national parks, national monuments, wilderness
15 areas, national wildlife refuges, and national seashores), and also in Washington, DC (Sisler,
16 1996). PM2.5 samples, collected from March 1992 through February 1995 twice weekly for 24-
17 hour duration at 43 sites (some co-located in the same rural park area), were analyzed for a suite
B of chemical constituents, including EC (using TOR). EC concentrations in these rural locations
19 may have EC source contributions quite different from those in the urban areas in which the
20 fraction of EC attributable to diesel exhaust has been reported. The lack of information
21 regarding EC sources in these rural locations makes the application of the EC surrogate highly
22 uncertain. It is noteworthy that annual average EC concentrations in the rural and remote
23 regions reported as part of the IMPROVE network range from 0.1 H-g/m3 for Denali National
24 Park, AK, to 0.9 |ig/m3 for the Lake Tahoe, CA, area. In Washington, DC, the annual average
25 EC concentration of 1.7 |ig/m3 is estimated as an annual average DPM concentration of 1.4
26 |ig/m3.
27 The annual average EC measurements in Washington, DC, suggest that the DPM
28 concentrations are in the range from 1.0 |ig/m3 to 2.2 p.g/m3, accounting for 5% to 12% of
29 ambient PM2.5. Seasonally averaged data for the Washington, DC, site indicates that EC
30 concentrations and, by extension, DPM concentrations at this site peak in the autumn and winter
31 (2.0 |ag/m3 and 0.9 |lg/m3 EC, respectively).
32 DPM concentrations reported recently as part of the MATES-II study at eight locations
33 ranged from 2.4 ng/m3 to 4.5 |ig/m3. DPM concentrations at Huntingdon Park and Pico Rivera,
34 CA, were higher than other DPM concentrations in the South Coast Air Basin, potentially due to
^P higher diesel truck traffic, proximity to nonroad diesel sources, or nondiesel sources of EC,
36 including gasoline vehicle traffic.
7/25/00 2-59 DRAFT—DO NOT CITE OR QUOTE
-------
1 2.4.2.3. Dispersion Modeling Results
2 Dispersion models estimate ambient levels of PM at a receptor site on the basis of
3 emission factors for the relevant sources and parameters that simulate atmospheric processes
4 such as the advection, mixing, deposition, and chemical transformation of compounds as they
5 are transported from the source to the receptor site(s). Cass and Gray (1995), Gray and Cass
6 (1998), and Kleeman and Cass (1998) have applied dispersion models to the South Coast Air
7 Basin to estimate DPM concentrations. The models used by these investigators applied emission
8 factors from 1982 and consequently are representative of concentrations prior to the
9 implementation of DPM emission controls. Dispersion modeling has also been performed as
10 part of the EPA National Air Toxics Assessment (NATA) using the Assessment System for
11 Population Exposure Nationwide (ASPEN) to estimate ambient concentrations of DPM from on-
12 road and nonroad sources. The results from this analysis will be available in 2000. In addition
13 to offering another approach for estimating ambient DPM concentrations, dispersion models can
14 provide the ability to distinguish on-highway from nonroad diesel source contributions and have
15 presented an approach for quantifying the concentrations of secondary aerosols from diesel
16 exhaust.
17 Cass and Gray (1995) used a Lagrangian particle-in-cell model to estimate the source
18 contributions to atmospheric fine carbon particle concentrations in the Los Angeles area,
19 including diesel emission factors from on-highway and off-highway sources. Then- dispersion
20 model indicates that for 1982, the annual average ambient concentrations of DPM ranged from
21 1.9 Hg/m3 in Azusa, CA, to 5.6 p-g/m3 in downtown Los Angeles (Table 2-26). The contribution
22 of on-highway sources to DPM ranged from 63.3% in downtown Los Angeles to 89% in west
23 Los Angeles. Of the on-highway diesel contribution, the model predicted that for southern
24 California, HD trucks made up the majority (85%) of the DPM inventory, and overall they
25 contributed 66% of the DPM in the ambient air. Nonroad sources of diesel exhaust include
26 pumping stations, construction sites, shipping docks, railroad yards, and heavy equipment repair
27 facilities. Cass and Gray (1995) also report that wintertime peaks in DPM concentrations can
28 reach 10 }J,g/m3.
29 Kleeman and Cass (1998) developed a Lagrangi^n model that examines the size and
30 chemical evolution of aerosols including gas-to-particle conversion processes during transport.
31 This model was applied to one well-characterized episode in Claremont, CA, on August
32 27-28, 1987. The model provided reasonable predictions of PM10 (overpredicting PM10 13%),
33 EC, and OC, and it adequately reconstructed the size distribution of the aerosols. The model
34 indicated that on August 27-28. 1987. the PM2.5 concentration was 76.7 }-Lg/:n3, 13.2% of which
35 (10.1 |J,g/m3) was attributable to diesel engine emissions Thic estimate includes secondary
36 aerosol formation for sulfate, ammonium, nitrate, and organic compounds, which accounted for
7/25/00 2-60 DRAFT—DO NOT CITE OR QUOTE
-------
4.9 Hg/m3 of the total estimated DPM mass. The secondary organic aerosol was estimated to be
2 1.1 jig/m3, or 31 % of the total secondary aerosol mass, with the remainder composed of nitrate,
3 ammonium, and sulfate aerosols.
4
5 2.4.3. Exposures to Diesel Exhaust
6 Ultimately, personal exposure determines health impacts. To understand the distribution
7 of risk in the population, the distribution of exposures among the general population and more
8 highly exposed groups needs to be assessed. An exposure assessment addresses the exposure
9 profile for the general public, including the distribution of ambient diesel exhaust exposures in
10 different geographic and demographic regions; the most highly exposed (90th percentile),
11 exposures in microenvironments for short and long durations, and the maximum exposure range
12 (98th percentile) and number of maximum exposed individuals. Because diesel exhaust is a
13 mixture of particles and gases, one must choose a measure of exposure (i.e., dosimeter); ng/m3
14 of DPM has historically been used in many studies as the dosimeter for the entire diesel exhaust
15 .mixture. Because of the lack of data, a comprehensive exposure assessment cannot be
16 completed at this point. Exposure models are under development at EPA and elsewhere and
17 personal exposure monitoring is being conducted that will improve this assessment of the
A distribution of exposures to DPM.
19 In the following sections, modeled average exposures and some information reflecting
20 potential exposures for those who spend a large portion of their time outdoors are presented.
21 Occupational exposures to DPM are summarized for the variety of workplaces in which diesel
22 engines are used. These occupational exposures are placed into context with equivalent
23 environmental exposures to understand the potential for overlap in average occupational and
24 average ambient exposures.
25
26 2.4.3.1. Occupational Exposure to Diesel Exhaust
27 The National Institute for Occupational Safety and Health (NIOSH, 1988) estimates that
28 approximately 1.35 million workers are occupationally exposed to diesel exhaust emissions.
29 Workers exposed to diesel exhaust emissions include mine workers, railroad workers, bus and
30 truck drivers, truck and bus maintenance garage workers, loading dock workers, fire fighters,
31 heavy equipment operators, and farm workers.
32 Measurements of DPM exposure in occupational environments have included respirable
33 particulate (<3.5 um), smoking-corrected respirable paniculate, combustible paniculate, and EC
among other methods. Occupational exposures to DPM as well as breathing zone concentrations
of DPM have been described in some detail by Watts (1995), Hammond (1998), the World
7/25/00 2-61 DRAFT—DO NOT CITE OR QUOTE
-------
1 Health Organization (1996), and Birch and Gary (1996) and are briefly, but not
2 comprehensively, summarized here.
3 The highest occupational exposures to DPM are for workers in coal mines and noncoal
4 mines using diesel-powered equipment. These exposures, reported by several investigators,
5 range from approximately 10 fig/m3 to 1,280 ng/m3 (Table 2-27 ). Rogers and Whelan (1999)
6 report exposures to specific DPM-associated PAHs (including naphthalene, fluorene,
7 phenanthrene, pyrene, and benz[a]anthracene) for mine workers using diesel fuels containing
8 low and high levels of sulfur, aliphatic, and aromatic compounds. Results of this study indicate
9 that the composition of DPM to which workers were exposed varies considerably based on
10 engine condition, fuel, and other operating parameters. Mine worker exposures to PAH
11 compounds were highest for naphthalene, ranging from 1,312 ng/g to 3,228 ng/g of organics
12 and exposures were lowest for benz[a]anthracene, ranging from less than 3 ng/g up to 18 ng/g of
13 organics.
14 Other investigators have reported DPM-associated PAH concentrations that do not
15 necessarily represent personal exposures but are a snapshot of short periods of elevated
16 concentration that make up a portion of a worker's daily exposure. Bagley et al. (1991, 1992)
17 reported levels of B[a]P ranging from below the detection limit of 0.05 ng/m3 to 61 ng/m3
18 collected only during periods of mining activity. Watts (1995) reported DPM concentrations in
19 four mines collected during significant diesel activity, which range from 850 p.g/m3 to 3,260
20 |J.g/m3. Heino (1978) reports DPM concentrations for locomotive engineers reaching 2,000
21 ^ig/m3.
22 In a study of four railroads, Woskie et al. (1988) reported concentrations of respirable
23 dust (corrected for cigarette smoke paniculate) that ranged from 39 ng/m3 for engineers/firers to
24 134 ng/m3 for locomotive shop workers and 191 ng/m3 for hostlers. Woskie et al. (1988) also
25 reported smoking-corrected respirable dust for railroad clerks (17 ng/m3) who are considered to
26 be not exposed to diesel exhaust. Although these exposures may have included nondiesel PM
27 (background respirable dust levels have been estimated to have contributed approximately 10
28 ng/m3 to 33 jig/m3 for this study), the majority of the respirable PM is believed to have
29 originated from the diesel locomotive emissions. DPM exposures reported for firefighters
30 operating diesel engine vehicles range from 4 ng/mj to 748 |ig/mj, which also encompasses the
31 range of DPM exposures reported for airport ground crew and public transportation system
32 personnel (7 ng/m3 to 98 ng/m3).
33 Studies reporting diesel exhaust exposure among fire station employees typically report
34 paniculate levels below 100 ng/m3 (ranging from 4 ng/m3 to 79 ng/m3) (NIOSH, 1992; Birch and
35 Gary 1996). In a study by Froines et al. (1987), DPM exposures for firefighters in two stations
36 ranged from 39 ng/m3 to 73 ng/m3. DPM were also reported for airport ground crew and public
7/25/00 2-62 DRAFT—DO NOT CITE OR QUOTE
-------
transit workers by Birch and Gary (1986), ranging from 7 |ag/m3 to 15 ng/m3 for airport ground
2 crews and 15 ng/m3 to 98 jig/m3 for public transit workers. Dock workers using diesel-powered
3 forklifts have been reported to have DPM exposures ranging from 6 |ig/m3 to 61 ng/m3 (NIOSH
4 1990; Zaebst et al., 1991). In studies by NIOSH (1990), and Fowler (1985), the organic material
5 measured accounted for about one-half to almost all of the carbonaceous DPM exposures,
6 providing evidence that some pieces of nonroad equipment (forklifts and construction
7 equipment) emitted DPM with a significant OC fraction in the 1980s and early 1990 timeframe.
8 Zaebst et al. (1991) also reported DPM exposures for mechanics, road drivers, and local
9 drivers for 8-hour shifts at each of six large hub truck terminals. Residential background and
10 highway background samples at fixed sites were also collected during warm-weather and cold-
11 weather periods, and the geometric mean for DPM concentrations ranged from 1 jJ-g/m3 to 5
12 pg/rn3. DPM exposures for road and local truckers in warm- and cold-weather periods ranged
13 from 2 |ig/m3 to 7 Hg/m3, while exposure levels for mechanics were reported between 5 (ig/m3
14 and 28 p.g/m3 (geometric means).
15 Kittelson et al. (2000) are measuring DPM exposures for bus drivers, parking garage
16 attendants, and mechanics. Personal exposures for bus drivers on four different routes range
17 from 1 ng/m3 to 3 |ig/m3, and exposure among parking ramp attendants averaged 2 |ig/m3.
B These results are preliminary, and data for the mechanics have not yet been analyzed. This study
19 will also characterize PAH compounds to which these workers are exposed.
20 Bus garage workers have also been assessed for exposure to diesel exhaust using
21 biomarkers for particle-associated benzene exposure (Muzyka et al., 1998) and urinary excretion
22 of 8-oxo-2'-deoxyguanosine (Loft et al., 1999). Other biomarkers of diesel exhaust exposure in
23 occupational workers have included measurements of urinary 1-hydroxypyrene, adducts of DNA
24 and hemoglobin, and 8-hydroxyguanosine hi lung tissue (Nielsen et al., 1996; Tokiwa et al.,
25 1999; Zwimer-Baier and Neumann, 1999; Kara et al., 1997).
26 To estimate an environmental exposure that is equivalent to an occupational lifetime
27 exposure, the fraction of lifetime worker inhalation exposure (calculated as the amount of air
28 breathed on the job multiplied by the typical amount of time spent on the job) is calculated
29 relative to 70-year lifetime inhalation exposure: (10 m3/shift/20m3/day) * (5 days/7days) * (48
30 weeks/52 weeks) * (45-year career/70-year lifetime) = 0.21. Using this calculation, 21% of an
31 annual average occupational lifetime exposure is roughly equivalent to a 70-year annual average
32 lifetime environmental exposure. The equivalent environmental exposures for the occupational
33 exposures presented in Table 2-28 range from 0.6 |ig/m3 to 14 ^ig/m3 for truckers, dock workers,
34 and mechanics, and from 2 ng/m3 to 269 ^g/m3 for miners. The low end of the range of
^P environmental equivalent exposures for several of the occupational settings overlap with average
36 exposures predicted from the HAPEM model for on-road sources (described below) and with
7/25/00 2-63 DRAFT—DO NOT CITE OR QUOTE
-------
1 ambient concentrations from all sources reported for DPM in urban areas in the 1990-96
2 timeframe. The overlap between some occupational exposures and environmental exposures as
3 well as the small difference between occupational environmental equivalent exposures and
4 environmental exposures are a significant concern and suggest the potential for significant risk in
5 the general population. The potential for cancer risk in the general population is discussed in
6 Section 8.3.
7
8 2.4.3.2. Ambient Exposure to Diesel Exhaust
9 Modeled estimates of population exposures to DPM integrate exposure hi various indoor
1 0 and outdoor environments and also account for the demographic distribution, tune-activity
1 1 patterns, and DPM concentrations in the various environments, including job-related exposures.
1 2 Two modeling efforts have been developed to determine DPM exposures in the general
1 3 population: the Hazardous Air Pollutant Exposure Model-Mobile Sources, version 3 (HAPEM-
1 4 MS3), and the California Population Indoor Exposure Model (CPIEM). EPA is currently
1 5 developing version 4 of the HAPEM, which will provide State-specific average exposures for
16 DPM.
17
1 8 2.4.3.2.1. The Hazardous Air Pollutant Exposure Model - Mobile Sources, version 3. To
1 9 estimate population exposures to DPM, EPA currently uses HAPEM-MS3 (U.S. EPA, 1999c).
20 This model provides national and urban-area specific exposures to DPM from on-road sources
21 only. HAPEM-MS3 is based on the CO probabilistic National Ambient Air Quality Standards
22 (NAAQS) exposure model (pNEM/CO), which is used to estimate the frequency distribution of
23 population exposure to CO and the resulting carboxyhemaglobin levels (Law et al., 1997).
24 HAPEM simulates the CO exposure scenario of individuals in 22 demographic groups for 37
25 microenvironments. CO concentrations are based on ambient measurements made in 1990 and
26 are related to exposures of individuals in a 10 km radius around the sampling site. DPM
27 exposures are calculated as in Equation 2-4, using a ratiometric approach to CO.
28
29 DPM^CO ICOgim )x DPMglm (2-5)
30 Data provided to the model includes CO monitoring data for 1990; time-activity
31 data collected in Denver, Washington, DC, and Cincinnati from 1982 to 1985;
32 microenvironmental data; and 1990 census population data. Motor vehicle DPM and CO
33 emission rates reported by EPA (1999c) are used to calculate mobile-source DPM exposures,
34 and exposures in future years are projected based on the increase in vehicle miles traveled.
35 EPA's PART5 model is used to estimate DPM emission rates (g/mi) for the fleet as a whole in
7/25/00 2-64 DRAFT— DO NOT CITE OR QUOTE
-------
)| any given calendar year. PARTS is currently being modified to account for deterioration, actual
2 in-use emissions, poor maintenance, and tampering effects, all of which increase emission
3 factors. As a result, HAPEM-MS3 exposure estimates based on PARTS emission factors may
4 underestimate true exposures from on-road sources. A comparison of PARTS HD diesel vehicle
5 emission factors with those presented earlier in this chapter suggest that PARTS may
6 underestimate HD diesel vehicle emissions by up to 50%.
7 HAPEM-MS3 assumes that the highway fleet (gasoline plus diesel) emissions ratio of
8 CO to DPM can be used as an adjustment factor to convert estimated CO personal exposure to
9 DPM exposure estimates. This assumption is supported by the observation that even though
10 gasoline vehicles emit the large majority of CO, gasoline and diesel highway vehicles travel on
11 the same roadways, albeit with somewhat different spatial and temporal patterns. DPM and CO
12 are both relatively long-lived atmospheric species (1-3 days) except under certain conditions;
13 therefore the model does not account for chemical and physical differences between the DPM
14 • and CO, and the model assumes that for the average person in a modeled air district, CO and
15 DPM are well mixed. Exposure in microscale environments in which these assumptions may not
16 be valid were not modeled.
17 A validation study conducted for the pNEM/CO model on which HAPEM-MS3 is based
nJ indicates that CO exposures for the population in the S^1 percentile were overestimated by
19 approximately 33%, while those with exposures in the 98^ percentile were underestimated by
20 about 30%. This validation study is considered applicable to the HAPEM-MS3 model. To
21 address the underestimate of exposures for the most highly exposed, Brodowicz (1999) used CO
22 concentrations relevant to the most highly exposed populations to determine DPM exposures for
23 different demographic groups within this population; the results are discussed below.
24 Annual average DPM exposures from on-road vehicles nationwide for the general
25 population, rural and urban population, outdoor workers, and urban children are reported for
26 1990 and 1996 and projected for 2007 and 2020 in Table 2-29. The modeled annual average
27 DPM exposure nationwide (urban and rural areas) in 1996 from on-road sources only was 0.7
28 [Lg/m3. The annual average exposure in urban areas for the same year was 0.7 }J,g/m3, and the
29 modeled exposure for rural areas was 0.3 jig/m3. Among the demographic groups modeled,
30 urban outdoor workers in general were found to have the highest average exposure to DPM,
31 averaging 0.8 p.g/m3 from on-road sources in 1996. DPM exposures attributable to on-road
32 sources are projected to decrease until approximately 2007 due to fleet turnover and the full
33 implementation of Federal regulations that are currently in place. After 2007, the increase in
34 vehicle miles traveled will offset reductions in PM emissions, and consequently, exposure to
> DPM will begin to increase. Nationally in 1996, 97% of DPM exposure from on-road vehicles
36 was attributable to HD diesel vehicles, and the rest was generated mainly by LD diesel trucks. If
7/25/00 2-65 DRAFT—DO NOT CITE OR QUOTE
-------
1 the modeled increase in diesels in the LD truck fleet occurs, projected DPM exposures in 2007
2 are expected to increase 38% on average over 1996 exposures.
3 Since diesel vehicle traffic, and therefore exposure fo DPM, varies for different urban
4 areas, HAPEM-MS3 was used to estimate annual average population exposures for 10 urban
5 areas. Modeled 1996 DPM exposures in the cities ranged from 0.5 Hg/m3 in Chicago and St.
6 Louis to 1.2 ng/m3 in Phoenix (Table 2-30). In 1996, estimated average DPM exposure from
7 on-road sources was higher than the national average in five cities: Atlanta, Minnneapolis, New
8 York, Phoenix, and Spokane.
9 Since HAPEM-MS3 is suspected to underestimate exposures in the highly exposed
10 populations, 1990 CO concentrations relevant to the most highly exposed populations were used
11 to determine 1990 DPM exposures for different demographic groups in this population. The
12 highest DPM exposures ranged from 0.8 ng/m3 for outdoor workers in St. Louis to 2.0 jig/m3 for
13 outdoor workers in Spokane and up to 4.0 ng/m3 for outdoor children in New York (Table 2-30).
14 The highest exposed demographic groups were those who spend a large portion of their time
15 outdoors. It is important to note that these exposure estimates are lower than the total exposure
16 to DPM since they reflect only DPM from on-road sources and do not yet reflect exposure to
17 non-road DPM emissions.
18 Exposure estimates provided by HAPEM-MS3 for the general population in 1996 are
19 lower than the ambient DPM concentrations reported for urban areas in Section 2 A..2.1 that
20 ranged from 1.2 (Jtg/m3 to 3.6 |lg/m3 in a similar timeframe. Since most people in the general
21 population spend a large portion of time indoors, average exposures to DPM are expected to be
22 lower than ambient concentrations.
23
24 2.4.3.2.2. Personal Exposures: Microenvironments/Hotspots Personal monitoring for DPM
25 exposure has focused on occupationally exposed groups, including railroad workers, mine
26 workers, mechanics, and truck drivers. While some studies have measured personal exposures
27 to ambient PM, none have conducted detailed chemical analysis to quantify the portion of PM
28 attributable to diesel exhaust (e.g., using extended species CMB, discussed above). Elemental
29 carbon concentrations have been reported for some microenvironments and are discussed in this
30 section. Microenvironmentai exposures of significant concern include in-vehicle exposures such
31 as school buses and passenger cars as well as near highways and in urban canyons. Since DPM
32 from mobile sources is emitted into the breathing zone of humans, this source has a greater
33 potential for human exposure (per kg of emissions) compared to combustion particulates emitted
34 from point sources.
35 Recent EC measurements.reported for enclosed vehicles driving on Sacramento
36 roadways ranged from below detection limits up to 10 |ig/m3 and from 3 (J-g/m3 to 40 |lg/m3 on
-7 n «/nn ->
-66 DRAFT—DO NOT CITE OR QUOTE
-------
(lp)l Los Angeles roadways. Elevated levels of PM2.5 and EC were observed when the vehicle being
2 followed was powered by a HD diesel truck or bus (Cal EPA, 1998b). These measurements are
3 likely to include some EC from gasoline vehicles. The SHEDS (Stochastic Human Exposure
4 and Dose Simulation) model for PM predicts that although the typical person only spends
5 approximately 5% of his or her time in a vehicle, this microenvironment can contribute on
6 average 20% and as much as 40% of a person's total PM exposure (Burke et ah, 2000).
7 The California Air Resources Board also collected EC near the Long Beach Freeway for
8 4 days in May 1993 and 3 days in December 1993 (Cal EPA, 1998a). Using emission estimates
9 from their EMFAC7G model and EC and OC composition profiles for diesel and gasoline
10 exhaust, tire wear, and road dust, CARB estimated the contribution of the freeway to DPM
11 concentrations. For the 2 days of sampling in December 1993, diesel exhaust from vehicles on
12 the nearby freeway were estimated to contribute from 0.7 p,g/m3 to 4.0 ^ig/rn3 excess DPM above
13 background concentrations, with a maximum of 7.5 ^ig/m3.
14 In 1986, EC concentrations were measured in Glendora, CA, during a carbonaceous
15 aerosol intercomparison study (Cadle and Mulawa, 1990; Hansen and Novakov, 1990). One
16 technique used during the study reported EC concentrations in 1-minute intervals, reflecting the
17 impact from diesel vehicles 50 m from the study site. The diesel vehicles were estimated to
contribute up to 5 [ig/m3 EC above the background concentration.
19 In a study designed to investigate relationships between diesel exhaust exposure and
20 respiratory health of children in the Netherlands, EC measurements were collected in 23 schools
21 located from 47 m to 377 m from a freeway and in 8 schools located at a distance greater than
22 400 m from a freeway (Brunekreef, 1999). EC concentrations in schools near freeways ranged
23 from 1.1 [ig/m3 to 6.3 {J,g/m3, with a mean of 3.4 |J,g/m3, and EC concentrations in schools more
24 than 400 m from freeways ranged from 0.8 jig/m3 to 2.1 ^Ag/rn3, with a mean of 1.4 |j,g/m3.
25 Brunekreef et ah (2000), using a reflectance method to report "soot" or carbonaceous particulate
26 concentrations as a surrogate for EC, found a statistically significant increase in carbonaceous
27 particle concentrations inside and outside of the schools with increasing truck traffic
28 (predominantly diesel), with decreasing distance between the school and the highway, and with
29 an increase in the percent of time the school was downwind of the highway.
30 While currently there is little quantitative information regarding personal exposure to
31 DPM, certain exposure situations are expected to result in higher than average exposures. Those
32 in the more highly exposed categories would generally include people living in urban areas in
33 which diesel delivery trucks, buses, and garbage trucks frequent the roadways, but also included
34 would be people living near freeways, bus stations, construction sites, train stations, marinas
frequented by diesel-powered vessels, and distribution hubs using diesel truck transport. One
36 study using the 1 -hydroxypyrene biomarker of diesel exhaust exposure reported diesel exhaust
7/25/00 2-67 DRAFT—DO NOT CITE OR QUOTE
-------
1 exposure among most (76%) of the 26 adolescents sampled in Harlem (Northridge et al., 1999).
2 In a follow-on study, Kinney et al. (1999) reported EC concentrations from personal monitors
3 worn by study staff on sidewalks at four Harlem intersections that ranged from 1 .5 |ig/m3 to 6
4 Hg/m3. The EC concentrations were found to be associated with diesel bus and truck counts
5 such that spatial variations in sidewalk concentrations of EC were attributed to local diesel
6 sources in Harlem.
7 In any situation in which diesel engines operate and a majority of time is spent outdoors,
8 personal exposures to diesel exhaust are expected to exceed average exposures. Since a large but
9 currently undefined portion of DPM is emitted during acceleration, those living and working hi
1 0 the vicinity of sources operating in this transient mode could experience highly elevated levels of
1 1 DPM. DPM enriched hi soluble organic material (as opposed to EC) is emitted from LD
1 2 vehicles, some nonroad equipment, on-road diesel engines during cold-start and motoring
1 3 conditions, and poorly maintained vehicles. The potential health effects of acute exposures to
1 4 elevated DPM levels as well as health effects resulting from chronic exposures are discussed in
1 5 subsequent chapters in this document.
16
1 7 2.4.3.2.3. The California Population Indoor Exposure ModeL CPIEM, developed under
1 8 contract to the CARB, estimates Califomians' exposure to DPM using distributions of input data
1 9 and a Monte Carlo approach (Cal EPA, 1998a). This model uses population-weighted outdoor
20 DPM concentrations in a mass balance model to estimate DPM concentrations hi four indoor
21 environments: residences, office buildings, schools, and stores/retail buildings. The model takes
22 into account air exchange rates, penetration factors, and a net loss factor for deposition/removal.
23 In four additional environments (industrial plants, restaurants/lounges, other indoor places, and
24 enclosed vehicles), assumptions were made about the similarity of each of these spaces to
25 environments for which DPM exposures had been calculated. Industrial plants and enclosed
26 vehicles were assumed to have DPM exposures similar to those in the outdoor environment;
27 restaurant/lounges were assumed to have DPM concentrations similar to stores; and other indoor
28 places were assumed to have DPM concentrations similar to offices. The estimated DPM
29 concentrations in the indoor and outdoor environments range from 1.6 Hg/m3 to 3.0 |ig/m3
3 1 The DPM concentrations reported in Table 2-3 1 were used as input to CPIEM, and time-
32 activity patterns for children and adults were used to estimate total indoor and total air exposures
33 to DPM. Overall, total indoor exposures were estimated to be 2.0 ± 0.7 u.g/m3, and total air
34 exposures (indoor and outdoor exposures) were 2.1 ± 0.7 p.g/m3 (Table 2-32). The South Coast
35 Air Basin and the San Francisco Bay Area were also modeled using CPIEM, where total air
36 exposures to DPM were estimated to be 2.5 ± 0.9 Jlg/m' and 1 .7 ± 0.9 Hg/mJ, respectively.
7/25/00 2-68
DRAFT—DO NOT CITE OR QUOTE
-------
Exposure estimates were also made by Cal EPA (1998a) for 1995, 2000, and 2010 using
2 a ratiometric approach to 1990 exposures. Total air exposures reported for 1995 and projected
3 for 2000 and 2010 were 1.5 |ig/m3,1.3 M-g/m3, and 1.2 Jig/m3, respectively.
4
5 2.5. SUMMARY
6 This chapter summarizes information regarding the history of the use of diesel engines,
7 technological developments and their impact on emissions over time, Federal standards on diesel
8 exhaust, the chemical and physical character of diesel exhaust, atmospheric transformations of
9 diesel exhaust, and ambient diesel exhaust concentrations and exposures. The aspects of each of
10 these topics that are most relevant to the discussion of health effects in later chapters of this
11 document are summarized here. Since the majority of information regarding the chemical
12 composition and historical changes hi diesel exhaust pertains to on-road diesel engines, these
13 data are discussed in greater detail than diesel emissions from nonroad equipment. Where
14 possible, nonroad emissions were discussed in Chapter 2 and are briefly summarized here.
15
16 2.5.1. History of Diesel Engine Use, Standards, and Technology
The use of diesel engines in the trucking industry began in the 1940s, and diesel engines
slowly displaced gasoline engines among heavy HD trucks comprising 36% of new HD truck
19 sales in 1960, 85% of sales in 1970, and almost 100% of sales in 1997. It is estimated that in
20 2000, HD diesel vehicles will travel over 224 billion miles (U.S. EPA, 2000b). In 1997, on-
21 highway HD diesel engines contributed 66% of the PM2.5 emitted by on-highway vehicles.
22 To understand changes in emissions over time, it is important to note the difference
23 between model year emission trends and calendar year emission trends. Emission trends by
24 model year refer to the year hi which an engine was made; the emission rate is specific to the
25 technology and regulations in effect for that year. Emissions in a specific calendar year refer to
26 aggregate emissions due to the mix of model year engines on the road. Due to the tune required
27 for fleet turnover, emission rates for the on-road fleet in any calendar year are not as low as the
28 most recent model year emission rate. In 1997,40% of the HD vehicles on the road were at least
29 10 years old arid traveled approximately 17% of HD vehicle miles traveled.
30 EPA set a smoke standard for on-road HD diesel engines beginning with the 1970 model
31 year. In the ensuing years, standards for PM from diesel engines for on-road applications
32 decreased from 0.6 g/bhp-hr in 1988 to 0.1 g/bhp-hr for trucks in 1994-1995 and 0.05 g/bhp-hr
for buses in 1996-1997. Calendar year emission contributions of PM from diesel engines to
national PM10 inventories reflect decreases expected to result from Federal regulations since the
35 emission factor models (MOBILES and PARTS) used to provide emission estimates for mobile
7/25/00 2-69 DRAFT—DO NOT CITE OR QUOTE
-------
1 sources largely use engine test data required for certification. The U.S. EPA Trends Report
2 estimates that PM10 emissions attributed to on-road diesel vehicles decreased 27% in the period
3 between 1980 and 1998. DPM emission factors (g/mi by model year) measured from in-use
4 vehicles have, on average, decreased by a factor of six in the period from the mid-1970s to the
5 mid-1990s.
6 It is important to note that in spite of the decreasing trend in DPM emission factors by
7 model year, a wide range in emission factors from in-use testing are reported, even for newer
8 model year HD vehicles (from less than 0.1 g/mi to over 1 g/mi for model year 1996 vehicles).
9 The high variability in DPM emissions within one model year has been attributed to
10 deterioration3 and differences in measurement methods and test conditions at the various testing
11 facilities. Studies in which consistent testing methods were used suggest that deterioration (even
12 for newer model year engines) causes some of the variability in emission factors, while other
13 studies clearly demonstrate the important influence of test conditions and driving protocols (e.g.,
14 aggressive driving) on DPM emission factors.
15 While significant reductions in DPM from diesel vehicle emissions for on-road
16 applications have been realized, diesel engines (nonroad and on-road combined) are still
17 significant contributors to 1998 inventories of particulate matter, contributing approximately
18 23% of PM2.5 emissions (not including the contribution from natural and miscellaneous
19 sources). As the result of fleet turnover and full implementation of Federal emission standards
20 currently in place, DPM emissions from on-road diesel engines are projected to decrease until
21 2007, at which time increases in vehicle miles traveled offset current emission standards and the
22 DPM emissions begin to increase.
23 Technology innovations that impact diesel engine emissions have occurred in the years
24 since 1960, in particular the advent of turbocharging with charge air cooling and direct injection
25 engines. The use of these new technologies tends to lower emissions from on-road diesel
26 engines; until the late 1970s, however, engines were optimized for performance rather than
27 emissions, so the effect on emissions prior to this time was small. Overall, it is expected that on-
28 road engines in the 1950 to 1970 timeframe would have DPM emissions similar to those of the
29 mid-1970 engines that were not yet controlled for particulates.
30 Limited data are available to assess the changes in emission rates from locomotive,
31 marine, or other nonroad diesel engine sources over time. It is expected that since the typical
Deterioration includes increase? in emission rates (g/bhp hr) due to normal wear as well as manufacturing defects
and malfunctions such as retarded timing, fuel injector malfunction, smoke limiting mechanism problems, clogged
air filter, wrong or worn turbocharger, clogged intercooler, engine mechanical failure, excess oil consumption, and
electronics that have been tampered with or have failed.
7/25/00 2-70 DRAFT—DO NOT CITE OR QUOTE
-------
lifespan of a locomotive engine is at least 40 years and since PM regulations for these engines do
2 not take effect until 2000, PM emission rates by model year from locomotives are not likely to
3 have changed substantially since the introduction of the diesel engine into the railroad industry in
4 the early 1950s.
5 Particulate matter regulations for nonroad diesel equipment are not as stringent as PM
6 regulations for on-road diesel engines. While PM emissions have declined for on-road trucks, it
7 is estimated that PM10 emissions from nonroad diesel engines increased 17% between 1980 and
8 1998. DPM emissions from nonroad diesel engines are expected to continue to increase from
9 current levels in the absence of new regulations. No information is available regarding changes
10 in the chemical composition of nonroad engine emissions over time.
11
12 2.5.2. Physical and Chemical Composition of Diesel Exhaust
13 Complete and incomplete combustion of fuel in the diesel engine results in the formation
14 of a complex mixture of hundreds of organic and inorganic compounds in the gas and particle
15 phases. Among the gaseous components of diesel exhaust, the aldehydes are particularly
16 important because of their health effects and because they comprise an important fraction of the
f gaseous emissions. Formaldehyde comprises the majority of the aldehyde emissions (65%-80%)
from diesel engines, with the next most abundant aldehydes being acetaldehyde and acrolein.
19 Other gaseous components of diesel exhaust that are notable for their health effects include
20 benzene, 1,3-butadiene, PAH, and nitro-PAH. Dioxin compounds have also been detected in
21 trace quantities in diesel exhaust and currently account for 1.2% of the national inventory.
22 Dioxin compounds are known to accumulate in certain foods, e.g., beef, vegetables, and diary
23 products. It is unknown whether deposition of diesel exhaust emissions has an impact on food
24 chains in local areas.
25 DPM contains EC, OC, and small amounts of sulfate, nitrate, metals, trace elements,
26 water, and unidentified compounds. DPM is typically composed of over 50% to approximately
27 75% EC depending on the age of the engine, deterioration, HD versus LD, fuel characteristics,
28 and driving conditions. The OC portion of DPM originates from unbumed fuel, engine
29 lubrication oil, and low levels of partial combustion and pyro lysis products and typically ranges
30 from approximately 19%-43%, although the range can be broader depending on many of the
31 same factors that influence the EC content of DPM. Polyaromatic hydrocarbons generally
32 comprise less than 1% of the DPM mass. Metal compounds and other elements in the fuel and
f engine lubrication oil are exhausted as ash and typically comprise l%-5% of the DPM mass.
Elements and metals detected in diesel exhaust include barium, calcium, chlorine, chromium,
35 copper, iron, lead, manganese, mercury, nickel, phosphorus, sodium, silicon, and zinc. The
7/25/00 2-71 DRAFT—DO NOT CITE OR QUOTE
-------
1 composition of DPM contrasts strongly with the typical chemical composition of ambient
2 DPM2.5 that is dominated by sulfate for aerosols measured in the eastern United States and by
3 nitrate, ammonium, and organic carbon in the western United States.
4 Approximately 1%-20% of the mass of diesel paniculate matter in diesel exhaust is in the
5 ultrafine size range (nuclei mode), with the majority of particles ranging in size from 0.005-0.05
6 microns and having a mean diameter of about 0.02 microns. These particles account for 50%-
7 90% of the number of particles. These ultrafine particles are largely composed of sulfate and/or
8 sulfate with condensed organic carbon.
9 Evidence regarding an increase in the number of ultrafine particles from new HD engines
10 is inconclusive. The dilution conditions used to measure the size distribution of diesel exhaust
11 have a large impact on the number of ultrafine particles quantified. To understand the size
12 distribution of DPM to which people are exposed will require measurements under conditions
13 that more closely resemble ambient conditions.
14 Approximately 80%-95% of the mass of particles in diesel exhaust is in the size range
15 from 0.05-1.0 microns, with a mean particle diameter of about 0.2 microns, and are therefore in
16 the fine PM size range. Diesel particles in the 0.05-1.0 micron range are aggregates of primary
17 spherical particles consisting of an EC core, adsorbed organic compounds, sulfate, nitrate and
18 trace elements. These particles have a very large surface area per gram of mass, which make
19 them an excellent carrier for adsorbed inorganic and organic compounds that can effectively
20 reach the lowest airways of the lung. The elemental carbon core has a high specific surface area
21 of approximately 30-90 m2/g.
22 Due to the potential toxicological significance of the organic components associated with
23 DPM, it is important to understand, to the extent possible, the historical changes in the amount
24 and composition of the DPM-associated organic fraction. The organic component of DPM has
25 typically been characterized by extraction with organic solvents, although other techniques such
26 as thermogravimetric methods have also been used. Results from studies using similar extraction
27 methods were compared to characterize historical changes in the SOF emission rates, the
28 percentage of DPM comprised by SOF. and the composition of SOF. Data from both engine and
29 chassis dynamometer tests suggest that SOF emission rates have decreased by model year from
30 1975 to 1995. When expressed as a percentage of total DPM, the contribution of SOF to total
31 DPM demonstrates a wide range of variability that may be attributed to different test cycles,
32 different engine types, and different deterioration rates among the vehicles tested. Currently, LD
oo -K^f^i „„,-.:—,,<. „„,;* r^nx* ,.~»u ~ u ;„;-.=,- A.o~<-;~~ ^--f c/"~nr +u~_ UT\ „__:_ —
*J*J lll\*OV*l WAAglllwb Wlllll. J-XA 1V1 Wllll CX HAgllWA JLAdWllWll WA Ot^TA IAXCL11 1 l±*f l>li££ll!C-^.
34 Chassis dynamometer tests demonstrate an overall decrease in the mass percentage
35 contribution of SOF to DPM ranging from 10%-60% in the 1980s and ~5%-20% in the 1990s.
7/25/00 2-72 DRAFT—DO NOT CITE OR QUOTE
-------
In contrast, engine dynamometer tests demonstrate that typically 10%-50% of DPM mass is
2 soluble organic matter for engines in model years 1980-1995. The higher SOF fraction of DPM
3 from 1990s model year engine dynamometer tests is attributed primarily to the differences in the
4 engine and chassis dynamometer driving cycles. The engine dynamometer testing includes high-
5 speed and low-load or low-speed lugging test modes in the engine Federal Test Procedure that
6 produce DPM with a high SOF fraction.
7 The chassis dynamometer data are considered to reflect real-world trends in emissions
8 from heavy HD vehicles by model year since vehicles from different model years, with different
9 mileage and different levels of deterioration, are represented. Thus it is expected that the
10 percentage of SOF from new (1990 or later) model year heavy HD diesel vehicles is lower than
11 that from older vehicles. This expectation is supported by data demonstrating an overall increase
12 in the fraction of EC in the carbonaceous component of DPM. The important observation from
13 the engine test data is that some driving modes occurring in real-world applications even with
14 new (post-1990) engines may produce DPM with a high SOF component (up to 50%).
15 PAH and nitro-PAH are present in DPM from both new and older engine exhaust. There
16 is no information to suggest that the overall PAH composition profile for DPM has changed.
There are too few data to speculate on the changes in emissions of total PAH, nitro-PAH, or
PAH and nitro-PAH components such as BaP and 1-NP. The data suggest that differences in a
19 vehicles' engine type and make, general engine condition, fuel composition, and test conditions
20 can influence the emissions levels of PAH. Some studies suggest that fuel composition is the
21 most important determinant of PAH emissions. There is limited evidence that gas-phase PAH
22 emission rates increase with higher fuel PAH content and that some particle-phase PAH emission
23 rates increase with higher fuel PAH content. These data suggest that during the period from
24 1960 to 1986, when the aromatic content of fuel increased, PAH emissions may have increased
25 until the aromatic content of diesel fuel was capped in 1993. The aromatic content of nonroad
26 diesel fuel is not federally regulated and is typically greater than 30%. PAH emissions from
27 nonroad equipment would also be expected to vary with the PAH content of the fuel.
28 Currently, information regarding emission rates, chemical composition, and relative
29 contribution of DPM from high-emitting HD diesel vehicles is not available and may
30 significantly change the current understanding of DPM composition to which people are
31 exposed. Some studies have reported a substantial number of smoking diesel trucks in the in-use
32 fleet. While the correlation between smoke and paniculate concentration varies with the driving
cycle and measurement method, the results of smoke opacity tests suggest that high-emitting HD
diesel vehicles may be important contributors to ambient diesel exhaust and DPM
35 concentrations.
7/25/00 2-73 DRAFT—DO NOT CITE OR QUOTE
-------
1 The chemical composition of DPM to which people are currently exposed is determined
2 by a combination of older and newer technology on-road and nonroad engines. Consequently,
3 the decrease in the SOF of DPM by model year does not directly translate into a proportional
4 decrease in DPM-associated organic material to which people are currently exposed. In addition,
5 . the impact from high-emitting and/or smoking diesel engines is not quantified at this time.
6 Because of these uncertainties, the changes in DPM composition over time cannot presently be
7 quantified. The data clearly indicate that toxicologically significant organic components of DE
8 (e.g., PAHs, PAH derivatives, nitro-PAHs) were present in DPM and DE in the 1970s and are
9 still present in DPM and DE as a whole.
10 While a significant fraction of ambient DPM (e.g., over 50% is possible) is also emitted
11 by nonroad equipment, there are no data available to characterize changes in the chemical
12 composition of DPM from nonroad equipment over time.
13 Some analysts project that diesel engines will increase substantially in the LD fleet in
14 coming years. While LD engines currently emit DPM with higher SOF than HD engines of the
15 same model year, recently promulgated Tier 2 standards will require control measures in the
16 2004-2007 timeframe that will reduce PM emissions from these vehicles. These control
17 measures provide some assurance that even if LD diesel use increases, DPM emitted from these
18 vehicles will likely have a smaller SOF component than they currently emit.
19
20 2.5.3. Atmospheric Transformation of Diesel Exhaust
21 An understanding of the physical as well as chemical transformations of diesel exhaust in
22 the atmosphere is necessary to fully understand the impact of this complex chemical mixture on
23 human health. In the past two decades, data acquired from laboratory and ambient experiments
24 have provided information regarding the atmospheric loss processes and transformation of diesel
25 exhaust, but knowledge concerning the products of these chemical transformations is still
26 limited. A recent study has suggested that DPM exposed to ambient levels of ozone is
27 sufficiently altered to increase the rat lung inflammatory effect compared with DPM not exposed
28 to ozone.
29 Studies investigating the chemical and physical changes of diesel exhaust emissions
30 suggest that there is little or no hygroscopic growth of primary diesel particles, but that as they
31 undergo oxidation, hygroscopicity increases. Increased solubility can increase the removal
32 efficiency of secondary diesel particles compared with their precursor compounds. Secondary
33 aerosols from diesel exhaust rosy also exhibit different biological reactivities than the primary
34. nartiMpc Fr«r *»vsimr»l<» th(=r<» ic e*\tir\f*rtr*f> fnr nitration nf cnn-io P AU r*f\mr\f\tmr\c- ,-o.-,,l+;.-.~ I— +U~
'^ A _ . — £,.~, «- — . - — -— — _~_ —»«.».».*... v . — »_ «. . ~_» « w...»f *^ *»*AMi; * ^kAAVAAA^ JAA UAW
35 formation of nitroarenes that are often more mutagenic than their precursors.
7/25/00 2-74 DRAFT—DO NOT CITE OR QUOTE
-------
2 2.5.4. Ambient Concentrations and Exposure to Diesel Exhaust
3 Due to changes in engine technology and DPM emissions over time, ambient
4 concentrations reported from studies before 1990 are compared here to those reported after 1990.
5 There are no studies in which direct comparisons can be made due to different analytical and
6 modeling tools used to assess DPM ambient levels.
7 DPM concentrations reported from CMB and dispersion modeling studies in the 1980s
8 suggest that in urban and suburban areas (Phoenix, AZ, and Southern California), annual average
9 DPM concentrations ranged from 2-13 M.g/m3, with possible maximum daily values in Phoenix,
10 AZ, of 22 ng/m3. In these studies, the average contribution of DPM in urban areas to total
11 ambient PM ranged from 7% in Pasadena, CA, to 36% in Los Angeles, CA.
12 In the 1990 timeframe, annual or seasonal average DPM concentrations reported in CMB
13 studies and from EC measurements for urban and suburban areas range from 1.2-4.5 ng/m3. The
14 contribution of DPM to ambient PM at these sites averaged 10%-15% on a seasonal or annual
15 basis, with contributions up to 38% on individual days (Brighton, CO). Dispersion modeling on
16 individual days in Southern California in the 1990s predicts DPM concentrations ranging from
^fc 1.9-4.4 ng/m3 (8%-12% of ambient PM). On individual days at a major bus stop in New York
18 City, DPM concentrations were reported to reach 46.7 |o.g/m3 and averaged 53% of ambient PM,
19 highlighting the important influence of diesel bus traffic in an urban street canyon.
20 In non-urban and rural areas in the 1980s, DPM concentrations reported range from 1.4 -
21 5 u.g/m3 and on average comprised 5%-12% of the ambient aerosol. In the 1990s, non-urban air
22 basins in California were reported to have DPM concentrations ranging from 0.2-2.6 ng/m3.
23 While estimates from emissions models suggest that DPM emissions from on-road
24 sources decreased during the 1990s, the atmospheric data available do not provide a clear
25 indication of trends in DPM concentrations but are likely to be more a reflection of the choice in
26 sampling sites, source apportionment methods, and modeling techniques. In general, from the
27 limited number of studies available it appears that DPM concentrations averaged over at least a
28 season in the 1990s typically ranged from 1-4 |ag/m3. These data can be used in model-monitor
29 comparisons and to provide an indication of long-term average exposures in some urban areas.
30 Additional work is needed to assess ambient DPM and DE concentrations in several urban
31 environments, to assess microenvironments, and to evaluate the relative impact of nonroad and
32 on-road sources on concentrations.
tA comprehensive exposure assessment cannot currently be conducted due to the lack of
data. Information regarding DPM in occupational environments suggests that exposure ranges
35 up to approximately 1,280 |ag/m3 for miners, with lower exposure measured for railroad workers
7/25/00 2-75 DRAFT—DO NOT CITE OR QUOTE
-------
1 (39-191 |ig/m3), firefighters (4-748 fig/m3), public transit personnel who work with diesel
2 equipment (7-98 ng/m3), mechanics and dockworkers (5-65 ng/m3), truck drivers (2-7
3 and bus drivers (1-3 ng/m3). Work area concentrations at fixed sites are often higher than
4 measured exposures, especially for mining operations or other enclosed spaces. For several
5 occupations involving DE exposure, an increased risk of lung cancer has been reported by
6 epidemiological studies (discussed in Chapter 7). An estimate of the 70-year lifetime
7 environmental exposure equivalent to these occupational exposures provides one means of
8 comparing the potential overlap between occupational exposures and exposures modeled for the
9 general public. The estimated 70-year lifetime exposures equivalent to those for the occupational
10 groups discussed above range from 0.4-2 fig/m3 on the low end to 2-269 on the high end.
11 The EPA has performed a national-scale exposure assessment for DPM from on-road
12 sources. Current national exposure modeling using the HAPEM-MS3 model suggests that in
13 1996 annual average DPM exposure from on-road diesel exhaust sources in urban areas was 0.7
14 Hg/m3, while in rural areas, exposures were 0.3 ng/m3. Among 10 urban areas in which DPM
15 exposures were modeled, 1996 annual average exposure from on-road diesel exhaust sources
16 ranged from 0.5 u.g/m3 to 1.2 jag/m3. Outdoor workers and children who spend a large amount of
17 time outdoors were estimated to have elevated DPM exposures in 1990, ranging up to 4.0 pg/m3
18 from on-road sources only. Based on the national inventory, DPM exposure that includes the
19 contribution from nonroad emission sources could contribute at least twice the on-road exposure.
20 Additional national-scale exposure modeling efforts have been initiated at EPA and are being
21 refined to initially provide information regarding average exposures and eventually to provide
22 more informed estimates of exposures for more highly exposed individuals.
23 Low-end exposures for many of the occupational groups overlap 1990 and 1996
24 exposures from on-road sources modeled for the general population (0.7-0.8 |ig/m3) and for the
25 more highly exposed groups. This potential overlap or small difference between occupational
26 and ambient exposures presents a concern that health effects observed in occupational groups
27 may also be evidenced in the general population. The potential magnitude of this risk is
28 discussed in Chapter 8
29 In different exposure environments, the types of diesel vehicles, their mode of operation,
30 maintenance, atmospheric transformation and many additional factors influence the chemical
31 nature and quantity of DPM to which people are exposed. The potential health consequences of
32 both short- and long-term exposures to diesel exhaust are discussed in the following chapters of
7/25/00 2-76 DRAFT—DO NOT CITE OR QUOTE
-------
Table 2-1. Vehicle classification and weights for on-road trucks
Class
1
2
3
4
5
6
7
8Aa
8Ba
Medium duty (MD)
Light-heavy duty (LHD)
Heavy-heavy duty (HHD)
Gross vehicle weight (Ib)
<6,000
6,001-10,000
10,001-14,000
14,001-16,000
16,001-19,500
19,501-26,000
26,001-33,000
33,001-60,000
>60,000
10,001-19,500 (same as Classes 3-5)
19,501-26,000 (same as Class 6)
>26,001 (same as Class 7-8)
'Class 8A and Class SB are often considered together.
Table 2-2. Total (gas and diesel) and diesel trucks in the fleet in 1992
Truck class
Class 1 and 2
(Light duty)
Class 3, 4, and 5
(Medium duty)
Class 6
(Light heavy-duty)
Class 7 and 8
(Heavy heavy-duty)
1992 gas and
diesel trucks
55,193,300
1,258,500
732,300
2,016,600
1992 diesel
trucks
1,387,600
326,300
273,800
1,725,300
% Diesels
3
26
37
86
Source: Census of Transportation (1995)
7/25/00
2-77
DRAFT—DO NOT CITE OR QUOTE
-------
Table 2-3. Typical chemical composition of fine participate matter
Elemental carbon
Organic carbon
Sulfate, nitrate,
ammonium
Minerals
Unknown
Eastern U.S.
4%
21%
48%
4%
23%
Western U.S.
15%
39%
35%
15%
—
Diesel PM2.5
75%
19%
1%
2%
3%
Source: U.S. EPA, 1999a.
Table 2-4. U.S. emission standards: HD highway diesel engines
Model
year
1970
1974
1979
1985C
1988
1990
1991
1993
1994
1996
1998
2004
Pollutant (g/bhp-hr)
HC
—
—
1.5
1.3
1.3
1.3
1.3
1.3
1.3
1.3
1.3
1.3
CO
—
40
25
15.5
15.5
15.5
15.5
15.5
15.5
15.5
15.5
15.5
NO,
— -
—
—
10.7
10.7
6.0
5.0
5.0
5.0
5.0
4.0
—
HC + NO,
—
16b
10b
—
—
—
—
—
—
—
—
2.4 NMHCd
Particulate (PM)
t=truck, b=bus,
ub=urban bus
—
—
—
—
0.60
0.60
0.25
0.25t,0.10b
O.lOt, 0.07 ub
O.lOt, 0.05 ub
O.lOt, 0.05 ub
O.lOt, 0.05 ub
Smoke*
A:40%; L:20%
A:20%; L:15%; P:50%
A:20%; L:15%; P:50%
A:20%; L:15%; P:50%
A:20%; L:15%; P:50%
A:20%; L:15%; P:50%
A:20%; L:15%; P:50%
A:20%; L:15%; P:50%
A:20%; L:15%; P:50%
A:20%; L:15%; P:50%
A:20%; L:15%; P:50%
A:20%; L:15%; P:50%
'Emissions measured in percent opacity during different operating modes: A=acceleration; L=lug; P=peaks
during either mode.
bT«»,i up
i VJV4&1 1. L\_"
cln 1985, test cycle changed from steady-state to transient operation for HC. CO. and NO., measurement and in
1988 for PM.
dOr 2.5 plus a limit of 0.5 nonmethane hydrocarbon (NMHC).
7/25/00
2-78
DRAFT—DO NOT CITE OR QUOTE
-------
Table 2-5. U.S. emission standards: locomotives (g/bhp-hr)
Line-haul
Switch
Line-haul
Switch
Line-haul
Switch
Year1
1973-2001 (Tier 0)
1973-2001 (Tier 0)
2002-2004 (Tier 1)
2002-2004 (Tier 1)
2005 + (Tier 2)
2005 + (Tier 2)
CO
5.0
8.0
2.2
2.5
1.5
2.4
HC
1.0
2.1
0.55
1.2
0.3
0.6
NO,
9.5
14.0
7.4
11.0
5.5
8.1
PM
0.6
0.72
0.45
0.54
0.20
0.24
'Date of engine manufacture.
7/25/00
2-79
DRAFT—DO NOT CITE OR QUOTE
-------
Table 2-6. U.S. emission standards for nonroad diesel equipment (g/bhp-hr)
Power rating
11 750hp
Model
year
2000
2005+
2000
2005+
2000
2005+
1998+
2004
2008+
1997+
2003
2007+
1996+
2003
2006+
2001
2006+
2002
2006+
2000+
2006+
Pollutant (g/bhp-hr)
HC
—
—
—
—
—
—
—
—
—
—
—
—
1.0
—
—
—
—
—
—
!.0
—
CO
6.0
6.0
4.9
4.9
4.1
4.1
—
3.7
3.7
—
3.7
3.7
8.5
2.6
2.6
2.6
2.6
2.6
2.6
8.5
2.6
NO,
—
—
—
—
—
—
6.9 (AST)
6.9 (ABT)
—
—
6.9 (ABT)
—
—
—
—
—
—
6.9 (ABT)
—
NMHC +
NO,
7.8 (ABT)
5.6 (ABT)
7.0 (ABT)
5.6 (ABT)
7.0 (ABT)
5.6 (ABT)
5.6 (ABT)
3.5 (ABT)
—
4.9 (ABT)
3.0 (ABT)
—
4.9 (ABT)
3.0 (ABT)
4.8 (ABT)
3.0 (ABT)
4.8 (ABT)
3.0 (ABT)
—
4.8 (ABT)
PM
0.74 (ABT)
0.60 (ABT)
0.60 (ABT)
0.60 (ABT)
0.60 (ABT)
0.44 (ABT)
—
0.30 (ABT)
—
—
0.22 (ABT)
—
0.4
0.1 5 (ABT)
—
0.1 5 (ABT)
—
0.1 5 (ABT)
—
0.4
0.1 5 (ABT)
Smoke %'
20/15/50
20/15/50
20/15/50
20/15/50
'Emissions measured in percent opacity during different operating modes: Accelcration/lug/peaks during either mode.
ABT=average banking and trading.
Note: The standards for engines less than 50 hp also apply to diesel marine engines.
7/25/00
2-80
DRAFT—DO NOT CITE OR QUOTE
-------
Table 2-7. Comparison of in-use truck fleet
with truck fleet tested on chassis dynamometer,
percent of total vehicles
Class
3
4&5
6&7
8A
8B
In-use trucks,
1995 census
17.7
13.3
25.0
20.9
23.1
Tested
trucks
1
0
17
52
30
7/25/00
2-81
DRAFT—DO NOT CITE OR QUOTE
-------
lYble 2-8. Diesel engine emissions data from engine dynamometer tests
L/i
Reference
Hare, 1977
Springer, 1979
Ferez, 1980
Martin, 1981a
Mirtin, 1981b
Ullma:i, 1984
Marin el: al., 1984
Bar-vet al.. 1985
Engine'
Cat 3208 (NA)
DDC 6 V71 (blower)
Mack ETAY(B)673A (DI,
TC.AC)
Cat 3208 (EGR, NA)
Ca 3406 (DI, TC, AC)
Cat 3406 (DI, TC, AC, EGR)
Cat 3406 (IDI, TC, AC)
DB OM-352A (DI, TC, AC)
DB OM-352A (DI, NA)
Cat (DI, NA)
Cat (DI, EGR)
Cat (DI, TC, AC)
Cat 3208
CumminsNTC350
DDC 6V92T (2S)
Cummins NTCC350
DDC8V71N(2S)
DDC6V92TA(2S)
IH DTI466B
Mack ETAY(B)673A
MackETSX676-01
Cummins VTB-903
Cut 3406
Cat 3406PCTA
Cummins BigCam NTC350
m DT466
DDC 6V92TA (2S)
DDC8V71TA(2S)
Cummins NTC290
CumminsNH-250
Cummins VTB-903
DDC8V71TA(2S)
IH DTI466B
DDAD6V-71(2S)
Cummins NTC300
DDC 8V-92 TA (2S)
Cat 3406B
Year
1976
1976
1977
1977
1977
1977
1977
1977
1977
1978
1978
1978
1978
1976
1978
1979
1978
1979
1979
1979
1980
1979
1979
1979
1979
1979
1979
1979
1979
1979
1980
1980
1980
1980
1981
1980
1985
Test"
SS
ss
SS
ss
ss
ss
ss
ss
ss
ss
ss
ss
T
T
T
T
T
T
T
T
T
T
T
T
T
T
T
T
T
T
T
T
T
T
T
T
T
NO,
g/bhp-
hr
7.98
10.24
6.613
3.747
9.79
5.49
5.14
8.93
7.46
8.12
5.16
7.66
7.83
11.41
9.55
6.58
7.15
7.80
7.46
9.01
6.90
8.10
11.28
7.24
9.97
7.91
11.66
9.81
11.10
10.87
5.59
7.91
4.41
6.09
8.13
8.15
6.58
PM
g/bhp-
hr
0.871
1.92
0.61
2.21
0.35
0.93
0.28
0.56
0.99
0.77
1.21
0.33
1.06
0.81
0.72
0.52
0.92
0.65
0.48
0.77
0.85
0.53
0.69
0.49
0.54
0.71
0.73
0.51
0.78
0.97
0.67
0.44
0.62.
0.56
0.45
0.45
0.48
CO
g/bhp-
hr
4.04
6.55
1.588
6.200
2.34
4.81
1.26
5.92
5.37
2.20
2.0
2.28
2.35
3.86
2.70
2.61
2.1
THC
g/bhp-
hr
1.11
0.71
0.476
1.163
0.35
0.17
0.12
0.77
0.57
0.27
2.23
0.73
0.87
1.42
1.36
0.53
0.5
SOF
g/bhp-
hr
0.103
0.937
0.098
0.063
0.181
0.031
0.190
0.287
0.19
0.079
0.037
0.228
0.176
0.186
0.298
—
—
0.061
SOF
Methc
c-hexane
c-hexane
Benz/cyc
Benz/cyc
Benz/cyc
Benz/cyc
Benz/cyc
Benz/cyc
DCM
DCM
DCM
DCM
DCM
DCM
DCM
DCM
Total B[aJP (PAH) 1-NP (NPAH)
Aldehyde, ug/bhp-hrd ug/bhp-hr*
me/bho-hr
• 0.76
0.24
65 2.23
161 1.72
73 0.15
80 0.08
80 0.11
280 0.87
280 1.07
1.08
4.34
0.34
23
70 1
-------
--J
K)
Table 2-8. Diesel engine emissions data from engine dynamometer tests (continued)
Reference
Enga et al.,
Baines et al.,
Wachter et al.
1985
1986
, 1990
McCarthy et al., 1992
Perez and Williams,
1989
Needham et al
, 1989
Kresoetal., 1998b
Bagley et al.,
1998
Graboski, 1998
(and references
Norbeck et al,,
Spreen et al.,
therein)
1998b
1995
Sienicki et al., 1990
Ullman et al..
1990
Engine*
DDC8V-71TAC(2S)
Cummins NTCC-400
Iveco 8460
Navistar DTA466ES2 10
Engine 1
Engine 2
Engine 3
Engine 4
Engine 5
Engine 6
Average of 16 engines
Average of 3 engines
Cummins LI 0-300
Cummins LI 0-3 10
Cummins M11-330E
Cat 3304 (IDI, NA) non-road
DDC6V-71N-77(MUI,2S)
DDC6V-92TA-91 (DDEC1I)
DDC-6V-92TA-87 (2S)
DDC-6V92TA-83 (MUI, 2S)
DDC 6V-92TA -88 (DDECII, 2S)
DDC 6V-92TA-91 (DDECII, 2S)
DDC6V-71N-77(MUI,2S)
DDC 6V-92TA-81/89 (MUI, 2S)
DDC 6V-92TA-91 (DDECII, 2S)
DDC 6V-92TA-89 (DDECII, 2S)
DDC Series 60-91 DDECII
Cummins L- 10-87 (MUI)
DDC Series 60-91 (DDECII)
Cummins N-14-87 (MUI)
DDC Series 60-89 (DDECII)
DDC Series 60-91 (DDECII)
Cummins B5.9
CumminsLlO
Navistar DTA466
DDC Series 60
Navistar DTA466
DDC Series 60
Year
1984
1985
1991
1993
1982
1982
1982
1982
1982
1982
1988
1991
1988
1991
1995
1983
1977
1991
1987
1983
1988
1991
1977
1981
1991
1989
1991
1987
1991
1987
1989
1991
1995
1991
1994
1994
1991
1991
Test"
SS
T
T
T
T
T
T
T
T
T
T
T
T
SS
SS
SS
SS
T
T
T
T
T
T
T
T
T
T
T
T
T
T
T
T
T
T
T
T
T
T
NO,
g/bhp-
hr
6.64
5.85
4.62
4.93
5.15
4.70
3.82
9.96
4.23
10.77
5.62
8.52
4.4
11.72
10.06
4.84
4.855
4.635
5.64
4.68
6.32
5.128
4.303
4.37
4.77
4.779
4.89
5.25
4.552
PM
g/bhp-
hr
0.36
1.26
0.55
0.22
0.082
0.93
0.86
0.59
0.96
1.06
0.88
0.37
0.24
0.103
0.035
0.037
0.56
0.83
0.197
0.59
0.265
0.2
0.276
0.282
0.268
0.227
0.338
0.300
0.309
0.220
0.369
0.252
0.182
0.106
0.224
0.090
0.112
0.22
0.188
CO
g/bhp-
hr
1.83
2.99
3.21
1.3
3.59
1.51
0.71
1.19
1.6
1.65
3.18
2.16
1.51
2.499
4.458
2.33
2.26
2.20
4.008
2.004
1.47
2.26
0.989
1.402
~
2.102
THC
g/bhp-
hr
0.38
1.48
0.53
0.28
0.26
0.067
0.16
2.01
0.72
-
0.435
0.6
0.42
0.86
0.42
0.44
0.526
0.164
0.89
0.08
0.58
0.154
0.392
0.30
0.53
0.181
0.065
0.23
0.508
SOF
g/bhp-
hr
0.0255
—
0.0957
0.0237
0.179
0.145
0.185
0.325
0.076
0.344
0.12
0.10
0.030
0.022
0.016
0.319
0.729
0.0788
—
0.133
0.116
0.07
0.212
0.144
—
—
—
«
0.066
0.100
~
0.061
0.05
—
0.035
0.043
0.05
~
SOF
Methc
?
SFE
DCM
DCM
DCM
DCM
DCM
DCM
DCM .
DCM
DCM
DCM
DCM
Benz/cyc
DCM
?
DCM
Tol/EtOH
Tol/EtOH
DCM
DCM
DCM
?
Tol/EtOH
DCM
DCM
DCM
DCM
Total B[a]P (PAH) 1-NP (NPAH)
Aldehyde, ug/bhp-hrd ug/bhp-hr*
me/bhp-hr
. ~
26 0.83
5.8
4.9 0.89
26 1.2
5.3
1.5(133) 2.2
..
..
..
—
..
0.24((1 8.5)
80 20(1725) 1.95(4.92)
26
17
-------
Table 2-8. Diesel engine emissions data from engine dynamometer tests (continued)
Reference
Kadoetal., 1998
Ullmaneial., 1988
Mitchell ctal, 1994
Tanakaetal, 1998
Rantanen et al, 1993
Engine'
Cat 3406E
Cummins NTCC400
DDC Series 60
' Navistar DTA466
Unknown
Scania
Valmet
Volvo
Volvo
Year
1997
1988
1994
1994
1994
1990
1990
1990
1995
Test1
T
T
T
T
SS
ss
SS
ss
ss
NO,
g/bhp-
hr
4.47
4.43
4.86
4.934
9.30
8.67
9.87
4.56
PM
g/bhp-
hr
0.42
0.111
0.099
0.143
0.157
0.157
0.262
0.135
CO THC
g/bhp- g/bhp-
hr hr
2.22 0.53
2.17 0.22
1.10 0.34
0.807 0.352
SOF
g/bhp-
hr
0.021
0.046
0.036
0.031
SOF
Methc
DCM
DCM
DCM
DCM
DCM
Total
Aldehyde,
me/bhp-hr
34
56
B(a|P (PAH)
ug/bhp-hr"
0.07(30)
(141)
0.11(242)
.076
1-NP (NPAH)
ug/bhp-hrc
0.34
0.04(0.12)
0.3(0.6)
o
o
to
oo
O
O
2
o
H
O
"NA=naturally aspirated. TC=turbochared (engines not designated as NA or TC are turbocharged). AC=aftercooled. DI=direct injection. IDI=indirect injection.
EGR=exhaust gas recirculation. 2S=two-stroke (engines not designated as 2S are four-stroke). MUl=mechanical unit injector (not electronically controlled).
DDEC=:Detroit Diesel Corporation's engine control module (electronic control).
bSS=vai ious single or multimode steady-state tests. T=heavy-duty FTP (transient test).
CSOF extraction method. SFE=Supercritical fluid extraction. All others by Soxhlet extraction using the indicated solvents (? for unreported).
DCM=dichloromethane. Tol/Et()H=toluene/ethanol mixture. Benz/cyc=benzene/cyclohexane mixture. C-hexane=cyclohexane.
dNumber in parentheses is the total PAH emission obtained by summing emissions of all PAHs reported.
'Number in parentheses is the total NPAH emission obtained by summing emissions of all NPAHs reported.
G
I
-------
Table 2-9. HD diesel emissions results from tunnel tests (adapted from Yanowitz et al., 1999)
to
(Si
o
o
Test
Pierson and
Brachaczeck, 1983
Rogaketal., 1998
Miguel etal., 1998
Weingartner et al.,
1997
Pierson et al., 1996
Pierson et al., 1996
Pierson et al., 1996
Kirchstetter et al.,
1999
Gertler,1999
Tunnel location,
year of study
Allegheny, 1974
Allegheny, 1975
Allegheny, 1976
Allegheny, 1976
Tuscarora, 1976
Tuscarora, 1976
Allegheny, 1977
Allegheny, 1979
Allegheny, 1979
Cassiar Tunnel,
1995, Vancouver
Caldecott Tunnel,
1996, San Francisco
Gubrist Tunnel,
1993, Zurich
Fort McHenry
Tunnel, downhill,
1992, Baltimore
Fort McHenry
Tunnel, uphill,
1992, Baltimore
Tuscarora Tunnel
1992, Pennsylvania
Caldecott Tunnel,
1997, San Francisco
Tuscarora Tunnel,
1999, Pennsylvania
Fuel
efficiency
(mi/gal)
5.42"
8.03"
5.42C
5.60'
11.46"
5.42b
6.44b
5.42C
NO,"
(g/mi)
19.50
±4.22
23.82 ±
4.17
9.66
±0.32
22.50
±1.00
19.46
±0.85
23.82 ±
2.98
NMHC
(g/mi)
-0.16
±0.88
0.92
±0.21
2.55
±1.05
0.68
±0.20
CO
(g/mi)
6.79
± 11.78
6.8
±1.5
14.3
±5.5
6.03
±1.61
DPM
(g/mi)
.90-1.80
1.75 ±0.19
l.5±0.10
1. 4 ±0.07
1.3±0.19
1.39 ±26
1.3 ±0.08
1.2 ±0.03
1.4 ±0.04
1.67
± 0.24d
0.62
± 0.02f
1.43
±0.12"
0.29
CO,
(g/mi)
1,280
±40
897
±48
1,897
±168
1,596
±78
NO/
(g/gal)
157
±34
129
±23
111
±4
122
±5
125
±5
129
±16
NMHC
(g/gal)
-1±7
11
±2
14
±6
4
±1
CO
(g/gal)
55
±95
78
±17
78
±30
39
±10
DPM
(g/gal)
4.9-9.8
9.49±1.03
8.1 ±0.54
7.6 ±0.4
7.0 ± 1.0
7.5 ±1.40
7.0 ±0.43
6.5 ±0.16
7.6 ±0.19
9.0
±1.3d
3.5
±o.r
7.7
±0.6"
to
I
00
o
o
z
o
H
O
H— I
tn
o
G
O
a
"NOX reported as N02.
bCalculated from observed C02 emissions assuming fuel density 7.1 Ib/gal
and C is 87% of diesel fuel by weight.
'Since C02 emissions not available, fuel efficiency assumed to be the same
as in slightly uphill tunnel (Fort McHenry).
dReported as black carbon, assumed that 50% of total PM emissions are BC.
'Slope of tunnel unknown, so used average fuel efficiency for the United
States.
rPM3.
«PM2S.
hUncertainry reported as ±1.0 standard deviation, except where literature
report did not specify standard deviation; in those cases uncertainty listed as
reported.
-------
Table 2-10. Remote sensing results for HD vehicles
NCX
CO
THC
Reference
Jimenez et al., 1998
Cohen etal., 1997
Countess et al., 1999
Bishop et al., 1996
Cohen etal., 1997
Countess et al., 1999
Bishop etal., 1996
Cohen et al., 1997
Year study
conducted
1997
1997
1998
1992
1997
1998
1992
1997
Emissions (g/gal)
150*-b-c
10gvb.c
187*-b-c
59b
54 b
85"
0.002 HC/CO2 mole ratio d
0.00073 HC/CO, mole ratio d
'Remote sensing measures NO. The reported value was corrected to a NOX (as NOJ value by
assuming 90% (mole fraction) of NOX is NO.
""Emissions in g/gal calculated by assuming that fuel density is 7.1 Ib/gal and C is 87% by weight
of fuel.
°No humidity correction factor is included.
dln order to calculate emissions in g/gal, an average molecular weight is needed.
Source: Yanowitz et al., 1999.
7/25/00
2-86
DRAFT—DO NOT CITE OR QUOTE
-------
Table 2-11. Summary of CDD/CDF emissions from diesel-fueled vehicles
O
o
NJ
i
00
Study
CARB, I987a; Lew, 1996
Marklund et al., 1990
Hagenmaier et al., 1990
Hagenmaier et al., 1994
Oehmeetal., 1991 (tunnel
study)
Schwind ct al., 1991
Hutzinger et al., 1992
Gertler et al., 1996 (tunnel
study)
Gullett and Ryan, 1997
Country
United States
Sweden
Germany
Germany
Norway
Germany
United States
United States
Vehicle tested
Diesel truck
Diesel truck
Diesel car
Diesel bus
Diesel car
Diesel truck
Diesel trucks
Diesel truck
Number
of test
vehicles
1
1
1
1
(b)
1
1
(d)
1
Emission factor
(pg TEQ/km driven)
663-1,300
not detected (< 18)'
2.4'
not detected (< 1 pg/L)
520C
38C
avg = 280
9,500"
720C
avg = 5, 100
5.0-13"
13-15"
mean= 172
mean - 29.0
Driving cycle; sampling location
6-hr dynamometer test at 50 km/hr
U.S. Federal mode 13 cycle; before muffler
Comparable to FTP-73 test cycle; in tailpipe
On-the-road testing
Cars moving uphill (3.5% incline) at 60 km/hr
Cars moving downhill (3.5% decline) at 70 km/hr
Trucks moving uphill (3.5% incline) at 60 km/hr
Trucks moving downhill (3.5% decline) at 70 km/hr
Various test conditions (i.e., loads and speeds)
Various test conditions (i.e., loads and speeds)
Mean of seven 12-hour samples
Mean of five sample routes
•Results reported were in units of pg TEQ/liter of fuel. For purposes of this table, the fuel economy factor used by Marklund et al. (1990), 10 km/L or 24 miles/gal, was used to convert the emission rates
into units of pg TEQ/km driven for the cars. For the diesel-fueled truck, the fuel economy factor reported in CARB (I987a) for a 1984 heavy-duty diesel truck, 5.5 km/L (or 13.2 miles/gal), was used.
"Tests were conducted over portions of 4 days, with traffic rates of 8,000-14,000 vehicles/day. Heavy-duty vehicles (defined as vehicles over 7 meters in length) ranged from 4% to 15% of total.
'Emission factors are reported in units of pg Nordic TEQ/km driven; the values in units of I-TEQ/km are expected to be about 3% to 6% higher.
JTcsts were conducted over 5 days with heavy-duty vehicle rates of 1,800-8,700 vehicles per 12-hour sampling event. Heavy-duty vehicles accounted for 21% to 28% of all vehicles.
o
z;
o
H
O
H-H
H
tn
o
?o
o
c
o
-------
Table 2-12. Baltimore Harbor Tunnel Study: estimated CDD/CDF emission factors for HD vehicles
Congener/congener group
2,3,7,8-TCDD
,2,3,7,8-PeCDD
,2,3,4,7,8-HxCDD
,2,3,6,7,8-HxCDD
,2,3,7,8,9-HxCDD
,2,3,4,6,7,8-HpCDD
OCDD
2,3,7,8-TCDF
1,2,3,7,8-PeCDF
2,3,4,7,8-PeCDF
1, 2,3,4,7,8- HxCDF
1, 2,3,6,7,8- HxCDF
1,2,3,7,8,9- HxCDF
2,3,4,6,7,8. HxCDF
1,2,3,4,6,7,8-HpCDF
1,2,3,4,7,8,9-HpCDF
OCDF
Total 2,3,7,8-CDD
Total 2,3,7,8-CDF
Total TEQ
Total TCDI)
Total PeCDD
Total HxCDD
Total HpCDD
Total OCDD
Total TCDF
Total 1'eCDF
Total HxCDF
Total HpCDF
Total OCDI-
Total CDD/CDF
HD vehicles as % of tolal
vehicles
Run-specific emission factors
Run no. 2
(pg/km)
24.5
40.2
18.2
37.5
53.6
0
0
0
0
24.5
15.4
0.3
27.7
15.2
12.6
0
0
174
95.7
73.8
245
110
677
0
0
0
124
136
0
0
1,291
21.2
Run no. 3
(pg/km)
61.6
20.6
25.2
28.2
56.5
401
3,361
94.3
48.9
75.7
139
75.1
14.8
82.5
280
58.5
239
3,954
1,108
175
0
21.9
0
802
3361
901
119
319
223
239
5,987
22.0
Run no. 5
(pg/km)
0.0
15.4
46.5
64.3
91.6
729
3,382
67.6
72.6
131
204
73.7
75.6
152
445
60.8
401
4,328
1,684
170
140
83.3
753
1,498
3,382
1,314
1,152
852
814
401
10,390
22.6
Run no. 6
(pg/km)
21.2
5.6
8.3
19.6
48.4
111
1,120
152.8
23.6
46.6
93.8
51.0
0
55.7
154
31.1
175
1,335
784
96
165
35.6
54.5
142
1,120
656
78.4
67.6
144
175
2,638
34.0
Run no. 8
(pg/km)
37.8
38.4
64.5
153
280
2,438
9,730
155.8
53.3
85.0
124
61.3
20.6
93.0
313
25.0
416
12,743
1,347
235
311
174
2,009
5,696
9,730
2,416
1,055
444
513
416
22,766
28.8
Run no. 9
(pg/km)
40.1
0.0
0.0
71.1
126
963
5,829
73.4
0.0
63.9
164
54.4
37.2
86.8
354
2.3
534
7,028
1,371
153
109
0.0
1,666
1,933
5,829
1,007
282
719
354
534
12,434
24.2
Run no. 10
(pg/km)
54.9
83.0
123
186
370
2,080
7,620
61.7
43.3
108
166
95.5
63.5
111
308
34.9
370
10,515
1,362
303
97.3
165
2,971
4,377
7,620
687
626
619
637
370
18,168
27.4
Mean
emission
factors
(pg/km)
34.3
29.0
40.8
80.0
147
960
4,435
86.5
34.5
76.4
129
58.8
34.2
85.2
267
30.4
305
5,725
1,107
172
152
84.2
1,162
2,064
4,435
997
491
451
384
305
10,525
25.7
O
o
to
I
00
oo
O
O
2
O
H
n
t— H
H.
m
o
Notes:
(I) Listed value> are based on (he difference between the calculated chemical mass entering the tunnel and the mass exiting the tunnel.
(2) All calculated negative emission factors were set equal to zero.
(3) All CDD/CDF emissions were assumed to result from heavy-duty diesel-fueled vehicles. The table presents in the last row the percent of total traffic that was heavy-duty vehicles.
Source: Gcrtler st al., 1996.
c
o
3
-------
Table 2-13. Organic and elemental carbon fractions of diesel and gasoline engine
PM exhaust
Engine type
HD diesel engines "
HD diesel engines fSPECIATE)1*
LD diesel engines0
LD diesel engines (SPECIATE)"
Gasoline engines (hot stabilized)*
Gasoline engines (smoker and high emitter)"-'
Gasoline engines (cold start)"
%
Organic
carbon
19±8
21-36
30 ±9
22-43
56±11
76 ±10
46 ±14
%
Elemental
carbon
75 ±10
52-54
61 ±16
51-64
25 ±15
7±6
42 ±14
' Fujita et al., 1998, and Watson et al., 1998.
"U.S. EPA SPECIATE database.
cNorbecketal., 1998c.
7/25/00
2-89
DRAFT—DO NOT CITE OR QUOTE
-------
Table 2-14. Emission rates of PAH (rag/mi) from LD and HD diesel vehicles
PAH
Naphthalene
2-Menaphthalene
1-Menaphthalene
Dimethylnaphthalenes
Biphenyl
2-Methylbiphenyl
3 -Methy Ibipheny 1
4-Methylbiphenyl
Trimethylnaphthalenes
Acenaphthylene
Acenaphthene
Phenanthrene
Fluorene
Methylfluorenes
Methylphenanthrenes
Dimethylphenanthrenes
Anthracene
9-Methylanthracene
Fluoranthene
Pyrene
Methyl(pyrenes/fluoranthenes)
Benzonaphthothiophene
Benz[a]anthracene
Chrysene
Benz[b+j+k]fluoranthene
Benzo[e]pyrene
Benzo[a]pyrene
Indeno[l ,2,3-cd]pyrene
Dibenzo[a]anthracene
Benzo [b] chry sene
Benzo [ghi] perlyne
Coronene
Light-duty
diesel
5.554 ± 0.282
3. 068 ±0.1 85
2.313 ±0.134
5.065 ± 0.333
0.743 ± 0.041
0.203 ± 0.015
1.048 ±0.063
0.447 ± 0.028
6.622 ± 0.563
0.422 ± 0.024
0.096 ± 0.008
1.411 ±0.072
0.442 ± 0.038
1.021 ±0.091
1.1 15 ±0.064
0.637 ± 0.047
0.246 ± 0.025
0.01 3 ±0.002
0.213 ±0.014
0.245 ± 0.020
0.548 ± 0.045
0.002 ± 0.002
0.020 ± 0.005
0.029 ± 0.005
0.056 ± 0.005
0.019 ±0.003
0.013 ± 0.004
0.010 ±0.003
0.002 ± 0.003
0.001 ± 0.002
0.01 8 ±0.004
0.006 ± 0.006
Heavy-duty
diesel
2.451 ±0.1 54
2.234 ±0.1 52
1.582 ±0.1 03
2.962 ± 0.488
0.505 ± 0.037
0.049 ± 0.024
0.401 ±0.036
0.1 44 ±0.021
1.940 ±0.221
0.059 ± 0.087
0.030 ± 0.040
0.084 ±0.0 11
0.066 ± 0.022
0.071 ± 0.055
0.124 ±0.069
0.090 ± 0.096
0.052 ±0.016
0.434 ± 0.082
0.044 ± 0.026
0.071 ±0.017
0.022 ± 0.082
0.001 ± 0.027
0.066 ± 0.046
0.009 ± 0.021
0.009 ± 0.022
0.010 ±0.014
0.01 3 ±0.044
0.001 ± 0.037
0.000 ± 0.053
0.001 ± 0.027
0.01 3 ±0.048
0.001 ±0.095
7/25/00
2-90
DRAFT—DO NOT CITE OR QUOTE
-------
Table 2-15. Polycyclic aromatic hydrocarbons identified in
extracts of diesel particles from LD diesel engine exhaust
Compound
Acenaphthylene
Trimethy (naphthalene
Fluorene
Dimethylbiphenyl
C4 -Naphthalene
Trimethylbiphenyl
Dibenzothiophene
Phenanthrene
Anthracene
Methyidibenzothiophene
Methylphenanthrene
Methylanthracene
Ethylphenanthrene
4H-
Cyclopentafrfe/Jphenanthrene
Ethyldibenzothiophene
2-Phenylnaphthalene
Dimethyl(phenanthrene/anth
racene)
Fluoranthene
Benzo[cfe/]dibenzothiophene
Benzacenaphlhylene
Pyrene
Ethylmethyl(phenanthrene/a
nthracene)
Methyl(fluoranthene/pyrene)
Benzo[
-------
Table 2-16. Emission rates of particle-bound PAH (jig/mi) from diesel and gasoline engines
PAH
Pyrene .
Fluoranthene
Benzo[a]pyrene
Benzo[e]pyrene
Diesel engines
HDD
(a)
71
44
13
10
(b)
17.6
27.2
<0.1
0.24
(c)
36.2
20.8
2.1
4.2
LDD
(a)
245
213
13
19
(d)
66
50
NA
NA
Gasoline engines
Noncatalyst
(c)
49.6
77.3
69.6
73.3
(e)
45
32
3.2
4.8
Catalyst
(a)
248
196
1.0
1.0
(c)
4.0
3.6
3.0
3.6
(a) Watson et al., 1998 included gas-phase PAH
(b) Westerholm etal., 1991.
(c) Rogge et al., 1993.
(d) Smith, 1989; 1986 Mercedes Benz.
(e) Alsberg et al., 1985.
7/25/00
2-92
DRAFT—DO NOT CITE OR QUOTE
-------
Table 2-17. Concentrations of nitro-PAHs identified in LD diesel participate extracts
Nitro-PAH1
4-nitrobiphenyl
2-nitrofluorene
2-nitroanthracene
9-nitroanthracene
9-nitrophenanthrene
3-nitrophenanthrene
2-methy 1- 1-nitroanthracene
1 -nitrofluoranthene
7-nitrofluoranthene
3-nitrofluoranthene
8-nitrofluoranthene
1-nitropyrene
6-nitrobenzo[a]pyrene
1 ,3-dinitropyreneb
1 ,6-dinitropyreneb
1 ,8-dinitropyreneb
2, 7-dinitrofluorenec
2,7-dinitro-9-fluorenonec
3 -nitrobenzanthroned
Concentration
(Hg/gof
particles)
2.2
-1.8
4.4
1.2
1.0
4.1
8.3
1.8
0.7
4.4
0.8
18.9; 75"
2.5
0.30
0.40
0.53
4.2; 6.0
8.6; 3.0
0.6 to 6.6
'From Campbell and Lee (1984) unless noted otherwise. Concentrations recalculated from Jig/g of
extract to |J.g/g of particles using a value of 44% for extractable material (w/w).
Tram Paputa-Peck et al, 1983.
Trom Schuetzle, 1983.
"From Enya et al., 1997 (Isuzu Model 6HEL 7127cc).
7/25/00
2-93
DRAFT—DO NOT CITE OR QUOTE
-------
Table 2-18. Average emission rates for potycyclic aromatic hydrocarbons for different fuel
types (units are mg/mi)
PAH
2,3,5-trimethyl naphthalene
Phenanthrene
Anthracene
Methylphenanthrenes/anthracenes
Fluoranthene
Pyrene
Benzo[c] phenanthrene
Benzo[ghi]fluoranthene
Cyclopenta[cd]pyrene
Benz[a]anthracene
Chrysene + triphenylene
Benzo[b+j+k]fluoranthene
Benzo[e]pyrene
Benzo[a]pyrene
Perylene
Indenof 1 ,2,3-cd]fluoranthene
Benzo[c]chrysene
Dibenz[aJ]anthracene
Indeno[ 1 ,2,3-cd]pyrene
Dibenz[a,h+a,c]anthracene
Benzo[b]chrysene
Benzo [ghi] pery lene
Coronene
Dibenzo[a,ljpyrene
Dibenzo[a,e]pyrsr.e
Dibenzo[a,i]pyrene
Dibenzo[a,h]pyrene
Pre-1993 diesel
fuel
Cetane No. >40
Aromatic 33% v.
PAH 8% wt.
283.68 ± 5.27
336.71 ±9.08
38.89 ±1.43
33 1.32 ±16.07
128.45 ± 7.60
193.03 ± 16.51
3. 03 ±0.24
24.84 ± 2.68
21.44±4.11
16.42 ±1.67
17.36 ±1.66
31.05 ±4.17
16.71 ±2.72
20.46 ±3. 27
4.32 ±0.88
0.34 ± 0.07
0.29 ± 0.05
0.93 ± 0.05
19.45 ±2.71
1.54 ±0.15
0.40 ±0.01
49. 17 ±9.63
9.49 ±3. 13
2.84 ± 0.45
1.10 ±0.29
0.91 ±0.21
1.33 ±0.25
Low aromatic
diesel fuel
Cetane No. >48
Aromatic 10% v.
PAH1.4%wt.
14.77 ±2.42
160.92 ±15.54
18.54 ±2.13
25.17 ±1.41
132.36 ±18.30
21 1.19 ±37.35
1.74 ±0.14
18.93 ±2.14
26.15 ±3.12
10.57 ±1.15
10.38 ±0.54
23.17 ±1.98
14.55 ±1.34
16.48 ± 1.56
3.71 ± 0.74
0.21 ± 0.02
0.18 ±0.05
0.55 ±0.10
14.04 ±1.99
0.87 ±0.12
0.15 ±0.05
39.81 ±7.22
4.93 ± 0.47
1.25 ±0.15
0.61 ± 0-06
0.27 ± 0.09
0.75 ± 0.07
Reformulated
diesel blend
Cetane No. 50-55
Aromatic 20%-25%
v. PAH 2%-5% wt.
56.21 ±2.82
220.73 ± 52.68
26.16 ±6.86
11 1.98 ±28.74
123. 07 ±26.21
206.82 ± 39.04
1.54 ±0.26
16.94 ±2.31
21.25 ±3.46
10.96 ±2.42
12.20 ±2.72
29.18 ±7.93
18.99 ±5.58
20.59 ± 5.75
4.18±1.16
0.17 ±0.00
0.14 ±0.04
0.67 ± 0.09
22.16±9.11
1.48 ±0.67
0.27 ± 0.05
60.74 ± 26.60
7.48 ±1.59
2.31 ±0.48
1.13 + 0.15
0.71 ±0.15
0.84 ± 0.20
Source: Norbeck et al., 1998c.
7/25/00
2-94
DRAFT—DO NOT CITE OR QUOTE
-------
Table 2-19. Classes of compounds in diesel exhaust
Particulate phase
Gas phase
Heterocyclics, hydrocarbons (C,4-C3s), and
PAHs and derivatives:
Acids
Alcohols
Alkanoic acids
n-Alkanes
Anhydrides
Aromatic acids
Cycloalkanes
Esters
Halogenated cmpds.
Ketones
Nitrated cmpds.
Sulfonates
Quinones
Elemental carbon
Inorganic sulfates and nitrates
Metals
Water
Heterocyclics, hydrocarbons (C,-C,0), and
derivatives:
Acids Cycloalkanes, Cycloakenes
Aldehydes Dicarbonyls
Alkanoic acids Ethyne
n-Alkanes Halogenated cmpds.
n-Alkenes Ketones
Anhydrides Nitrated cmpds.
Aromatic acids Sulfonates
Quinones
Acrolein
Ammonia
Carbon dioxide, carbon monoxide
Benzene
1,3-Butadiene
Formaldehyde
Formic acid
Hydrogen cyanide, hydrogen sulflde
Methane, methanol
Nitric and nitrous acids
Nitrogen oxides, nitrous oxide
Sulfur dioxide
Toluene
Water
Sources: Mauderly, 1992, which summarized the work of Lies et al., 1986; Schuetzle and Frazier,
1986; Carey, 1987; Zaebst et al., 1988, updated from recent work by Johnson, 1993; McDonald,
1997;Schaueretal., 1999.
7/25/00
2-95
DRAFT—DO NOT CITE OR QUOTE
-------
Table 2-20. Calculated atmospheric lifetimes for gas-phase reactions of
selected compounds present in automotive emissions with important reactive
species
Compound
NO2
NO
HNO3
S02
NH3
Propane
n-Butane
n-Octane
Ethylene
Propylene
Acetylene
Formaldehyde
Acetaldehyde
Benzaldehyde
Acrolein
Formic acid
Benzene
Toluene
m-Xylene
Phenol
Naphthalenef
2-Methylnaphthalenef
1 -Nitronaphthalenef
Acenaphthenef
Acenaphthylenef
Phenanthrenef
Anthracenef
Fluoranthenef
Pyrenef
Atmospheric lifetime resulting from reaction with:
OH1
1.3 days
2.5 days
1 10 days
16 days
90 days
12 days
5.6 days
1.9 days
1.9 days
7h
19 days
1.9 days
0.6 day
1.2 days
0.6 day
31 days
1 1 days
2.5 days
7h
6h
6.8 h
2.8 h
2.3 days
1.5 h
1.3 h
11.2h
8.6 h
-2.9 h
-2.9 h
03"
12 h
1 min
—
>200 years
—
>7,000 years
>4,500 years
—
9 days
1 .5 days
6 years
>2- 104 years
>7 years
—
60 days
—
600 years
300 years
75 years
—
>80 days
>40 days
>28 days
>30 days
~43 min
41 days
—
—
—
NO3C
24 min
1.2 min
—
> 1.4x10" years
—
—
3.6 years
1.2 years
1.2 years
6 days
>5.6 years
84 days
20 days
24 days
—
—
>6.4 years
3.6 years
0.8 years
8 min
1 .5 years
180 days
1 8 years
1.2 h
6 min
4.6 h
—
-1 year
~ 120 days
H600 years
—
—
—
—
—
—
—
23 days
—
—
—
—
—
—
—
—
—
—
—
—
—
—
—
hv«
2 min
—
—
—
—
—
—
—
—
—
—
4h
60 h
—
—
—
—
—
—
—
—
—
1.7 h
—
—
—
—
—
—
' For 12-h average concentration of OH radical of 1.6><106 molecule/cm3 (Prinn et al., 1992).
b For 24-h average O3 concentration of 7x 10" molecule/cm3.
°For 12-h average NO3 concentration of 5*10" molecule/cm3 (Atkinson, 1991).
d For 12-h average HO: concentration of 108 molecule/cm3.
e For solar zenith angle ot 0°.
f Lifetimes from Arey (1998), for 12-h concentration of OH radical of 1.9* 106 molecule/cm3.
Source: Winer and Busby, 1995, unless noted otherwise.
7/25/00
2-96
DRAFT—DO NOT CITE OR QUOTE
-------
Table 2-21. Major components of gas-phase diesel engine emissions, their known
atmospheric transformation products, and the biological impact of the reactants and
products
Gas-phase emission
component
Carbon dioxide
Carbon monoxide
Oxides of nitrogen
Sulfur dioxide
Atmospheric reaction
products
—
—
Nitric acid, ozone
Sulfuric acid
Biological impact
Major contributor to global
warming.
Highly toxic to humans; blocks
oxygen uptake.
Nitrogen dioxide is a respiratory
tract irritant and major ozone
precursor. Nitric acid contributes to
acid rain.
Respiratory tract irritation.
Contributor to acid rain.
Hydrocarbons:
Alkanes (£C18)
Alkenes(<;C4)
(e.g., 1,3-butadiene)
Aldehydes, alkyl nitrates,
ketones
Aldehydes, ketones
Respiratory tract irritation. Reaction
products are ozone precursors (in the
presence of NOX).
Respiratory tract irritation. Some
alkenes are mutagenic and
carcinogenic. Reaction products are
ozone precursors (in the presence of
NOJ.
Aldehydes:
Formaldehyde
Higher aldehydes (e.g.,
acetaldehyde, acrolein)
Monocyclic aromatic
compounds (e.g., benzene,
toluene)
PAHs (^4 rings) (e.g.,
phenanthrene, fluoranthene)b
Nitro-PAHs (2 and 3 rings)
(e.g., nitronaphthalenes)
Carbon monoxide,
hydroperoxyl radicals
Peroxyacyl nitrates
Hydroxylated and
hydroxylated-nitro derivatives'
Nitro-PAHs (4 rings)'
Quinones and hydroxylated-
nitro derivatives
Formaldehyde is a probable human
carcinogen and an ozone precursor
(in the presence of NOX).
Respiratory tract and eye irritation;
causes plant damage.
Benzene is toxic and carcinogenic in
humans. Some reaction products are
mutagenic in bacteria (Ames assay).
Some of these PAHs and nitro-PAHs
are known mutagens and
carcinogens.
Some reaction products are
mutagenic in bacteria (Ames assay).
"Some reaction products expected to partition into the particle phase.
bNitro-PAHs with more than two rings will partition into the particle phase.
°PAHs containing four rings are usually present in both the vapor and particle phases.
Source: Health Effects Institute, 1995.
7/25/00
2-97
DRAFT—DO NOT CITE OR QUOTE
-------
Table 2-22. Major components of particle-phase diesel engine emissions, their known
atmospheric transformation products, and the biological impact of the reactants and
products
Particle-phase
emission component
Elemental carbon
Inorganic sulfate and
nitrate
Hydrocarbons (CM-Cj5)
PAHs (;>4 rings) (e.g.,
pyrene, benzo[a]pyrene)
Nitro-PAHs (zl rings)
(e.g., nitropyrenes)
Atmospheric reaction
products
—
—
Little information;
possibly aldehydes, ketones,
and alkyl nitrates
Nitro-PAHs (^4 rings)1
Nitro-PAH lactones
Hydroxylated-nitro derivatives
Biological impact
Nuclei adsorb organic compounds;
size permits transport deep into the
lungs (alveoli)
Respiratory tract irritation
Unknown
Larger PAHs are major contributors
of carcinogens in combustion
emissions. Many nitro-PAHs are
potent mutagens and carcinogens.
Many nitro-PAHs are potent
mutagens and carcinogens. Some
reaction products are mutagenic in
bacteria (Ames assay).
"Nitro-PAHs with more than two rings will partition into the particle phase.
Source: Health Effects Institute, 1995.
7/25/00
2-98
DRAFT—DO NOT CITE OR QUOTE
-------
Table 2-23. Ambient DPM concentrations reported from chemical mass balance modeling
K3
LH
O
O
Ref.
a
b
c
d
e
f
Location
West LA, CA
Pasadena, CA
Rubidoux, CA
Los Angeles, CA
West Phoenix, AZ
Central Phoenix, AZ
South Scottsdale, AZ
Estrella Park, AZ
Gunnery Park, AZ
Pinnacle Peak, AZ
California, 6 air basins
California, 9 air basins
Manhattan, NY
Phoenix, AZ
Welby, CO
Brighton, CO
Year of sampling
1982, annual average (-60
samples at each site)
1989-90, winter
1 1 days at each site
1988-92, annual
1993, spring 3 days
1994-95, winter 12 days
1996-97, winter 60 days
Location type
Urban
Urban
Urban
Urban
Urban
Urban
Urban
Nonurban
Nonurban
Nonurban
Urban*!*
Nonurban^
Urban
Urban
Urban
Suburban
Diesel PM2 5
ug/m3 mean,
(range)
4.4
5.3
5.4
11.6
13 (max. 22)
13 (max. 16)
10 (max. 12)
5
3
2
1.8-3.6*
0.2-2.6*
29.2(13.2-46.7)*
2.4 (0-5.3)
1.7(0-7.3)
1.2(0-3.4)
Average
DPM % of
total PM
(range)
18
19
13
36
18
20
17
9
10
12
t
t
53(31-68)
15 (0-27)
10(0-26)
10(0-38)
Source profile used
EC, OCS, elements
EC, OCT, MI, elements
EC, OCT, MI, elements
EC, OCT, MI, elements
EC, OCT, MI, elements
EC, OCS, MI, elements
EC, OCS, MI, elements
K)
O
O
2
O
H
n
t—t
a
O
w
O
c
O
*PM10. fNot available.
EC'.Elemental carbon; OCT:Organic carbon total; OCS: Organic carbon species; MI: Major ions including nitrate, sulfate, chloride and, in some cases,
ammonium, sodium, potassium
*!*Urban air basins are qualitatively defined as those areas that are moderately or largely urbanized, and nonurban air basins are those areas that are largely
nonurban, but may have one or more densely populated areas.
" Schauer et al., 1996.
bChowetal., 1991.
c California Environmental Protection Agency, 1998.
dWittorffetal., 1994.
' Maricopa Association of Governments, 1999.
fFujitaetal., 1998.
-------
Table 2-24. Comparison of DPM concentrations reported by CMB and EC
surrogate calculation
Location
Brighton, CO
Welby, CO
Phoenix, AZ
Number
of
samples
67
61
12
Average DPM,
ug/m3
(CMB)
1.1 ±0.8
1.7±1.4
2.4 ±1.6
DPM, ug/m3
EC surrogate*
(lower-upper
bound estimate)
0.8-1.6
1.4-3.1
1.1-2.4
Lower-bound estimate: DPM=EC*0.62; Upper bound estimate: DPM=EC* 1.31.
7/25/00
2-100
DRAFT—DO NOT CITE OR QUOTE
-------
Table 2-25. Ambient diesel participate matter concentrations from elemental
carbon measurements in urban locations
Ref.
a
b
Year of sampling
1995, annual
1992-1995, annual
c
1995-6, annual
Location
Boston, MA
Rochester, NY
Washington, DC
MATES II f
Anaheim, CA
Burbank, CA
Los Angeles, CA
Fontanta, CA
Huntington Park, CA
Long Beach, CA
Pico Rivera, CA
Rubidoux, CA
DPMi5
ug/m3 lower-
upper bound
range (point
estimate)*
0.8-1.7(1.1)
Diesel PM2 5
^m3
avg± std dev.
2.4 ±1.8
3.3 ±1.9
3.5 ±1.9
3.4 ±2.3
4.5 ± 2.4
2.5 ±1.7
4.4 ± 2.2
3.4 ±2.0
DPM %
of total
PM
6-12
0.4-0.8
1.0-2.2
J
t
t
J
i
t
t
t
"Lower-bound range: DPM=EC*0.62; upper-bound range: DPM=EC* 1.31;
point estimate: DPM=EC*0.89
t Not available.
JThe Multiple Air Toxics Exposure Study in the South Coast Air Basin reported DPM calculated from EC
concentrations as DPM=EC* 1.04. Standard deviations are reported.
a Salmon etal., 1997.
b Sisler, 1996.
c South Coast Air Quality Management District, 1999.
7/25/00
2-101
DRAFT—DO NOT CITE OR QUOTE
-------
Table 2-26. Ambient diesel participate matter concentrations from dispersion modeling
Ref.
a
b
c
Location
Azusa, CA
Anaheim, CA
Lennox, CA
Pasadena, CA
Long Beach, CA
Downtown LA, CA
West LA, CA
Claremont, CA
Long Beach, CA
Fullerton, CA
Riverside, CA
Year of sampling
1982, annual
1982, annual
1982, annual
1982, annual
1982, annual
1982, annual
1982, annual
18-19 Aug 1987
24 Sept 1996
24 Sept 1996
25 Sept 1996
Location
type
Nonurban
Nonurban
Nonurban
Urban
Urban
Urban
Urban
Nonurban
Urban
Nonurban
Suburban
DPM^
jig/m3
(mean)
1.4a
2.7a
3.8a
2.0a
3.5a
3.5a
3.8a
2.4 (4.0)a-b
1.9(2.6)b
2.4(3 .9)b
4.4(1 3.3)b
DPM %
of total
PM
5
12
13
7
13
11
16
8(6)b
8(7)b
9(8)b
12(13)b
On-road diesel vehicles only; all other values are for on-road plus nonroad diesel emissions.
bValue in parenthesis includes secondary DPM (nitrate, ammonium, sulfate and hydrocarbons) due to atmospheric
reactions of primary diesel emissions of NO,, SO2 and hydrocarbons. For the traction of ambient PM attributable to
DPM, the value in parenthesis reports total DPM (primary plus secondary) as a fraction of total ambient PM
(primary plus secondary).
Refences:
' Cass and Gray, 1995.
b Kleeman and Cass, 1999.
e Kleeman et al., 1999.
7/25/00
2-102 DRAFT—DO NOT CITE OR QUOTE
-------
Table 2-27. Occupational exposure to DPM
Author
Gangal and Dainty,
1993s
Saverin, 1999
Rogers and
Whelan, 1999
Haney, 1990"
Ambs, 199 la1
Woskie et al., 1988
Froines et al., 1987
NTOSH 1992a
Birch and Gary,
1996
Birch and Gary,
1996
NIOSH, 1990
Zaebstetal., 1991
Kittelson et al.,
2000
Year of
sample
NA
1992
1990-99
1980s
NA
3 -year
period in
mid-1980s
1985
NA
NA
NA
NA
NA
1990
1990
1999-2000
Location/job type
Typical work schedule of 8
hours
Noncoal mine workers
Noncoal mine workers
Coal mine workers
Coal mine workers
(five mines)
Coal mine workers
(four mines)
Railroad engineer/frier
Railroad broker/conductor
Railroad shop workers
Firefighters (two stations)
Firefighters
(three stations)
Firefighters
Fire station employees (four
stations)
Airport ground crew
Public transit workers
Diesel forklift dockworkers
Dockworkers
Mechanics
Long- and short-haul
truckers
Bus drivers
Parking ramp attendants
n
-200
255"
>1,30
0
NA
NA
128
158
176
238
18
NA
NA
NA
NA
24
75
80
128
39
12
Sample
type
RCD
RTC
DPSM
M
SJ1
PDEAS
ARP
ARP
ARP
TSP
EC
EC
EC
EC
EC
EC
EC
EC
EC
EC
EC
Range in
DPM jig/m3
100-900
38-1,280
10-640
180-1,000
750-780
39-73
52-191
114-134
63-748
6-70
20-79
4-52
7-15
15-98
12-61
9-20
5-28
2-7
1-3
2 ±0.4
"Cited in Watts (1995). NA: not available.
b Personal exposure and area samples were not reported separately for this study.
RCD: respirable combustible dust; RTC: respirable total carbon SPM: submicrometer PM; DPSMM: diesel
paniculate submicron mass (two-stage impaction sampler used to separate PM by size); EC: elemental carbon; SJI:
single-jet impactor agreed within 10% with simultaneous PDEAS measurements; PDEAS: personal diesel exhaust
aerosol sampler collects DPM <0.8um, McCartney and Cantrell (1992); SPM: paniculate matter; ARP: respirable
paniculate adjusted to remove the influence of cigarette smoke; TSP: total suspended paniculate matter.
7/25/00
2-103
DRAFT—DO NOT CITE OR QUOTE
-------
Table 2-28. Ranges of occupational exposure to diesel particulate matter by job category
with estimates of equivalent environmental exposures
Year of
sampling
1980s and
1990s
1980s
1985 and later
NA
1990
1990
Occupations
Miners
Railroad workers
Firefighters
Airport crew, public transit workers
Dockworkers, mechanics .
Long- and short-haul truckers
Occupational
DPM, ug/m3
10-1,280
39-191
4-748
7-98
5-61
2-7
Environmental
equivalent*
exposure, u£/m3
2-269
8^0
1-157
2-21
1-13
0.4-2
"Environmental equivalent exposure is calculated as the occupational exposure * (lOmVshift / 20m3/day) * (5 days /
7days) * (48 weeks / 52 weeks) * (45 year career / 70 year lifetime), or occupational exposure * 0.21 (discussed in
section 2.4.3.1.
7/25/00
2-104
DRAFT—DO NOT CITE OR QUOTE
-------
Table 2-29. Annual average nationwide DPM exposure estimates (p-g/m3) from on-road
sources for rural, urban, and urban demographic groups in 1990,1996,2007, and 2020
using HAPEM-MS3
Demographic group
50-State population
Rural population
Urban population
Urban outdoor workers
Urban children (0-1 7)
1990
0.8
0.5
0.9
1.1
0.9
1996
0.7
0.3
0.7
0.8
0.7
2007
0.4
0.2
0.4
0.5
0.4
2020
0.4
0.2
0.4
0.5
0.4
Source: U.S. EPA, 1999c, adjusted to reflect HDDV VMT described in U.S. EPA, 2000b.
7/25/00
2-105
DRAFT—DO NOT CITE OR QUOTE
-------
Table 2-30. Annual average DPM exposures for 1990 and 1996 in the general
population and among the highest exposed demographic groups in nine urban
areas and nationwide from (on-road sources only) using HAPEM-MS3
Urban area
Nationwide
Atlanta, GA
Chicago, IL
Denver, CO
Houston, TX
Minneapolis, MN
New York, NY
Philadelphia, PA
Phoenix, AZ
Spokane, WA
St. Louis, MO
1990
Population
average
exposure,
lig/m3
0.8
0.8
0.8
0.7
0.6
1.0
1.6
0.7
1.4
1.3
0.6
1996
Population
average
exposure,
Hg/m3
0.7
0.8
0.5
0.7
0.7
0.8
1.0
0.6
1.2
1.0
0.5
Highest DPM exposure in
1990, Hg/m3 (demographic
group experiencing this
exposure)
NA
NA
1 .3 (outdoor workers)
1 .2 (outdoor workers)
0.8 (outdoor workers)
1 .5 (outdoor workers)
4.0 (outdoor children)
1 .2 (outdoor children)
2.4 (nonworking men 18-44)
2.0 (outdoor workers)
0.8 (outdoor workers)
NA - Not available.
Source: U.S. EPA, 1999c, adjusted to reflect HDDV VMT described in U.S. EPA, 2000b.
7/25/00
2-106 DRAFT—DO NOT CITE OR QUOTE
-------
Table 2-31. Modeled and estimated concentrations of
DPM in microenvironments for California for all sources
Microenvironment
Residences
Offices
Schools
Stores/public/retail bldgs
Outdoor places
Industrial plants3
Restaurants/lounges*
Other indoor places3
Enclosed vehicles3
Estimated mean DPM
(stdev), Hg/m3
1.9(0.9)
1.6(0.7)
1.9(0.8)
2.1 (0.9)
3.0(1.1)
3.0(1.1)
2.1 (0.9)
1.6(0.7)
3.0(1.1)
"Concentrations assumed based on similarity with modeled environments.
Source: California EPA, 1998a.
7/25/00
2-107 DRAFT—DO NOT CITE OR QUOTE
-------
Table 2-32. Estimated indoor air and total air exposures to DPM in California in 1990
Exposed population
All Califomians
South Coast Air Basin
San Francisco Bay Area
Total indoor
exposure (stdev),
Hg/m3
2.0 (0.7)
2.4 (0.9)
1.7(0.9)
Total air
exposure, (stdev),
Hg/m3
2.1 (0.8)
2.5 (0.9)
1.7(0.9)
Source: California EPA, 1998a.
7/25/00
2-108
DRAFT—DO NOT CITE OR QUOTE
-------
200 000
180 000
160 000
140 000
? 120 000
(0
j; 100 000
.a
E 80 000
60 000 -
40,000 -
20,000 -
•
•" .1
•• m\Uf
mm
m mm mm
mm • "•
Tfc »
•• • " o °
_ • O&w °
•" C^f
o V ^
_-^°*f*£te^°°£-'- "*-•••— iT
1930 1940 1950 1960 1970 1980 1990 2000
« Class 5
. Class 6
o Class 7
BClass 8
Figure 2-1. Diesel truck sales (domestic) for the years 1939-1997.
Source: AAMA Motor Truck Facts.
7/25/00
2-109
DRAFT—DO NOT CITE OR QUOTE
-------
100.0%
0.0%
1930 1940 1950 1960 1970 1980 1990 2000
Figure 2-2. Diesel truck sales as a percentage of total truck sales for the years 1939-199'
Source: AAMA Motor Truck Facts.
7/25/00
2-110 DRAFT—DO NOT CITE OR QUOTE
-------
Figure 2-3. Percentage of truck miles attributable to diesel trucks. VMT = vehicle miles
traveled.
Source: U.S. Bureau of the Census, 1999b.
7/25/00
2-111
DRAFT—DO NOT CITE OR QUOTE
-------
Q
I
«
o
W
O
Q.
45%
40%
35%
30%
25%
20%
15%
10%
5%
0%
N<»
N*-
^
N<»
N*
Model Year
Figure 2-4. Model year distribution of in-use heavy HD truck fleet in 1997.
Source: U.S. Bureau of the Census, 1999b.
7/25/00
2-112
DRAFT—DO NOT CITE OR QUOTE
-------
18%
16% -
g 14% .
H 12% -
0
X 10% .
X
"o 8% .
*•*
£ 6% -
0
* 4%
Q_
2%
0%
I V 1
|. ,]
?&
V
-, :
V'
v
' '•
>„:
s
''^,/
^ '"^'
,'A,
:.';
^ 'j
','.'
,';,
^v
%•
'?•}•
•• /:
'/',;•
;'','
$'•••
;•' .
"$
s®
s^v
,j.,,.
P
i/ , '
. '$
•«,i
v>v.
•;v,;
', f'
'?'!-
',-
"y ^
' %'
,*
•i^i
,;-,::
1 * •
;£:
'/«
„;,:
vT
'^"C'
;||
i'l'Xv
'vv
' .
'i> %'
,'v.V''.
1
I
i
n
m
m
m
m
m
1998 1997 1996 1995 1994 1993 1992 1991 1990 1989 1988 Pre-
1988
ModelYear
Figure 2-5. Model year distribution of vehicle miles traveled by the in-use heavy HD truck
fleet in 1997.
Source: U.S. Bureau of the Census, 1999b.
7/25/00
2-113 DRAFT—DO NOT CITE OR QUOTE
-------
Single Hole
Injection Nozzle
Precombustion
Chamber
Four Hole
Injection Nozzle
Figure 2-6. A comparison of IDI (A) and DI (B) combustion systems of high-speed HD
diesel truck engines. DI engines almost completely replaced IDI engines for these
applications by the early 1980s.
7/25/00
2-114
DRAFT—DO NOT CITE OR QUOTE
-------
Solid Carbonaceous/Ash Particle
with adsorbed hydrocarbon/sulfate layer
Sulfuric Acid Particles
Hydrocarbon/Sulfate Particles
0.2
Figure 2-7. Schematic diagram of diesel engine exhaust particles.
Source: Modified from Kittelson, 1998.
7/25/00
2-115 DRAFT—DO NOT CITE OR QUOTE
-------
Diesel PM2.5 Chemical Composition
SnUate, Nitrate
1%
(1-4%)
Organic Carbon
19%
(7-49%)
M e tall & E le menti
2%
(1-5%)
Other
3%
(1-10%)
Elemental Carbon
75%
(33-90%)
Figure 2-8. Typical chemical composition for diesel participate matter (PM2.5) from new
(post-1990) HD diesel vehicle exhaust.
7/25/00
2-116
DRAFT—DO NOT CITE OR QUOTE
-------
Figure 2-9. Trends in PM10 emissions from on-road and nonroad diesel engines from 1970
to 1998 and projections of emissions to 2007 and 2030.
Source: U.S. EPA, 2000a, National air pollutant emission trends, 1900-1998.
7/25/00
2-117 DRAFT—DO NOT CITE OR QUOTE
-------
Figure 2-10. Trends in NOX emissions from on-road and nonroad diesel engines from 1970
to 1998.
Source: U.S. EPA, 2000a, National air pollutant emission trends, 1900-1998.
7/25/00
2-118
DRAFT—DO NOT CITE OR QUOTE
-------
o eon
•
|
Figure 2-11. Trends in SO2 emissions from on-road diesel engines from 1970 to 1998 and
nonroad diesel engines from 1990 to 1998.
Source: U.S. EPA, 2000a, National air pollutant emission trends, 1900-1998.
7/25/00
2-119 DRAFT—DO NOT CITE OR QUOTE
-------
•5 Joo
Figure 2-12. Trends in VOC emissions from on-road and nonroad diesel engines from 1970
to 1998.
Source: U.S. EPA, 2000a, National air pollutant emission trends, 1900-1998.
7/25/00
2-120 DRAFT—DO NOT CITE OR QUOTE
-------
0 O n -r« ad Q N o n ro ad
o 1000
-5
19*0 1990
C»l«nd«r Y«»r
Figure 2-13. Trends in CO emissions from on-road and nonroad diesel engines from 1970
to 1998.
Source: U.S. EPA, 2000a, National air pollutant emission trends, 1900-1998.
7/25/00
2-121 DRAFT—DO NOT CITE OR QUOTE
-------
£
o
u
on
«
Figure 2-14. Percentage of total motor fuel use that is on-road diesel fuel since 1949.
Source: Federal Highway Administration, 1995.
7/25/00
2-122 DRAFT—DO NOT CITE OR QUOTE
-------
30,000,000
25,000,000
«)
c
_o
co 20,000,000
0
«4-
o
» 15.000,000
c
CO
CO
g 10,000,000
£
1-
5nnn nnn .
,UUU,UUU
0
III
Illllllllll
CDCDCDCDCD
COCO
-------
o
o
o
CM
m
o>
o>
O m
3 £
0)
«n -o
co X
o 2.
o
CO
O)
o
o
o
CM
_ o>
11
LO
o
CO
o>
10
r-
o>
o
o
o
o
(9|!UJ/6)
XQN
(e|!UJ/6)
QO
I
an
"o
ce
V
O
B
e
I
03
a
u
o
u
o
o
o
CM
o>
o>
Jl
o
00
C35
h-
O5
o
o
o
CM
in
O m
I I
13
m T3
o
- CO
in
O
z
•e
V O
i o
T^ o
«
•o
o
«•*
UD
(3|;uj/6) suojssjLug
(e|!UJ/6) suojssitug
N
"I
o
-------
14
12 -
10
I 8
_OJ
1
4 -
t •
• •
' i
1975
1980
1995
2000
1985 1990
Engine Model Year
Figure 2-17. Diesel engine certification data for NOx emissions as a function of model year.
Source: Data are from the transient test results provided in Table 2-8.
€
o>
1.50
1.25
1.00 -
o 0.75 -
09
at
| 0.50
a.
0.25
0.00
•:.
1975
1980
1995
2000
1985 1990
Engine Model Year
Figure 2-18. Diesel engine certification data for PM emissions as a function of model year.
Source: Data are from the transient test results provided in Table 2-8.
7/25/00
2-125 DRAFT—DO NOT CITE OR QUOTE
-------
1 90
1 70
1 50
1 30
1 1.10 -
0.90 .
0.70
0.50
0.30 -
0 10
19
70
* Allegheny/Tuscarora - PA
g Fort McHenry - MD
+ 4 Caldocott • CA
• ^
•
A
•
•
^
w
1975 1980 1985 1990 1995 2000
Year of Measurement
Figure 2-19. Emission factors from HD diesel vehicles from tunnel studies.
Source: Data from Pierson and Brachaczek, 1976; Szkarlat and Japar, 1983; Pierson et al., 1983;
Kirchstetter et al., 1999; Gertler et al., 1995, 1996; Gertler, 1999.
7/25/00
2-126 DRAFT—DO NOT CITE OR QUOTE
-------
15
0 10
5
(
40
30
£ 20
X
.
o
0
Line-Haul Cycle Emissions Data
MOx and PM (g/bhp-hr)
o £ o
o
n
a'ijf ' 0 O 0
I I I I
) 0.1 0.2 0.3 0.4 0.
PM
Switch Cycle Emissions Data
NO* and PM (g/bhp-hr)
.
^3
a
Oa ao 0
: 1 1 !
0.2 0.4 0.6 0.8
PM
5
1
!
i
Figure 2-20. Line-haul and switch emissions data.
Source: U.S. EPA, 1998a.
7/25/00
2-127 DRAFT—DO NOT CITE OR QUOTE
-------
130 -I
120 -
110
100
90 -
80 -
70
Naturally Aspirated
Turbocharged/AftereooJed
60 65 70 75 80 85 90 95 100 105 110 115
NOX %
Figure 2-21. Effect of turbocharging and aftercooling on NO, and PM.
Source: Mori, 1997.
"
M
M
• Turfaocharged
O Naturally Aspirated
I •
1974 1976 1378 1980
Engine Model Year
1S82
1984
Figure 2-22. Comparison of diesel engine dynamometer PM emissions for 4-stroke,
naturally aspirated and turbocharged engines.
Source: Data are from Table 2-8.
7/25/00
2-128 DRAFT—DO NOT CITE OR QUOTE
-------
Engine
(Exhaust
Air Intake
Ports
\
Positive
Displacement
Blower
Figure 2-23. An example of uniflow scavenging of a two-stroke diesel engine with a positive
displacement blower. Scavenging is the process of simultaneously emptying the cylinder of
exhaust and refilling with fresh air.
Source: Adapted from Taylor, 1990.
9 -
1
"5>
i 3-
M
E 2
UJ
* 1-
n -
• 4-Stroke Engines
o 2-Stroke Engines
o
o
o
e
s a
n *
O
i
Q <
U ,
S !
81
8 8
8 § J
Q
•
» •
I
! • o
PS i
i ••
1978 1980 1982 1984 1986 1988 1990 1992 1994 1996 1998
Engine Emissions Model Year
Figure 2-24. Comparison of two- and four-stroke vehicle diesel PM emissions from chassis
dynamometer studies.
Source: Yanowitz et al., 2000.
7/25/00
2-129
DRAFT—DO NOT CITE OR QUOTE
-------
JO
o>
0)
• 4-Stroke
o 2-Stroke
• •
0 0 •
8 * oo
*'i.
JJL
1975 1980 1985 1990
Engine Model Year
1995
2000
Figure 2-25. Comparison of two- and four-stroke engine diesel PM emissions from engine
dynamometer studies.
Source: Data are from Table 2-8.
i.uu -
j= 0.75 -
a.
"&
i °-50 '
CD
111
to 0.25 -
n nn .
O
• 4-Stroke
o 2-Stroke
o
. o '•
8 • o •
0 • o
* • • • * 8
1975
1S80
1995
2000
1985 1990
Engine Model Year
Figure 2-26. Diesel engine dynamometer SOF emissions from two- and four-stroke
engines. SOF obtained by dichloromethane extraction in most studies.
Source: Data are from Table 2-8.
7/25/00
2-130 DRAFT—DO NOT CITE OR QUOTE
-------
1000
= 800
E
"
600
w 400
o
I
2 200
6
• 4-stroke
o 2-stroke
1975
1980
1995
2000
1985 1990
Engine Model Year
Figure 2-27. Diesel engine aldehyde emissions measured in chassis dynamometer studies.
Source: Data are from Warner-Selph and Dietzmann, 1984; Schauer et al., 1999; Unnasch et al.,
1993.
300
250
200
jj 150
iu
•o 100
2
<
50
• 4-stroke
o 2-stroke
1976 1978 1980 1982 1984 1986 1988 1990 1992 1994 1996
Engine Model Year
Figure 2-28. Diesel engine aldehyde emissions from engine dynamometer studies.
Source: Data from Table 2-8.
7/25/00
2-131
DRAFT—DO NOT CITE OR QUOTE
-------
1.4
1.2
I"
fe 0.8
O
U. 0.6
O
V)
0.4
0.2
0.0
Wamer-Selph etal., 1984
Dietzmann et al.. 1980
Graboski et al., 1998
Rogge et al., 1993
1975
1980
1985
1990
1995
2000
Engine Model Year
Figure 2-29. Trend in SOF emissions based on chassis dynamometer testing of heavy-duty
diesel vehicles. Warner-Selph and coworkers: dichloromethane for 8 hours. Dietzman and
coworkers: hexane followed by dichloromethane, extraction times not reported. Graboski
and coworkers: VOF by vacuum sublimation at 225° C for 2.5 to 3 hours. Rogge and
coworkers: cyclohexane followed by a benzene/2-propanol mixture that may extract
significantly more organic matter.
0.8
0.6 -
f 0-4 A
0.2 -
O
0
o.o
O
A
A
Ullmanetal., 1984
McCarthy etal., 1992
Perez and Williams, 1989
Needham etal., 1989
Graboski, 1998
Spreenetal., 1995
Sienickietal., 1990
Martin, 1981
Mitchell et al., 1994
Barry etal., 1985
„
1976 1978 1980
1982 1984 1986 1988 1990 1992 1994
Engine Model Year
1996
Figure 2-30. Trend in SOF emissions for transient engine dynamometer testing of HD
diesel engines. Various extraction methods used; see Table 2-8.
7/25/00
2-132
DRAFT—DO NOT CITE OR QUOTE
-------
100
80 -
3 60 -
o
o
s*
uT
O
to
40
20
• Wamer-Selph, 1984
O Dietzmann, et al., 1980
T Graboski, et al., 1998
V Rogge, et al., 1993
1975 1980 1985 1990
Engine Model Year
1995
2000
Figure 2-31. Trend in SOF emissions as a percent of total PM based on chassis
dynamometer testing of HD diesel vehicles. Warner-Selph and coworkers:
dichloromethane for 8 hours. Dietzman and coworkers: hexane followed by
dichloromethane, extraction times not reported. Graboski and coworkers: VOF by
vacuum sublimation at 225° C for 2.5 to 3 hours. Rogge and coworkers: cyclohexane
followed by a benzene/2-propanol mixture, that may extract significantly more organic
matter.
7/25/00
2-133
DRAFT—DO NOT CITE OR QUOTE
-------
100
80
I
~3 60 •
o
35
OT
40 -
20 -
o
o
A •
Q
1976 1978 1980 1982 1984 1986 1988 1990 1992 1994 1996
Engine Model Year
Figure 2-32. Trend in SOF emissions as a percentage of total PM from engine
dynamometer testing. Data are from Table 2-8.
7/25/00
2-134 DRAFT—DO NOT CITE OR QUOTE
-------
I 2 -
I
at
tn
0
Ul
0 -
O
Zielinska, 1998
Schauer, 1999
Rogge, 1993
Norbeck, 1998 (pickup trucks)
I
I
80
82
84
94
96
98
86 88 90 92
Engine Model Year
Figure 2-33. EC emission rates for diesel vehicles. All studies employed TOR for
measurement of EC. Vehicles tested by Norbeck and co-workers (1998) were all light and
medium HD pickup trucks.
0
4"
CO
o
1
s
s?
o"
UJ
IUU
90 -
80 -
70 -
60 -
50 -
40 -
30 -
20 -
10 -
n
v
• V ^7
. • * *
v m
V •
• * v
v •
^7
™
• Zielinska, etal., 1998
0 Schauer, 1999
T Rogge, 1993
v Norbeck, 1998
80 82 84 86 88 90 92
Engine Model Year
94 96
98
Figure 2-34. EC content as percent of total carbon content for DPM samples obtained in
chassis dynamometer studies.
7/25/00
2-135
DRAFT—DO NOT CITE OR QUOTE
-------
1
a
o
"5
(O
E
UJ
a
'ts
a
5
a.
T-
1<» '
12 -
10 -
8 -
6 -
4 -
2 -
0 -
0 • Benzo(a)pyrene
o 1-Nitropyrene
"
o o
o
8
o
o ° *
• 0 0 O
0 p o o •
1975 1980 1985 1990
Engine Model Year
1995
2000
Figure 2-35. Diesel engine emissions of benzo[a]pyrene and 1-nitropyrene measured in
chassis dynamometer studies.
Source: Schuetzle and Perez, 1983; Zielinska et al., 1988; Kado et al., 1996; Dietzmann et al.,
1980; Wamer-Selph and Dietzmann, 1984; Rogge et al., 1993; Schauer et al., 1999.
7/25/00
2-136
DRAFT—DO NOT CITE OR QUOTE
-------
30
S 25
"a
_o
» 20
|
I 11H
O 10 !
I
&
S 5
o
• Transient Test B(a)P
o Steady-State Test B(a)P
T Transient Test 1-NP
v Steady-State Test 1-NP
1970 1975 1980 1985 1990
Engine Model Year
1995
2000
Figure 2-36. Diesel engine dynamometer measurements of benzo[a]pyrene and 1-
nitropyrene emissions from HD diesel engines.
Source: Data are from Table 2-8.
7/25/00
2-137 DRAFT—DO NOT CITE OR QUOTE
-------
Fine PBrticJBS ^ |
Dp < 2.5 nm I
Nanopartides
Dp < 50 run
*I Uteafine Particles
Op< 100 nm
H ,,
/ Accumulatkjn \
Mode
PM10
Op<10(im
Coarse
Mode
0.001
0.010
0.100
1.000
•Mass Weighting Numbar W€Mghting|
10.000
Figure 2-37. Particle size distribution in diesel exhaust.
Source: Kittelson, 1998.
7/25/00
2-138 DRAFT—DO NOT CITE OR QUOTE
-------
2.6. REFERENCES
Abbass, MK; Andrews, GE; Ishaq, RB; et al. (1991) A comparison of the paniculate composition between
turbocharged and naturally aspirated DI diesel engines. SAE Technical Paper Ser. No. 910733.
5
6 Abdul-Khalek, IS; Kittelson, DB; Graskow, BR; et al. (1998) Diesel exhaust particle size: measurement issues and
7 trends. SAE Technical Paper Ser. No. 980525.
8
9 Abdul-Khalek, IS; Kittelson, DB; Brear, F. (1999) The influence of dilution conditions on diesel exhaust particle
10 size distribution measurements. SAE Technical Paper Ser. No. 1999-01-1142.
11
12 Alkidas, AC. (1984) Relationships between smoke measurements and paniculate measurements. SAE Technical
13 Paper Ser. No. 840412.
14
15 American Automobile Manufacturers Association (AAMA). (1975-1998) Motor vehicle facts and figures.
16
17 Andrews, GE; Ishaq, RB; Farrar-Khan, JR; et al. (1998) The influence of speciated diesel fuel composition on
18 speciated paniculate SOF emissions. SAE Technical Paper Ser. No. 980527 (and references therein).
19
20 Arey, J. (1998) Atmospheric reactions of PAHs including formation of nitroarenes. In: The handbook of
21 environmental chemistry, Vol. 3, Part I. PAHs and related compounds. Nielsen, AH, ed. Berlin/Heidelberg,
22 Germany: Springer-Verlag, pp. 347-385.
23
24 Arey, J; Zielinska, B; Atkinson, R; et al. (1986) The formation of nitro-PAHs from the gas-phase reactions of
25 fluoranthene and pyrene with the OH radical in the presence of NOX. Atmos Environ 20:2339-2345.
26
27 Arey, J; Zielinska, B; Atkinson, R; et al. (1987) Polycyclic aromatic hydrocarbon and nitroarene concentrations in
2 8 ambient air during a wintertime high-NO, episode in the Los Angeles Basin. Atmos Environ 21:143 7-1444.
^F Arey, J; Zielinska, B; Atkinson, R; et al. (1988) Formation of nitroarenes during ambient high-volume sampling.
31 Environ Sci Technol 22:457-462.
32
33 Arey, J; Zielinska, B; Atkinson, R; et al. (1989) Nitroarene products from the gas-phase reactions of volatile
34 polycyclic aromatic hydrocarbons with OH radical and N2O5. Int J Chem Kinet 21 -.715-199.
35
36 Atkinson, R. (1986) Kinetics and mechanisms of the gas-phase reactions of the hydroxyl radical with organic
37 compounds under atmospheric conditions. Chem Rev 86:69-201.
38
39 Atkinson, R. (1988) Atmospheric transformation of automotive emissions. In: Air pollution, the automobile and
40 public health. Watson, AY; Bates, RR; Kennedy, D, eds. Washington, DC: National Academy Press, pp. 99-132.
41
42 Atkinson, R. (1991) Kinetics and mechanisms of the gas-phase reactions of the NO} radical with organic
43 compounds. J Phys Chem Ref Data 20:459-507.
44
45 Atkinson, R; Arey, J. (1994) Atmospheric chemistry of gas-phase polycyclic aromatic hydrocarbons: formation of
46 atmospheric mutagen. Environ Health Perspect 102(Suppl. 4):117-126.
47
48 Atkinson, R; Arey, J; Zielinska, B; et al. (1987) Kinetics and products of the gas-phase reactions of OH radicals and
49 N,O5 with naphthalene and biphenyl. Environ Sci Technol 21:1014-1022.
50
51 Atkinson, R; Arey, J; Winer, AM; et al. (1988) A survey of ambient concentrations of selected polycyclic aromatic
52 hydrocarbons (PAH) at various locations in California. Final report prepared under contract no. A5-185-32 for
53 California Air Resources Board, Sacramento, CA.
54
» Atkinson, R; Arey, J; Zielinska, B; et al. (1990) Kinetics and nitro-products of the gas-phase OH and NO3 radical-
initiated reactions of naphthalene-d8 fluoranthene-d,0 and pyrene. Int J Chem Kinet 22:999-1014.
58 Automobile Manufacturers Association (AMA). (1927-1974) Motor truck facts. Motor Truck Committee.
7/25/00 2-139 DRAFT—DO NOT CITE OR QUOTE
-------
1 Bagley, ST; Baumgard, KJ; Gratz, LD. (1991) Comparison of in-rain and laboratory-generated diesel paniculate
2 matter, biological activity, and polynuclear aromatic hydrocarbon levels. In: Proceedings of the Third Symposium
3 of Respirable Dust in the Mineral Industries. Franz, RL; Ramini, RV, eds. Littleton, CO: Society for Mining,
4 Metallurgy, and Exploration, pp. 61-70.
5
6 Bagley, ST; Baumgard, KJ; Gratz, LD. (1992) Polynuclear aromatic hydrocarbons and biological activity associated
7 with diesel paniculate matter collected in underground coal mines. IC9324. Pittsburgh, PA: Bureau of Mines, pp.
8 40-48.
9
10 Bagley, ST; Gratz, LD; Leddy, DG; et al. (1993) Characterization of panicle and vapor phase organic fraction
11 emissions from a HD diesel engine equipped with a particle trap and regeneration controls. Health Effects Institute,
12 Research Report No. 56. Cambridge, MA: HEI.
13
14 Bagley, ST; Baumgard, KJ; Gratz, LD; et al. (1996) Characterization of fuel and aftertreatment effects on diesel
15 emissions. Health Effects Institute, Research Report No. 76. Cambridge, MA: HEI.
16
17 Bagley, ST; Gratz, LD; Johnson, JH; et al. (1998) Effects of an oxidation catalytic converter and a biodiesel fuel on
18 the chemical, mutagenic, and particle size characteristics of emissions from a diesel engine. Environ Sci Technol
19 32:1183-1191.
20
21 Baines, TM. (1986) Correlation testing of Cummins NTCC-400. EPA-AA-SDSB-87-1.
22
23 Ballschmiter, K.; Buchert, H.; Niemczyk, R.; et al. (1986) Automobile exhausts versus municipal waste incineration
24 as sources of the polychloro-dibenzodioxins (PCDD) and -furans (PCDF) found in the environment. Chemosphere
25 15(7):901-915.
26
27 Baranescu, RA. (1988) Influence of fuel sulfur on diesel paniculate emissions. SAE Technical Paper Ser. No.
28 881174.
29
30 Barry, EG; McCabe, LJ; Gerke, DH; et al. (1985) Heavy-duty diesel engine/fuels combustion performance and
31 emissions—a cooperative research program. SAE Technical Paper Ser. No. 852078.
32
33 Bata, R; Wang, W; Gautam, M; et al. (1992) Fleet-site measurements of exhaust gas emissions from urban buses.
34 ASME, ICE. New Dev Off-Highway Eng 18:185-196.
35
36 Baumgard, KJ; Johnson, JH. (1992) The effect of low sulfur fuel and a ceramic particle filter on diesel exhaust
37 particle size distributions. SAE Technical Paper Ser. No. 920566.
38
39 Baumgard, KJ; Johnson, JH. (1996) The effect of fuel and engine design on diesel exhaust particle size distributions.
40 SAE Technical Paper Ser. No. 960131.
41
42 Behymer, TD; Kites, RA. (1985) Photolysis of polycyclic aromatic hydrocarbons adsorbed on simulated
43 atmospheric particulates. Environ Sci Technol 19:1004-1006.
44
45 Behymer, TD; Hites. RA. (1988) Photolysis of polycyclic aromatic hydrocarbons adsorbed on fly ash. Environ Sci
46 Technol 22:1311-1319.
47
48 Benner, BA; Gordon, GE; Wise, SA. (1989) Mobile sources of atmospheric PAH: a roadway runnel study. Environ
49 Sci Technol 23:1269-1278.
50
51 Birch, ME. (1998) Analysis of carbonaceous aerosols: interlaboratory comparison. Analyst 123:851-857.
52
53 Birch, ME; Gary, RA. (1996) Elemental carbon-based method for monitoring occupational exposures to particulate
54 diesel exhaust. Aerosol Sci Technol 25:221-241.
55
56 Bishop, GA; McLaren, SE; Stedman, DH; et al. (1996) Method comparisons of vehicle emission* measurements in
57 the Fort McHenry and Tuscarora Mountain Tunnels. Atmos Environ 30:2307.
58
7/25/00 2-140 DRAFT—DO NOT CITE OR QUOTE
-------
1 Black, F; High, L. (1979) Methodology for determining paniculate and gaseous diesel hydrocarbon emissions. SAE
"1 Technical Paper Ser. No. 790422.
>3
4 Brodowicz, P. (1999) Determination of the demographic groups with the highest annual average modeled diesel PM
5 exposure. Memorandum. U.S. Environmental Protection Agency, National Fuel and Vehicle Emissions Laboratory.
6
7 Brown, KF. (1998) Diesel aftertreatment technologies. In: Proceedings of the SAE Topical Technical Seminar
8 Diesel Technology for the New Millennium. Schaumburg, IL, April 22, 1998.
9
10 Brown, KF; Rideout, GR; Turner, JE. (1997) Urban driving cycle results of retrofitted diesel oxidation catalysts on
11 heavy duty vehicles: one year later. SAE Technical Paper Ser. No. 970186.
12
13 Brown, DF; Rideout, GR. (1996) SAE Technical Paper Ser. No. 960134.
14
15 Brunekreef, B. (1999) Environmental diesel exhaust exposure and respiratory health of children in the Netherlands.
16 Presented at the Health Effects Institute Diesel Workshop: Building a Research Strategy to Improve Risk
17 Assessment. Stone Mountain, GA, March 7-9, 1999.
18
19 Brunekreef, B; Janssen, NAH; van Vliet, PHN; et al. (2000) Paniculate matter concentrations in relation to degree of
20 urbanization and proximity to highways in the Netherlands. Presented at: PM2000: Paniculate Matter and Health -
21 The Scientific Basis for Regulatory Decision-Making. Specialty Conference and Exhibition. January 24-28, 2000,
22 Charleston, SC.
23
24 Burke, JB; Aufall, M; Ozkaynak, H; et al. (2000) Predicting population exposures to PM: The importance of
25 microenvironmental concentrations and human activities. Presented at: PM2000: Paniculate Matter and Health -
26 The Scientific Basis for Regulatory Decision-Making. Specialty Conference and Exhibition. January 24-28,2000,
27 Charleston, SC.
28
Cadle, SH; Mulawa, PA. (1990) Atmospheric carbonaceous species measurement methods comparison study:
General Motors results. Aerosol Sci Technol 12(1): 128-141.
31
32 California Air Resources Board (CARB). (1987) Determination of PCDD and PCDF emissions from motor
33 vehicles. Draft report. October 1987. Test report no. C-86-029.
34
35 California Air Resources Board and Energy and Environmental Analysis Inc. (1997) Regulatory Amendments to
36 California's heavy-duty inspection program (HDVIP) and periodic smoke inspection program (PSIP). August 1997.
37
38 California Environmental Protection Agency. (1998a) Report to the California Air Resources Board on the proposed
39 identification of diesel exhaust as a toxic air contaminant. Appendix III, part A: Exposure assessment. April 1998.
40 http://www.arb.ca.gov/regact/diesltac/diesltac.htm.
41
42 California Environmental Protection Agency. (1998b) Measuring concentrations of selected air pollutants inside
43 California vehicles. Final report.
44
45 Calvert, JG; Stockwell, WR. (1983) Acid generation in the troposphere by gas-phase chemistry. Environ Sci Technol
46 17:428A-443A.
47
48 Campbell, RM; Lee, ML. (1984) Capillary column gas chromatographic determination of nitro polycyclic aromatic
49 compounds in paniculate extracts. Anal Chem 56:1026-1030.
50
51 Cantrell, BK; Salas, LJ; Johnson, WB; et al. (1988) Phase distributions of low volatility organics in ambient air.
52 Prepared for the U.S. Environmental Protection Agency, Washington, DC, under EPA contract no. 68-02-3748.
53
54 Carey, PM. (1987) Air toxics emissions from motor vehicles. U.S. Environmental Protection Agency, Washington,
DC, Office of Mobile Sources, Report No. EPA/AA-TSS-PA-86-5.
7/25/00 2-141 DRAFT—DO NOT CITE OR QUOTE
-------
1 Cass, GR; Gray, HA. (1995) Regional emissions and atmospheric concentrations of diesel engine paniculate matter:
2 Los Angeles as a case study. In: Diesel exhaust: a critical analysis of emissions, exposure, and health effects. A
3 special report of the Institute's Diesel Working Group. Cambridge, MA: Health Effects Institute, pp. 125-137.
4
5 Cautreels, W; Van Cauwenberghe, K. (1978) Experiments on the distribution of organic pollutants between airborne
6 paniculate matter and the corresponding gas phase. Atmos Environ 12:1133-1141.
7
8 Chow, JC; Watson, JG; Richards, LW; et al. (1991) The 1989-1990 Phoenix PM10 study. Volume II: source
9 apportionment. Final report. DRI Document No. 8931.6F1. Prepared for Arizona Department of Environmental Air
10 Quality, Phoenix, AZ, by Desert Research Institute, Reno, NV.
11
12 Ciccioli, P; Cecinato, A; Brancaleoni, E; et al. (1989) Evaluation of nitrated PAH in anthropogenic emission and air
13 samples: a possible means of detecting reactions of carbonaceous particles in the atmosphere. Aerosol Sci Technol
14 10:296-310.
15
16 Ciccioli, P; Cecinato, A; Cabella, R; et al. (1993) The contribution of gas-phase reactions to the nitroarene fraction
17 of molecular weight 246 present in carbon particles sampled in an urban area of northern Italy. Atmos Environ
18 27A:1261.
19
20 Clark, N; Messer, JT; McKain, DL; et al. (1995) Use of the West Virginia University truck test cycle to evaluate
21 emissions from Class 8 trucks. SAE Technical Paper Ser. No. 951016.
22
23 Clark, N; Gautam, M; Lyons, D; et al. (1997) Natural gas and diesel transit bus emissions: review and recent data.
24 SAE Technical Paper Ser. No. 973203.
25
26 Clark, NN; Gautam, M; Rapp, BL; et al. (1999) Diesel and CNG transit bus emissions characterization by two
27 chassis dynamometer laboratories: results and issues. SAE Technical Paper Ser. No. 1999-01-1469.
28
29 CFR, Code of Federal Regulations, 40 CFR Subpart N; Vol. 40.
30
31 CFR, Code of Federal Regulations, Title 40, Section 80.29.
32
33 Cohen, LH; Countess, RJ; Countess, SJ. (1997) Advanced remote sensing technology demonstration. Mobile
34 Sources Air Pollution Review Committee.
35
36 Coifrnan, B. (1994) Evolution of the diesel locomotive in the United States.
37 http://www.cyberus.ca/~yardlimit/guide/locopaper.html
38
39 Cooper, C. (1999) Heavy-duty diesel smoke enforcement program in the Northeast. Presented at the Mobile Sources
40 Technical Review Subcommittee, July 1999.
41
42 Cooper, JA. et al. and NEA, Inc. (1987) PM10 source composition library for the South Coast Air Basin, volumes I
43 and II. Prepared for the South Coast Air Quality Management District, El Monte, CA. July 15, 1987.
44
45 Countess, R; Countess, S; Cohen, R. (1999) Remote sensing of heavy duty diesel trucks revisited. 9th Coordinating
46 Research Council On-Road Vehicle Emissions Workshop, San Diego, CA.
47
48 Coutant, RW; Grown, L; Chuang, JC; et al. (1988) Phase distribution and artifact formation in ambient air sampling
49 for polynuclear aromatic hydrocarbons. Atmos Environ 22:403-409.
50
51 Crebelii, R; Conti, L; Crochi, B; et al. (1995) Tne effect of fuel composition on the mutagenicity of diesel engine
52 exhaust. Mutat Res 346:167-172.
53
54 Cunningham, LF; Henly, TJ; Kulinowski, AM. (1990) The effects of diesel ignition improver in low-sulfur fuels on
55 heavy-duty diesel emissions. SAE Technical Paper Ser. No. 902173.
no
57 Dec, JE; Espey, C. (1995) Ignition and early soot formation in a diesel engine using multiple 2-D imaging
58 diagnostics. SAE Technical Paper Ser. No. 950456.
59
7/25/00 2-142 DRAFT—DO NOT CITE OR QUOTE
-------
I
•
Dickson, CL; Sturm, GP. (1994) Diesel fuel oils, 1994. National Institute for Petroleum and Energy Research.
NIPER-187PPS94/5.
Dietzmann, HE; Pamess, MA; Bradow, RL. (1980) Emissions from trucks by chassis version of 1983 transient
5 procedure. SAE Technical Paper Ser. No. 801371.
6
7 Dua, SK; Hopke, PK; Raunemaa, T. (1999) Hygroscopicity of diesel aerosols. Water Air Soil Pollut 112:247-257.
8
9 Dunlap, LS; Pellegrin, V; Ikeda, R; et al. (1993) Chassis dynamometer emissions testing results for diesel and
10 alternative-fueled transit buses. SAE Technical Paper Ser. No. 931783.
11
12 Dunstan, TDJ; Mauldin, RF; Jinxian, Z; et al. (1989) Adsorption and photodegradation of pyrene on magnetic,
13 carbonaceous, and mineral subtractions of coal stack ash. Environ Sci Technol 23:303-308.
14
15 Engler, BH. (1998) Catalytic reduction of NOX with hydrocarbons under lean diesel exhaust conditions. SAE
16 Technical Paper Ser. No. 980189.
17
18 Enya, T; Suzuki, H; Watanabe, T; et al. (1997) 3-Nitrobenzanthrone, a powerful bacterial mutagen and suspected
19 human carcinogen found in diesel exhaust and airborne particles. Environ Sci Technol 31:2772-2776.
20
21 Federal Highway Administration (FHA). (1995) Motor fuel use, 1919-1995. U.S. Department of Transportation:
22 www.fhwa.dotgov/ohim/Summary95/sectionl.html; Vol. MF221.
23
24 Federal Highway Administration. (1997) Highway statistics summary to 1995. U.S. Department of Transportation.
25 FHWA-PL-97-009.
26
27 Feilberg, A; Kamens, RM; Strommen, MR; et al. (1999) Modeling the formation, decay, and partitioning of
28 semivolatile nitro-polycyclic aromatic hydrocarbons (nitronaphthalenes) in the atmosphere. Atmos Environ
33:1231-1243.
31 Ferguson, DH; Gautam, M; Wang, WG; et al. (1992) Exhaust emissions from in-use heavy duty vehicles tested on a
32 transportable transient chassis dynamometer. SAE Technical Paper Ser. No. 922436.
33
34 Finlayson-Pitts, BJ; Pitts, JN, Jr. (1986) Atmospheric chemistry: fundamentals and experimental techniques. New
3 5 York: John Wiley & Sons.
36
37 Foster, DE; Tree, DR. (1994) Optical measurements of soot particle size, number density and temperature in a direct
38 injection diesel engine as a function of speed and load. SAE Technical Paper Ser. No. 940270.
39
40 Fowler, DP. (1985) Industrial hygiene/environmental sampling program to develop qualitative and quantitative data
41 for diesel exhaust emission exposure. Final report. Coordinating Research Council-APRAC Project No. CAPM-24-
42 78.
43
44 Frey, JW; Corn, M. (1967) Diesel exhaust particulates. Nature 216:615-616.
45
46 Friones, JR; Hinds, WC; Duffy, RM; Lafuente, EJ; Liu, WV. (1987) Exposure of firefighters to diesel emissions in
47 fire stations. A.m Ind. Hyg. Assoc. J. 48:202-207.
48
49 Fritz, SG. (1995) Emission measurements-locomotives. Prepared by Southwest Research Institute under U.S. EPA
50 Contract no. 68-C2-0144.
51
52 Fujita, EM; Lu, Z; Sheetz, L; et al. (1997) Determination of mobile source emission source fraction using ambient
53 field measurements. Final report. Prepared for Coordinating Research Council, Atlanta, GA, by Desert Research
54 Institute, Reno, NV, under CRC Project No. E-5-1.
» Fujita, E; Watson, JG; Chow, JC; et al. (1998) Northern Front Range Air Quality Study, volume C: source
apportionment and simulation methods and evaluation. Prepared for Colorado State University, Cooperative
58 Institute for Research in the Atmosphere, by Desert Research Institute, Reno, NV.
59
7/25/00 2-143 DRAFT—DO NOT CITE OR QUOTE
-------
1 Gautam, M; Ferguson, D; Wang, WG; et al. (1992) In-use emissions and performance monitoring of heavy duty
2 vehicles using a transportable transient chassis test facility. SAE Technical Paper Ser. No. 921751.
3
4 Gertler, AW. (1999) Real-world measurements of diesel paniculate matter. Presentation to CASAC-Mobile Sources
5 Technical Review Subcommittee, October 13, 1999.
6
7 Gertler, AW; Sagebiel, JC; Dippel, WA; et al. (1996) A study to quantify on-road emissions of dioxins and furans
8 from mobile sources: phase 2. Reno, NV: Desert Research Institute.
9
10 Gertler, AW; Sagebiel, JC; Dippel, WA; et al. (1998) Measurements of dioxin and furan emission factors from HD
11 diesel vehicles. J Air Waste Manage Assoc 48:276-278.
12
13 Ohio, AJ; Richards, JH; Dailey, LA; et al. (2000) The effects of oxidants on modifying the toxicity of ambient PM.
14 Society of Toxicology Annual Meeting.
15
16 Gorse, RA, Jr.; Riley, TL; Ferris, FC; et al. (1983) 1-Nitropyrene concentration and bacterial mutagenicity in on-
17 road vehicle paniculate emissions. Environ Sci Technol 17:198-202.
18
19 Graboski, MS; McCormick, RL. (1996) Effect of diesel fuel chemistry on emissions at high altitude. SAE Technical
20 Paper Ser. No. 961947.
21
22 Graboski, MS; McCormick, RL; Alleman, T. (1998a) Testing of natural gas and diesel buses for comparison with
23 WVU mobile dynamometer. Colorado Institute for Fuels and High Altitude Engine Research, Colorado School of
24 Mines, Golden, CO.
25
26 Graboski, MS; McCormick, RL; Yanowitz, J; et al. (1998b) HD diesel testing for the Northern Front Range Air
27 Quality Study. Colorado Institute for Fuels and Engine Research, Colorado School of Mines, Golden, CO.
28 http://nfraqs.cira.colostate.edu/index2.html.
29
30 Gray, HA. (1986) Control of atmospheric fine primary carbon panicle concentrations. Ph.D. thesis, California
31 Institute of Technology.
32
33 Gray, HA; Kuklin, A. (1996) Benefits of mobile source NOX related paticulate matter reductions. Final report.
34 Prepared by Systems Applications International for U.S. Environmental Protection Agency. SYSAPP-96/61.
35
36 Gray, HA; Cass, GR. (1998) Source contributions to atmospheric fine carbon panicle concentrations. Atmos
37 Environ 32(22):3805-3825.
38
39 Gullett, BK; Ryan, JV. (1997) On-road sampling of diesel engine emissions of polychlorinated dibenzo-p-dioxin
40 and polychlorinated dibenzofuran. Organohalogen Compounds 32:451-456.
41
42 Hagenmaier, H. (1994) Contributions of diesel-powered vehicles and wood burning to overall PCDD/PCDF
43 emissions. Organo Comp 20:267-270.
44
45 Hagenmaier, H; Dawidowsky, V; Weber, UB; et a!. (1990) Emission of polyhalogenated dibenzcdicxins and
46 dibenzofurans from combustion-engines. Organo Comp 2:329-334.
47
48 Hames, RJ; Straub, RD; Amann, RW. (1985) DDEC Detroit Diesel electronic control. SAE Technical Paper Series
49 No. 850542.
50
51 Hammerle, RH; Schuetzle, D; Adams, W. (1994) A perspective on the potential development of environmentally
52 acceptable light-duty diesel engines. Environ Health Perspect (Suppl.) 102:25-30.
53
54 Hammond, SK. (1998) Occupational exposure to diesel exhaust. Presented at the Scientific Review Panel Meeting,
55 California Air Resources Board, March 11, 1998.
56
57 Hansen, ADA; Novakov, T. (1990) Real-time measurement of aerosol black carbon during the carbonaceous species
58 methods comparison study. Aerosol Sci Technol 12(1): 194-199.
7/25/00 2-144 DRAFT—DO NOT CITE OR QUOTE
-------
Hansen, KF; Bak, F; Andersen, EM; et al. (1994) The influence of an oxidation catalytic converter on the chemical
and biological characteristics of diesel exhaust emissions. SAE Technical Paper Ser. No. 940241.
4 Hara, K; Hanaoka, T; Yamano, Y; et al. (1997) Urinary 1-hydroxypyrene levels of garbage collectors with low-level
5 exposure to polycyclic aromatic hydrocarbons. Sci Total Environ 199:159-164.
6
7 Hardenberg, H; Fraenkle, G. (1978) The effect of charge air cooling on exhaust emissions and power output of
8 turbocharged engines. Institute of Mechanical Engineers Paper No. C71/78.
9
1 0 Hare, CT. (1977) Characterization of diesel gaseous and particulate emissions. Final report. Prepared by Southwest
1 1 Research Institute under contract no. 68-02-1777.
12
1 3 Hare, CT. (1985) Study of the effects of fuel composition, and injection and combustion system type and adjustment,
14 on exhaust emissions from light-duty diesels. Final report. Prepared by Southwest Research Institute for Project
1 5 CAPE-32-80.
16
1 7 Hare, CT; Springer, KJ. (1972) Exhaust emissions from uncontrolled vehicles and related equipment using internal
1 8 combustion engines. Part I, locomotive diesel engines and marine counterparts. Southwest Research Institute under
1 9 contract no. EHS 70- 1 08.
20
2 1 Harvey, GD; Baumgard, KJ; Johnson, JH; et al. (1994) Effects of a ceramic particle trap and copper fuel additive on
22 HD diesel emissions. SAE Technical Paper Ser. No. 942068.
23
24 Hawker, P; Myers, N; Huethwohl, G; et al. (1997) Experience with a new particulate trap technology in Europe.
2 5 SAE Technical Paper Ser. No. 970 1 82.
26
27 Hawker, P; Huethwohl, G; Henn, J; et al. (1998) Effect of a continuously regenerating diesel particulate filter on
28 non-regulated emissions and particle size distribution. SAE Technical Paper Ser. No. 980 1 89.
m
U Health Effects Institute. (1995) Diesel exhaust: a critical analysis of emissions, exposure, and health effects.
31 Cambridge, MA.
32
33 Heino, M. (1978) Work conditions and health of locomotive engineers. Scand J Work Environ Health 4:3 14.
34
35 Henry, RC; Lewis, CW; Collins, JF. (1994) Vehicle-related hydrocarbon source compositions from ambient data:
36 the GRACE/SAFER method. Environ Sci Technol 28:823-832.
37
38 Heywood, JB. Internal combustion engine fundamentals. New York: McGraw-Hill, Inc., 1988.
39
40 Hildemann, LM; Markowski, GR; Cass, GR. (1991) Chemical composition of emissions from urban sources of fine
41 organic aerosol. Environ Sci Technol 25:744-759.
42
43 Hori, S; Narusawa, K. (1998) Fuel composition effects on SOF and PAH exhaust emissions from a DI diesel engine.
44 SAE Technical Paper Ser. No. 980507.
45
46 Horvath, H; Kreiner, I; Norek, C; et al. (1988) Diesel emissions in Vienna. Atmos Environ 22:1255-1269.
47
48 Hutzinger, O; Essers, U.; Hagenmaier, H. (1992) Untersuchungen zur emission halogenierter dibenzodioxine und
49 dibenzofurane aus verbrennungsmortoren beim betrieb mit handelsiiblichen betriebsstoffen. Universities of
50 Bayreuth, Stuttgart and Tubingen, Germany. GSF-Forschungszentrum, Munich, Germany, ISSN 0937-9932.
51
52 Jang, M; McDow, SR. (1995) Benz[a]anthracene photodegradation in the presence of known organic constituents of
53 atmospheric aerosols. Environ Sci Technol 29:2654.
54
Japar, SM; Szkarlat, AC; Gorse, RA; et al. (1984) Comparison of solvent extraction and thermal-optical carbon
analysis methods: application to diesel vehicle exhaust aerosol. Environ Sci Technol 18:23 1-234.
58 Jimenez, J; McRae, G; Nelson, D; et al. (1998) Remote sensing of HD diesel truck NOX emissions using TILDAS.
59 8th Coordinating Research Council On-Road Vehicle Emissions Workshop: San Diego, CA, pp. 7-45-7-61.
7/25/00 2-145 DRAFT— DO NOT CITE OR QUOTE
-------
1 Johnson, JE. (1993) Hydrocarbon oxidation in a diesel oxidation catalytic converter. MS thesis, University of
2 Minnesota.
3
4 Johnson, JE; Kittelson, DB. (1996) Deposition, diffusion and adsorption in the diesel oxidation catalyst Appl Catal
5 B: Environ 10:117-137.
6
7 Johnson, JH; Bagley, ST; Gratz, LD; et al. (1994) A review of diesel paniculate control technology and emissions
8 effects-1992 Homing Memorial Award Lecture. SAE Technical Paper Ser. No. 940233.
9
10 Kado, NY; Okamoto, RA; Kuzmicky, PA. (1996) Chemical and bioassay analyses of diesel and biodiesel paniculate
11 matter: pilot study. Final report to Montana State Department of Environmental Quality and the U.S. Department of
12 Energy by Department of Environmental Toxicology, University of California, Davis.
13
14 Kado, NY; Kuzmicky, PA; Kiefer, KL; et al. (1998) Emissions from biodiesel fuel combustion: bioassay and
1 5 chemical analyses of the particle and semi-volatile emissions from hydrogenated biodiesel fuels. Final report to U.S.
1Q Department of Energy by Department of Environmental Toxicology, University of California, Davis.
17
18 Kakoi, Y; Tsutsui, Y; Ono, N; et al. (1998) Emission reduction technologies applied to high-speed direct injection
19 diesel engine. SAE Technical Paper Ser. No. 980173.
20
21 Kamens, RM; Guo, Z; Fulcher, JN; et al. (1988) Influence of humidity, sunlight, and temperature on the daytime
22 decay of polyaromatic hydrocarbons on atmospheric soot particles. Environ Sci Technol 22:103-108.
23
24 Kantola, TC; Bagley, ST; Gratz, LD; et al. (1992) The influence of a low sulfur fuel and a ceramic particle trap on
25 the physical, chemical, and biological character of HD diesel emissions. SAE Technical Paper Ser. No. 920565.
26
27 Katragadda, S; Bata, R; Wang, WG; et al. (1993) A correlation study between two HD vehicle chassis dynamometer
28 emissions testing facilities. SAE Technical Paper Ser. No. 931788.
29
30 Katz, M; Chan, C; Tosine, H; et al. (1979) Relative rates of photochemical and biological oxidation (in vitro) of
31 polycyclic aromatic hydrocarbons. In: Polynuclear aromatic hydrocarbons: Third International Symposium on
32 Chemistry and Biology-Carcinogenesis and Mutagenesis, October 1978, Columbus, OH. Jones, PW; Leber, P, eds.
33 Ann Arbor, MI: Ann Arbor Science Publishers, Inc., pp. 171-189.
34
35 Kawatani, T, Mori, K, Fukano, I, Sugawara, K, Koyama, T. (1993) Technology for meeting the 1994 USA exhaust
36 emission regulations on heavy-duty diesel engine. SAE Technical Paper Ser. No. 932654.
37
38 Khatri, NJ; Johnson, JH; Leddy, DG. (1978) The characterization of the hydrocarbon and sulfate fractions of diesel
3 9 paniculate matter. SAE Technical Paper Ser. No. 780111.
40
41 Kirchstetter, TW; Harley, RA; Kreisberg, NM; et al. (1999) On-road measurement of fine particle and nitrogen
42 oxide emissions from light- and HD motor vehicles. Atmos Environ 33:2955-2968.
43
44 Kirkland, JF. (1983) Dawn of the diesel age. Glendale, CA: Interurban Press.
45
46 Kittelson, DB. (1998) Engines and nanoparticles: a review. J Aerosol Sci 29:575-588.
47
48 Kittelson, DB. (2000) What do we know about the nature of the nanoparticles emitted by diesel engines?
49 Coordinating Research Council, January 2000.
50
51 Kitteison, DB; Pipho, ML; Ambs, JL; et ai. (1986) Particle concentrations in a diesel cylinder: comparison of theory
52 and experiment. SAE Technical Paper Ser. No. 861569.
53
54 Kittelson, DB; Winthrop, W; Baltensperger, R; et al. (1999) Diesel aerosol sampling methodology. University of
55 Minnesota Center for Diesel Research. http://www.me.umn.edu/centers/cdr/Proj_EPA.html.
55
7/25/00 2-146 DRAFT—DO NOT CITE OR QUOTE
-------
1 Kittelson, D; Watts, W; Ramachandran, G. (2000) Exposures to diesel exhaust aerosol: relationship between 6-hour
k2 time weighted average personal exposures to EC/OC and real-time variations in aerosol properties. Health Effects
Institute Annual Conference 2000: exploring exposure and risk issues. April 9-11,2000, Atlanta, GA.
4
5 Kleeman, MJ; Cass, GR. (1998) Source contributions to the size and composition distribution of urban particulate air
6 pollution. Atmos Environ 32(16):2803-2816.
7
8 Kleeman, MJ; Schauer, JJ; Cass, GR. (2000) Size and composition distribution of fine particulate matter emitted
9 from motor vehicles. Environ Sci Technol 34:1132-1142.
10
11 Klein, M. (1991) The diesel revolution. Am Heritage Invention Technol 6.
12
13 Kleindienst, TE; Shepson, PB; Edney, EO; et al. (1985) Peroxyacetyl nitrate: measurement of its mutagenic activity
14 using the Salmonella/mammalian microsome reversion assay. Mutat Res 157:123-128.
15
16 Kreso, AM; Johnson, JH; Gratz, LD; et al. (1998a) A study of the vapor- and particle-phase sulfur species in the HD
17 diesel engine EGR cooler. SAE Technical Paper Ser. No. 981423.
18
19 Kreso, AM; Johnson, JH; Gratz, LD; et al. (1998b) A study of the effects of exhaust gas recirculation on HD diesel
20 engine emissions. SAE Technical Paper Ser. No. 981422.
21
22 Ladomatos, N; Abdelhalim, SM; Zhao, H; et al. (1996-1997) The dilution, chemical, and thermal effects of exhaust
23 gas recirculation on diesel engine emissions-parts 1, 2, 3, and 4. SAE Technical Paper Ser. No.
24
25 Last, RJ; Krueger, M; Duernholz, M. (1995) Emissions and performance characteristics of a 4-stroke, direct injected
26 diesel engine fueled with blends of biodiesel and low sulfur diesel fuel. SAE Technical Paper Ser. No. 950054.
27
28 Law, PL; Lioy, PJ; Zelenka, MP; et al. (1997) Evaluation of a probabilistic exposure model applied to carbon
^fi monoxide (pNEM/CO) using Denver personal exposure monitoring data. J Air Waste Manage Assoc 47:491-500.
^B»
31 Lee, R; Hobbs, CH; Pedley, JF. (1998) Fuel quality impact on heavy duty diesel emissions: literature review. SAE
3 2 Technical Paper Ser. No. 982649.
33
34 Leet, JA; Friesen, T; Shadboume, A. (1998) EGR's effect on oil degradation and intake system performance. SAE
3 5 Technical Paper Ser. No. 980179.
36
37 Levson, K. (1988) The analysis of diesel particulate. Fresenius Z Anal Chem 331:467-478.
38
39 Lew, G. (1996) Letter to G. Schweer (Versar, Inc.) dated January 11, 1996. Sacramento, CA: State of California
40 Air Resources Board, Engineering and Laboratory Branch. Available for inspection at: U.S. Environmental
41 Protection Agency, Office of Research and Development, National Center for Environmental Assessment,
42 Washington, DC.
43
44 Lies, KH; Hartung, A; Postulka, A; et al. (1986) Composition of diesel exhaust with particular reference to particle
45 bound organics including formation of artifacts. In: Carcinogenic and mutagenic effects of diesel engine exhaust.
46 London: Elsevier Science Limited, pp. 65-82.
47
48 Ligocki, MP; Pankow, JF. (1989) Measurements of the gas/particle distributions of atmospheric organic compounds.
49 Environ Sci Technol 23:75-83.
50
51 Loft, S; Poulsen, HE; Vistisen, K; et al. (1999) Increased urinary excretion of 8-oxo-2'-deoxyguanosine, a biomarker
52 of oxidative DNA damage, in urban bus drivers. Mutat Res 441:11 -19.
53
54 Lowenthal, DH; Chow, JC; Watson, JG; et al. (1992) The effects of collinearity on the ability to determine aerosol
» contributions from diesel- and gasoline-powered vehicles using the chemical mass balance model. Atmos Environ
26A(13):2341-2351.
58 Lowenthal, DH, Zielinska, B., Chow, JC, Watson, JG, Gautam, M., Ferguson, DH, Neuroth, GR, Stevens, KD
59 (1994) Characterization of heavy-duty diesel vehicle emissions. Atmos. Environ. 28:731-743.
7/25/00 2-147 DRAFT—DO NOT CITE OR QUOTE
-------
1 Lucas, A; Duran, A; Carmona, M; et al. (1999) SAE Technical Paper Ser. No. 1999-01-3532.
2
3 Lilders, H; Stommel, P; Backes, R. (1997) Applications for the regeneration of diesel participate traps by combining
4 different regeneration systems. SAE Technical Paper Ser. No. 970470.
5
6 Lilders, H, Kruger, M, Strommel, P, Luers, B (1998) The role of sampling conditions in particle size distribution
7 measurements. SAE Technical Paper Ser. No. 981374.
8
9 Maricopa Association of Governments ( 1 999) The 1 999 Brown Cloud Project for the Maricopa Association of
1 0 Governments Area, Final Report, December 1999.
11
1 2 Marklund, S.; Andersson, R.; Tysklind, M.; Rappe, C; Egeback, K.E.; Bjorkman, E.; Grigoriadis, V. (1990)
1 3 Emissions of PCDDs and PCDFs in gasoline and diesel fueled cars. Chemosphere 20(5):553-561 .
14
1 5 Martin, S. (198 la) Emissions from heavy-duty engines using the 1984 transient test procedure. Volume II-diesel.
16 EPA/460/3-81/031.
17
1 8 Martin, S. (1981b) Diesel paniculate by 1986 HD transient Federal test procedure. Final report. Prepared by
1 9 Southwest Research Institute under contract no. 68-03-2603 .
20
2 1 Martin, S. (1984) Emissions from two in-use heavy-duty diesel engines. Final report. Prepared by Southwest
22 Research Institute. EPA/460/3-84/017.
23
24 Mauderly, J. (1992) Diesel exhaust. In: Environmental toxicants: human exposures and their health effects.
25 Lippmann, M, ed. New York: Van Nostrand Reinhold, pp. 1 19-162.
26
27 McCarthy, CI; Slodowske, WJ; Sienicki, EJ; et al. (1992) Diesel fuel property effects on exhaust emissions from a
28 heavy duty diesel engine that meets 1994 emissions requirements. SAE Technical Paper Ser. No. 922267.
29
30 McClure, BT; Bagley ST; Gratz, LD. (1992) The influence of an oxidation catalytic converter and the fuel
3 1 composition on the chemical and biological characteristics of diesel exhaust emissions. SAE Technical Paper Ser.
32 No. 920854.
33
34 McCormick, RL; Ryan, LBA; Daniels, TL; et al. (1998) Comparison of chassis dynamometer in-use emissions with
35 engine dynamometer FTP emissions for three HD diesel vehicles. SAE Technical Paper Ser No. 982653.
36
37 McCormick, RL; Graboski, MS; Alleman, TL; et al. (1999) In-use emissions from natural gas fueled HD vehicles.
38 SAE Technical Paper Ser. No. 1999-01-1507.
39
40 McCormick, RL; Alleman, TL; Graboski, MS; et al. (2000) Quantifying the emissions benefit of opacity testing and
41 repair of heavy-duty diesel vehicles. Proceedings of the 10* Coordinating Research Council On-Road Vehicle
42 Emissions Workshop, March 27-29, 2000, San Diego, CA.
43
44 McDonald, JF. (1997) The emissions and combustion characteristics of a soy-methyl-ester biodiesel fuel in a
45 naturally aspirated indirect injection diesel engine. MS thesis, University of Minnesota.
46
A "7 N/i^r^«^M T- i>it*>s*ir- fwr j;^Coi —,;«;—« ~*r*\;~~*.\~r*r. T-.
•1 * j.*jL.W*~rw*>**i%»9 w, A bu^WAA, wu9 *«Awvkl^iw, & A. . \iSS~rj J.**.w*i*jr> &MW1 X/^JT fy wJtifctvwti kbAwltJ ivA t^L^Owi lliliAUl.£ C4.L/L/4 iwilllO Iftt7. *-**.
48 Proceedings of the 7th U.S. Mine Ventilation Symposium, Society for Mining, Metallurgy, and Exploration, Inc.
49
50 McDow, SR; Sun, Q; Vartiainen, M; et al. (1994) Effect of composition and state of organic components on
b 1 polycyclic aromatic hydrocarbon decay in atmospheric aerosols. Environ Sci Technol 28:2 147.
52
53 McGuckin, J; Rykowski, R. (1981) Correlation between smoke and paniculate emissions from HD diesel engines.
54 Technical Report. Office of Mobile Sources, U.S. Environmental Protection Agency. Washington. DC.
55
_ _ __ . __ _ __ ___ _„_ — ---- ---- _ -- „ --- ---- -- --- ,
N^»W ivi^rvuiic^ , i JL*. \i.77 i ) ivivruwiLiig, ounuiiv awiu wsiiucudaiiim in uiw^&i canine UVJAX wuuicia. ^r^.£L 1 CWlUUCcil
57 Ser. No. 970636.
58
7/25/00 2- 1 48 DRAFT— DO NOT CITE OR QUOTE
-------
Miguel, AH; Kirchstetter, TW; Harley, RA; et al. (1998) On-road emissions of particulate polycyclic aromatic
hydrocarbons and black carbon from gasoline and diesel vehicles. Environ Sci Technol 32:450-455.
*4 Mitchell, K; Steere, DE; Taylor, JA; et al. (1994) Impact of diesel fuel aroman'cs on particulate, PAH, and nitro-
5 PAH emissions. SAE Technical Paper Ser. No. 942053.
6
7 Mori, K. (1997) Worldwide trends in HD diesel engine exhaust emission legislation and compliance technologies.
8 SAE Technical Paper Ser. No. 970753.
9
10 Muzyka, V; Veimer, S; Schmidt, N. (1998) On the carcinogenic risk evaluation of diesel exhaust: benzene in
11 airborne particles and alterations of heme metabolism in lymphocytes as markers of exposure. Sci Total Environ
12 217:103-111.
13
14 National Institute for Occupational Safety and Health (NIOSH). (1988) Carcinogenic effects of exposure to diesel
15 exhaust. Current Intelligence Bulletin 50. No. 88-116. U.S. Department of Health and Human Services, Public
16 Health Service, Centers for Disease Control, National Institute for Occupational Safety and Health, Cincinnati, OH.
17
18 National Institute for Occupational Safety and Health (NIOSH). (1990) Health hazard evaluation report. Prepared by
19 Yellow Freight System, Inc. HETA 90-088-2110.
20
21 National Institute for Occupational Safety and Health (NIOSH). (1992) Health Hazard Evaluation Report: City of
22 Lancaster, Division of Fire, Lancaster, OH. NIOSH Report No. HHE HETA 92-0160-2360, NTIA No. PB-
23 94140142. U.S. Department of Health and Human Services, Public Health Service, Centers for Disease Control,
24 National Institute for Occupational Safety and Health, Cincinnati, OH.
25
26 National Institute for Petroleum and Energy Research (NIPER). (1986 and earlier years) Diesel fuel oils, annual
27 survey.
28
• National Research Council. (1982) Diesel cars: benefits, risks and public policy. Final report of the Diesel Impacts
Study Committee. Washington, DC: National Academy Press.
31
32 Needham, JR; Doyle, DM; Faulkner, SA; et al. (1989) Technology for 1994. SAE Technical Paper Ser. No. 891949.
33
34 Nielsen, T; Ramdahl, T. (1986) Discussion on determination of 2-nitrofluoranthene and 2-nitropyrene in ambient
35 particulate matter: evidence for atmospheric reactions. Atmos Environ 20:1507.
36
37 Nielsen, PS; Andreassen, A; Farmer, PB; et al. (1996) Biomonitoring of diesel exhaust-exposed workers. DNA and
38 hemoglobin adducts and urinary 1-hydroxypyrene as markers of exposure. Toxicology Letters 86:27-37.
39
40 Noorkhoek, J; Bos, RP. (1995) 1-Nitropyrene as a marker for the mutagenicity of diesel exhaust derived particulate
41 matter in workplace atmospheres. Environ Mol Mutagen 25:134.
42
43 Norbeck, JM; Truex, TJ; Smith, MR; et al. (1998a) Evaluation of factors that affect diesel exhaust toxicity. Final
44 report, prepared under Contract No. 94-312, for California Air Resources Board, by the Center for Environmental
45 Research and Technology, College of Engineering, University of California, Riverside, CA.
46
47 Norbeck, JM; Durbin, TD; Truex, TJ; et al. (1998b) Characterizing particulate emissions from medium- and light
48 heavy-duty diesel-fueled vehicles. Final report. Prepared for South Coast Air Quality Management District by for
49 California Air Resources Board by the Center for Environmental Research and Technology, College of Engineering,
50 University of California, Riverside, C A, under contract no. 97031.
51
52 Norbeck, JM; Durbin, TD; Truex, TJ. (1998c) Measurement of primary particulate matter emissions from light-duty
53 motor vehicles. Prepared for Coordinating Research Council, Inc., and South Coast Air Quality Management
54 District, by Center for Environmental Research and Technology, College of Engineering, University of California,
Riverside, CA, 1998.
Norris, GA; Birch, ME; Tolocka, MP; et al. (2000) Comparison of particulate organic and elemental carbon
58 measurements made with the IMPROVE and NIOSH method 5040 protocols. Aerosol Sci Technol, in press.
7/25/00 2-149 DRAFT—DO NOT CITE OR QUOTE
-------
1 Northridge, ME; Yankura, J; Kinney, PL; et al. (1999) Diesel exhaust exposure among adolescents in Harlem: a
2 community-driven study. Am J Public Health 89:998-1002.
3
4 Odum, JR; McDow, SR; Kamens, RM. (1994) Mechanistic and kinetic studies of the photodegradation of
5 benz[a]anthracene in the presence of methoxyphenols. Environ Sci Technol 28: 1285.
6
7 Oehme, M; Larssen, S; Brevik, EM. (1991) Emission factors of PCDD/CDF for road vehicles obtained by a tunnel
8 experiment. Chemosphere 23 : 1 699- 1 708 .
9
1 0 Opris, CN; Grate, LD; Bagley, ST; et al. (1993) The effects of fuel sulfur concentration on regulated and unregulated
1 1 HD diesel emissions. SAE Technical Paper Ser. No. 930730.
12
1 3 Pagan, J. (1999) Study of particle size distributions emitted by a diesel engine. SAE Technical Paper Ser. No. 1999-
14 01-1141.
15
1 6 Paputa-Peck, MC; Marano, RS; Schuetzle, D; et al. (1983) Determination of nitrated polynuclear aromatic
1 7 hydrocarbons in paniculate extracts by capillary column gas chromatography with nitrogen selective detection. Anal
18 Chem 55: 1946- 1954.
19
20 Pataky, GM; Baumgard, KJ; Gratz, LD; et al. (1994) Effects of an oxidation catalytic converter on regulated and
2 1 unregulated diesel emissions. SAE Technical Paper Ser. No. 940243.
22
23 Perez, JM. (1980) Measurement of unregulated emissions-some heavy-duty diesel engine results. Health Effects of
24 Diesel Engine Emissions, EPA/600/9-80/05 7a.
25
26 Perez, JM; Williams, RL. (1989) A study of particulate extracts from 1980s heavy duty diesel engines run on steady-
27 state and transient cycles. SAE Technical Paper Ser. No. 89249 1 .
28
29 Peters, A; Langer, H-J; Joki, B; et al. ( 1 998) Catalytic NOX reduction on a passenger car diesel common rail engine.
30 SAE Technical Paper Ser. No. 980 1 9 1 .
31
32 Pierson, WR; Brachaczek, WW. (1976) Particulate matter associated with vehicles on the road. SAE Trans 85:209-
33 227.
34
35 Pierson, WR; Brachaczek, WW. (1983) Particulate matter associated with vehicles on the road. Aerosol Sci Tech
36 2:1-40.
37
38 Pierson, WR; Gertler, AW; Robinson, NF; et al. (1996) Real-world automotive emissions-summary of studies in the
39 Fort McHenry and Tuscarora mountain tunnels. Atmos Environ 30:2233-2256.
40
41 Pischinger, R; Cartellieri, W. (1972) Combustion system parameters and their effect on diesel engine exhaust
42 emissions. SAE Technical Paper Ser. No. 720756.
43
44 Pitts, JN, Jr; Van Cauwenberghe, KA; Grosjean, D; et al. (1978) Atmospheric reactions of polycyclic aromatic
45 hydrocarbons: facile formation of mutagenie ni'ao derivatives. Science 202:515-519.
46
4/ Pitts, JN, Jr; Lokensgsrd, DM; iiipley, PS; et s!. (1980) Atmospheric epoxiriaHon of b?n7o[?i]pyren<; by O70"§'
48 formation of the metabolite benzo[a]pyrene-4,5-oxide. Science 210:1347-1349.
49
50 Pitts, JN, Jr; Sweetman, JA; Zielinska, B; et al. (1985a) Determination of 2-nitrofluoranthene and 2-nitropyrene in
b 1 ambient particulate matter: evidence for atmospheric reaction. Atmos Environ 19: 160 1- 1 608.
52
53 Pitts, JN, Jr; Zielinska, B; Sweetman, JA; et al. (1985b) Reaction of adsorbed pyrene and perylene with gaseous
54 N,O5 under simulated atmospheric conditions. Atmos Environ 19:91 1-915.
55
.- _. ----- --- .
w/w l llLO, Ji"*) J* j *J vvvwuiiem, ^m., ^
57 aromatic hydrocarbons with dinitrogen pentoxide. Environ Sci Technol 19:1 1 15-1 121.
58
7/25/00 2-150 DRAFT— DO NOT CITE OR QUOTE
-------
Pitts, JN, Jr; Paur, HR; Zielinska, B; et al. (1986) Factors influencing the reactivity of polycyclic aromatic
hydrocarbons adsorbed on filters and ambient POM with ozone. Chemosphere 15:675-685.
Prinn, R; Cunnold, D; Simmonds, P; et al. (1992) Global average concentration and trend for hydroxyl radicals
5 deduced from ALE/GAGE trichloroethane (methyl chloroform data for 1978-1990). J Geophys Res 97:2445-2461.
6
7 Rantanen, L; Mikkonen, S; Nylund, L; et al. (1993) Effect of fuel on the regulated, unregulated and mutagenic
8 emissions of DI diesel engines. SAE Technical Paper Ser. No. 932686.
9
10 Rideout, G; Kirshenblatt, M; Prakash, C. (1994) Emissions form methanol, ethanol, and diesel powered urban transit
11 buses. SAE Technical Paper Ser. No. 942261.
12
13 Rogak, SN; Pott, U; Dann, T; et al. (1998) Gaseous emissions from vehicles in a traffic tunnel in Vancouver, British
14 Columbia. J Air Waste Manage Assoc 48:604-615.
15
16 Rogers, A; Whelan, B. (1999) Exposures in Australian mines. Health Effects Institute Number 7. March 7-9, 1999
17 Stone Mountain, GA.
18
19 Rogge, WF; Hildemann, LM; Mazurek, MA; et al. (1993) Sources of fine organic aerosol. 2. Noncatalyst and
20 catalyst-equipped automobiles and HD diesel trucks. Environ Sci Technol 27:636-651.
21
22 Salmon, LG; Cass, GR; Pedersen, DU; et al. (1997) Determination of fine particle concentration and chemical
23 composition in the northeastern United States, 1995. Progress Report to Northeast States for Coordinated Air Use
24 Management (NESCAUM), October 1997.
25
26 Sampson, PJ. (1988) Atmospheric transport and dispersion of air pollutants associated with vehicular emissions. In:
27 Air pollution, the automobile and public health. Watson, AY; Bates, RR; Kennedy, D, eds. Washington, DC:
28 National Academy Press, pp. 77-97.
SSverin, R. (1999) German potash miners: cancer mortality. Health Effects Institute Number 7. March 7-9, 1999
31 Stone Mountain, GA. pp. 220-229.
32
33 Sawyer, RF; Johnson, JH. (1995) Diesel emissions and control technology. In: Diesel exhaust: a critical analysis of
34 emissions, exposure, and health effects. Cambridge, MA: Health Effects Institute, pp. 65-81.
35
36 Schauer, JJ; Rogge, WF; Hildemann, LM; et al. (1996) Source apportionment of airborne paniculate matter using
37 organic compounds as tracers. Atmos Environ 30(22):3837-3855.
38
39 Schauer, JJ; Kleeman, MJ; Cass, GR; et al. (1999) Measurement of emissions from air pollution sources. 2. C,
40 through C30 organic compounds from medium duty diesel trucks. Environ Sci Technol 33:1578-1587.
41
42 Scholl, JP; Bagley, ST; Leddy, DG; et al. (1982) Study of aftertreatment and fuel injection variables for paniculate
43 control in heavy-duty diesel engines. EPA-460/3-83-002.
44
45 Schuetzle, D. (1983) Sampling of vehicle emissions of chemical analysis and biological testing. Environ Health
46 Perspect 47:65-80.
47
48 Schuetzle, D; Frazier, JA. (1986) Factors influencing the emission of vapor and paniculate phase components from
49 diesel engines. In: Carcinogenic and mutagenic effects of diesel engine exhaust. London: Elsevier Science Limited,
50 pp. 41-63.
51
52 Schuetzle, D; Perez, JM. (1983) Factors influencing the emissions of nitrated-polynuclear aromatic hydrocarbons
53 (nitro-PAH) from diesel engines. J Air Pollut Control Assoc 33:751-755.
54
» Schuetzle, D; Jensen, TE; Ball, JC. (1985) Polar polynuclear aromatic hydrocarbon derivatives in extracts of
particulates: biological characterization and techniques for chemical analysis. Environ Int 11:169-181.
7/25/00 2-151 DRAFT—DO NOT CITE OR QUOTE
-------
1 Schwind, K-H; Thoma, H; Hutzinger, O; et al. (1991) Emission halogenierter dibenzodioxine (PXDD) und
2 dibenzofurane (PXDF) aus verbrennungsmotoren. UWSF-Z. Umweltchem. Oekotox.J, 291-298. [English
3 translation]
4
5 Seinfeld, JH; Pandis, SN. (1998) Atmospheric chemistry and physics: from air pollution to climate change. New
6 York: John Wiley & Sons, Inc.
7
8 Shelton, EM. (1977) Diesel fuel oils. BERC/PPS-77/5.
9
10 Shelton, EM. (1979) Diesel fuel oils. BETC/PPS-79/5.
11
12 Shi, JP; Harrison, RM. (1999) Investigation of ultrafine particle formation during diesel exhaust dilution. Environ
13 Sci Technol 33:3730-3736.
14
15 Shi, JP; Mark, D; Harrison, RM. (2000) Characterization of particles from a current technology heavy-duty diesel
16 engine. Environ Sci Technol 34:748-755.
17
18 Sienicki, EJ; Jass, RE; Slodowske, WJ; et al. (1990) Diesel fuel aromatic and cetane number effects on combustion
19 and emissions from a prototype 1991 diesel engine. SAE Technical Paper Ser. No. 902172.
20
21 Sisler, JF. (1996) Spatial and seasonal patterns and long term variability of the composition of the haze hi the United
22 States: an analysis of data from the IMPROVE network. Cooperative Institute for Research in the Atmosphere.
23 Colorado State University. ISSN: 0737-5352-32.
24
25 South Coast Air Quality Management District (SCAQMD). (2000) Multiple air toxics exposure study (MATES-II).
26 Final Report and Appendices, March 2000.
27
28 Spreen, KB; Ullman, TL; Mason, RL. (1995) Effects of fuel oxygenates, cetane number, and aromatic content on
29 emissions from 1994 and 1998 prototype heavy-duty diesel engines. Final report. Prepared by Southwest Research
30 Institute under Contract No. VE-10.
31
32 Springer, KJ. (1979) Characterization of sulfates, odor, smoke, POM and particulates from light and heavy duty
33 engines-part IX. Prepared by Southwest Research Institute. EPA/460/3-79/007.
34
35 Springer, KJ; White, JT; Domke, CJ. (1974) Emissions from in-use 1970-1971 diesel-powered trucks and buses.
3 6 SAE Technical Paper Ser. No. 741006.
37
38 Stone, R. (1995) Introduction to internal combustion engines. Warrendale, PA: Society of Automotive Engineers.
39
40 Sweetman, JA; Zielinska, B; Atkinson, R; et al.(1986) A possible formation pathway for the 2-nitrofluoranthene
41 observed in ambient paniculate organic matter. Atmos Environ 20:235-238.
42
43 Tamanouchi, M; Morihisa, H; Araki, H; et al. (1998) Effects of fuel properties and oxidation catalyst on exhaust
44 emissions for heavy duty diesel engines and diesel passenger cars. SAE Technical Paper Ser. No. 980530.
45
46 Tanaka, S; Takizawa, H; Shimizu, T; et ai. (1998) Effect of fuel compositions on PAH in paniculate matter from DI
47 dje«el ens'.nes SAF, Technical Paper Ser No. 982648.
48
49 Tancell, PJ; Rhead, MM; Trier, CJ; et al. (1995) The sources of benzo[a]pyrene in diesel exhaust emissions. Sci
50 Total Environ 162:179-186.
51
52 Taylor, CF. (1990) The internal combustion engine in theory and practice - volume 1. Cambridge, MA: MIT Press.
53
54 Thrane, KE; Mikalsen, A. (1981) High-volume sampling of airborne polycyclic aromatic hydrocarbons using glass
55 fibre filters and polyurethane foam. Atmos Environ 15:909-918.
56
57 Tokiwa, H; Sera, N; Nakanishi, Y; et al. (1999) 8-Hydroyguanosine formed in human lung tissues and the
58 association with diesel exhaust particles. Free Rad Biol Med 27:1251-1258.
7/25/00 2-152 DRAFT—DO NOT CITE OR QUOTE
-------
1 Tong, HY; Sweetman, JA; Karasek, FW; et al. (1984) Quantitative analysis of polycyclic aromatic compounds in
2 diesel exhaust paniculate extracts by combined chromatographic techniques. J Chromatogr 312:183-202.
3
4 Ullman, TL; Hare, CT; Baines, TM. (1984) Preliminary paniculate trap tests on a 2-stroke diesel bus engine. SAE
5 Technical Paper Ser. No. 840079.
6
7 Ullman, TL. (1988) Emissions evaluation of two heavy-duty diesel engines on two low-sulfur fuels with variation in
8 aromatics. Final report. Prepared by Southwest Research Institute for California Air Resources Board under contract
9 no. A4-132-32.
10
11 Ullman, TL; Mason, RL; Montalvo, DA. (1990) Effects of fuel aromatics, cetane number, and cetane improver on
12 emissions from a 1991 prototype heavy-duty diesel engine. SAE Technical Paper Ser. No. 902171.
13
14 Ullman, TL; Spreen, KB; Mason, RL. (1995) Effects of cetane number on emissions from a prototype 1998 HD
15 diesel engine. SAE Technical Paper Ser. No. 950251.
16
17 U.S. Bureau of the Census. (1995) Truck inventory and use survey. 1992 Census of Transportation, TC92-T-52.
18
19 U.S. Bureau of the Census. (1999a) 1997 Economic census, vehicle inventory and use survey. EC97TV-US.
20
21 U.S. Bureau of the Census. (1999b) Truck inventory and use survey. 1997 Census of Transportation, EC97-TV-US.
22
23 U.S. EPA. (1992) Highway vehicle emission estimates. Office of Mobile Sources.
24
25 U.S. EPA. (1995) Highway vehicle emissions estimates. Office of Mobile Sources.
26
27 U.S. EPA. (1998a) National air pollutant emission trends update, 1970-1997. EPA/454/E-98-007.
28
U.S. EPA. (1998b) Locomotive emission standards. Office of Mobile Sources Regulatory Support Document.
U.S. EPA. (1999a) Air quality criteria for paniculate matter. External review draft, October 1999.
32
33 U.S. EPA. (1999b) Fleet characterization data for MOBILE6: development and use of age distributions, average
34 annual mileage accumulation rates and projected vehicle counts for use in MOBILE6. EPA/420/P-99-011.
35
36 U.S. EPA. (1999c) Analysis of the Impacts of control programs on motor vehicle toxic emissions and exposure in
37 urban areas and nationwide: volume I and II. Prepared for EPA by Sierra Research, Inc. and Radian International
38 Corporation/Eastern Research Group, November 30, 1999. Report Nos. EPA/420/R-99/029 and
39 EPA/420/R-99/030.
40
41 U.S. EPA. (2000a) National air pollutant emission trends, 1900-1998. EPA-454/R-00-002, March 2000.
42
43 U.S. EPA. (2000b) Proposed heavy-duty engine and vehicle standards and highway diesel fuel sulfur control
44 requirements. Draft regulatory impact analysis. Office of Transportation and Air Quality. EPA420-D-00-001.
45
46 Unnasch, S; Lowell, D; Lonyai, F; et al. (1993) Performance and emissions of clean fuels in transit buses with
47 Cummins L10 engines. SAE Technical Paper Ser. No. 931782.
48
49 Van Vaeck, L; Van Cauwenberghe, K. (1984) Conversion of polycyclic aromatic hydrocarbons on diesel paniculate
50 matter upon exposure to ppm levels of ozone. Atmos Environ 18:323-328.
51
52 Van Vaeck, L; Van Cauwenberghe, K; Janssens, J. (1984) The gas-particle distribution of organic aerosol
53 constituents: measurements of the volatilisation artifact in hi-vol cascade impactor sampling. Atmos Environ 18:417-
54 430.
Voss, JR; Vanderpoel, RE. (1977) The shuttle distributor for a diesel fuel injection pump. SAE Technical Paper Ser.
No. 770083.
58
7/25/00 2-153 DRAFT—DO NOT CITE OR QUOTE
-------
1 Wachter, WF. (1990) Analysis of transient emission data of a model year 1991 heavy duty diesel engine. SAE
2 Technical Paper Ser. No. 900443.
3
4 Wall, JC. (1998) Diesel fuel composition for future emissions regulations. Cummins Engine Co. panel discussion,
5 SAE International Fall Fuels and Lubricants Meeting and Exposition, October 21, 1998.
6
7 Wall, JC; Shimpi, SA; Yu, ML. (1987) Fuel sulfur reduction for control of diesel paniculate emissions. SAE
8 Technical Paper Ser. No. 872139.
9
10 Wang, W; Gautam, M; Sun, X; et al. (1993) Emissions comparisons of twenty-six HD vehicles operated on
11 conventional and alternative fuels. SAE Technical Paper Ser. No. 932952.
12
13 Wang, WG; Lyons, D; Bata, R; et al. (1994) In-use emissions tests of alternatively fueled HD vehicles by a chassis
14 dynamometer testing facility. SAE Technical Paper Ser. No. 945124.
15
16 Wamer-Selph, MA; Dietzmann, HE. (1984) Characterization of HD motor vehicle emissions under transient driving
17 conditions. Prepared by Southwest Research Institute. EPA/600/3-84/104.
18
19 Watson, N; Janota, MS. (1982) Turbocharging the internal combustion engine. New York: John Wiley and Sons.
20
21 Watson, JG; Fujita, EM; Chow, JC; et al. (1998) Northern Front Range Air Quality Study final report. Prepared by
22 Desert Research Institute for Colorado State University, Cooperative Institute for Research in the Atmosphere, 1998.
23
24 Watts, FW, Jr. (1995) Assessment of occupational exposure to diesel emissions. In: Diesel exhaust: a critical
25 analysis of emissions, exposure, and health effects. Cambridge, MA: Health Effects Institute, pp. 107-123.
26
27 Weaver, CS, Klausmeier, RF. (1988) Heavy-duty diesel vehicle inspection and maintenance study. Final Report
28 Vol. 1: Summary Report to California Air Resources Board, Contract No. A4-151-32.
29
30 Weingartner, E; Burtscher, H; Baltensperger, U. (1997a) Hygroscopic properties of carbon and diesel soot particles. -
31 Atmos. Environ. 31:2311 -2327.
32
33 Weingartner, E; Keller, C; Stahel, WA; et al. (1997b) Aerosol emissions in a road tunnel. Atmos Environ 31(3):451-
34 462.
35
36 Wevers, M; De Fre, R; Rymen, T. (1992) Dioxins and dibenzofurans in tunnel air. In: Sources of exposure.
37 Extended Abstracts, vol. 9. OrganoComp 9:321-324.
38
39 Whitby, KT; Sverdrup, GM. (1980) California aerosols: their physical and chemical characteristics. In: Hidy, GM;
40 Mueleeer, PD; Grosjean, D; et al., eds. The character and origins of smog aerosols: a digest of results from the
41 California Aerosol Characterization Experiment (ACHEX). New York, NY: John Wiley & Sons, Inc., pp. 477-517.
42 (Advances in environmental science and technology: v.9).
43
44 Whitfield, JK; Harris, DB. (1998) Comparison of HD diesel emissions from engine and chassis dynamometers and
45 on-road testing. 8th Coordinating Research Council On-Road Vehicle Emissions Workshop: San Diego, CA.
46
47 Williams. PT. Andrews. GE: Bartle. KD. (1987) The role of lubricating oil in diesel particulate and particulate PAH
48 emissions. SAE Technical Paper Ser. No. 872084.
49
50 Williams, DJ; Milne, JW; Quigley, SM; et al. (1989) Particulate emissions from in-use motor vehicles. II. Diesel
51 vehicles. Atmos Environ 23(12):2647.
52
53 Wittorff, DN; Gertler, AW; Chow, JC; et al. (1994) The impact of diesel particulate emissions on ambient particulate
54 loadings. Air & Waste Management Association 87th Annual Meeting, Cincinnati, OH, June 19-24, 1994.
55
56 ^/incr A.N^' Bust?^7 WF V fiQQ^ A.tmosnheric transnort snd transformation of dicse' emissions ^p** O^c^s'
57 exhaust: a critical analysis of emissions, exposure, and health effects. Cambridge, MA: Health Effects Institute, pp.
58 84-105.
7/25/00 2-154 DRAFT—DO NOT CITE OR QUOTE
-------
Woskie, SR; Smith, TJ; Hammond, SK; et al. (1988) Estimation of the diesel exhaust exposures of railroad workers.
I. Current exposures. Am JIndMed 13:381-394.
4
5 Yamasaki, H; Kuwata, K; Miyamoto, H. (1982) Effects of ambient temperature on aspects of airborne polycyclic
6 aromatic hydrocarbons. Environ Sci Technol 16:189-194.
7
8 Yanowitz, J; Graboski, MS; Ryan, LBA; et al. (1999) Chassis dynamometer study of emissions from 21 in-use HD
9 diesel vehicles. Environ Sci Technol 33:209-216.
10
11 Yanowitz, J; McCormick, RL; Graboski, MS. (2000) Critical review: in-use emissions from HD diesel vehicles.
1 2 Environ Sci Technol. 34:729-740.
13
14 Yokely, RA; Garrison, AA; Wehry, EL; et al. (1986) Photochemical transformation of pyrene and benzo[a]pyrene
15 vapor-deposited on eight coal stack ashes. Environ Sci Technol 20:86-90.
16
17 Zaebst, DD; Blad, LM; Morris, JA; et al. (1988) Elemental carbon as a surrogate index of diesel exhaust exposure.
18 In: Proceedings of the American Industrial Hygiene Conference, May 15-20, 1988, San Francisco, CA.
19
20 Zaebst, DD; Clapp, DE; Blake, LM; et al. (1991) Quantitative determination of trucking industry workers' exposures
21 to diesel exhaust particles. Am Ind Hyg Assoc J 52:529-541.
22
23 Zelenka, P; Kriegler, W; Herzog, PL; et al. (1990) Ways toward the clean HD diesel. SAE Technical Paper Ser. No.
24 900602.
25
26 Zielinska, B. (1999) Changes in diesel engine emissions over the last two decades. Diesel workshop: building a
27 research strategy to improve risk assessment. Health Effects Institute Number 7. March 7-9, 1999, Stone Mountain,
28 GA.
Zielinska, B; Arey, J; Atkinson, R; et al. (1986) Reaction of dinitrogen pentoxide with fluoranthene. J Am Chem Soc
31 108:4126-4132.
32
33 Zielinska, B; Arey, J; Atkinson, R; et al. (1988) Nitration of acephenanthrylene under simulated atmospheric
34 conditions and in solution and the presence of nitroacephenanthrylene(s) in ambient air. Environ Sci Technol
35 22:1044-1048.
36
37 Zielinska, B; Arey, J; Atkinson, R; et al. (1989a) Formation of methylnitronaphthalenes from the gas-phase reactions
38 of 1- and 2-methylnaphthalene with OH radicals and N2O5 and their occurrence in ambient air. Environ Sci Technol
39 23:723-729.
40
41 Zielinska, B; Arey, J; Atkinson, R; et al. (1989b) The nitroarenes of molecular weight 247 in ambient paniculate
42 samples collected in southern California. Atmos Environ 23:223-229.
43
44 Zielinska, B; Arey, J; Atkinson, R. (1990) The atmospheric formation of nitroarenes and their occurrence in ambient
45 air. In: Proceedings of the Fourth International Conference on N-Substituted Aryl Compounds: Occurrence,
46 Metabolism and Biological Impact of Nitroarenes. Cleveland, OH, July 1989.
47
48 Zielinska, B; McDonald, J; Hayes, T; et al. (1998) Northern Front Range Air Quality Study, volume B: source
49 measurements. Desert Research Institute.
50
51 Zwirner-Baier, I; Neumann, H-G. (1999) Polycyclic nitroaarenes (nitro-PAHs) as biomarkers of exposure to diesel
52 exhaust. Mut Res 441:135-144.
7/25/00 2-155 DRAFT—DO NOT CITE OR QUOTE
-------
3. DOSIMETRY OF DIESEL PARTICULATE MATTER
1 3.1. INTRODUCTION
2 Clearly, animals and humans receive different internal doses when breathing the same
3 external concentrations of airborne materials such as diesel paniculate matter (DPM) (Brain and
4 Mensah, 1983; Schlesinger, 1985). The dose received in different species differs from the
5 aspects of the total amount deposited within the respiratory tract, the relative distribution of the
6 dose to specific regions in the respiratory tract, and the residence time of these materials within
7 the respiratory tract, i.e., clearance. Using an external concentration breathed by laboratory
8 animals as a basis for any guidance for human exposure to DPM would then be an inadequate
9 approximation of the total and regional dose that humans may receive. The objective of this
10 chapter is to evaluate and address this issue of interspecies dosimetric differences through:
11
12 A general overview of what is known about how particles like DPM are deposited,
13 transported to, and cleared from the respiratory tract. Information on both
14 laboratory animals (mainly rodents) and humans will be considered and interspecies
15 similarities and differences highlighted.
• An overview of what is known about the bioavailability of the organic compounds
adsorbed onto DPM from information in humans, animals, and in vitro studies, and
18 from model predictions.
19 • An evaluation of the suitability of available dosimetric models and procedures for
20 DPM to perform interspecies extrapolations whereby an exposure scenario,
21 conditions, and outcome in laboratory animals are adjusted to an equivalent
22 outcome in humans via calculation of an internal dose.
23
24 The focus in this chapter will be on the paniculate fraction of diesel emissions, i.e, DPM.
25 Although diesel engine exhaust consists of a complex mixture of typical combustion gases,
26 vapors, low-molecular-weight hydrocarbons, and panicles, it is the particle phase that is
27 considered to be of major health concern. The major constituents of diesel exhaust and their
28 atmospheric reaction products are described here (Chapter 2).
29 As will be deduced in Chapter 5, pulmonary toxiciry and carcinogenicity is the major
30 focal point of diesel toxicity and of DPM deposition. Therefore, dosimetric considerations are
31 limited to the lung. Aspects of respiratory tract dosimetry to be considered in this chapter
32 include the characteristics of DPM, deposition of DPM in the conducting airways and alveolar
regions, normal DPM clearance mechanisms and rates of clearance in both these regions,
7/25/00 3-1 DRAFT—DO NOT CITE OR QUOTE
-------
1 clearance rates during lung overload (in rats), elution of organics from DPM, transport of DPM
2 to extra-alveolar sites, and the interrelationships of these factors.
3 The overall goal in this chapter follows from the objective—to judge the feasibility and
4 suitability of procedures allowing for derivation of an internal dose estimate of DPM for humans,
5 i.e., of a human equivalent concentration to exposure concentrations and conditions used in
6 animal studies. This goal is of significance especially in the quantitative dose-response analysis
7 of DPM effects proposed in Chapter 6.
8
9 3.2. CHARACTERISTICS OF INHALED DPM
10 The formation, transport, and characteristics of DPM are considered in detail in Chapter
11 2. DPM consists of aggregates of spherical carbonaceous particles (typically about 0.2 um mass
12 median aerodynamic diameter [MMAD]) to which significant amounts of higher-molecular-
13 weight organic compounds are adsorbed. DPM has an extremely large surface area that allows
14 for the adsorption of organic compounds. The organic carbon portion of DPM can range from at
15 least 19% to 43% from highway diesel engines; no data are available to characterize the organic
16 content of DPM from nonroad engines. The lexicologically relevant organic chemicals include
17 high-molecular-weight hydrocarbons such as the polycyclic aromatic hydrocarbons (PAHs) and
18 their derivatives (Section 2.2.8).
19
20 3.3. REGIONAL DEPOSITION OF INHALED DPM
21 This section discusses the major factors controlling the disposition of inhaled particles.
22 Note that disposition is defined as encompassing the processes of deposition, absorption,
23 distribution, metabolism, and elimination. The regional deposition of paniculate matter in the
24 respiratory tract is dependent on the interaction of a number of factors, including respiratory tract
25 anatomy (airway dimensions and branching configurations), ventilatory characteristics (breathing
26 mode and rate, ventilatory volumes and capacities), physical processes (diffusion, sedimentation,
27 impaction, and interception), and the physicochemical characteristics (particle size, shape,
28 density, and electrostatic attraction) of the inhaled particles. Regional deposition of paniculate
29 material is usually expressed as deposition fraction of the total particles or mass inhaled and may
30 be represented by the ratio of the particles or mass deposited in a specific region to the number or
31 mass of particles inspired. The factors affecting deposition in these various regions and their
32 importance in understanding the fate of inhaled DPM are discussed in the following sections.
33 It is beyond the scope of this document to present a comprehensive account of the
34 complexities of respiratory mechanics; physiology, and toxicology, and only a brief review will
3R hp nrpspntprl here The reader is referred tn mihlieations that rvrovirle a more in-depth treatment
7/25/00 3-2 DRAFT—DO NOT CITE OR QUOTE
-------
of these topics (Weibel, 1963; Brain and Mensah, 1983; Raabe et al., 1988; Stober et al., 1993;
U.S. EPA, 1996).
3 The respiratory tract in both humans and experimental mammals can be divided into three
4 regions on the basis of structure, size, and function (International Commission on Radiological
5 Protection, 1994): the extrathoracic (ET), the tracheobronchial (TB), and the alveolar (A). In
6 humans, inhalation can occur through the nose or mouth or both (oronasal breathing). Many
7 animal models used in respiratory toxicology studies are, however, obligate nose breathers.
8
9 3.3.1. Deposition Mechanisms
10 This section provides an overview of the basic mechanisms by which inhaled particles
11 deposit within the respiratory tract. Details concerning the aerosol physics that explain both how
12 and why particle deposition occurs as well as data on total human respiratory tract deposition are
13 presented in detail in the earlier PM Criteria Document (U.S. EPA, 1996) and will only be briefly
14 summarized here. For more extensive discussions of deposition processes, refer to reviews by
15 Morrow (1966), Raabe (1982), U.S. EPA (1982), Phalen and Oldham (1983), Lippmann and
16 Schlesinger (1984), Raabe et al. (1988), and Stober et al. (1993).
17 Particles may deposit by five major mechanisms (inertial impaction, gravitational settling,
Brownian diffusion, electrostatic attraction, and interception). The relative contribution of each
deposition mechanism to the fraction of inhaled particles deposited varies for each region of the
20 respiratory tract.
21 It is important to appreciate that these processes are not necessarily independent but may,
22 in some instances, interact with one another such that total deposition in the respiratory tract may
23 be less than the calculated probabilities for deposition by the individual processes (Raabe, 1982).
24 Depending on the particle size and mass, varying degrees of deposition may occur in the
25 extrathoracic or ET (or nasopharyngeal), tracheobronchial (TB), and alveolar regions of the
26 respiratory tract.
27 Upon inhalation of particulate matter such as that found in diesel exhaust, particle
28 deposition will occur throughout the respiratory tract. Because of high airflow velocities and
29 abrupt directional changes in the ET and TB regions, inertial impaction is a primary deposition
30 mechanism, especially for particles >2.5 urn dae (aerodynamic equivalent diameter). Although
31 inertial impaction is a prominent process for deposition of larger particles in the tracheobronchial
32 region, it is of minimal significance as a determinant of regional deposition patterns for DPM,
33 which have a dae < 1 urn.
34 All aerosol particles are continuously influenced by gravity, but particles with a
^B dae > 0.5 ^m are affected to the greatest extent. A spherical compact particle will acquire a
36 terminal settling velocity when a balance is achieved between the acceleration of gravity acting
7/25/00 3-3 DRAFT—DO NOT CITE OR QUOTE
-------
1 on the particle and the viscous resistance of the air; it is this velocity that brings the particle into
2 contact with airway surfaces. Both sedimentation and inertia! impaction cause the deposition of
3 many particles within the same size range. These deposition processes act together in the ET and
4 TB regions, with inertial impaction dominating in the upper airways and sedimentation becoming
5 increasingly dominant in the lower conducting airways, especially for the largest particles that
6 can penetrate into the smaller bronchial airways.
7 As particle diameters become <1 um, the particles are increasingly subjected to diffusive
8 deposition because of random bombardment by air molecules, which results in contact with
9 airway surfaces. A dae of 0.5 um is often considered a boundary between diffusion and
10 aerodynamic (sedimentation and impaction) mechanisms of deposition. Thus, instead of having
11 a dae, diffusive particles of different shapes can be related to the diffusivity of a thermodynamic
12 equivalent size based on spherical particles (Heyder et al., 1986). Diffusive deposition of
13 particles is favored in the A region of the respiratory tract as particles of this size are likely to
14 penetrate past the ET and TB regions.
15 Electrostatic precipitation is deposition related to particle charge. The electrical charge
16 on some particles may result in an enhanced deposition over what would be expected from size
17 alone. This is due to image charges induced on the surface of the airway by these particles, or to
18 space-charge effects whereby repulsion of particles containing like charges results in increased
19 migration toward the airway wall. The effect of charge on deposition is inversely proportional to
20 particle size and airflow rate. A recent study employing hollow airway casts of the human
21 tracheobronchial tree that assessed deposition of ultrafine (0.02 um) and fine (0.125 um)
22 particles found that deposition of singly charged particles was 5-6 times that of particles having
23 no charge, and 2-3 times that of particles at Boltzmann equilibrium (Cohen et al., 1998). This
24 suggests that within the TB region of humans, electrostatic precipitation may be a significant
25 deposition mechanism for ultrafine and some fine particles, the latter of which are inclusive of
26 DPM. Thus, although electrostatic precipitation is generally a minor contributor to overall
27 particle deposition, it may be important for DPM.
28 Interception is deposition by physical contact with airway surfaces and is most important
29 for fiber deposition (U.S. EPA, 1996).
30
31 3.3.1.1. Biological Factors Modifying Deposition
32 The available experimental deposition data in humans are commonly derived using
33 healthy adult Caucasian males. Various factors can act to alter deposition patterns from those
34 obtained in this group. The effects of different biological factors, including gender, age, 2nd
3R reTMratorv tract disea^? <"*r! riarticle dcsiticp have been reviewed r»«M/ir>"...»_- 1. . -* .-.- ^.~.VB.1UM-M*yy t^ . fc_* . 4—fJ. i J •* ' *"' J
7/25/00 3-4 DRAFT—DO NOT CITE OR QUOTE
-------
Section 10.4.1.6). In general, there appears to be an inverse relationship between airway
resistance and total deposition.
3 The various species that serve as the basis for dose-response assessment in inhalation
4 toxicology studies do not receive identical doses in a comparable respiratory tract region (ET,
5 TB, or A) when exposed to the same aerosol or gas (Brain and Mensah, 1983). Such interspecies
6 differences are important because the adverse toxic effect is likely more related to the
7 quantitative pattern of deposition within the respiratory tract than to the exposure concentration;
8 this pattern determines not only the initial respiratory tract tissue dose but also the specific
9 pathways by which the inhaled material is cleared and redistributed (Schlesinger, 1985).
10 Differences in patterns of deposition between humans and animals have been summarized (U.S.
11 EPA, 1996; Schlesinger et al., 1997). Such differences in initial deposition must be considered
12 when relating biological responses obtained in laboratory animal studies to effects in humans.
13 The deposition of inhaled diesel particles in the respiratory tract of humans and
14 mammalian species has been reviewed (Health Effects Institute, 1995). Schlesinger (1985)
15 showed that physiological differences in the breathing mode for humans (nasal or oronasal
16 breathers) and laboratory rats (obligatory nose breathers), combined with different airway
17 geometries, resulted in significant differences in lower respiratory tract deposition for larger
particles (>1 ^m dae). In particular, a much lower fraction of inhaled larger particles is deposited
in the alveolar region of the rat compared with humans. However, relative deposition of the
20 much smaller diesel exhaust particles was not affected as much by the differences among species,
21 as was demonstrated in model calculations by Xu and Yu (1987). These investigators modeled
22 the deposition efficiency of inhaled DPM in rats, hamsters, and humans on the basis of
23 calculations of the models of Schum and Yeh (1980) and Weibel (1963). These simulations
24 (Figure 3-1) indicate relative deposition patterns in the lower respiratory tract (trachea =
25 generation 1; alveoli = generation 23) and are similar among hamsters, rats, and humans.
26 Variations in alveolar deposition of DPM over one breathing cycle in these different species were
27 predicted to be within 30% of one another. Xu and Yu (1987) attributed this similarity to the fact
28 that deposition of the submicron diesel particles is dominated by diffusion rather than
29 sedimentation or impaction. Although these data assumed nose-breathing by humans, the results
30 would not be very different for mouth-breathing because of the low filtering capacity of the nose
31 for particles in the 0.1 to 0.5 urn range.
32 For dosimetric calculations and modeling, it would be of much greater importance to
33 consider the actual dose deposited per unit surface area of the respiratory tract rather than the
34 relative deposition efficiencies per lung region. Table 3-1 compares the predicted deposited
^P doses of DPM inhaled in 1 min for the three species, based on the total lung volume, the surface
36 area of all lung airways, or the surface area of the epithelium of the alveolar region only. In
7/25/00 3-5 DRAFT—DO NOT CITE OR QUOTE
-------
1 Table 3-1, the deposited dose, expressed as either mass/lung volume (M) or mass/surface area(s)
2 (M,), or mass/alveolar surface area (M2) is lower in humans than in the two rodent species as a
3 result of the greater respiratory exchange rate in rodents and smaller size of the rodent lung.
4 Such differences in the deposited dose in relevant target areas are important and have to be
5 considered when extrapolating the results from DPM/DE exposure studies in animals to humans.
6 Table 3-1 indicates that the differences (between humans to animals) are less on a surface area
7 basis (=3-fold) than on a lung volume basis (~ 14-fold). This is due to larger alveolar diameters
8 and concomitant lower surface area per unit of lung volume in humans.
9 Particle deposition will initiate particle redistribution processes (e.g., clearance
10 mechanisms, phagocytosis) that transfer the particles to various subcompartments, including the
11 alveolar macrophage pool, pulmonary interstitium, and lymph nodes. Over time, therefore, only
12 small amounts of the original particle intake would be associated with the alveolar surface.
13
14 3.3.2. Particle Clearance and Translocation Mechanisms
15 This section provides an overview of the mechanisms and pathways by which particles
16 are cleared from the respiratory tract. The mechanisms of particle clearance as well as clearance
17 routes from the various regions of the respiratory tract have been considered in the PM Criteria
18 Document (U.S. EPA, 1996) and reviewed by Schlesinger et al. (1997).
19 Particles that deposit upon airway surfaces may be cleared from the respiratory tract
20 completely, or be translocated to other sites within this system, by various regionally distinct
21 processes. These clearance mechanisms can be categorized as either absorptive (i.e., dissolution)
22 or nonabsorptive (i.e., transport of intact particles) and may occur simultaneously or with
23 temporal variations. Particle solubility in terms of clearance refers to solubility within the
24 respiratory tract fluids and cells. Thus, a poorly soluble particle is one whose rate of clearance by
25 dissolution is insignificant compared to its rate of clearance as an intact particle (as is the case
26 with DPM). The same clearance mechanisms act on specific particles to different degrees, with
27 their ultimate fate being a function of deposition site, physicochemical properties (including any
28 toxicity), and sometimes deposited mass or number concentration.
29
30 3.3.2.1. Extrathoracic Region
31 The clearance of poorly soluble particles deposited in the nasal passages occurs via
32 mucociiiary transport, and the general flow of mucus is backwards, i.e., towards the nasopharynx.
33 Mucus flow in the most anterior portion of the nasal passages is forward, clearing deposited
34 particle^ tn the vestihular region where removal is by sneezing, wiping, or blowing.
35 Soluble material deposited on the nasal epithelium is accessible to underlying cells via
36 diffusion through the mucus. Dissolved substances may be subsequently translocated into the
7/25/00 3-6 DRAFT—DO NOT CITE OR QUOTE
-------
bloodstream. The nasal passages have a rich vasculature, and uptake into the blood from this
region may occur rapidly.
3 Clearance of poorly soluble particles deposited in the oral passages is by coughing and
4 expectoration or by swallowing into the gastrointestinal tract.
5
6 3.3.2.2. Tracheobronchial Region
7 The dynamic relationship between deposition and clearance is responsible for
8 determining lung burden at any point in time. Clearance of poorly soluble particles from the TB
9 region is mediated primarily by mucociliary transport, a more rapid process than those operating
10 in alveolar regions. Mucociliary transport (often referred to as the mucociliary escalator) is
11 accomplished by the rhythmic beating of cilia that line the respiratory tract from the trachea
12 through the terminal bronchioles. This movement propels the mucous layer containing deposited
13 particles (or particles within alveolar macrophages [AMs]) toward the larynx. Clearance rate by
14 this system is determined primarily by the flow velocity of the mucus, which is greater in the
15 proximal airways and decreases distally. These rates also exhibit interspecies and individual
16 variability. Considerable species-dependent variability in tracheobronchial clearance has been
17 reported, with dogs generally having faster clearance rates than guinea pigs, rats, or rabbits
(Felicetti et al., 1981). The half-time (t,/2) values for tracheobronchial clearance of relatively
insoluble particles are usually on the order of hours, as compared to alveolar clearance, which is
20 on the order of hundreds of days in humans and dogs. The clearance of particulate matter from
21 the tracheobronchial region is generally recognized as being biphasic or multiphasic (Raabe,
22 1982). Some studies have shown that particles are cleared from large, intermediate, and small
23 airways with t,/2 of 0.5, 2.5, and 5 h, respectively. However, reports have indicated that clearance
24 from airways is biphasic and that the long-term component for humans may take much longer for
25 a significant fraction of particles deposited in this region, and may not be complete within 24 h as
26 generally believed (Stahlhofen et al., 1990; ICRP, 1994).
27 Although most of the particulate matter will be cleared from the tracheobronchial region
28 towards the larynx and ultimately swallowed, the contribution of this fraction relative to
29 carcinogenic potential is unclear. With the exception of conditions of impaired bronchial
30 clearance, the desorption t,/2 for particle-associated organics is generally longer than the
31 tracheobronchial clearance times, thereby making uncertain the importance of this fraction
32 relative to toxicity in the respiratory tract (Pepelko, 1987). However, Gerde et al. (199la)
33 showed that for low-dose exposures, particle-associated PAHs were released rapidly at the site of
^4 deposition. The relationship between the early clearance of poorly soluble particles of 4 jam
aerodynamic diameter from the tracheobronchial regions and their longer-term clearance from
36 the alveolar region is illustrated in Figure 3-2.
7/25/00 3-7 DRAFT—DO NOT CITE OR QUOTE
-------
1 Cuddihy and Yeh (1986) reviewed respiratory tract clearance of particles inhaled by
2 humans. Depending on the type of particle (ferric oxide, Teflon discs, or albumin microspheres),
3 the technique employed, and the anatomic region (midtrachea, trachea, or main bronchi), particle
4 velocity (moved by mucociliary transport) ranged from 2.4 to 21.5 mm/min. The highest
5 velocities were recorded for midtracheal transport, and the lowest were for main bronchi. In one
6 study, an age difference was noted for tracheal mucociliary transport velocity (5.8 mm/min for
7 individuals less than 30 years of age and 10.1 mm/min for individuals over 55 years of age).
8 Cuddihy and Yeh (1986) described salient points to be considered when estimating
9 particle clearance velocities from tracheobronchial regions: these include respiratory tract airway
10 dimensions, calculated inhaled particle deposition fractions for individual airways, and thoracic
11 (A + TB) clearance measurements. Predicted clearance velocities for the trachea and main
12 bronchi were found to be similar to those experimentally determined for inhaled radiolabeled
13 particles, but not those for intratracheally instilled particles. The velocities observed for
14 inhalation studies were generally lower than those of instillation studies. Figure 3-3 illustrates a
15 comparison of the short-term clearance of inhaled particles by human subjects and the model
16 predictions for this clearance. However, tracheobronchial clearance via the mucociliary escalator
17 is of limited importance for long-term clearance.
18 Exposure of F344 rats to whole DPM at concentrations of 0.35, 3.5, or 7.1 mg/m3 for up
19 to 24 mo did not significantly alter tracheal mucociliary clearance as assessed by clearance of
20 99mTc-macroaggregated albumin instilled into the trachea (Wolff et al., 1987). The authors stated
21 that measuring retention would yield estimates of clearance efficiency comparable to measuring
22 the velocity for transport of the markers in the trachea. The results of this study were in
23 agreement with similar findings of unaltered tracheal mucociliary clearance in rats exposed to
24 DPM (0.21, 1.0, or 4.4 mg/m3) for up to 4 mo (Wolff and Gray, 1980). However, the 1980 study
25 by Wolff and Gray, as well as an earlier study by Battigelli et al. (1966), showed that acute
26 exposure to high concentrations of diesel exhaust soot (1.0 and 4.4 mg/m3 in the study by Wolff
27 and Gray [1980] and 8 to 17 mg/m3 in the study by Battigelli et al. [1966]) produced transient
28 reductions in tracheal mucociliary clearance. Battigelli et al. (1966) also noted that the
29 compromised tracheal clearance was not observed following cessation of exhaust exposure.
30 That tracheal clearance does not appear to be significantly impaired or is impaired only
31 transiently following exposure to high concentrations of DPM is consistent with the absence of
32 pathological effects in the tracheobronchial region of the respiratory tract in experimental
33 animals exposed to DPM. The apparent retention of a fraction of the deposited dose in
34 the airways could be cause for some concern regarding possible effects in tills region, especially
35 in light of the results from simulation studio Ky Gerde et al. (1991b) suggesting mat release of
36 PAHs from particles may occur within minutes and therefore at the site of initial deposition.
7/25/00 3-8 DRAFT—DO NOT CITE OR QUOTE
-------
1 However, the absence of effects in the TB areas in long-term DPM studies and experimental
2 evidence that particle-associated PAHs are released at the site of particle deposition together
3 suggest that these PAHs and other organics may be of lesser importance in tumorigenic responses
4 of rats than originally suspected. On the other hand, however, a larger fraction of particles are
5 translocated to the interstitium of the respiratory tract in primates (and therefore presumably in
6 humans) than in rats, including the interstitium of the respiratory bronchioles, an anatomical site
7 absent in rats (Section 3.6) (Nikula et al., 1997a,b). Moreover, eluted PAHs in the TB region are
8 retained longer than those in the alveoli (Gerde et al., 1999), allowing time for activation. Thus
9 PAHs may have a role in human response to diesel exhaust that cannot be evaluated with the rat
10 model.
11 Also, impairment of mucociliary clearance function as a result of exposure to
12 occupational or environmental respiratory tract toxicants or to cigarette smoke may significantly
13 enhance the retention of particles in the TB region. For example, Vastag et al. (1986)
14 demonstrated that not only smokers with clinical symptoms of bronchitis but also symptom-free
15 smokers have significantly reduced mucociliary clearance rates. Although impaired
16 tracheobronchial clearance could conceivably have an impact on the effects of deposited DPM in
17 the conducting airways, it does not appear to be relevant to the epigenetic mechanism likely
responsible for diesel exhaust-induced rat pulmonary tumors.
T9 Poorly soluble particles such as DPM that are deposited within the TB region are cleared
20 predominantly by mucociliary transport towards the oropharynx, followed by swallowing.
21 Poorly soluble particles may also be cleared by traversing the epithelium by endocytotic
22 processes, and enter the peribronchial region. Clearance may occur following phagocytosis by
23 airway macrophages, located on or beneath the mucous lining throughout the bronchial tree, or
24 via macrophages that enter the airway lumen from the bronchial or bronchiolar mucosa
25 (Robertson, 1980).
26
27 3.3.2.3. A Region
28 A number of investigators have reported on the alveolar clearance kinetics of human
29 subjects. Bohning et al. (1980) examined alveolar clearance in eight humans who had inhaled
30 <0.4 mg of 85Sr-labeled polystyrene particles (3.6 ±1.6 um diam.). A double-exponential model
31 best described the clearance of the particles and provided t1/2 values of 29 ± 19 days and 298 ±
32 114 days for short-term and long-term phases, respectively. It was noted that of the particles
33 deposited in the alveolar region, 75% ± 13% were cleared via the long-term phase. Alveolar
34 retention t,/2 values of 330 and 420 days were reported for humans who had inhaled
01 aluminosilicate particles of MMAD 1.9 and 6.1 um (Bailey et al., 1982). In a comprehensive
36 study Bailey et al. (1985) followed the long-term retention of inhaled particles in a human
7/25/00 3-9 DRAFT—DO NOT CITE OR QUOTE
-------
1 respiratory tract. The retention of 1 and 4 ^m fused aluminosilicate particles labeled with
2 strontium-85 and yttrium-88, respectively, was followed in male volunteers for about 533 days.
3 Approximately 7% of the initial lung deposit of 1 um particles and 40% of the 4 urn particles
4 were associated with a rapid clearance phase corresponding to the calculated tracheobronchial
5 deposits. Retention of the remaining material followed a two-component exponential function,
6 with phases having half-times of the order of tens of days and several hundred days, respectively.
7 Quantitative data on clearance rates in humans having large lung burdens of particulate
8 matter are lacking. Bohning et al. (1982) and Cohen et al. (1979), however, did provide evidence
9 for slower clearance in smokers, and Freedman and Robinson (1988) reported slower clearance
10 rates in coal miners who had mild pneumoconiosis with presumably high lung burdens of coal
11 dust. Although information on particle burden and particle overload relationships in humans is
12 much more limited than in experimental animal models, inhibition of clearance does seem to
13 occur. Stober et al. (1967) estimated a clearance t,/2 of 4.9 years in coal miners with nil or slight
14 silicosis, based on postmortem lung burdens. The lung burdens and estimated exposure histories
15 ranged from 2 to 50 mg/g of lung or more, well above the value at which clearance impairment is
16 observed in the rat. Furthermore, impaired clearance resulting from smoking or exposure to
17 other respiratory toxicants may increase the possibility of an enhanced particle accumulation
18 effect resulting from exposure to other particle sources such as DPM.
19 Normal alveolar clearance rates in laboratory animals exposed to DPM have been
20 reported by a number of investigators (Table 3-2). Because the rat is, historically, the species for
21 which experimentally induced lung cancer data are available and for which most clearance data
22 exist, it is the species most often used for assessing human risk, and reviews of alveolar clearance
23 studies have been generally limited to this species.
24 Chan et al. (1981) subjected 24 male F344 rats to nose-only inhalation of DPM (6 mg/m3)
25 labeled with 13lBa or 14C for 40 to 45 min and assessed total lung deposition, retention, and
26 elimination. Based on radiolabel inventory, the deposition efficiency in the respiratory tract was
27 15% to 17%. Measurement of 131Ba label in the feces during the first 4 days following exposure
28 indicated that 40% of the deposited DPM was eliminated via mucociliary clearance. Clearance
29 of the particles from the lower respiratory tract followed a two-phase elimination process
30 consisting of a rapid (t/2 of 1 day) elimination by mucociliary transport and a slower (ti/2 of
31 62 days) macrophage-mediated alveolar clearance. This study provided data for normal alveolar
32 clearance rates of DPM not affected by prolonged exposure or particle overloading.
33 Several studies have investigated the effects of exposure concentration on the alveolar
34 clearance of DPM by laboratory animals. Wolff et a). (1986, 1987) provided clearance data (t;/)
35 and lun? burden values for F344 rats evpn-spH to Hi^?pl exh?.ust for 7 h/day, 5 days/week for 2/!
36 mo. Exposure concentrations of 0.35, 3.5, and 7.1 mg of DPM/m3 were employed in this whole
7/25/00 3-10 DRAFT—DO NOT CITE OR QUOTE
-------
body-inhalation exposure experiment. Intermediate (hours-days) clearance of "Ga^ particles
(30 min, nose-only inhalation) was assessed after 6, 12,18, and 24 mo of exposure at all of the
3 DPM concentrations. A two-component function described the clearance of the administered
4 radiolabel:
5
6 ' Fm = ^exp(-0.693 t/r,)+ 5exp(-0.693t/r2), (3-1)
7
8 where F(t) was the percentage retained throughout the respiratory tract, A and B were the
9 magnitudes of the two components (component A included nasal, lung, and gastrointestinal
10 clearance, while component B represented intermediate lung clearance) and T, and T2 were the
11 half-times for the A and B components, respectively. The early clearance half-times (T,), were
12 similar for rats in all exposure groups at all time points except hi the high-exposure (7.1 mg/m3)
13 group following 24 mo of exposure, which was faster than the controls. Significantly longer B
14 component retention half-times, representing intermediate clearance probably from nonciliated
15 structures such as alveolar ducts and alveoli, were noted after as little as 6 mo exposure to DPM
16 at 7.1 mg/m3 and 18 mo exposure to 3.5 mg/m3.
17 Nose-only exposures to |j4Cs fused aluminosilicate particles (FAP) were used to assess
^fc long-term (weeks-months) clearance. Following 24-mo exposure to DPM, long-term clearance
19 of I34Cs-FAP was significantly (pO.Ol) altered in the 3.5 (cumulative exposure [C x T] of
20 11,760 mg-h/m3) and 7.1 mg/m3 C x T = 23,520 mg-h/m3) exposure groups (t,/2 of 264 and 240
21 days, respectively) relative to the 0.35 mg/m3 and control groups (ti/2 of 81 and 79 days,
22 respectively). Long-term clearance represents the slow component of particle removal from the
23 alveoli. The decreased clearance correlated with the greater particle burden in the lungs of the
24 3.5 and 7.1 mg/m3 exposure groups. Based on these findings, the cumulative exposure of
25 > 11,760 mg-h/m3 (or 3.5 mg/m3 for a lifetime exposure) represented a particle overload
26 condition resulting in compromised alveolar clearance mechanisms; the clearance rate at the
27 lowest concentration (0.35 mg/m3; cumulative exposure of 118 mg-h/m3) was not different from
28 control rates (Figure 3-4).
29 Heinrich et al. (1986) exposed rats 19 h/day, 5 days/week for 2.5 years to DPM at a
30 particle concentration of about 4 mg/m3, equal to a C x T of 53,200 mg-h/m3. The deposition in
31 the alveolar region was estimated to equal 60 mg. The lung particle burden was apparently
32 sufficient to result in a "particle overload" condition (Section 3.4). With respect to the organic
33 matter adsorbed onto the particles, the authors estimated that over the 2.5-year period, 6-15 mg of
particle-bound organic matter had been deposited and was potentially available for biological
7/25/OD 3-11 DRAFT—DO NOT CITE OR QUOTE
-------
1 effects. This estimation was based on the analysis of the diesel exhaust used in the experiments,
2 values for rat ventilatory functions, and estimates of deposition and clearance.
3 Accumulated burden of DPM in the lungs following an 18-mo, 7 h/day, 5 days/week
4 exposure to diesel exhaust was reported by Griffis et al. (1983). Male and female F344 rats
5 exposed to 0.15, 0.94, or 4.1 mg DPM/m3 were sacrificed at 1 day and 1, 5, 15, 33, and 52 weeks
6 after exposure, and DPM was extracted from lung tissue dissolved in tetramethylammonium
7 hydroxide. Following centrifugation and washing of the supernatant, DPM content of the tissue
8 was quantitated using spectrophotometric techniques. The analytical procedure was verified by
9 comparing results to recovery studies using known amounts of DPM with lungs of unexposed
10 rats. Lung burdens were 0.035, 0.220, and 1.890 mg/g lung tissue, respectively, in rats exposed
11 to 0.15, 0.94, and 4.1 mg DPM/m3. Long-term retention for the 0.15 and 0.94 mg/m3 groups had
12 estimated half-times of 87 ± 28 and 99 ± 8 days, respectively. The retention t,/2 for the
13 4.1-mg/m3 exposure group was 165 ± 8 days, which was significantly (pO.OOOl) greater than
14 those of the lower exposure groups. The 18-mo exposures to 0.15 or 0.96 mg/m3 levels of DPM
15 C x T equivalent of 378 and 2,368 mg'h/m3, respectively) did not affect clearance rates, whereas
16 the exposure to the 4.1 mg/m3 concentration C x T = 10,332 mg-h/m3) resulted in impaired
17 clearance.
18 Lee et al. (1983) described the clearance of DPM (7 mg/m3 for 45 min or 2 mg/m3 for 140
19 min) by F344 rats (24 per group) and Hartley guinea pigs exposed by nose-only inhalation with
20 no apparent particle overload in the lungs as being in three distinct phases. The exposure
21 protocols provided comparable total doses based on a 14C radiolabel. I4CO2 resulting from
22 combustion of 14C-labeled diesel fuel was removed by a diffusion scrubber to avoid erroneous
23 assessment of 14C intake by the animals. Retention of the radiolabeled particles was determined
24 up to 335 days after exposure and resulted in a three-phase clearance with retention tI/2 values of
25 1,6, and 80 days. The three clearance phases are taken to represent removal of tracheobronchial
26 deposits by the mucociliary escalator, removal of particles deposited in the respiratory
27 bronchioles, and alveolar clearance, respectively. Species variability hi clearance of DPM was
28 also demonstrated because the Hartley guinea pigs exhibited negligible alveolar clearance from
29 day 10 to day 432 following a 45-mtn exposure to a DPM concentration of 7 nig/m3. Initial
30 deposition efficiency (20% ± 2%) and short-term clearance were, however, similar to those for
31 rats.
32 Lung clearance in male F344 rats preexposed to DPM at 0.25 or 6 mg/m3 20 h/day,
33 7 days/week for periods lasting from 7 to 112 days was studied by Chan et al. (1984). Following
34 this nreexposure protocol, rats were subjected to 45-rnin uusc-oniy exposure to :4C-L)ii, and
35 alveolar H^arorice of radiclsbe! %vas monitored for Uf> io I year. Two models were proposed: a
36 normal biphasic clearance model and a modified lung retention model that included a slow-
7/25/00 3-12 DRAFT—DO NOT CITE OR QUOTE
-------
clearing residual component to account for sequestered aggregates of macrophages. The first
model described a first-order clearance for two compartments: R(t) = Ae"ult + Be""2'. This yielded
3 clearance t1/2 values of 166 and 562 days for rats preexposed to 6.0 mg/m3 for 7 and 62 days,
4 respectively. These values were significantly (p<0.05) greater than the retention t1/2 of 77 ± 17
5 days for control rats. The same retention values for rats of the 0.25 mg/m3 groups were 90 ± 14
6 and 92 ± 15 days, respectively, for 52- and 112-day exposures and were not significantly
7 different from controls. The two-compartment model represents overall clearance of the tracer
8 particles, even if some of the particles were sequestered in particle-laden macrophages with
9 substantially slower clearance rates. For the second model, which excluded transport of the
10 residual fractions in sequestered macrophage aggregates, slower clearance was observed in the
11 group with a lung burden of 6.5 mg (exposed to 6.0 mg/m3 for 62 days), and no clearance was
12 observed in the 11.8 mg group (exposed to 6.0 mg/m3 for 112 days). Clearance was shown to be
13 dependent on the initial burden of particles, and therefore the clearance t1/2 would increase in
14 higher exposure scenarios. This study emphasizes the importance of particle overloading of the
15 lung and the ramifications on clearance of particles; the significant increases in half-times
16 indicate an increasing impairment of the alveolar macrophage mobility and subsequent transition
17 into an overload condition as is discussed further in Section 3.4.
§ Long-term alveolar clearance rates of particles in various laboratory animals and humans
have been reviewed by Pepelko (1987). Although retention t,/2 varies both among and within
20 species and is also dependent on the physicochemical properties of the inhaled particles, the
21 retention t,/2 for humans is much longer (>8 mo) than the average retention t1/2 of 60 days for rats.
22 Clearance from the A region occurs via a number of mechanisms and pathways, but the
23 relative importance of each is not always certain and may vary between species. Particle removal
24 by macrophages comprises the main nonabsorptive clearance process in this region. Alveolar
25 macrophages reside on the epithelium, where they phagocytize and transport deposited material,
26 which they contact by random motion or via directed migration under the influence of local
27 chemotactic factors (Warheit et al., 1988).
28 Particle-laden macrophages may be cleared from the A region along a number of
29 pathways (U.S. EPA, 1996). Uningested particles or macrophages in the interstitium may
30 traverse the alveolar-capillary endothelium, directly entering the blood (Raabe, 1982; Holt,
31 1981); endocytosis by endothelial cells followed by exocytosis into the vessel lumen seems,
32 however, to be restricted to particles <0.1 urn diameter, and may increase with increasing lung
33 burden (Lee et al., 1985; Oberdorster, 1988). Once in the systemic circulation, transmigrated
34 macrophages, as well as uningested particles, can travel to extrapulmonary organs.
^P Alveolar macrophages constitute an important first-line cellular defense mechanism
36 against inhaled particles that deposit in the alveolar region of the lung. It is well established that
7/25/00 3-13 DRAFT—DO NOT CITE OR QUOTE
-------
1 a host of diverse materials, including DPM, are phagocytized by AMs shortly after deposition
2 (White and Garg, 1981; Lehnert and Morrow, 1985) and that such cell-contained particles are
3 generally rapidly sequestered from both the extracellular fluid lining in the alveolar region and
4 the potentially sensitive alveolar epithelial cells. In addition to this role in compartmentalizing
5 particles from other lung constituents, AMs are prominently involved in mediating the clearance
6 of relatively insoluble particles from the air spaces (Lehnert and Morrow, 1985). Although the
7 details of the actual process have not been delineated, AMs with their particle burdens gain
8 access and become coupled to the mucociliary escalator and are subsequently transported from
9 the lung via the conducting airways. Although circumstantial, numerous lines of evidence
10 indicate that such AM-mediated particle clearance is the predominant mechanism by which
11 relatively insoluble particles are removed from the alveolar region of the lungs (Gibb and
12 Morrow, 1962; Ferin, 1982; Harmsen et al., 1985; Lehnert and Morrow, 1985; Powdrill et al.,
13 1989).
14 The removal characteristics for particles deposited in the alveolar region of the lung have
15 been descriptively represented by numerous investigators as a multicompartment or
16 multicomponent process in which each component follows simple first-order kinetics (Snipes
17 and Clem, 1981; Snipes et al., 1988; Lee et al., 1983). Although the various compartments can
18 be described mathematically, the actual physiological mechanisms determining these differing
19 clearance rates have not been well characterized.
20 Lehnert et al. (1988, 1989) performed studies using laboratory rats to examine
21 particle-AM relationships over the course of alveolar clearance of low to high lung burdens of
22 noncytotoxic microspheres (2.13 |im diam.) to obtain information on potential AM-related
23 mechanisms that form the underlying bases for kinetic patterns of alveolar clearance as a function
24 of particle lung burdens. The intratracheally instilled lung burdens varied from 1.6 x 107
25 particles (about 85 ug) for the low lung burden to 2.0 x 108 particles (about 1.06 mg) for the mid-
26 dose and 6.8 x 108 particles (about 3.6 mg) for the highest lung burden. The lungs were lavaged
27 at various times postexposure and the numbers of spheres in each macrophage counted.
28 Although such experiments provide information regarding the response of the lung to particulate
29 matter, intratracheal instillation is not likely to result in the same depositional characteristics as
30 inhalation of particles. Therefore, it is unlikely that the response of alveolar macrophages to
31 these different depositional characteristics will be quantitatively similar.
32 The t]/2 values of both the early and later components of clearance were virtually identical
33 following deposition of the low and medium lung burdens. For the highest lung burden,
34 significant prolongations were found in both the early, mere rapid, as well as the slower
35 component of alveolar clearance The percentages of the particle burden associated with Lhe
36 earlier and later components, however, were similar to those of the lesser lung burdens. On the
7/25/00 3-14 DRAFT—DO NOT CITE OR QUOTE
-------
basis of the data, the authors concluded that translocation of AMs from alveolar spaces by way of
the conducting airways is fundamentally influenced by the particle burden of the cells so
3 translocated. In the case of particle overload that occurred at the highest lung burden, the
4 translocation of AMs with the heaviest cellular burdens of particles (i.e., greater than about
5 100 microspheres per AM) was definitely compromised.
6 On the other hand, analysis of the disappearance of AMs with various numbers of
7 particles indicates that the particles may not exclusively reflect the translocation of AMs from the
8 lung. The observations are also consistent with a gradual redistribution of retained particles
9 among the AMs in the lung concurrent with the removal of particle-containing AMs via the
10 conducting airways. Experimental support suggestive of potential processes for such particle
11 redistribution comes from a variety of investigations involving AMs and other endocytic cells
12 (Heppleston and Young, 1973; Evans et al., 1986; Aronson, 1963; Sandusky et al., 1977;
13 Heppleston, 1961; Riley and Dean, 1978).
14
15 3.3.3. Translocations of Particles to Extra-Alveolar Macrophage Compartment Sites
16 Although the phagocytosis of particles by cells free within the lung and the mucociliary
17 clearance of the cells with their particulate matter burdens represent the most prominent
mechanisms that govern the fate of particles deposited in the alveolar region, other mechanisms
exist that can affect both the retention characteristics of relatively insoluble particles in the lung
20 and the lung clearance pathways for the particles. One mechanism is endocytosis of particles by
21 alveolar lining (Type I) cells (Sorokin and Brain, 1975; Adamson and Bowden, 1978, 1981) that
22 normally provide >90% of the cell surface of the alveoli in the lungs of a variety of mammalian
23 species (Crapo et al., 1983). This process may be related to the size of the particles that deposit
24 in the lungs and the numbers of particles that are deposited. Adamson and Bowden (1981) found
25 that with increasing loads of carbon particles (0.03 nm diam.) instilled in the lungs of mice, more
26 free particles were observed in the alveoli within a few days. The relative abundance of particles
27 endocytosed by Type I cells also increased with increasing lung burdens of the particles, but
28 instillation of large particles (1.0 um) rarely resulted in their undergoing endocytosis. A 4 mg
29 burden of 0.1 (am diameter latex particles is equivalent to 8 x 1012 particles, whereas a 4 mg
30 burden of 1.0 um particles is composed of 8 * 109 particles. Regardless, DPM with volume
31 median diameters between 0.05 and 0.3 um (Frey and Corn, 1967; Kittleson et al., 1978) would
32 be expected to be within the size range for engulfment by Type I cells should suitable encounters
33 occur. Indeed, it has been demonstrated that DPM is endocytosed by Type I cells in vivo (White
34 and Garg, 1981).
^P Unfortunately, information on the kinetics of particle engulfment (endocytosis) by Type I
36 cells relative to that by AMs is scanty. Even when relatively low burdens of particulate matter
7/25/00 . 3-15 DRAFT—DO NOT CITE OR QUOTE
-------
1 are deposited in the lungs, some fraction of the particles usually appears in the regional lymph
2 nodes (Ferin and Fieldstein, 1978; Lehnert, 1989). As will be discussed, endocytosis of particles
3 by Type I cells is an initial, early step in the passage of particles to the lymph nodes. Assuming
4 particle phagocytosis is not sufficiently rapid or perfectly efficient, increasing numbers of
5 particles would be expected to gain entry into the Type I epithelial cell compartment during
6 chronic aerosol exposures. Additionally, if particles are released on a continual basis by AMs
7 that initially sequestered them after lung deposition, some fraction of the "free" particles so
8 released could also undergo passage from the alveolar space into Type I cells.
9 The endocytosis of particles by Type I cells represents only the initial stage of a process
10 that can lead to the accumulation of particles in the lung's interstitial compartment and the
11 subsequent translocation of particles to the regional lymph nodes. As shown by Adamson and
12 Bowden (1981), a vesicular transport mechanism in the Type I cell can transfer particles from the
13 air surface of the alveolar epithelium into the lung's interstitium, where particles may be
14 phagocytized by interstitial macrophages or remain in a "free" state for a poorly defined period
15 that may be dependent on the physicochemical characteristics of the particle. The lung's
16 interstitial compartment accordingly represents an anatomical site for the retention of particles in
17 the lung, especially so for primates. Whether or not AMs, and perhaps polymorphonuclear
18 neutrophils (PMNs) that have gained access to the alveolar space compartment and phagocytize
19 particles there, also contribute to the particle translocation process into the lung's interstitium
20 remains a controversial issue.
21 Translocation of paniculate matter to the various interstitial spaces within the lung is a
22 prominent phenomenon occurring at least at high (occupational) exposures that has been
23 examined extensively for both DPM and coal dust in a species comparison between rats and
24 primates (Nikula et al., 1997a,b). Detailed pulmonary morphometry conducted on F344 rats and
25 cynomolgus monkeys that had been exposed for 24 months to occupational levels of DPM (1.95
26 mg/mj; see Lewis et al., 1989) showed major differences in the pulmonary sites of particulate
27 deposition. In rats about 73% of DPM was present in the alveolar ducts/alveoli and 27% in
28 interstitial compartments; for monkeys the corresponding figures were markedly different at 43%
29 and 57%. The corresponding pulmonary histopathology confirmed that both species were
30 affected, although rats are more sensitive, as incidence and severity scores for alveolar effects
31 ranged from 15 of 15 with severity scores from 1 -4 (minimal to moderate), whereas for monkeys
32 the corresponding values were only 4 of 15 at a range of 0-2 (not observed to minimal).
33 Similarly, both species exhibited histopathology at the interstitial sites of deposition but with
34 effects in monkeys being slightly mere severe (1 of 15 graded as slight, 14 of 15 graded as
35 minimal) than those in rats (M of 15 graded as sli^lu, 1 of 15 graded as minimal). The basis for
36 this interspecies difference may be due to any number of clear contrasts that exist between rat
7/25/00 3-16 DRAFT—DO NOT CITE OR QUOTE
-------
and primate lungs, including anatomical (primates and humans have respiratory bronchioles
whereas rats do not), kinetic (primates and human clearance processes allow more residence time
3 of particles in the lung than do those in rats), or morphological (primates and humans have more
4 interstitial tissue, more and thicker pleura, and wider interstitial spaces than do rats). The
5 analysis of Kuempel (2000) using human occupational data clearly showed that models require
6 an interstitialization process to provide adequate fits to the empirical human (miners') lung
7 deposition data discussed in that study. Hypotheses about possible mechanisms for the
8 interstitialization process are scant, although Harmsen et al. (1985) provided some evidence in
9 dogs that migration of AMs may contribute to the passage of particles to the interstitial
10 compartment and also may be involved in the subsequent translocation of particles to draining
11 lymph nodes. Translocation to the extrapulmonary regional lymph nodes apparently can involve
12 the passage of free particles as well as particle-containing cells via lymphatic channels in the
13 lungs (Harmsen et al., 1985; Ferin and Fieldstein, 1978; Lee et al., 1985). Further, it has been
14 noted that particles accumulate both more rapidly and more abundantly in lymph nodes that
15 receive lymphatic drainage from the lung (Ferin and Feldstein, 1978; Lee et al., 1985). As a final
16 point, it should be stressed that further investigation is required to confirm the character and even
17 existence of the interstitialization process in the lungs of humans with exposures to particles at
lower environmental concentrations, or to submicrometer particles such as DPM.
20 3.3.3.1. Clearance Kinetics
21 The clearance kinetics of PM have been reviewed in the PM CD (U.S. EPA, 1996) and by
22 Schlesinger et al. (1997), the results of which indicate that clearance kinetics may be profoundly
23 influenced by several factors. The influence of time, for example, is definitively showed by the
24 work of Bailey et al. (1985; discussed above), who showed that the rate of clearance from the
25 pulmonary region to the GI tract decreased nearly fourfold from initial values to those noted at
26 200 days and beyond after particle inhalation.
27
28 3.3.3.2. Interspecies Patterns of Clearance
29 The inability to study the retention of certain materials in humans for direct risk
30 assessment requires the use of laboratory animals. Adequate toxicological assessment
31 necessitates that interspecies comparisons consider aspects of dosimetry including knowledge of
32 clearance rates and routes. The basic mechanisms and overall patterns of clearance from the
33 respiratory tract are similar in humans and most other mammals. Regional clearance rates,
34 however, can show substantial variation between species, even for similar particles deposited
^P under comparable exposure conditions (U.S. EPA, 1996; Schlesinger et al., 1997; Snipes et al.,
36 1989).
7/25/00 3-17 DRAFT—DO NOT CITE OR QUOTE
-------
1 In general, there are species-dependent rate constants for various clearance pathways.
2 Differences in regional and total clearance rates between some species are a reflection of
3 differences in mechanical clearance processes. For consideration in assessing particle dosimetry,
4 the end result of interspecies differences in clearance is that the retained doses in the lower
5 respiratory tract can differ between species, which may result in differences in response to similar
6 particulate exposures.
7
8 3.3.3.3. Clearance Modifying Factors and Susceptible Populations
9 A number of host and environmental factors may modify clearance kinetics and may
10 consequently make individuals exhibiting or afflicted with these factors particularly susceptible
11 to the effects resulting from exposure to DPM. These include age, gender, physical activity,
12 respiratory tract disease, and inhalation of irritants (U.S. EPA, 1996, Section 10.4.2.5).
13 Respiratory tract clearance appears to be prolonged in a number of pathophysical conditions in
14 humans, including chronic sinusitis, chronic bronchitis, ashthma, chronic obstructive lung
15 disease, and various acute respiratory infections.
16
17 3.3.3.4. Respiratory Tract Disease
18 Earlier studies reviewed in the PM CD (U.S. EPA, 1996) noted that various respiratory
19 tract diseases are associated with alterations in overall clearance and clearance rates. Prolonged
20 nasal mucociliary clearance in humans is associated with chronic sinusitis or rhinitis, and cystic
21 fibrosis. Bronchial mucus transport may be impaired in people with bronchial carcinoma,
22 chronic bronchitis, asthma, and various acute infections. In certain of these cases, coughing may
23 enhance mucus clearance, but it generally is effective only if excess secretions are present.
24 The rates of A region particle clearance are reduced in humans with chronic obstructive
25 lung disease and in laboratory animals with viral infections, whereas the viability and functional
26 activity of macrophages are impaired in human asthmatics and in animals with viral-induced lung
27 infections (U.S. EPA, 1996). However, any modification of functional properties of
28 macrophages appears to be injury specific, reflecting the nature and anatomic pattern of disease.
29
30 3.4. PARTICLE OVERLOAD
31 3.4.1. Introduction
32 Some experimental studies using laboratory rodents employed high exposure
33 concentrations of relatively nontoxic, poorly soluble particles. These particle loads interfered
34 with normal clearance mechanisms, producing clearance rates different from those that would
35 occur at lower exposure levels. Prolonged exposure to high particle concentrations is associated
36 with what is termed particle overload. This is defined as the overwhelming of macrophage-
7/25/00 3-18 DRAFT—DO NOT CITE OR QUOTE
-------
mediated clearance by the deposition of particles at a rate exceeding the capacity of that
clearance pathway. Aspects and occurrence of this phenomenon have already been alluded to in
3 earlier portions of this chapter on alveolar clearance (Section 3.3.2.3). The relevance of this
4 phenomenon for human risk assessment has long been the object of scientific inquiry. A
5 monograph on this matter and many others relevant to DPM has appeared (ILSI, 2000), and the
6 results, opinions, and judgments put forth therein are used extensively in this chapter and in this
7 assessment.
8 Wolff et al. (1987) used 134Cs-labeled fused aluminosilicate particles to measure alveolar
9 clearance in rats following 24-mo exposure to low, medium, and high concentrations of diesel
10 exhaust (targeted concentrations of DPM of 0.35, 3.5 and 7.1 mg/m3). The short-term
11 component of the multicomponent clearance curves was similar for all groups, but long-term
12 clearance was retarded in the medium and high exposure groups (Figure 3-4). The half times of
13 the long-term clearance curves were 79, 81, 264, and 240 days, respectively, for the control, low-,
14 medium-, and high-exposure groups. Clearance was overloaded at the high and medium but not
15 at the low exposure level. Lung burdens of DPM were measured after 6, 12, 18, and 24 mo of
16 exposure. The results (Figure 3-5) indicate that the lung burden of freshly deposited particles
17 was appreciably increased in the two highest exposures post 6 mo., whereas the lung burden at
the low-exposure level remained the same throughout all time periods examined.
Morrow (1988) has proposed that the condition of particle overloading in the lungs is
20 caused by a loss in the mobility of particle-engorged AMs and that such an impediment is related
21 to the cumulative volumetric load of particles in the AMs. Morrow (1988) has further estimated
22 that the clearance function of an AM may be completely impaired when the particle burden in the
23 AM is of a volumetric size equivalent to about 60% of the normal volume of the AM. Morrow's
24 hypothesis was the initial basis for the physiology-oriented multicompartmental kinetic (POCK)
25 model derived by Stober et al. (1989) for estimating alveolar clearance and retention of relatively
26 insoluble, respirable particles in rats.
27 A revised version of this model refines the characterization of the macrophage pool by
28 including both the mobile and immobilized macrophages (Stober et al., 1994). Application of
29 the revised version of the model to experimental data suggested that lung overload does not cause
30 a dramatic increase in the total burden of the macrophage pool but results in a great increase in
31 the particle burden of the interstitial space, a compartment that is not available for macrophage-
32 mediated clearance. The revised version of the POCK model is discussed in greater detail in the
33 context of other dosimetry models below.
34 Oberdorster and co-workers (1992) assessed the alveolar clearance of smaller (3.3 um
^A diam.) and larger (10.3 um diam.) polystyrene particles, the latter of which are volumetrically
36 equivalent to about 60% of the average normal volume of a rat AM, after intratracheal instillation
7/25/00 3-19 DRAFT—DO NOT CITE OR QUOTE
-------
1 into the lungs of rats. Even though both sizes of particles were found to be phagocytized by AMs
2 within a day after deposition, and the smaller particles were cleared at a normal rate, only
3 minimal lung clearance of the larger particles was observed over an approximately 200-day
4 postinstillation period, thus supporting the volumetric AM overload hypothesis.
5 It has been hypothesized that when the retained lung burden approaches 1 mg particles/g
6 lung tissue, overloading will begin in the rat (Morrow, 1988); at 10 mg particles/g lung tissue
7 macrophage-mediated clearance of particles would effectively cease. Overloading appears to be
8 a nonspecific effect noted in experimental studies, generally in rats, using many different kinds of
9 poorly soluble particles (including TiO2, volcanic ash, DPM, carbon black, and fly ash) and
10 results in A region clearance slowing or stasis, with an associated inflammation and aggregation
11 of macrophages in the lungs and increased translocation of particles into the interstitium (Muhle
12 et al., 1990; Lehnert, 1990; Morrow, 1994). Following overloading, the subsequent retardation
13 of lung clearance, accumulation of particles, chronic inflammation, and the interaction of
14 inflammatory mediators with cell proliferative processes and DNA may lead to the development
15 of fibrosis, epithelial cell mutations, and fibrosis in rats (Mauderly, 1996). The phenomenon of
16 overload has been discussed in greater detail in the previous PM CD (U.S. EPA, 1996).
17
18 3.4.2. Relevance to Humans
19 The relevance of lung overload to humans, and even to species other than laboratory rats
20 and mice, is not clear. Although likely to be of little relevance for most "real world" ambient
21 exposures of humans, this phenomenon is of concern in interpreting some long-term
22 experimental exposure data and perhaps for human occupational exposure. In addition,
23 relevance to humans is clouded by the suggestion that macrophage-mediated clearance is
24 normally slower and perhaps less important in humans than in rats (Morrow, 1994), and that
25 there can be significant differences in macrophage loading between species. Particle overload
26 appears to be an important factor in the pulmonary carcinogenicity observed in rats exposed to
27 DPM. Studies described in this section provide additional data showing a particle overload
28 effect. A study by Griffis et al. (1983) demonstrated that exposure (7 h/day, 5 days/week) of rats
29 to DPM at concentrations of 0.15, 0.94, or 4.1 mg/m3 for 18 mo resulted in lung burdens of
30 U.035, 0.220, and 1.89 mg/g of lung tissue, respectively. The alveolar clearance of those rats
31 with the highest lung burden (1.89 mg/g of lung) was impaired, as determined by a significantly
32 greater (pO.OOOl) retention t,/2 for DPM. Impaired clearance was reflected in the greater lung
33 burden/exposure concentration ratio at the highest exposure level. Similarly, in the study by
34 Chan et al. (1984), rats exposed for 20 h/day, 7 days/week to DPM (6 rng/rn3) for 112 days had
35 an extraordinarilv high lung particle burden nf 1i g rng; with nc alveolar particle clearance being
36 detected over 1 year.
7/25/00 3-20 DRAFT—DO NOT CITE OR QUOTE
-------
Muhle et al. (1990) indicated that overloading of rat lungs occurred when lung particle
burdens reached 0.5 to 1.5 mg/g of lung tissue and that clearance mechanisms were totally
3 compromised at lung particle burdens £ 10 mg/g for particles with a specific density close to 1,
4 observations that are concordant with those of Morrow (1988).
5 Pritchard (1989), utilizing data from a number of diesel exhaust exposure studies,
6 examined alveolar clearance in rats as a function of cumulative exposure. The resulting analysis
7 noted a significant increase in retention t1/2 values at exposures above 10 mg/m3'h/day and also
8 showed that normal lung clearance mechanisms appeared to be compromised as the lung DPM
9 burden approached 0.5 mg/g of lung.
10 Animal studies have revealed that impairment of alveolar clearance can occur following
11 chronic exposure to DPM (Griffis et al., 1983; Wolff et al., 1987; Vostal et al., 1982; Lee et al.,
12 1983) or a variety of other diverse poorly soluble particles of low toxicity (Lee et al., 1986, 1988;
13 Ferin and Feldstein, 1978; Muhle et al., 1990). Because high lung burdens of relatively
14 insoluble, biochemically inert particles result in diminution of normal lung clearance kinetics or
15 in what is now called particle overloading, this effect appears to be more related to the mass
16 and/or volume of particles in the lung than to the nature of the particles per se. Particle overload
17 relates only to poorly soluble particles of low toxicity. It must be noted, however, that some
* types of particles may be cytotoxic and impair clearance at lower lung burdens (e.g., crystalline
silica may impair clearance at much lower lung burdens than DPM). Regardless, as pointed out
20 by Morrow (1988), particle overloading in the lung modifies the dosimetry for particles in the
21 lung and thereby can alter toxicologic responses.
22 Although quantitative data are limited regarding lung overload associated with impaired
23 alveolar clearance in humans, impairment of clearance mechanisms appears to occur, and at a
24 lung burden generally in the range reported to impair clearance in rats, i.e., approximately 1 mg/g
25 lung tissue. Stober et al. (1967), in their study of coal miners, reported lung particle burdens of 2
26 to 50 mg/g lung tissue, for which estimated clearance t,/2 values were very long (4.9 years).
27 Freedman and Robinson (1988) also reported slower alveolar clearance rates in coal miners,
28 some of whom had a mild degree of pneumoconiosis. It must be noted, however, that no lung
29 cancer was reported even among those miners with apparent particle overload.
30
31 3.4.3. Potential Mechanisms for an AM Sequestration Compartment for Particles During
32 Particle Overload
33 Several factors may be involved in the particle-load-dependent retardations in the rate of
34 particle removal from the lung and the corresponding functional appearance of an abnormally
^A slow clearing or particle sequestration compartment. As previously mentioned, one potential site
36 for particle sequestration is the containment of particles in the Type I cells. Information on the
7/25/00 3-21 DRAFT—DO NOT CITE OR QUOTE
-------
1 retention kinetics for particles in the Type I cells is not currently available. Also, no
2 morphometric analyses have been performed to date to estimate what fraction of a retained lung
3 burden may be contained in the Type I cell population of the lung during lung overloading.
4 Another anatomical region in the lung that may be a slow clearing site is the interstitial
5 compartment (Kuempel, 2000). Little is known about the kinetics of removal of free particles or
6 particle-containing macrophages from the interstitial spaces, or what fraction of a retained burden
7 of particles is contained in the lung's interstitium during particle overload. The gradual
8 accumulation of particles in the regional lymph nodes and the appearance of particles and cells
9 with associated particles in lymphatic channels and in the peribronchial and perivascular
10 lymphoid tissue (Lee et al., 1985; White and Garg, 1981) suggest that the mobilization of
11 particles from interstitial sites via local lymphatics is a continual process.
12 Indeed, it is clear from histologic observations of the lungs of animals chronically
13 exposed to DPM that Type I cells, the interstitium, the lymphatic channels, and pulmonary
14 lymphoid tissues could collectively comprise subcompartments of a more generalized slow
15 clearing compartment.
16 Although these sites must be considered potential contributors to the increased retention
17 of particles during particle overload, a disturbance in particle-associated AM-mediated clearance
18 is undoubtedly the predominant cause, inasmuch as, at least in animals, the AMs are the primary
19 reservoirs of deposited particles. The factors responsible for a failure of AMs to translocate from
20 the alveolar space compartment in lungs with high paniculate matter burdens remain uncertain,
21 although a hypothesis concerning the process involving volumetric AM burden has been offered
22 (Morrow, 1988).
23 Other processes also may be involved in preventing particle-laden AMs from leaving the
24 alveolar compartment under conditions of particle overload in the lung. Clusters or aggregates of
25 particle-laden AMs in the alveoli are typically found in the lungs of laboratory animals that have
26 received large lung burdens of a variety of types of particles (Lee et al., 1985), including DPM
27 (White and Garg, 1981; McClellan et al., 1982). The aggregation of AMs may explain, in part,
28 the reduced clearance of particle-laden AM during particle overload. The definitive
29 mechanism(s) responsible for this clustering of AMs has not been elucidated to date. Whatever
30 the underlying mechanism(s) for the AM aggregation response, it is noteworthy that AMs
31 lavaged from the lungs of diesel exhaust-exposed animals continue to demonstrate a propensity
32 to aggregate (Strom, 1984). This observation suggests that the surface characteristics of AMs are
33 fundamentally altered in a manner that promotes their adherence to one another in the alveolar
34 region, and that AM aggregation may not simply he directly caused by their abundant
35 accumulation as a result of immobilization hy large particle loads. Furthermore, even though
36 overloaded macrophages may redistribute particle burden to other AMs, clearance may remain
7/25/00 3-22 DRAFT—DO NOT CITE OR QUOTE
-------
inhibited (Lehnert, 1988). This may, in part, be because attractants from the overloaded AMs
cause aggregation of those that are not carrying a particle burden.
3
4 3.5. BIO AVAILABILITY OF ORGANIC CONSTITUENTS PRESENT ON DIESEL
5 EXHAUST PARTICLES
6 Because it has been shown that DPM extract is not only mutagenic but also contains
7 known carcinogens, the organic fraction was originally considered to be the primary source of
8 carcinogenicity in animal studies. Since then evidence has been presented that carbon black,
9 lacking an organic component, is capable of inducing lung cancer at exposure concentrations
10 sufficient to induce lung particle overload. This suggested that the relatively insoluble carbon
11 core of the particle may be of greater importance for the pathogenic and carcinogenic processes
12 observed in the rat inhalation studies conducted at high exposure concentrations. (See Chapter 7
13 for a discussion of this issue.) However, lung cancer reported in epidemiology studies was
14 associated with diesel exposure levels far below those inducing particle overload in lifetime
15 studies in rats. It is therefore reasoned that compounds in the organic fraction of DPM may have
16 some role in the etiology of human lung cancers.
17 The bioavailability of toxic organic compounds adsorbed to DPM can be influenced by a
variety of factors. Although the agent may be active while present on the particle, most particles
are taken up by AMs, a cell type not generally considered to be a target site. In order to reach the
20 target site, elution from the particle surface is necessary followed by diffusion and uptake by the
21 target cell. Metabolism to an active form by either the phagocytes or the target cells is also
22 required for activity of many of the compounds present.
23
24 3.5.1. In Vivo Studies
25 3.5.1.1. Laboratory Investigations
26 Several studies reported on the retention of particle-adsorbed organics following
27 administration to various rodent species. In studies reported by Sun et al. (1982, 1984) and Bond
28 et al. (1986), labeled organics were deposited on DPM following heating to vaporize away the
29 organics originally present. Sun et al. (1982) compared the disposition of either pure or diesel
30 particle-adsorbed benzo[a]pyrene (BaP) following nose-only inhalation by F344 rats. About
31 50% of particle-adsorbed BaP was cleared with a half-time of 1 h, predominantly by mucociliary
32 clearance. The long-term retention of particle-adsorbed 3H-BaP of 18 days was approximately
33 230-fold greater than that for pure 3H-BaP (Sun et al., 1982). At the end of exposure, about 15%
34 of the 3H label was found in blood, liver, and kidney. Similar results were reported in a
^^ companion study by Bond et al. (1986), and by Sun et al. (1984) with another PAH, 1-
36 nitropyrene, except the retention half-time was 36 days.
7/25/00 3-23 DRAFT—DO NOT CITE OR QUOTE
-------
1 Ball and King (1985) studied the disposition and metabolism of intratracheally instilled
2 I4C-labeled 1-NP (>99.9% purity) coated onto DPM. About 50% of the I4C was excreted within
3 the first 24 h; 20% to 30% of this appeared in the urine, and 40% to 60% was excreted in the
4 feces. Traces of radiolabel were detected in the trachea and esophagus. Five percent to 12% of
5 the radiolabel in the lung co-purified with the protein fraction, indicating some protein binding.
6 The corresponding DNA fraction contained no I4C above background levels.
7 Bevan and Ruggio (1991) assessed the bioavailability of BaP adsorbed to DPM from a
8 5.7-L Oldsmobile diesel engine. In this study, exhaust particles containing 1.03 ug BaP/g
9 particles were supplemented with exogenous 3H-BaP to provide 2.62 ug BaP/g of exhaust
10 particles. In vitro analysis indicated that the supplemented BaP eluted from the particles at the
11 same rate as the original BaP. Twenty-four hours after intratracheal instillation in Sprague-
12 Dawley rats, 68.5% of the radiolabel remained in the lungs. This is approximately a 3.5-fold
13 greater proportion than that reported by Sun et al. (1984), possibly because smaller amounts of
14 BaP adsorbed on the particles resulted in stronger binding or possibly because of differences
15 between inhalation exposure and intratracheal exposure. At 3 days following administration,
16 more than 50% of the radioactivity remained in the lungs, nearly 30% had been excreted into the
17 feces, and the remainder was distributed throughout the body. Experiments using rats with
18 cannulated bile ducts showed that approximately 10% of the administered radioactivity appeared
19 in the bile over a 10-h period and that less than 5% of the radioactivity entered the feces via
20 mucociliary transport. Results of these studies showed that when organics are adsorbed to DPM
21 the retention of organics in the lungs is increased considerably. Because retention time is very
22 short following exposure to pure compounds not bound to particles, it can be concluded that the
23 increased retention time is primarily the result of continued binding to DPM. The detection of
24 labeled compounds in blood, systemic organs, urine, and bile as well as the trachea, however,
25 provides evidence that at least some of the organics are eluted from the particles following
26 deposition in the lungs and would not be available as a carcinogenic dose to the lung. As
27 discussed in Section 3.6.3, most of the organics eluted from particles deposited in the alveolar
28 region, especially PAHs. are predicted to rapidly enter the bloodstream and thus not to contribute
29 to potential induction of lung cancer.
30
31 3.5.1.2. Studies in Occupationally Exposed Humans
32 DNA adducts in the lungs of experimental animals exposed to diesel exhaust have been
33 measured in a number of animal experiments (World Health Organization, 1996). Such studies,
34 however, provide limited information regarding bicavailability of organics, as positive results
35 may well have heen related to factors associated with lung particle overload, a circumstance
36 reported by Bond et al. (1990), who found carbon black, a substance virtually devoid of organics,
7/25/00 3-24 DRAFT—DO NOT CITE OR QUOTE
-------
1 to induce DNA adducts in rats at lung overload doses. These authors showed that levels of DNA
fadducts present in pulmonary type II cells from the lungs of rats (n=15) exposed to equivalent
conditions of either carbon black or diesel exhaust (each at 6.2 mg/m3) were nearly the same and
4 4- to 5-fold more than air-exposed controls. This similarity was noted despite a difference of
5 nearly three orders of magnitude in solvent-extractable organic content between diesel exhaust
6 (30%) and carbon black (0.04%). None of the diesel exhaust or carbon black adducts comigrated
7 with BPDE (BaP diol epoxide).
8 On the other hand, DNA adduct formation and/or mutations in blood cells following
9 exposure to DPM, especially at levels insufficient to induce lung overload, can be presumed to be
10 the result of organics diffusing into the blood. Hemminki et al. (1994) reported increased levels
11 of DNA adducts in lymphocytes of bus maintenance and truck terminal workers. Osterholm
12 et al. (1995) studied mutations at the hprt-locus of T-lymphocytes in bus maintenance workers.
13 Although they were unable to identify clear-cut exposure-related differences in types of
14 mutations, adduct formation was significantly increased in the exposed workers. Nielsen et al.
15 (1996) reported significantly increased levels of lymphocyte DNA adducts, hydroxyvaline
16 adducts in hemoglobin, and 1-hydroxypyrene in urine of garage workers exposed to diesel
17 exhaust.
18
40 3.5.2. In Vitro Studies
20 3.5.2.1. Extraction of Diesel Particle-Associated Organics by Biological Fluids
21 In vitro extraction of mutagenic organics by biological fluids can be estimated by
22 measurement of mutagenic activity in the particular fluid. Using this approach, Brooks et al.
23 (1981) reported extraction efficiencies of only 3% to 10% that of dichloromethane following
24 DPM incubation in lavage fluid, serum, saline, albumin, or dipalmitoyl lecithin. Moreover,
25 extraction efficiency did not increase with incubation time up to 120 h. Similar findings were
26 reported by King et al. (1981), who also reported that lung lavage fluid and lung cytosol fluid
27 extracts of DPM were not mutagenic. Serum extracts of DPM did exhibit some mutagenic
28 activity, but considerably less than that of organic solvent extracts. Furthermore, the mutagenic
29 activity of the solvent extract was significantly reduced when combined with serum or lung
30 cytosol fluid, suggesting protein binding or biotransformation of the mutagenic components.
31 Siak et al. (1980) assessed the mutagenicity of material extracted from DPM by bovine serum
32 albumin in solution, simulated lung surfactant, fetal calf serum (PCS), and physiological saline.
33 Only .PCS was found to extract some mutagenic activity from the DPM. Keane et al. (1991),
34 however, reported positive effects for mutagenicity in Salmonella and sister chromatid exchange
3^^ in V79 cells exclusively in the supernatant fraction of DPM dispersed in aqueous mixtures of
36 dipalmitoyl phosphatidyl choline, a major component of pulmonary surfactant, indicating that
7/25/00 3-25 DRAFT—DO NOT CITE OR QUOTE
-------
1 pulmonary surfactant components can extract active components of DPM and result in
2 bioavailability.
3 The ability of biological fluids to extract organics in vitro and their effectiveness in vivo
4 remains equivocal because of the character of the particular fluid. For example, extracellular
5 lung fluid is a complex mixture of constituents that undoubtedly have a broad range of
6 hydrophobicity (George and Hook, 1984; Wright and Clements, 1987), which is fundamentally
7 different from serum in terms of chemical composition (Gurley et al., 1988). Moreover,
8 assessments of the ability of lavage fluids, which actually represent substantially diluted
9 extracellular lung fluid, to extract mutagenic activity from DPM clearly do not reflect the in vivo
10 condition. Finally, except under very high exposure concentrations, few particles escape
11 phagocytosis and possible intracellular extraction. In this respect, Hiura et al. (1999) have
12 shown that whole DPM, but not carbon black or diesel particles devoid of organics, induces
13 apoptosis, apparently through generation of oxygen radicals. This study implicates organic
14 compounds present on DPM. It also indicates the bioavailability of organics for generation of
15 radicals from reaction with particle-associated organics or following elution from DPM.
16
17 3.5.2.2. Extraction of DPM-Associated Organics by Lung Cells and Cellular Components
18 A more likely means by which organics may be extracted from DPM and metabolized in
19 the lung is either through particle dissolution or extraction of organics from the particle surface
20 within the phagolysosomes of AMs and other lung cells. This mechanism presupposes that the
21 particles are internalized. Specific details about the physicochemical conditions of the
22 intraphagolysosomal environment, where particle dissolution in AMs presumably occurs in vivo,
23 have not been well characterized. It is known that phagolysosomes constitute an acidic (pH 4 to
24 5) compartment in macrophages (Nilsen et al., 1988; Ohkuma and Poole, 1978). The relatively
25 low pH in the phagolysosomes has been associated with the dissolution of some types of
26 inorganic particles (some metals) by macrophages (Marafante et al., 1987; Lundborg et al.,
27 1984), but few studies provide quantitative information concerning how organics from DPM may
28 be extracted in the phagolysosomes (Bond et al., 1983). Whatever the mechanism, assuming
29 elution occurs, the end result is a prolonged exposure of the respiratory epithelium to DPM
30 organics, which include low concentrations of carcinogenic agents such as PAH.
31 Early studies by King et al. (1981) found that when pulmonary alveolar macrophages
32 were incubated with DPM, amounts of organic compounds and mutagenic activity decreased
33 measurably from the amount originally associated with the particles, suggesting that organics
34 were removed from the phagocytized particles. Leung et a! (1988) studied the ability cf rat lung
35 and liver microsomes to facilitate transfer and metabolism of BaP from diese! particles. 14C Ba?
36 coated diesel particles, previously extracted to remove the original organics, were incubated
5/00 3-26 DRAFT—DO NOT CITE OR QUOTE
-------
directly with liver or lung microsomes. About 3% of the particle-adsorbed BaP was transferred
to the lung microsomes within 2 h. Of this amount about 1.5% was metabolized, for a total of
3 about 0.05% of the BaP originally adsorbed to the DPM. Although transformation is slow, the
4 long retention of particles, including DPM, in humans may cause the fraction eluted and
5 metabolized to be considerably higher than this figure.
6 In analyzing phagolysosomal dissolution of various ions from particles in the lungs of
7 Syrian golden hamsters, however, Godleski et al. (1988) demonstrated that solubilization did not
8 necessarily result in clearance of the ions (and therefore general bioavailability) in that binding of
9 the solubilized components to cellular and extracellular structures occurred. It is reasonable to
10 assume that phagocytized DPM particles may be subject to similar processes and that these
11 processes would be important in determining the rate of bioavailability of the particle-bound
12 constituents of DPM.
13 Alveolar macrophages or macrophage cell lines that were exposed to high concentrations
14 of DPM in vitro were observed to undergo apoptosis, which was attributed to the generation of
15 reactive oxygen radicals (ROR) (Hiura et al. 1999). Further experimentation showed that DPM
16 with the organic constituents extracted was no longer able to induce apoptosis or generate ROR.
17 The organic extracts alone, however, were able to induce apoptosis as well as the formation of
stress-activated protein kinases that play definitive roles in cellular apoptotic pathways. The
injurious effects of nonextracted DPM or of DPM extracts were observed to be reversible by the
20 antioxidant radical scavenger N-acetyl cysteine. These data suggest strongly that, at least at high
21 concentrations of DPM, the organic constituents contained on DPM play a central role in cellular
22 toxicity and that this toxicity may be attributable to the generation of ROR.
23
24 3.5.3. Modeling Studies
2 5 Gerde et al. (1991 a,b) described a model simulating the effect of particle aggregation and
26 PAH content on the rate of PAH release in the lung. According to this model, particle
27 aggregation will occur with high exposure concentrations, resulting in a slow release of PAHs
28 and prolonged exposure to surrounding tissues. However, large aggregates of particles are
29 unlikely to form at doses typical of human exposures. Inhaled particles, at low concentrations,
30 are more likely to deposit and react with surrounding lung medium without interference from
31 other particles. The model predicts that under low-dose exposure conditions, more typical in
32 humans, particle-associated organics will be released more rapidly from the particles because
33 they are not aggregated. Output from this model suggests strongly that sustained exposure of
34 target tissues to PAHs will result from repeated exposures, not from increased retention due to
association of PAHs with carrier particles. This distinction is important because at low doses
7/25/00 3-27 DRAFT—DO NOT CITE OR QUOTE
-------
1 PAH exposure and lung tumor formation would be predicted to occur at sites of deposition rather
2 than retention, as occurs with high doses.
3 The site of release of PAHs influences effective dose to the lungs because, as noted
4 previously, at least some free organic compounds deposited in the lungs are rapidly absorbed into
5 the bloodstream. Gerde et al. (1991b) predicted PAHs would be retained in the alveoli less than
6 1 min, whereas they may be retained in the conducting airways for hours. These predictions were
7 based on an average diffusion distance to capillaries of only about 0.5 um in the alveoli, as
8 compared to possibly greater than 50 (am in the conducting airways such as the bronchi. An
9 experimental study by Gerde et al. (1999) provided support for this prediction. Beagle dogs were
10 exposed to 3H-BaP adsorbed on the carbonaceous core of DPM at a concentration of 15 |j.g
11 BaP/gm particles. A rapidly eluting fraction from DPM deposited in the alveoli was adsorbed
12 into the bloodstream and metabolized in the liver, whereas the rapidly eluting fraction from DPM
13 deposited in the conducting airways was to a large extent retained and metabolized in situ in the
14 airway epithelium. Thus, organics eluting from DPM depositing in the conducting airways (i.e.,
15 the TB region) would have a basis for a longer residence time in the tissues (and for consequent
16 biological activity) than would organics eluting from DPM depositing in the pulmonary
17 parenchyma. And, given the same overall deposited dose of DPM to the total pulmonary system,
18 a deposited dose with a higher proportion in the TB region would incur a higher probability of
19 tissue interactions with any eluted organics. This may be the case when comparing regional
20 doses of DPM to humans as compared to rats for two reasons. First, one deposition model
21 (Freijer et al., 1999) projects that for air concentrations of DPM at either 0.1 or 1.0 mg/m3, a
22 higher proportion of the total DPM dose to the pulmonary system would be deposited in the TB
23 area for humans at 31 % (TB/Total; 0.098 / 0.318) than for rats at only 16% (0.04 / 0.205).
24 Second, comparative morphometry data of DPM from chronically exposed rats and primates
25 showed higher levels of DPM adjacent to conducting airways in primates (i.e., the interstitium of
26 the respiratory bronchioles) than were present in parallel regions in the rat (interstitium of the
27 alveolar ducts) (Nikula et al., 1997a,b). The focal nature of this deposition could give rise to
28 localized high concentrations of any organics eluted.
29 Overall, the results of studies presented in Section 3.6 provide evidence that at least some
30 of the organic matter adsorbed to DFM deposited in the rcspiratui> tract is eluted. The
31 percentage taken up and metabolized to an active form by target cells is, however, uncertain.
32 Organics eluted from particles deposited in alveoli are likely to rapidly enter the bloodstream via
33 translocation across endothelial cells, where they may undergo metabolism by enzymes such as
34 cytochromes P-450 that are capable of producing reactive species. Organics eluted from particles
35 deposited in the conducting airways (the bronchioles, bronchi, and trachea) may also undergo
36 metabolism in other cell types such as the Clara cells with constituent or inducible cytochrome P-
7/25/00 3-28 DRAFT—DO NOT CITE OP-QUOTE
-------
450 species. Risk of harmful effects for particles deposited in the conducting airways is
predicted to be greater because solubilized organic compounds will be retained in the thicker
3 tissue longer, allowing for metabolism by epithelial cells lining the airways. Furthermore, since
4 some deposition conducting airways occurs primarily at bifurcations, localized higher
5 concentrations may occur. At present, unfortunately, the available data are insufficient to
6 accurately model the effective dose of organics in the respiratory tract of humans or animals
7 exposed to DPM.
8
9 3.5.4. Unavailability/Deposition of Organics
10 Using the data presented by Xu and Yu (1987), it is possible to calculate the total mass of
11 DPM, as well as the total organic mass and specific carcinogenic PAHs deposited in the lungs of
12 an individual exposed to DPM. For example, the annual deposition of DPM in the lungs of an
13 individual exposed continuously to 1 ug/m3 DPM can be estimated to be about 420 ng based on
14 total lung volume (see Table 3-1). About 0.7% of particle mass consists of PAHs (see Section
15 2.2.6.2, Chapter 2) for a total of 2.94 ^g. Of this amount, the deposited mass of nitro-polycyclic
16 aromatic compounds, based on data by Campbell and Lee (1984), would equal 37 ng, while the
17 deposited mass of 7 PAHs that tested positive in cancer bioassays (U.S. EPA, 1993), and
measured by Tong and Karasek (1984), would range from 0.16 to 0.35 ng. Exercises similar to
this have been carried out by others, e.g., Valberg and Watson (1999). However, the possibility
20 that high concentrations of DPM may result in localized areas of deposition (such as the
21 conducting airways), the fact that human exposures may be considerably greater than those
22 presupposed in the exercise (e.g., 1 ng/m3), the nature of the assays (i.e., in vitro in Chapter 4 vs.
23 actual inhalation exposures), and the findings that DNA adducts may result from other known
24 noncarcinogens such as carbon black (Bond et al., 1990) make the interpretation of such
25 exercises problematic and their meaning unclear.
26
27 3.6. MODELING THE DEPOSITION AND CLEARANCE OF PARTICLES IN THE
28 RESPIRATORY TRACT
29 3.6.1. Introduction
30 The biological effects of inhaled particles are a function of their disposition, i.e., their
31 deposition and clearance. This, in turn, depends on their patterns of deposition (i.e., the sites
32 within which particles initially come into contact with airway epithelial surfaces and the amount
33 removed from the inhaled air at these sites) and clearance (i.e., the rates and routes by which
34 deposited materials are removed from the respiratory tract). Removal of deposited materials
A involves the competing processes of macrophage-mediated clearance and dissolution-absorption.
36 Over the years, mathematical models for predicting deposition, clearance and, ultimately,
7/25/00 3-29 DRAFT—DO NOT CITE OR QUOTE
-------
1 retention of particles in the respiratory tract have been developed. Such models help interpret
2 experimental data and can be used to make predictions of deposition for cases where data are not
3 available. A review of various mathematical deposition models was given by Morrow and Yu
4 (1993) and in U.S. EPA (1996).
5 Currently available data for long-term inhalation exposures to poorly soluble particles
6 (e.g., TiO2, carbon black, and DPM) show that pulmonary retention and clearance of these
7 particles are not adequately described by simple first-order kinetics and a single compartment
8 representing the alveolar macrophage particle burden. Several investigators have developed
9 models for deposition, transport, and clearance of poorly soluble paniculate matter in the lungs.
10 All of these models identify various compartments and associated transport rates, but empirically
11 derived data are not available to substantiate many of the assumptions made in these models.
12
13 3.6.2. Dosimetry Models for DPM
14 3.6.2.1. Introduction
15 The extrapolation of tdxicological results from laboratory animals to humans, the goal of
16 this chapter, requires the use of dosimetry models for both species that include, first, the
17 deposition of DPM in various regions of the respiratory tract, and second, the transport and
18 clearance of the particles, including adsorbed constituents, from their deposited sites. Therefore
19 the ideal model structure would incorporate both deposition and clearance in animals and
20 humans.
21 Deposition of particles in the respiratory tract, as described above, can be by impaction,
22 sedimentation, interception, and diffusion, with the contribution from each mechanism a
23 function of particle size, lung structure, and size and breathing parameters. Because of the size
24 of diesel particles, under normal breathing conditions most of this deposition takes place by
25 diffusion, and the fraction of the inhaled mass that is deposited in the thoracic region (i.e., TB
26 plus A regions) is substantially similar for rats and humans.
27 Among deposition models that include aspects of lung structure and breathing dynamics,
28 the most widely used have been typical-path or single-path models (Yu, 1978; Yu and Diu,
29 1983). The single-path models are based on an idealized symmetric geometry of the lung,
30 assuming regular dichotomous branching of the airways and alveolar ducts (Weibei, 1963). They
31 lead to modeling the deposition in an average regional sense for a given lung depth. Although
32 the lower airways of the lung may be reasonably characterized by such a symmetric
33 representation, there are major asymmetries in the upper airways of the tracheobronchial tree that
34 in turn lead to different apportionment of airflow and particulate burden to the different lung
35 lobes. The rat lung structure is highly asymmetric because of its monopodial nature, leading to
36 significant errors in a single-path description. This is rectified in the multiple-path model of the
-3 1f\
J-JV/
-------
lung, which incorporates asymmetry and heterogeneity in lung branching structure and calculates
deposition at the individual airway level. This model has been developed for the rat lung
3 (Anjilvel and Asgharian, 1995; Freijer et al., 1999) and, in a limited fashion because of
4 insufficient morphometric data, for the human lung (Subramaniam et al., 1998; Yeh and Schum,
5 1980). Such models are particularly relevant for fine and ultrafine particles such as occur in
6 DPM. However, models for clearance have not yet been implemented in conjunction with the
7 use of the multiple-path model.
8 Clearance of particles in the respiratory tract takes place (1) by mechanical processes:
9 mucociliary transport in the ciliated conducting airways and macrophage phagocytosis and
10 migration in the nonciliated airways, and (2) by dissolution. The removal of material such as the
11 carbonaceous core of DPM is largely by mechanical clearance, whereas the clearance of the
12 organics adsorbed onto the carbon core is principally by dissolution.
13 Several clearance models currently exist, some specifically for humans and others
14 specific for laboratory animals. They differ significantly in the level of physiological detail that
15 is captured in the model and in the uncertainties associated with the values of the parameters
16 used. All of these models identify various compartments and associated transport rates, but
17 empirically derived data are not available to validate many of the assumptions made in the
« models. A review of the principal human and animal deposition/clearance models, including
candidate models for use in animal-to-human extrapolation in this assessment, are considered
20 below.
21
22 3.6.2.2. Human Models
23 The International Commission on Radiological Protection (ICRP) recommends specific
24 mathematical dosimetry models as a means to calculate the mass deposition and retention by
25 different parts of the human respiratory tract and, if needed, tissues beyond the respiratory tract.
26 The latest ICRP-recommended model, ICRP66 (1994), considers the human respiratory tract as
27 four general anatomical regions: the ET region, which is divided into two subregions; the TB
28 region, which is also subdivided into two regions; and the gas-exchange tissues,- which are
29 further defined as the alveolar-interstitial (Al) region but are exactly comparable to the
30 pulmonary or A region. The fourth region is the lymph nodes. Deposition in the four regions is
31 given as a function of particle size with two different types of particle size parameters: activity
32 median thermodynamic diameter (AMTD) for deposition of particles ranging in size from 0.0005
33 to 1.0 um and the activity median aerodynamic diameter (AMAD) for deposition of particles
34 from 0.1 to lOOum. Reference values of regional deposition are provided and guidance is given
^P for extrapolating to specific individuals and populations under different levels of activity. This
36 model also includes consideration of particle inhalability, a measure of the degree to which
7/25/00 3-31 DRAFT—DO NOT CITE OR QUOTE
-------
1 particles can enter the respiratory tract and be available for deposition. After deposition occurs in
2 a given region, two different intrinsic clearance processes act competitively on the deposited
3 particles: particle transport, including mucociliary clearance from the respiratory tract and
4 physical clearance of particles to the regional lymph nodes; and absorption, including movement
5 of material to blood and both dissolution-absorption and transport of ultrafine particles. Rates of
6 particle clearance derived from studies with human subjects are assumed to be the same for all
7 types of particles. The ICRP model provides average concentration or average number values on
8 a regional basis, i.e., mass or number deposited or retained in the ET, TB, or A regions.
9 Additionally, while the ICRP66 model was developed primarily for use with airborne radioactive
10 particles and gases in humans, its use for describing the dosimetry of inhaled mass of
11 nonradioactive substances in humans is also appropriate.
12 An alternative new human respiratory tract dosimetry model that developed concurrently
13 with the new ICRP model is being proposed by the National Council on Radiation Protection
14 (NCRP). This model was described in outline by Phalen et al. (1991). As with the 1994 ICRP66
15 model (ICRP66, 1994), the proposed NCRP model addresses (1) inhalability of particles, (2) new
16 subregions of the respiratory tract, (3) dissolution-absorption as an important aspect of the
17 model, and (4) body size (and age). The proposed NCRP model defines the respiratory tract in
18 terms of a naso-oro-pharyngo-laryngeal (NOPL) region, a TB region, a pulmonary (P) region,
19 and the lung-associated lymph nodes (LN). The rates of dissolution-absorption of particles and
20 their constituents are derived from clearance data from humans and laboratory animals. The
21 effect of body growth on particle deposition is also considered in the model, but particle
22 clearance rates are assumed to be independent of age. The NCRP model does not consider the
23 fate of inhaled materials after they leave the respiratory tract. Although the proposed NCRP
24 model describes respiratory tract deposition, clearance, and dosimetry for radioactive substances
25 inhaled by humans, the model can also be used for evaluating inhalation exposures to all types of
26 particles. Graphical outputs of regional deposition fractions from both the ICRP66 (1994) and
27 draft NCRP models presented in U.S. EPA (1996) indicate approximately 15% would be
28 deposited in the alveolar region at the MMAD of DPM, 0.2 jim.
29
30 3.6=2.3. Animal Models
31 Strom et al. (1988) developed a multicompartmental model for particle retention that
32 partitioned the alveolar region into two compartments on the basis of the physiology of clearance.
33 The alveolar region has a separate compartment for sequestered macrophages, corresponding to
34 phagocytic macrophages that are heavily laden \vith particles and clustered, and consequently
35 have significantly !cv/ered mobility. The rncde! has the following compartments:
36 (1) tracheobronchial tree. (2) free paniculate on the alveolar surface, (3) mobile phagocytic
7/25/00 3-32 DRAFT—DO NOT CITE OR QUOTE
-------
alveolar macrophages, (4) sequestered particle-laden alveolar macrophages, (5) regional lymph
nodes, and (6) gastrointestinal tract. The model is based on mass-dependent clearance (the rate
3 coefficients reflect this relationship), which dictates sequestration of particles and their eventual
4 transfer to the lymph nodes. The transport rates between various compartments were obtained by
5 fitting the calculated results to lung and lymph node burden experimental data for both exposure
6 and postexposure periods. Because the number of fitted parameters was large, the model is not
7 likely to provide unique solutions that would simulate experimental data from various sources
8 and for different exposure scenarios. For the same reason, it is not readily possible to use this
9 model for extrapolating to humans.
10 Stober and co-workers have worked extensively in developing models for estimating
11 retention and clearance of relatively insoluble respirable particles (as DPM) in the lung. Their
12 most recent work (1994), a revised version of the POCK model, is a rigorous attempt to
13 incorporate most of the physiologically known aspects of alveolar clearance and retention of
14 inhaled relatively insoluble particles. Their multicompartmental kinetics model has five
15 subcompartments. The transfer of particles between any of the compartments within the alveolar
16 region is macrophage mediated. There are two compartments that receive particles cleared from
17 the alveolar regions: the TB tract and the lymphatic system. The macrophage pool includes both
mobile and particle-laden immobilized macrophages. The model assumes a constant maximum
volume capacity of the macrophages for particle uptake and a material-dependent critical
20 macrophage load that results in total loss of macrophage mobility. Sequestration of those
21 macrophages heavily loaded with a particle burden close to a volume load capacity is treated in a
22 sophisticated manner by approximating the particle load distribution in the macrophages. The
23 macrophage pool is compartmentalized in terms of numbers of macrophages that are subject to
24 discrete particle load intervals. Upon macrophage death, the phagocytized particle is released
25 back to the alveolar surface; thus phagocytic particle collection competes to some extent with
26 this release back to the alveolar surface. This recycled particle load is also divided into particle
27 clusters of size intervals defining a cluster size distribution on the alveolar surface. The model
28 yields a time-dependent frequency distribution of loaded macrophages that is sensitive to both
29 exposure and recovery periods in inhalation studies.
30 The POCK model also emphasizes the importance of interstitial burden in the particle
31 overload phenomenon and indicates that particle overload is a function of a massive increase in
32 particle burden of the interstitial space rather than total burden of the macrophage pool. The
33 relevance of the increased particle burden in the interstitial space lies with the fact that this
34 compartmental burden is not available for macrophage-mediated clearance and, therefore,
persists even after cessation of exposure.
7/25/00 3-33 DRAFT—DO NOT CITE OR QUOTE
-------
1 Although the POCK model is the most sophisticated in the physiological complexity it
2 introduces, it suffers from a major disadvantage. Experimental retention studies provide data
3 only on total alveolar and lymph node mass burdens of the particles as a function of time. The
4 relative fraction of the deposition between the alveolar subcompartments in the Stober model
5 therefore cannot be obtained experimentally; the model thus uses a large number of parameters
6 that are simultaneously fit to experimental data. Although the model predictions are tenable,
7 experimental data are not currently available to substantiate the proposed compartmental burdens
8 or the transfer rates associated with these compartments. Thus, overparameterization in the
9 model leads to the possibility that the model may not provide a unique solution that may be used
10 for a variety of exposure scenarios, and for the same reason, cannot be used for extrapolation to
11 humans. Stober et al. have not developed an equivalent model for humans; therefore the use of
12 their model in our risk assessment for diesel is not attempted.
13
14 3.6.2.4. Combined Models (for Interspecies Extrapolation)
15 Currently available data for long-term inhalation exposures to poorly soluble particles
16 (e.g., TiO2, carbon black, and DPM) show that pulmonary retention and clearance of these
17 particles are not adequately described by simple first-order kinetics and a single compartment
18 representing the alveolar macrophage particle burden. A two-compartment lung model that
19 could be applied to both humans and animals was developed by Smith (1985) and includes
20 alveolar and interstitial compartments. For uptake and clearance of particles by alveolar surface
21 macrophages and interstitial encapsulation of particles (i.e., quartz dust), available experimental
22 data show that the rate-controlling functions followed Michaelis-Menton type kinetics, whereas
23 other processes affecting particle transfer are assumed to be linear. The model was used in an
24 attempt to estimate interstitial dust and fibrosis levels among a group of 171 silicon carbide
25 workers; the levels were then compared with evidence of'fibrosis from chest radiographs. A
26 significant correlation was found between estimated fibrosis and profusion of opacities on the
27 radiographs. This model provides as many as seven different rate constants derived by various
28 estimations and under various conditions from both animal and human sources. The model was
29 intended for estimation of generalized dust described only as respirable without any other regard
30 to sizing for establishing the various particle-related rate constants. As most of the described
31 functions could not be validated with experimental data, the applicability of this model,
32 especially for particulates in the size range of DPM, was unclear.
33 Yu et al. (1991; also reported as Yu and Yoon, 1990) have developed a three-
34 compartment lung model that consists of trachenbrnnchial (T), alveolar (A), and lymph node (L)
35 compartments (Appendix A. Figure A-l) and: in aHHitinn^ ^opsidered filtration by a
36 nasopharyngeal or head (H) compartment. Absorption by the blood (B) and gastrointestinal (G)
7/25/00 3-34 DRAFT—DO NOT CITE OR QUOTE
-------
compartments was also considered. Although the treatment of alveolar clearance is
physiologically less sophisticated than that of the Stober et al. model, the Yu model provides a
3 more comprehensive treatment of clearance by including systemic compartments and the head,
4 and including the clearance of the organic components of DPM in addition to the relatively
5 insoluble carbon core.
6 The tracheobronchial compartment is important for short-term considerations, whereas
7 long-term clearance takes place via the alveolar compartment. In contrast to the Stober and
8 Strom approaches, the macrophage compartment in the Yu model contains all of the
9 phagocytized particles; that is, there is no separate (and hypothetical) sequestered macrophage
10 subcompartment. Instead, in order to progress beyond the classical human ICRP66 retention
11 model, Yu has addressed the impairment of long-term clearance (the overload effect) by using a
12 set of variable transport rates for clearance from the alveolar region as a function of the mass of
13 DPM in the alveolar compartment. A functional relationship for this was derived mathematically
14 (Yu et al., 1989) based upon Morrow's hypothesis for the macrophage overload effect discussed
15 earlier in the section on pulmonary overload. The extent of the impairment depends on the initial
16 particle burden, with greater paniculate concentration leading to slower clearance.
17 Within this model DPM is treated as being composed of three material components: a
« relatively insoluble carbonaceous core, slowly cleared organics (10% particle mass), and fast-
cleared organics (10% particle mass). Such a partitioning of organics was based on observations
20 that the retention of particle-associated organics in lungs shows a biphasic decay curve (Sun et
21 al., 1984; Bond et al., 1986). For any compartment, each of these components has a different
22 transport rate. The total alveolar clearance rate of each material component is the sum of
23 clearance rates of that material from the alveolar to the tracheobronchial, lymph, and blood
24 compartments. In the Strom and Stober models discussed above, the clearance kinetics of DPM
25 were assumed to be entirely dictated by those of the relatively insoluble carbonaceous core. For
26 those organic compounds that get dissociated from the carbon core, clearance rates are likely to
27 be very different, and some of these compounds may be metabolized in the pulmonary tissue or
28 be absorbed by blood.
29 The transport rates for the three components were derived from experimental data for rats
30 using several approximations. The transport rates for the carbonaceous core and the two organic
31 components were derived by fitting to data from separate experiments. Lung and lymph node
3 2 burdens from the experiment of Strom et al. (1988) were used to determine the transport rate of
33 the carbonaceous core. The Yu model incorporates the impairment of clearance by including a
34 mass dependency in the transport rate. This mass dependency is easily extracted because the
animals in the experiment were killed over varying periods following the end of exposure.
7/25/00 3-35 DRAFT—DO NOT CITE OR QUOTE
-------
1 It was assumed that the transport rates from the alveolar and lymph compartments to the
2 blood were equal and independent of the paniculate mass in the alveolar region. The clearance
3 rates of particle-associated organics for rats were derived from the retention data of Sun et al.
4 (1984) for benzo [ajpyrene and the data of Bond et al. (1986) for nitropyrene adsorbed on diesel
5 particles.
6 In their model Yu et al. (1991) make two important assumptions to carry out the
7 extrapolation in consideration of inadequate human data. First, the transport rates of organics in
8 the DPM do not change across species. This is based upon lung clearance data of inhaled
9 lipophilic compounds (Schanker et al., 1986), where the clearance was seen to be dependent on
10 the lipid/water partition coefficient. In contrast, the transport rate of the carbonaceous core is
11 considered to be significantly species dependent (Bailey et al., 1982). DPM clearance rate is
12 determined by two terms in the model (see Equation A-82 in Appendix A). The first,
13 corresponding to macrophage-mediated clearance, is a function of the lung burden and is
14 assumed to vary significantly across species. The second term, a constant, corresponding to
15 clearance by dissolution, is assumed to be species independent. The mass-dependent term for
16 humans is assumed to vary in the same proportion as in rats under the same unit surface
17 particulate dose. The extrapolation is then achieved by using the data of Bailey et al. (1982) for
18 the low lung burden limit of the clearance rate. This value of 0.0017/day was lower than the rat
19 value by a factor of 7.6. This is elaborated further in Appendix A. Other transport rates that
20 have lung burden dependence are extrapolated in the same manner.
21 The Bailey et al. (1982) experiment, however, used fused monodisperse aluminosilicate
22 particles of 1.9 and 6.1 um aerodynamic diameters. Yu and co-workers have used the longer of
23 the half-times obtained in this experiment; in using such data for DPM 0.2 um in diameter, they
24 have assumed the clearance of relatively insoluble particles to be independent of size over this
25 range. This appears to be a reasonable assumption because the linear dimensions of an alveolar
26 macrophage are significantly larger, roughly 10 um (Yu et al., 1996). However, Snipes (1979)
27 has reported a clearance rate (converted here from half-time values) of 0.0022/day for 1 and
28 2 (am particles but a higher value of 0 0039/day for 0.4 um particles. In the absence of reliable
29 data for 0.2 um particles, clearance rate pertaining to this much larger particle size is being used.
30 Although such a choice may underestimate the correct clearance rate for DPM, the resulting error
31 in the output (i.e., a human equivalent concentration) is likely to be only more protective of
32 human health. Long-term clearance rates for particle sizes more comparable to DPM are
33 available, e.g., iron oxide and polystyrene spheres (Waite and Ramsden, 1971; Jammet et al.,
34 1978). but these data show a large range in the values obtained for half lives or are based upon a
35 very small number nf trial-:, anH thprpfnre compare unfavorably v/ith the quality of data fium luc
36 Bailey experiment.
7/25/00 3-36 DRAFT—DO NOT CITE OR QUOTE
-------
The deposition fractions of paniculate matter in the pulmonary and tracheobronchial
regions of the human lung remain relatively unchanged over the particle size range between
3 0.2 and 1.0 um, on the basis of the analysis done with the ICRP66 (1994) model as documented
4 in the PMCD (U.S. EPA, 1996). As the clearance of relatively insoluble particles is also likely to
5 remain the same over this range, the dosimetry results in this report for the carbonaceous core
6 component of DPM could also be extended to other particles in this size range within the PM, 5
7 For respirable particles with diameters larger than this range, e.g., between 1.0 and 3.5 um, the
8 extent of the fraction deposited in the pulmonary region is unclear. Results from the ICRP66
9 (1994) model predict little change in human deposition for this diameter range, whereas the
10 earlier model of Yu and Diu (1983) predicts a significant increase. It is therefore unclear if either
11 model would be applicable for particles in this range without changing the value for the
12 deposition fractions. As mentioned above, however, regional deposition fractions from both the
13 ICRP66 (1994) and draft NCRP models presented in U.S. EPA (1996) indicate approximately
14 15% would be deposited in the alveolar region at the MMAD of DPM, 0.2 [im. These values
15 compare favorably with the human alveolar deposition in humans specific for DPM, which has
16 been extimated with the Yu model to be 7% to 13% (Yu and Xu, 1986).
17 Although there was good agreement between experimental and modeled results, this
« agreement follows a circular logic (as adequately pointed out by Yu and Yoon [1990]) because
the same experimental data that figured into the derivation of transport rates were used in the
20 model. Nevertheless, even though this agreement is not a validation, it provides an important
21 consistency check on the model. Further experimental data and policy definitions on what
22 constitutes validation would be necessary for a more formal validation.
23 The model showed that at low lung burdens, alveolar clearance is dominated by
24 mucociliary transport to the tracheobronchial region, and at high lung burdens, clearance is
25 dominated by transport to the lymphatic system. The head and tracheobronchial compartments
26 showed quick clearance of DPM by mucociliary transport and dissolution. Lung burdens of both
27 the carbonaceous core and organics were found to be greater in humans than in rats for similar
28 periods of exposure.
29 The Yu and Yoon (1990) version of the model provides a parametric study of the
30 dosimetry model, examining variation over a range of exposure concentrations, breathing
31 scenarios, and ventilation parameters; particle mass median aerodynamic diameters; and
32 geometric standard deviations of the aerosol size distribution. It examines how lung burden
33 varies with age for exposure over a lifespan, provides dosimetry extrapolations to children, and
34 examines changes in lung burden with lung volume. The results showed that children would
^^ exhibit more diminished alveolar clearance of DPM at high lung burden than adults when
36 exposed to equal concentrations of DPM. These features make the model easy to use in risk
7/25/00 3-37 DRAFT—DO NOT CITE OR QUOTE
-------
1 assessment studies. The reader is referred to Appendix A for further details on the model and for
2 analyses of the sensitivity of the model to change in parameter values.
3 The Yu model presents some uncertainties in addition to those discussed earlier in the
4 context of particle size dependence of clearance rate. The reports of Yu and Yoon (1990) as well
5 as Yu et al. (1991) underwent extensive peer review; we list below the most important among the
6 model uncertainties discussed by the review panel. The experimental data used by the Yu model
7 for adsorbed organics used passively adsorbed radiolabeled compounds as surrogates for
8 combustion-derived organics. These compounds may adhere differently to the carbon core than
9 do those formed during combustion. Yu has estimated that slowly cleared organics represent 10%
10 of the total particle mass; the actual figure could be substantially less; the reviewers estimate that
11 the amount of tightly bound organics is probably only 0.1% to 0.25% of the particle mass.
12 The model was based upon the experimental data of Strom et al. (1988), where
13 Fischer-344 rats were exposed to DPM at a concentration of 6.0 mg/m3 for 20 h/day and 7
14 days/week for periods ranging from 3 to 84 days. Such exposures lead to particle overload effects
15 in rats, whereas human exposure patterns are usually to much lower levels at which overload will
16 not occur. Parameters obtained by fitting to data under the conditions of the experimental
17 scenario for rats may not be optimal for the human exposure and concentration of interest.
18 The extrapolation of retained dose from rats to humans assumed that the macrophage-
19 mediated mechanical clearance of the DPM varies with the specific particulate dose to the
20 alveolar surface in the same proportion in humans and in rats, whereas clearance rates by
21 dissolution were assumed to be invariant across species. This assumption has not been validated.
22 It should also be noted that the Yu et al. (1991) model does not possess an interstitial
23 compartment. The work of Nikula et al. (1997a,b) and of Kuempel (2000) provide compelling
24 information on the significance of an extensive interstitilization process in primates and in
25 humans. Kuempel (2000) developed a lung dosimetry model to describe the kinetics of particle
26 clearance and retention in coal miners' lungs. Models with overloading of lung clearance, as
27 observed in rodent studies, were found to be inadequate to describe the end-of-life lung dust
28 burdens in those miners. The model that provided the best fit to the human data included a
29 sequestration process representing the transfer of particles to the interstitium. These findings are
30 consistent with a study showing reduced lung clearance of particles in retired coal miners
31 (Freedman and Robinson, 1988) and with studies showing increased retention of particles in the
32 lung interstitium of humans and nonhuman primates compared to rodents exposed to coal dust
33 and/or diesel exhaust (Nikula et al., 1997a,b). Because the Yu model has not been validated on
34 human data and does not include an interstitial compartment, it is acknowledged that this mode!
35 may therefore underpredict the lung dust burdens in humans exposed tc occupational levels cf
36 dust. However, it is also not known whether the model based on coal miner data (Kuempel,
7/25/00 3-38 DRAFT—DO NOT CITE OR QUOTE
-------
2000) would also describe the clearance and retention processes in the lungs of humans with
exposures to particles at lower environmental concentrations, or to submicrometer particles such
~3 as diesel exhaust particulate. Further investigation of these issues is needed.
4
5 3.6.2.5. Use of the Yu et al (1991) Model for Interspecies Extrapolation
6 In addressing the objectives of this chapter, i.e., consideration of what is known and
7 applicable to DPM concerning particle disposition and the bioavailability of adsorbed organics
8 on DPM, it is apparent that the database is considerable for both the processes involved in
9 particle dosimetry and for DPM. This information makes the goal of predicting a human internal
10 dose from animal data through a model utilizing this database both feasible and appropriate.
11 In their charge to EPA through "Science and Judgment in Risk Assessment" (NRC,
12 1995), the National Research Council opines that EPA should have principles for judging when
13 and how to depart from default options. The extensive data presented in this chapter (including
14 the model of Yu), their scientific validity, and the limitations of the current default procedures
15 provide a basis for departing from the default options currently identified by the Agency for
16 extrapolating from animals to humans. The default option of assuming external concentrations
17 of DPM in animal studies as being representative of a human concentration (and an equivalent
18 internal dose) is clearly not adequate given the vast differences in the basic processes of
^P deposition and clearance between animals and humans documented by these data. Use of an
20 alternate default option, the Agency's dosimetric adjustment procedures for inhaled particles in
21 animal-to-human scenarios (described in U.S. EPA, 1994), is also inadequate as only deposition
22 is predicted and then only down to an MM AD of 0.5 um, whereas the MMAD of DPM is
23 typically 0.2 um or smaller. Models have been described in this section that consider both
24 deposition and retention specifically for DPM in both laboratory animals and in humans. These
25 points provide justification for moving away from default options and utilizing the best scientific
26 information available (i.e., that integrated into deposition/clearance models) in performing the
27 animal-to-human extrapolation.
28 Of the models evaluated in this chapter, that of Yu et al. (1991) is uniquely equipped to
29 perform animal-to-human extrapolation for DPM. The model structure is parsimonious, with
30 three lung compartments (tracheobroncial, pulmonary, lymph node). Design of the model
31 incorporated both human and animal information, utilizing empirical clearance data from both
32 rats and humans. In addition to DPM, this model considers deposition and clearance of two
33 classes of organics adsorbed onto DPM. The model does have limitations, such as a lack of
34 definitive information on variability of the results and absence of a lung compartment
(interstitial) that could well be of importance to humans. It is, however, considered that the
7/25/00 3-39 DRAFT—DO NOT CITE OR QUOTE
-------
1 attributes considerably outweigh the detractions in choosing this model as a means to perform
2 animal-to-human extrapolation for DPM.
3
4 3.7. SUMMARY
5 The most consistent historical measure of exposure for diesel exhaust is DPM in units of
6 fig or mg particles/m3, with the underlying assumption that all components of diesel emissions
7 (e.g., organics in the form of volatilized liquids or gases) are present in proportion to the DPM
8 mass. DPM is used as the basic dosimeter for effects from various scenarios such as chronic and
9 acute exposures as well as for different endpoints such as irritation, fibrosis, or even cancer.
10 There is, however, little evidence currently available to prove or refute DPM as being the most
11 appropriate dosimeter.
12 DPM dose to the tissue is related to the extent of the deposition and clearance of DPM.
13 DPM may deposit throughout the respiratory tract via sedimentation or diffusion, with the latter
14 being prevalent in the alveolar region. Particles that deposit upon airway surfaces may be cleared
15 from the respiratory tract completely or may be translocated to other sites by regionally distinct
16 processes that can be categorized as either absorptive (i.e., dissolution) or nonabsorptive (i.e.,
17 transport of intact particles via mucociliary transport). With poorly soluble particles such as
18 DPM, clearance by dissolution is insignificant compared to the rate of clearance as an intact
19 particle. Other mechanisms that can affect retention of DPM include endocytosis by alveolar
20 lining cells and interstitialization, which lead to the accumulation of DPM in the interstitial
21 compartment of the lung and subsequent translocation of DPM to lymph nodes; interstitialization
22 of poorly soluble particles is prominent in primates and humans as compared to rodents. For
23 poorly soluble particles such as DPM, species-dependent rate constants exist for the various
24 clearance pathways that can be modified by factors such as respiratory tract disease.
25 In rats, prolonged exposure to high concentrations of particles may be associated with
26 particle overload, a condition that is defined as the overwhelming of macrophage-
27 mediated clearance by the deposition of particles at a rate exceeding the capacity of that
28 clearance pathway. This condition seems to begin to occur in rats when the pulmonary dust
29 burden exceeds about 1 mg particles/g lung tissue. On the other hand, there is no clear evidence
30 for particle overload in humans. Macrophage-mediated clearance appears to be slower and
31 perhaps less important in humans than in rats, and interstitialization of poorly soluble particulate
32 matter may be of greater consequence in humans than in rats.
33 The degree of bioavailability of the organic fraction of DPM is still somewhat uncertain.
34. However, reports of F)NA alteration.? in Qcr.iipatirmaJly exposed workers, as well as results of
"3R animal <5tiiHi<=»«: iisincr raHmlahplf(\ nroanir*: HfnncitpH rm F)PM inHir*alY» that at l(=act a fr->ftir\n r»f
— - ^ • t^f - t _.~ ...._-..., .- . _..»* - « ...«._ .—-
36 the organics present are eluted prior to particle clearance. Carcinogenic organics eluted in
7/25/00 3-40 DRAFT—DO NOT CITE OR QUOTE
-------
regions where diffusion may be a relatively long process, such as in the conducting airways vs
the alveolar region, may remain in the lung long enough to be metabolized to an active form or to
3 interact directly with vital cellular components. The current information suggests that DPM-
4 associated organics could be involved in a carcinogenic process, although the quantitative data
5 are far from adequate to make any firm conclusions.
6 Use of laboratory animal data in an assessment meant to be applied to humans obligates
7 some form of interspecies extrapolation. Review and evaluation of the considerable, specific
8 database in humans and animals on disposition of DPM, its adsorbed organics, and other poorly
9 soluble particles led to the judgment that default options available for interspecies dosimetry
10 adjustment could be set aside for more scientifically valid, DPM-specific processes. Refinement
11 of this evaluation led to the identification and choice of the Yu et al. (1991) model to conduct
12 interspecies extrapolation. This model has a three-compartment lung consisting of
13 tracheobronchial, alveolar, and lymph node compartments. It treats DPM as being composed of
14 the insoluble carbonaceous core, slowly cleared organics, and fast-cleared organics, and
15 considers in an integrative manner the simultaneous processes of both deposition and clearance
16 through empirical data derived from both laboratory animals and humans. Also, the model has
17 some limited consideration of model variability in its outputs describing dose to the lung. Major
assumptions made in this model include that transport rates of organics in DPM do not change
across species and that the transport rate of the carbonaceous core is species dependent, with the
20 clearance rate varying with the dose to the alveolar surface in the same proportion in humans as
21 in rats. Limitations of the model include the lack of definitive information on variability and the
22 lack of a biological compartment (the interstitium) that may be of consequence in humans. The
23 basis of this model is to derive an internal dose from an external DPM concentration by utilizing
24 species-specific physiological and pharmacokinetic parameters and, as such, is considered to
25 have addressed the pharmacokinetic aspects of interspecies dosimetry. This aspect of the model
26 addresses some of the critical data needs for the quantitative analysis of noncancer effects from
27 DPM, the subject of Chapter 6.
28 As parallels have been drawn between DPM and PM2 5 in other chapters, it is perhaps
29 appropriate to compare them also from the aspect of dosimetry. Obvious comparisons include
30 the nature of the particle distribution, defined artificially for PM25 as compared with the thorough
31 characterization of DPM for both MMAD (which, at around 0.2 flm, is typically more than an
32 order of magnitude less than the PM25 cutoff) and geometric standard deviation. It is clear that a
33 larger portion of PM25 particles than DPM would be above the aerodynamic equivalent diameter
34 (dae) of 0.5 (1m, which is often considered as a boundary between diffusion and aerodynamic
^ft mechanisms of deposition. This would imply that a somewhat larger portion of DPM may pass
36 on to the lower respiratory tract than would PM2 5. Alveolar depostion in humans specific for
7/25/00 3-41 DRAFT—DO NOT CITE OR QUOTE
-------
1 DPM has been estimated with the Yu model to be 7%-13% (Yu and Xu, 1986). This fractional
2 deposition may be compared to one calculated for PM2S and reported in U.S. EPA (1996a);
3 assuming a MMAD of 2.25 \Lm and a geometric standard deviation of 2.4, a fractional alveolar
4 deposition of 10.2% was reported. This value is within the range and quite comparable to that
5 obtained by Yu and Xu (1986), indicating that little difference may exist in alveolar deposition
6 between DPM and PM2 5, at least for this assumed geometric standard deviation.
7/25/UU 3-42 DRAFT—DO NOT CITE OR QUOTE
-------
Table 3-1. Predicted doses of inhaled DPM per minute based on total lung volume
(M), total airway surface area (M,), or surface area in alveolar region (M2)
M , M, M2
Species (IP'3 ug/min/cm3) (IP"6 yig/min/cm2) (10"6 ug/min/cm2)
Hamster
Fischer rat
Human
3.548
3.434
0.249
3.088
3.463
1.237
2.382
2.608
0.775
M = mass DPM deposited in lung per minute
total lung volume
M = mass DPM deposited in lung per minute
total airway surface area
M. = mass DPM deposited on the unciliated airways per minute
surface area of the unciliated airways
Based on the following conditions: (1) mass median aerodynamic diameter (MMAD) = 0.2 urn; geometric standard
deviation (og) = 1.9; packing density (4>) = 0.3; and particle mass density (p) = 1.5 g/cm3; (2) particle concentration =
1 mg/m3; and (3) nose-breathing. For humans, total lung volume = 3200 cm3, total airway surface area = 633,000
cm3, surface area of the unciliated airways = 627,000 cm3.
Source: Xu and Yu, 1987.
7/25/00 3-43 DRAFT—DO NOT CITE OR QUOTE
-------
^
to
O
O
Table 3-2. Alveolar clearance in laboratory animals exposed to DPM
D
O
O
2
3
O
HH
3
Speci.'s/sex
Rats, 7-344, M
Rats, --344
Rats
Rats, F-344, MF
Rats, F-344;
Guinea pigs,
Hartley
Rats, F-344
Exposure
technique
Nose only;
Radioiabelec! DPM
Whole body;
assessed effect
on clearance of
67Ga2Oj particles
Whole body
Whole body
Nose-only;
Radiolabeled I4C
Exposure
duration
40-45 min
1 h/day
5 days/week
24 mo
19 h/day
5 days/week
2.5 years
1 h/day
5 days/week
18 mo
45 min
140 min
45 min
20 h/day
7 days/week
7- 112 days
Particles
mg/m3
6
0.35
3.5
7.1
4
0.15
0.94
4.1
7
2
7
0.25
6
Observed effects
Four days after exposure, 40% of DPM eliminated by
mucociliary clearance. Clearance from lower RT was in
2 phases. Rapid mucociliary (t|/2 = 1 day); slower
macrophage-mediated (t,/2 = 62 days).
T, significantly higher with exposure to 7.1 mg/m3 for
24 mo; t2 significantly longer after exposure to 7.1 mg/m3
for 6 mo and to 3.5 mg/m3 for 18 mo.
Estimated alveolar deposition = 60 mg; particle burden
caused lung overload. Estimated 6-15 mg particle-bound
organics deposited.
Long-term clearance was 87 ± 28 and 99 ± 8 days for
0.15 and 0.94 mg/m3 groups, respectively; t,/2 = 165 days
for 4. 1 mg/m3 group.
Rats demonstrated 3 phases of clearance with t1/2 = 1, 6,
and 80 days, representing tracheobronchial, respiratory
bronchioles, and alveolar clearance, respectively. Guinea
pigs demonstrated negligible alveolar clearance from
day 10 to 432.
Monitored rats for a year. Proposed two clearance models.
Clearance depends on initial particle burden; tl/2 increases
with higher exposure. Increases in tl/2 indicate increasing
impairment of AM mobility and transition into overload
condition.
Reference
Chan etal.( 1981)
Wolff etal.( 1986,
1987)
Heinrich et al.
(1986)
Griffisetal.(I983)
Lee etal.( 1983)
Chan etal. (1984)
RT = iespi:-atory tract.
AM = alveolar macrophage.
T, = clearance from primary, ciliated airways.
T2 = cl .-arajice from nonciliated passages.
-------
c 10
-2
05
L
o 10
Q.
O
Q
10
-4L
Hamster
Human
Fischer rat
4 8 12 16 20
Generation Number
24
Figure 3-1. Modeled deposition distribution patterns of inhaled diesel exhaust particles
in the airways of different species. Generation 1-18 are TB; >18 are A.
Source: Xu and Yu, 1987.
7/25/00
3-45
DRAFT—DO NOT CITE OR QUOTE
-------
1.0-1
Tracheobronchial
Deposition
Alveolar Deposition
40 60
Hours After Inhalation
Figure 3-2. Modeled clearance of poorly soluble 4-um particles deposited
in tracheobronchial and alveolar regions in humans.
Source: Cuddihy and Yeh, 1986.
7/25/00
3-46
DRAFT—DO NOT CITE OR QUOTE
-------
1.0
V,
Range of Three
Measu reme nts
(0
i. 0.8-
0)
Q
g 0.6-
o
ig 0.4-
'E
^r — ~r — — -—— -— — — — ._,
^^~ ~ — ~r~~~~^
i
*~ Model Projection
Same as Lower
Limit of Range
i
*o
ts
2
0
I
! i i i i 1 1 1 1 1 ! 1
D 20 40 60 80 100 120
Hours After Inhalation
Figure 3-3. Short-term thoracic clearance of inhaled particles as determined by model
prediction and experimental measurement.
Source: Cuddihy and Yeh, 1986 (from Stahlhofen et al., 1980).
7/25/00
3-47
DRAFT—DO NOT CITE OR QUOTE
-------
100
o
s
High
40
60
80 100 120
Time (Days)
140
160
i i
180
200
Figure 3-4. Clearance from lungs of rats of 134Cs-FAP fused aluminosilicate tracer
particles inhaled after 24 months of diesel exhaust exposure at concentrations of 0
(control), 0.35 (low), 3.5 (medium), and 7.1 (high) mg DPM7m3.
Source: Wolff etal., 1987.
3-48
DRAFT—DO NOT CITE OR QUOTE
-------
20 -i
O)
Q.
Q.
c
10-
i
6
i
12
18
24
Months Of Exposure
Figure 3-5. Lung burdens of DPM within rats exposed to 0.35 (low) (•), 3.5 (medium)
(A), and 7.1 (high) mg ppm/m3 (•).
Source: Wolff etal., 1987.
7/25/00
3-49
DRAFT—DO NOT CITE OR QUOTE
-------
1 3.8. REFERENCES
2
3 Adamson, IYR; Bowden, DH. (1978) Adaptive responses of the pulmonary macrophagic system to carbon: II.
4 Morphologic studies. Lab Invest 38:430-438.
5
6 Adamson, IYR; Bowden, DH. (1981) Dose response of the pulmonary macrophagic system to various particulates
7 and its relationship to transepithelial passage of free particles. Exp Lung Res 2:165-175.
8
9 Anjilvel, S; Asgharian, B. (1995) A multiple-path model of particle deposition in the rat lung. Fundam Appl Toxicol
10 28:41-50.
11
12 Aronson, M. (1963) Bridge formation and cytoplasmic flow between phagocytic cells. J Exp Med 118:1083-1088.
13
14 Bailey, MR; Fry, FA; James, AC. (1982) The long-term clearance kinetics of insoluble particles from the human
15 lung. Ann Occup Hyg 26:273-290.
16
17 Bailey, MR; Fry, FA; James, AC. (1985) Long-term retention of particles in the human respiratory tract. J Aerosol
18 Sci 16:295-305.
19
20 Ball, LM; King, LC. (1985) Metabolism, mutagenicity, and activation of 1-nitropyrene in vivo and in vitro. Environ
21 Int 11:355-361.
22
23 Battigelli, MC; Hengstenberg, F; Mannela, RJ; et al. (1966) Mucociliary activity. Arch Environ Health 12:460-466.
24
25 Bevan, DR; Ruggio, DM. (1991) Bioavailability in vivo of benzo[a]pyrene adsorbed to diesel particulate. Toxicol
26 Ind Health 7:125-139.
27
28 Bohning, DE; Conn, SH; Lee, HD; et al. (1980) Two-phase deep-lung clearance in man. In: Pulmonary toxicology of
29 respirable particles: proceedings of the nineteenth annual Hanford life sciences symposium; October 1979; Richland,
30 WA. Sanders, CL; Cross, FT; Dagle, GE; et al., eds. Washington, DC: U.S. Department of Energy; pp. 149-161.
31 Available from: NTIS, Springfield, VA; CONF-791002.
32
33 Bohning, DE; Atkins, HL; Conn, SH. (1982) Long-term particle clearance in man: normal and impaired. In: Inhaled
34 particles V: proceedings of an international symposium; September 1980; Cardiff, Wales. Walton, WH, ed. Ann
35 Occup Hyg 26:259-271.
36
37 Bond, JA; Mitchell, CE; Li, AP. (1983) Metabolism and macromolecular covalent binding of benzo[a]pyrene in
38 cultured Fischer-344 rat lung type II epithelial cells. Biochem Pharmacol 32:3771-3776.
39
40 Bond, JA; Sun, JD; Medinsky, MA; et al. (1986) Deposition, metabolism, and excretion of l-['4C]nitropyrene and
41 l-['4C]nitropyrene coated on diesel exhaust particles as influenced by exposure concentration. Toxicol Appl
42 Pharmacol 85:102-117.
43
44 Bond, JA; Johnson, NF; Snipes, MB; et al. (1990) DNA adduct formation in rat alveolar type II cells: cells
45 potentially at risk for inhaled diesel exhaust. Environ Mol Mutagen 16:64-69.
46
47 Brain, JD; Mensah, GA. (1983) Comparative toxicology of the respiratory tract. Am Rev Respir Dis 128:S87-S90.
48
49 Brooks, AL; Wolff, RK; Royer, RE; et al. (i98i) Deposition and biological availability of diesel particles and their
50 associated mutagenic chemicals. Environ Int 5:263-267.
51
52 Campbell, RM; Lee, ML. (1984) Capillary column gas chromatographic determination of nitro polycyclic aromatic
53 compounds in particulate extracts. Anal Chem 56:1026-1030.
54
7/25/00 3-50 DRAFT—DO NOT CITE OR QUOTE
-------
Chan, TL; Lee, PS; Hering, WE. (1981) Deposition and clearance of inhaled diesel exhaust particles in the
respiratory tract of Fischer rats. J Appl Toxicol 1:77-82.
Chan, TL; Lee, PS; Hering, WE. (1984) Pulmonary retention of inhaled diesel particles after prolonged exposures to
5 diesel exhaust. Fundam Appl Toxicol 4:624-631.
6
7 Cohen, D; Arai, SF; Brain, JD. (1979) Smoking impairs long-term dust clearance from the lung. Science
8 (Washington, DC) 204:514-517.
9
10 Cohen, BS; Xiong, JQ; Fang, CP; et al. (1998) Deposition of charged particles on lung airways. Health Phys
11 74:554-560.
12
13 Crapo, JD; Young, SL; Fram, EK; et al. (1983) Morphometric characteristics of cells in the alveolar region of
14 mammalian lungs. Am Rev Respir Dis 128:S42-S46.
15
16 Cuddihy, RG; Yeh, HC. (1986) Model analysis of respiratory tract clearance of particles inhaled by people. In:
17 Annual report of the Inhalation Toxicology Research Institute. Muggenburg, BA; Sun, JD, eds. Albuquerque, MM:
18 Lovelace Biomedical and Environmental Research Institute; report no. LMF-115; pp. 140-147.
19
20 Evans, MJ; Shami, SG; Martinez, LA. (1986) Enhanced proliferation of pulmonary alveolar macrophages after
21 carbon instillation in mice depleted of blood monocytes by Strontium-89. Lab Invest 54:154-159.
22
23 Felicetti, SA; Wolff, RK; Muggenburg, BA. (1981) Comparison of tracheal mucous transport in rats, guinea pigs,
24 rabbits, and dogs. J Appl Physiol: Respir Environ Exercise Physiol 51:1612-1617.
25
26 Ferin, J. (1982) Pulmonary alveolar pores and alveolar macrophage-mediated particle clearance. Anat Rec
27 203:265-272.
Ferin, J; Feldstein, ML. (1978) Pulmonary clearance and hilar lymph node content in rats after particle exposure.
Environ Res 16:342-352.
31
32 Freedman, AP; Robinson, SE. (1988) Noninvasive magnetopneumographic studies of lung dust retention and
33 clearance in coal miners. In: Respirable dust in the mineral industries: health effects, characterization, and control:
34 proceedings of the international symposium on respirable dust in the mineral industries; October 1986; University
35 Park, PA. Frantz, RL; Ramani, RV, eds. University Park, PA: Pennsylvania State University Press; pp. 181-186.
36
37 Freijer, JI; Cassee, FR; Subramaniam, R; et al. (1999) Multiple path particle deposition model developed by the
38 Chemical Industry Institute of Toxicology (CUT) and the Rijksinstituut Voor Volksgezondheid en Milieu (RIVM),
39 RIVM Publication No. 65001019.
40
41 Frey, JW; Corn, M. (1967) Physical and chemical characteristics of particulates in a diesel exhaust. Am Ind Hyg
42 AssocJ 28:468-478.
43
44 George, G; Hook, GER. (1984) The pulmonary extracellular lining. Environ Health Perspect 55:227-237.
45
46 Gerde, P; Medinsky, MA; Bond, JA. (199la) Particle-associated polycyclic aromatic hydrocarbons - a reappraisal of
47 their possible role in pulmonary carcinogenesis. Toxicol Appl Pharmacol 108:1-13.
48
49 Gerde, P; Medinsky, MA; Bond, JA. (1991b) The retention of polycyclic aromatic hydrocarbons in the bronchial
50 airways and in the alveolar region - a theoretical comparison. Toxicol Appl Pharmacol 107:239-252.
51
52 Gerde, P; Muggenberg, BA; Dahl, AR. (1999) Bioavailability, absorption and metabolism of diesel soot-absorbed
53 benzo(a)pyrene after single-breath exposures in dogs. In: Relationships between acute and chronic effects of air
pollution: 7th international inhalation symposium; February; Hannover, Federal Republic of Germany; p. 76.
7/25/00 3-51 DRAFT—DO NOT CITE OR QUOTE
-------
1 Gibb, FR; Morrow, PE. (1962) Alveolar clearance in dogs after inhalation of an iron 59 oxide aerosol. J Appl
2 Physiol 17:429-432.
3
4 Godleski, JJ; Steams, RC; Katler, MR; et al. (1988) Particle dissolution in alveolar macropbages assessed by
5 electron energy loss analysis using the Zeiss CEM902 electron microscope. J Aerosol Med 1:198-199.
6
7 Griffis, LC; Wolff, RK; Henderson, RF; et al. (1983) Clearance of diesel soot particles from rat lung after a
8 subchronic diesel exhaust exposure. Fundam Appl Toxicol 3:99-103.
9
10 Gurley, LR; Spall, WD; Valdez, JG; et al. (1988) An HPLC procedure for the analysis of proteins in lung lavage
11 fluid. Anal Biochem 172:465-478.
12
13 Harmsen, AG; Muggenburg, BA; Snipes, MB; et al. (1985) The role of macrophages in particle translocation from
14 lungs to lymph nodes. Science (Washington, DC) 230:1277-1280.
15
16 Health Effects Institute. (1995) Diesel exhaust: a critical analysis of emissions, exposure, and health effects: a special
1 7 report of the Institute's Diesel Working Group. Cambridge, MA: Health Effects Institute.
18
19 Heinrich, U; Muhle, H; Takenaka, S; et al. (1986) Chronic effects on the respiratory tract of hamsters, mice, and rats
20 after long-term inhalation of high concentrations of filtered and unfiltered diesel engine emissions. J Appl Toxicol
21 6:383-395.
22
23 Hemminki, K; Soderling, J; Ericson, P; et al. (1994) DNA adducts among personnel servicing and loading diesel
24 vehicles. Carcinogenesis 15:767-769.
25
26 Heppleston, AG. (1961) Observations on the disposal of inhaled dust by means of the double exposure technique. In:
27 Inhaled particles and vapours: proceedings of an international symposium; March-April 1960; Oxford, United
28 Kingdom. Davies, CN, ed. New York: Pergamon Press; pp. 320-326.
29
30 Heppleston, AG; Young, AE. (1973) Uptake of inert particulate matter by alveolar cells: an ultrastructural study. J
31 Pathol 111:159-164.
32
33 Heyder, J; Gebhart, J; Rudolf, G; et al. (1986) Deposition of particles in the human respiratory tract in the size range
34 0.005-15 urn. J Aerosol Sci 17:811-825.
35
36 Hiura, TS; Kaxzubowski, MP; Li, N; et al. (1999) Chemicals in diesel exhaust particles generate oxygen radicals
37 and induce apoptosis in macrophages. J Immunol 163:5582-5591.
38
39 Holt, PF. (1981) Transport of inhaled dust to extrapulmonary sites. J Pathol 133:123-129.
40
41 International Commission on Radiological Protection (ICRP). (1979) Limits for intakes of radionuclides by workers.
42 Oxford, United Kingdom: Pergamon Press; ICRP publication 30, part 1.
43
44 ICRP. (1994) Human respiratory tract model for radiological protection: a report of a task group of the International
45 Commission on Radiological Protection. Oxford, United Kingdom: Elsevier Science Ltd. (ICRP publication 66;
46 Annais of the ICRP: v. 24, r.os. 1-3).
47
48 International Life Sciences Institute (ILSI). (2000) ILSI Risk Science Institute Workshop: the relevance of the rat
49 lung response to particle overload for human risk assessment. Gardner, DE, ed. Inha! Toxicol I2(l-2):l-148.
50
51 Jammet, H; Drutel, P; Parrot, R; et al. (1978) Etude de 1'epuration pulmonaire chez 1'homme apres administration
52 d'aerosols de particules radioactives [Study of pulmonary function in man after administration of radioactive
53 particulate aerosols] P.adioprotecticn !3:143-166.
54
7/25/00 3-52 DRAFT—DO NOT CITE OR QUOTE
-------
Keane, MJ; Xing, S-G; Harrison JC; et al. (1991) Genotoxicity of diesel-exhaust particles dispersed in simulated
pulmonary surfactant. Mutat Res 260:233-238.
4 King, LC; Kohan, MJ; Austin, AC; et al. (1981) Evaluation of the release of mutagens from diesel particles in the
5 presence of physiological fluids. Environ Mutagen 3:109-121.
6
7 Kittelson, DB; Dolan, DF; Verrant, JA. (1978) Investigation of a diesel exhaust aerosol. Warrendale, PA: Society of
8 Automotive Engineers, Inc.; technical paper series no. 78109.
9
10 Kuempel, ED. (2000). Comparison of human and rodent lung dosimetry models for particle clearance and retention.
11 DrugChemToxicol23(l):203-222.
12
13 Lee, PS; Chan, TL; Hering, WE. (1983) Long-term clearance of inhaled diesel exhaust particles in rodents. J Toxicol
14 Environ Health 12:801-813.
15
16 Lee, KP; Trochimowicz, HJ; Reinhardt, CF. (1985) Transmigration of titanium dioxide (TiOJ particles in rats after
17 inhalation exposure. Exp Mol Pathol 42:331-343.
18
19 Lee, KP; Henry, NW, III; Trochimowicz, HJ.; et al. (.1986) Pulmonary response to impaired lung clearance in rats
20 following excessive TiO2 dust deposition. Environ Res 41:144-167.
21
22 Lee, KP; Ulrich, CE; Geil, RG; et al. (1988) Effects of inhaled chromium dioxide dust on rats exposed for two years.
23 Fundam Appi Toxicol 10:125-145.
24
25 Lehnert, BE. (1988) Distributions of particles in alveolar macrophages during lung clearance. J Aerosol Med
26 1:206-207.
27
Lehnert, BE. (1989) Rates of disappearance of alveolar macrophages during lung clearance as a function of
phagocytized paniculate burden [abstract]. Am Rev Respir Dis 139(suppl.):A161.
30
31 Lehnert, BE. (1990) Alveolar macrophages in a particle "overload" condition. J Aerosol Med 3(suppl. 1):S9-S30.
32
33 Lehnert, BE; Morrow, PE. (1985) Association of 59iron oxide with alveolar macrophages during alveolar clearance.
34 Exp Lung Res 9:1-16.
35
36 Lehnert, BE; Valdez, YE; Bomalaski, SH. (1988) Analyses of particles in the lung free cell, tracheobronchial lymph
37 nodal, and pleural space compartments following their deposition in the lung as related to lung clearance
38 mechanisms. In: Inhaled particles VI: proceedings of an international symposium and workshop on lung dosimetry;
39 September 1985; Cambridge, United Kingdom. Dodgson, J: McCallum, RI; Bailey, MR; et al., eds. Ann Occup Hyg
40 32(suppl. 1): 125-140.
41
42 Lehnert, BE; Cline, A; London, JE. (1989) Kinetics of appearance of polymorphonuclear leukocytes and their
43 particle burdens during the alveolar clearance of a high lung burden of particles. Toxicologist 9:77.
44
45 Leung, HW; Henderson, RF; Bond, JA; et al. (1988) Studies on the ability of rat lung and liver microsomes to
46 facilitate transfer and metabolism of benzo[a]pyrene from diesel particles. Toxicology 51:1-9.
47
48 Lippmann, M; Schlesinger, RB. (1984) Interspecies comparisons of particle deposition and mucociliary clearance in
49 tracheobronchial airways. J Toxicol Environ Health 13:441-469.
50
51 Lundborg, M; Lind, B; Camner, P. (1984) Ability of rabbit alveolar macrophages to dissolve metals. Exp Lung Res
52 7:11-22.
53
Marafante, E; Lundborg, M; Vahter, M; et al. (1987) Dissolution of two arsenic compounds by rabbit alveolar
macrophages in vitro. Fundam Appl Toxicol 8:382-388.
7/25/00 3-53 DRAFT—DO NOT CITE OR QUOTE
-------
1 Mauderly, JL. (1996) Lung overload: the dilemma and opportunities for resolution. In: Particle overload in the rat
2 lung and lung cancer: implications for human risk assessment Proceedings of a conference; March 1995;
3 Cambridge, MA. Mauderly, JL; McCunney, RJ; eds. New York: Taylor & Francis.
4
5 McClellan, RO; Brooks, AL; Cuddihy, RG; et al. (1982) Inhalation toxicology of diesel exhaust particles. In:
6 Toxicological effects of emissions from diesel engines: proceedings of the EPA diesel emissions symposium;
7 October 1981; Raleigh, NC. Lewtas, J, ed. New York: Elsevier Biomedical; pp. 99-120. (Developments in
8 toxicology and environmental science: v. 10).
9
10 Morrow, PE. (1966) International Commission on Radiological Protection (ICRP) task group on lung dynamics,
11 deposition and retention models for internal dosimetry of the human respiratory tract. Health Phys 12:173.
12
13 Morrow, PE. (1988) Possible mechanisms to explain dust overloading of the lungs. Fundam Appl Toxicol
14 10:369-384.
15
16 Morrow, PE. (1994) Mechanisms and significance of "particle overload." In: Toxic and carcinogenic effects of solid
1 7 particles in the respiratory tract: [proceedings of the 4th international inhalation symposium]; March 1993;
18 Hannover, Germany. Mohr, U; Dungworth, DL; Mauderly, JL; et al., eds. Washington, DC: International Life
19 Sciences Institute Press; pp. 17-25.
20
21 Morrow, PE; Yu, CP. (1993) Models of aerosol behavior in airways and alveoli. In: Aerosols in medicine:
22 principles, diagnosis and therapy. 2nd rev. ed. Moren, F; Dolovich, MB; Newhouse, MT; et al., eds. Amsterdam:
23 Elsevier; pp. 157-193.
24
25 Muhle, H; Bellman, B; Creutzenberg, O; et al. (1990) Dust overloading of lungs after exposure of rats to particles of
26 low solubility: comparative studies. J Aerosol Sci 21:374-377.
27
28 National Research Council (NRC). (1995) Science and judgment in risk assessment. Washington, DC: National
29 Research Council.
30
31 Nielsen, PS; Andreassen, A; Farmer, PB; et al. (1996) Biomonitoring of diesel exhaust-exposed workers. DNA and
32 hemoglobin adducts and urinary 1-hydroxypyrene as markers of exposure. Toxicol Lett 86:27-37.
33
34 Nikula, KJ; Avila, KJ; Griffith, WC; et al. (1997a) Sites of particle retention and lung tissue responses to chronically
35 inhaled diesel exhaust and coal dust in rats and cynomolgus monkeys. In: Proceedings of the sixth international
36 meeting on the toxicology of natural and man-made fibrous and non-fibrous particles; September 1996; Lake Placid,
37 NY. Driscoll, KE; Oberdorster, G, eds. Environ Health Perspect Suppl 105(5):1231-1234.
38
39 Nikula, KJ; Avila, KJ; Griffith, WC; et al. (1997b) Lung tissue responses and sites of particle retention differ
40 between rats and cynomolgus monkeys exposed chronically to diesel exhaust and coal dust. Fundam Appl Toxicol
41 37:37-53.
42
43 Nilsen, A; Nyberg, K; Camner, P. (1988) Intraphagosomal pH in alveolar macrophages after phagocytosis in vivo
44 and in vitro of fluorescein-labeled yeast particles. Exp Lung Res 14:197-207.
45
46 Oberdorster, G. (1988) Lung clearance of inhaled insoluble and soluble particles. J Aerosol Med 1:289-330.
47
48 Oberdorster, G; Ferin, J; Morrow, PE. (1992) Volumetric loading of alveolar macrophages (AM): a possible basis
49 for diminished AM-mediated particle clearance. Exp Lung Res 18:87-104.
50
51 Ohkuma, S; Poole, B. (1978) Fluorescence probe measurement of the intralysosomal pH in living cells and the
52 perturbation of pH by various agents. Proc Nat. Acad Sci U S A 75:3327-3331.
53
7/25/00 3-54 DRAFT—DO NOT CITE OR QUOTE
-------
Osterholm, AM; Fait, S; Lambert, B; et al. (1995) Classification of mutations at the human hprt-locus in
T-lymphocytes of bus maintenance workers by multiplex-PCR and reverse transriptase-PCR analysis. Carcinogenesis
16:1909-1995.
4
5 Pepelko, WE. (1987) Feasibility of dose adjustment based on differences in long-term clearance rates of inhaled
6 paniculate matter in humans and laboratory animals. Regul Toxicol Pharmacol 7:236-252.
7
8 Phalen, RF; Oldham, MJ. (1983) Tracheobronchial airway structure as revealed by casting techniques. Am Rev
9 RespirDis 128:S1-S4.
10
11 Powdrill, J; Buckley, C; Valdez, YE; et al. (1989) Airway intra-luminal macrophages: origin and role in lung
12 clearance. Toxicologist 9:77.
13
14 Pritchard, JN. (1989) Dust overloading causes impairment of pulmonary clearance: evidence from rats and humans.
15 Exp Pathol 37:39-42.
16
17 Raabe, OG. (1982) Deposition and clearance of inhaled aerosols. In: Mechanisms in respiratory toxicology. Witschi,
18 H, ed. Boca Raton, FL: CRC Press; pp. 27-76.
19
20 Raabe, OG; Al-Bayati, MA; Teague, SV; et al. (1988) Regional deposition of inhaled monodisperse, coarse, and fine
21 aerosol particles in small laboratory animals. In: Inhaled particles VI: proceedings of an international symposium and
22 workshop on lung dosimetry; September 1985; Cambridge, United Kingdom. Dodgson, J; McCallum, RI; Bailey,
23 MR; et al., eds. Ann Occup Hyg 32(suppl. l):53-63.
24
25 Riley, PA; Dean, RT. (1978) Phagocytosis of latex particles in relation to the cell cycle in 3T3 cells. Exp Cell Biol
26 46:367-373.
27
Robertson, B. (1980) Basic morphology of the pulmonary defence system. Eur J Respir Dis 61(suppl. 107):21-40.
30 Sandusky, CB; Cowden, MW; Schwartz, SL. (1977) Effect of particle size on regurgitative exocytosis by rabbit
31 alveolar macrophages. In: Pulmonary macrophage and epithelial cells: proceedings of the sixteenth annual Hanford
32 biology symposium; September 1976; Richland, WA. Sanders, CL; Schneider, RP; Dagle, GE; et al., eds.
33 Washington, DC: Energy Research and Development Administration; pp. 85-105. (ERDA symposium series: no.
34 43). Available from: NTIS, Springfield, VA; CONF-760927.
35
36 Schanker, LS; Mitchell, EW; Brown, RA, Jr. (1986) Species comparison of drug absorption from the lung after
37 aerosol inhalation or intratracheal injection. Drug Metab Dispos 14:79-88.
38
39 Schlesinger, RB. (1985) Comparative deposition of inhaled aerosols in experimental animals and humans: a review.
40 J Toxicol Environ Health 15:197-214.
41
42 Schlesinger, RB; Ben-Jebria, A; Dahl, AR; et al. (1997) Disposition of inhaled toxicants. In: Handbook of human
43 toxicology. Massaro, EJ, ed. Boca Raton, FL: CRC Press; pp. 493-550.
44
45 Schum, M; Yeh, HC. (1980) Theoretical evaluation of aerosol deposition in anatomical models of mammalian lung
46 airways. Bull Math Biol 42:1-15.
47
48 Siak, JS; Chan, TL; Lee, PS. (1980) Diesel paniculate extracts in bacterial test systems. In: Health effects of diesel
49 engine emissions: proceedings of an international symposium, v. 1; December 1979; Cincinnati, OH. Pepelko, WE;
50 Danner, RM; Clarke, NA, eds. Cincinnati, OH: U.S. EPA, Health Effects Research Laboratory; pp. 245-262; EPA
51 report no. EPA/600/9-80/057b. Available from: NTIS, Springfield, VA; PB81-173809.
52
§ Smith, TJ. (1985) Development and application of a model for estimating alveolar and interstitial dust levels. Ann
Occup Hyg 29:495-516.
7/25/00 3-55 DRAFT—DO NOT CITE OR QUOTE
-------
1 Snipes, MB. (1979) Long-term retention of monodisperse and polydisperse particles inhaled by beagle dogs, rats and
2 mice. Albuquerque, NM: Lovelace Biomedical and Environmental Research Institute; Inhalation Toxicology
3 Research Institute annual report LF-69; pp. 420-423.
4
5 Snipes, MB; Clem, MF. (1981) Retention of microspheres in the rat lung after intratracheal instillation. Environ Res
6 24:33-41.
7
8 Snipes, MB; Olson, TR; Yeh, HC. (1988) Deposition and retention patterns for 3-, 9-, and 15-um latex microspheres
9 inhaled by rats and guinea pigs. Exp Lung Res 14:37-50.
10
11 Snipes. MB; McClellan, RO; Mauderly, JL; et al. (1989) Retention patterns for inhaled particles in the lung:
12 comparisons between laboratory animals and humans for chronic exposures. Health Phys 57(suppl. l):69-78.
13
14 Sorokin, SP; Brain, JD. (1975) Pathways of clearance in mouse lungs exposed to iron oxide aerosols. Anat Rec
15 181:581-625.
16
17 Stahlhofen, W; Gebhart, J; Heyder, J. (1980) Experimental determination of the regional deposition of aerosol
18 particles in the human respiratory tract. Am Ind Hyg Assoc J 41:385-398a.
19
20 Stahlhofen, W; Koebrich, R; Rudolf, G; et al. (1990) Short-term and long-term clearance of particles from the upper
21 human respiratory tract as function of particle size. J Aerosol Sci 21 (suppl. 1 ):S407-S410.
22
23 Stb'ber, W; Einbrodt, HJ; Klosterko'tter, W. (1967) Quantitative studies of dust retention in animal and human lungs
24 after chronic inhalation. In: Inhaled particles and vapours II: proceedings of an international
25 symposium; September-October 1965; Cambridge, United Kingdom. Davies, CN, ed. Oxford, United Kingdom:
26 Pergamon Press; pp. 409-418.
27
28 Stb'ber, W; Morrow, PE; Hoover, MD. (1989) Compartmental modeling of the long-term retention of insoluble
29 particles deposited in the alveolar region of the lung. Fundam Appl Toxicol 13:823-842.
30
31 Stober, W; McClellan, RO; Morrow, PE. (1993) Approaches to modeling disposition of inhaled particles and fibers
32 in the lung. In: Toxicology of the lung. Gardner, DE; Crapo, JD; McClellan, RO, eds. New York: Raven Press; pp.
33 527-601.
34
35 Stober, W; Morrow, PE; Koch, W; et al. (1994) Alveolar clearance and retention of inhaled insoluble particles in
36 rats simulated by a model inferring macrophage particle load distributions. J Aerosol Sci 25:975-1002.
37
38 Strom, KA. (1984) Response of pulmonary cellular defenses to the inhalation of high concentrations of diesel
39 exhaust. J Toxicol Environ Health 13:919-944.
40
41 Strom, KA; Chan, TL; Johnson, JT. (1988) Pulmonary retention of inhaled submicron particles in rats: diesel exhaust
42 exposures and lung retention model. In: Inhaled particles VI: proceedings of an international symposium and
43 workshop on lung dosimetry; September 1985; Cambridge, United Kingdom. Dodgson, J; McCallum, RI; Bailey,
44 MR; et al., eds. Ann Occup Hyg 32(suppl. l):645-657.
45
46 Subramaniam, RP; Miller, FJ; Asgarian, B. (1998) Numerical predictions of regional deposition of fine particles in
47 the human lung using a multiple-path model. Proceedings of the American Association of Aerosol Research, annual
48 meeting; Denver, CO.
49
50 Sun, JD; Wolff, RK; Kanapilly, GM. (1982) Deposition, retention, and biological fate of inhaled benzo(a)pyrene
51 adsorbed onto ultrafine particles and as a pure aerosol. Toxicol Appl Pharmacol 65:231-244.
52
63 Sun, JD; Wolff, RK; Kanapilly, GM; et al. (1984) Lung retention and metabolic fate of inhaled benzo(a)pyrene
R4 associated ""th diese! exhaust particles. Tcxicc! App! Pharrnaco! 73:48-59.
55
7/25/00 3-56 DRAFT—DO NOT CITE OR QUOTE
-------
Tong, HY; Karasek, FW. (1984) Quantitation of polycyclic aromatic hydrocarbons in diesel exhaust paniculate
matter by high-performance liquid chromatography fractionation and high-resolution gas chromatography. Anal
Chem 56:2129-2134.
4
5 U.S. Environmental Protection Agency (U.S. EPA). (1982) Air quality criteria for paniculate matter and sulfur
6 oxides. Research Triangle Park, NC: Office of Health and Environmental Assessment, Environmental Criteria and
7 Assessment Office; EPA report no. EPA/600/8-82-029aF-cF. 3v. Available from: NTIS, Springfield, VA;
8 PB84-156777.
9
10 U.S. EPA. (1993) Provisional guidance for quantitative risk assessment of polycyclic aromatic hydrocarbons. Office
11 of Health and Environmental Assessment; EPA report no. EPA/600/R-93/089.
12
13 U.S. EPA. (1996) Air quality criteria for paniculate matter. Research Triangle Park, NC: National Center for
14 Environmental Assessment-RTP Office; report nos. EPA/600/P-95/001aF-cF. 3v. Available from: NTIS, Springfield,
15 VA;PB96-168224.
16
17 Valberg, PA; Watson, AY. (1999). Comparative mutagenic dose of ambient diesel engine exhaust. Inhal Toxicol
18 11(3):215-228.
19
20 Vastag, E; Matthys, H; Zsamboki, G; et al. (1986) Mucociliary clearance in smokers. Eur J Respir Dis 68:107-113.
21
22 Vostal, JJ; Schreck, RM; Lee, PS; et al. (1982) Deposition and clearance of diesel particles from the lung. In:
23 Toxicological effects of emissions from diesel engines: proceedings of the 1981 EPA diesel emissions symposium;
24 October 1981; Raleigh, NC. Lewtas, J, ed. New York: Elsevier Biomedical; pp. 143-159. (Developments in
25 toxicology and environmental science: v. 10).
26
27 Waite, DA; Ramsden, D. (1971) The inhalation of insoluble iron oxide particles in the sub-micron range. Part I.
Chromium-51 labelled aerosols. Winfrith, Dorchester, Dorset, United Kingdom: Atomic Energy Establishment;
report no. AEEW-R740.
30
31 Warheit. DB; Overby, LH; George, G; et al. (1988) Pulmonary macrophages are attracted to inhaled particles
32 through complement activation. Exp Lung Res 14:51-66.
33
34 Weibel, ER. (1963) Morphometry of the human lung. New York: Academic Press, Inc.
35
36 White, HJ; Garg, BD. (1981) Early pulmonary response of the rat lung to inhalation of high concentration of diesel
37 particles. J Appl Toxicol 1:104-110.
38
39 Wolff, RK; Gray, RL. (1980) Tracheal clearance of particles. In: Inhalation Toxicology Research Institute annual
40 report: 1979-1980. Diel, JH; Bice, DE; Martinez, BS, eds. Albuquerque, NM: Lovelace Biomedical and
41 Environmental Research Institute; p. 252; report no. LMF-84.
42
43 Wolff, RK; Henderson, RF; Snipes, MB; et al. (1986) Lung retention of diesel soot and associated organic
44 compounds. In: Carcinogenic and mutagenic effects of diesel engine exhaust: proceedings of the international
45 satellite symposium on toxicological effects of emissions from diesel engines; July; Tsukuba Science City, Japan.
46 Ishinishi, N; Koizumi, A; McClellan, R; et al. eds. Amsterdam: Elsevier Science Publishers BV.; pp. 199-211.
47 (Developments in toxicology and environmental science: v. 13).
48
49 Wolff, RK; Henderson, RF; Snipes, MB; et al. (1987) Alterations in particle accumulation and clearance in lungs of
50 rats chronically exposed to diesel exhaust. Fundam Appl Toxicol 9:154-166.
51
52 World Health Organization. (1996) Diesel fuel and exhaust emissions. Geneva, Switzerland: World Health
« Organization, International Programme on Chemical Safety. (Environmental health criteria 171).
Wright, JR; Clements, JA. (1987) Metabolism and turnover of lung surfactant. Am Rev Respir Dis 136:426-444.
7/25/00 3-57 DRAFT—DO NOT CITE OR QUOTE
-------
1 Xu, GB; Yu, CP. (1987) Deposition of diesel exhaust particles in mammalian lungs: a comparison between rodents
2 and man. Aerosol Sci Technol 7:117-123.
3
4 Yeh, HC; Schum, GM. (1980) Models of human lung airways and their application to inhaled particle deposition.
5 Bull Math Biol 42:461-480.
6
7 Yu, CP. (1978) Exact analysis of aerosol deposition during steady breathing. Powder Technol 21:55-62.
8
9 Yu, CP; Diu, CK. (1983) Total and regional deposition of inhaled aerosols in humans. J Aerosol Sci 5:599-609.
10
11 Yu, CP; Xu, GB. (1986) Predictive models for deposition of diesel exhaust particulates in human and rat lungs.
12 Aerosol Sci Technol 5:337-347.
13
14 Yu, CP; Yoon, KJ. (1990) Retention modeling of diesel exhaust particles in rats and humans. Amherst, NY: State
15 University of New York at Buffalo (Health Effects Institute research report no. 40).
16
17 Yu, CP; Chen, YK; Morrow, PE. (1989) An analysis of alveolar macrophage mobility kinetics at dust overloading of
18 the lungs. Fundam Appl Toxicol 13:452-459.
19
20 Yu, CP; Yoon, KJ; Chen, YK. (1991) Retention modeling of diesel exhaust particles in rats and humans. J Aerosol
21 Med4(2):79-115.
22
23 Yu, CP; Ding, YJ; Zhang, L; et al. (1996) A clearance model of refractory ceramic fibers in the rat lung including
24 fiber dissolution and breakage. J Aerosol Sci 27:151-160.
7/25/00 3-58 DRAFT—DO NOT CITE OR QUOTE
-------
4. MUTAGENICITY
1 Since 1978, more than 100 publications have appeared in which genotoxicity assays were
2 used with diesel emissions, the volatile and paniculate fractions (including extracts), or
3 individual chemicals found in diesel emissions. Although most of the studies deal with the
4 question of whether gas-phase or paniculate extracts from diesel emissions possess mutagenic
5 activity in microbial and mammalian cell assays, a number of studies have employed bioassays
6 (most commonly Salmonella TA98 without S9) to evaluate (1) extraction procedures, (2) fuel
7 modifications, (3) bioavailability of chemicals from diesel paniculate matter (DPM), and (4)
8 exhaust filters or other modifications and variables associated with diesel emissions. As
9 indicated in Chapter 2, the number of chemicals in diesel emissions is very large. Many of these
10 have been determined to exhibit mutagenic activity in a variety of assay systems (see Table II in
11 Claxton, 1983). Although a detailed discussion of those data is beyond the scope of this
12 document, some of the mutagenically active compounds found in the gas phase are ethylene,
13 benzene, 1,3-butadiene, acrolein, and several polycyclic aromatic hydrocarbons (PAHs) (see
14 Table 2-21). Of the particle-associated chemicals, several PAHs and nitro-PAHs have been the
15 focus of mutagenic investigations both in bacteria and in mammalian cell systems (see Table 2-
^ft 22). Several review articles, some containing more detailed descriptions of the available studies,
17 are available (IARC, 1989; Claxton, 1983; Pepelko and Peirano, 1983; Shirname-More, 1995).
18 Discussions of genotoxicity are also found in the proceedings of several symposia on the health
19 effects of diesel emissions (U.S. EPA, 1980; Lewtas, 1982; Ishinishi et al., 1986).
20
21 4.1. GENE MUTATIONS
22 Huisingh et al. (1978) demonstrated that dichloromethane extracts from DPM were
23 mutagenic in strains TA1537, TA1538, TA98, and TA100 of S. typhimurium, both with and
24 without rat liver S9 activation. This report contained data from several fractions as well as DPM
25 from different vehicles and fuels. Similar results with diesel extracts from various engines and
26 fuels have been reported by a number of investigators using the Salmonella frameshift-sensitive
27 strains TA1537, TA1538, and TA98 (Siak et al., 1981; Claxton, 1981; Dukovich et al., 1981;
28 Brooks et al., 1984). Similarly, mutagenic activity was observed in Salmonella forward mutation
29 assays measuring 8-azaguanine resistance (Claxton and Kohan, 1981) and in E. coli mutation
30 assays (Lewtas, 1983).
31 One approach to identifying significant mutagens in chemically complex environmental
t samples such as diesel exhaust or ambient particulate extracts is the combination of short-term
bioassays with chemical fractionation (Scheutzle and Lewtas, 1986). The analysis is most
34 frequently carried out by sequential extraction with increasingly polar or binary solvents.
7/25/00 4-1 DRAFT—DO NOT CITE OR QUOTE
-------
1 Fractionation by silica-column chromatography separates compounds by polarity or into acidic,
2 basic, and neutral fractions. The resulting fractions are too complex to characterize by chemical
3 methods, but the bioassay analysis can be used to determine fractions for further analysis. In
4 most applications of this concept, Salmonella strain TA98 without the addition of S9 has been
5 used as the. indicator for mutagenic activity. Generally, a variety of nitrated polynuclear aromatic
6 compounds have been found that account for a substantial portion of the mutagenicity (Liberti et
7 al., 1984; Schuetzle and Frazer, 1986; Schuetzle and Perez, 1983). However, not all bacterial
8 mutagenicity has been identified in this way, and the identity of the remainder of the mutagenic
9 compounds remains unknown. The nitrated aromatics thus far identified in diesel exhaust were
10 the subject of review in the I ARC monograph on diesel exhaust (IARC, 1989).
11 In addition to the simple qualitative identification of mutagenic chemicals, several
12 investigators have used numerical data to express mutagenic activity as activity per distance
1 3 driven or mass of fuel consumed. These types of calculations have been the basis for estimates
14 that the nitroarenes (both mono- and dinitropyrenes) contribute a significant amount of the total
1 5 mutagenic activity of the whole extract (Nishioka et al., 1 982; Salmeen et al., 1 982; Nakagawa et
16 al., 1983). In a 1983 review, Claxton discussed a number of factors that affected the mutagenic
1 7 response in Salmonella assays. Citing the data from the Huisingh et al. (1 978) study, the author
1 8 noted that the mutagenic response could vary by a factor of 100 using different fuels in a single
1 9 diesel engine. More recently, Crebelli et al. (1995) used Salmonella to examine the effects of
20 different fuel components. They reported that although mutagenicity was highly dependent on
21 aromatic content, especially di- or triaromatics, there was no clear effect of sulfur content of the
22 fuel. Later, Sjogren et al. (1996), using multivariate statistical methods with ten diesel fuels,
23 concluded that the most influential chemical factors in Salmonella mutagenicity were, sulfur
24 contents, certain PAHs (1-nitropyrene), and naphthenes.
25 Matsushita et al. (1986) tested particle-free diesel exhaust gas and a number of benzene
26 nitro-derivatives and PAHs (many of which have been identified as components of diesel exhaust
27 gas). The particle-free exhaust gas was positive in both TA100 and TA98, but only without S9
28 activation. Of the 94 nitrobenzene derivatives tested, 61 were mutagenic, and the majority
29 showed greatest activity in TA1GO without S9. Twenty-eight of 50 PAHs tested were mutagenic,
30 all required the addition of S9 for detection, and most appeared to show a stronger response in
31 TA100. When 1 ,6-dinitropyrene was mixed with various PAHs or an extract of heavy-duty (HD)
32 diesel exhaust, the mutagenic activity in TA98 was greatly reduced when S9 was absent but was
33 increased significantly when S9 was present. These latter results suggest that caution should be
used in estimating mutagenicity (or other toxic effects) of complex mixtures from the specific
of individual components.
•O A
7/25/00 4-2 DRAFT— DO NOT CITE OR QUOTE
-------
Mitchell et al. (1981) reported mutagenic activity of DPM extracts of diesel emissions in
the mouse lymphoma L5178Y mutation assay. Positive results were seen both with and without
3 S9 activation in extracts from several different vehicles, with mutagenic activity only slightly
4 lower in the presence of S9. These findings have been confirmed in a number of other
5 mammalian cell systems using several different genetic markers. Casto et al. (1981), Chescheir
6 et al. (1981), Li and Royer (1982), and Brooks et al. (1984) all reported positive responses at the
7 HPRT locus in Chinese hamster ovary (CHO) cells. Morimoto et al. (1986) used the APRT and
8 Ouar loci in CHO cells; Curren et al. (1981) used Ouar in BALB/c 3T3 cells. In all of these
9 studies, mutagenic activity was observed without S9 activation. Liber et al. (1981) used the
10 thymidine kinase (TK) locus in the TK6 human lymphoblast cell line and observed induced
11 mutagenesis only in the presence of rat liver S9 when testing a methylene chloride extract of
12 diesel exhaust. Barfknecht et al. (1982) also used the TK6 assay to identify some of the
13 chemicals responsible for this activation-dependent mutagenicity. They suggested that
14 fluoranthene, 1-methylphenanthrene, and 9-methylphenanthrene could account for more than
15 40% of the observed activity.
16 Morimoto et al. (1986) injected DPM extracts (250 to 4,000 mg/kg) into pregnant Syrian
17 hamsters and measured mutations at the APRT locus in embryo cells cultivated 11 days after i.p.
f injection. Neutral fractions from both light-duty (LD) and HD tar samples resulted in increased
mutation frequency at 2,000 and 4,000 mg/kg. Belisario et al. (1984) applied the Ames test to
20 urine from Sprague-Dawley rats exposed to single applications of DPM administered by gastric
•21 intubation, i.p. injection, or s.c. gelatin capsules. In all cases, dose-related increases were seen in
22 TA98 (without and with S9) from urine concentrates taken 24 h after particle administration.
23 Urine from Swiss mice exposed by inhalation to filtered exhaust (particle concentration 6 to 7
24 mg/m3) for 7 weeks (Pereira et al., 198la), or Fischer 344 rats exposed to DPM (2 mg/m3) for 3
25 months to 2 years was negative in Salmonella strains.
26 Schuler and Niemeier (1981) exposed Drosophila males in a stainless steel chamber
27 connected to the 3-m3 chamber used for the chronic animal studies at EPA (see Hinners et al.,
28 1980, for details). Flies were exposed for 8 h and mated to untreated females 2 days later.
29 Although the frequency of sex-linked recessive lethals from treated males was not different from
30 that of controls, the limited sample size precluded detecting less than a threefold increase over
31 controls. The authors noted that, because there were no signs of toxicity, the flies might tolerate
32 exposures to higher concentrations for longer time periods.
33 Driscoll et al. (1996) exposed Fischer 344 male rats to aerosols of carbon black (1.1,7.1
34 and 52.8 mg/m3) or air for 13 weeks (6 h/day, 5 days/week) and measured hprt mutations in
^P alveolar type II cells in animals immediately after exposure and at 12 and 32 weeks after the end
36 of exposure. Both the two higher concentrations resulted in significant increases in mutant
7/25/00 4-3 DRAFT—DO NOT CITE OR QUOTE
-------
1 frequency. Whereas the mutant frequency from the 7.1 mg/m3 group returned to control levels by
2 12 weeks, the mutant frequency of the high-exposure group was still higher than controls even
3 after 32 weeks. Carbon black particles have very little adsorbed PAHs, hence a direct chemically
4 induced mechanism is highly unlikely. Induction ofhprt mutations were also observed in rat
5 alveolar epithelial cells after intratracheal instillation with carbon black, a-quartz and titanium
6 dioxide (Driscoll et al., 1997). All three types of particles elicited an inflammatory response, as
7 shown by significant increases of neutrophils in bronchalveolar lavage (BAL) fluid. Culturing
8 the BAL from exposed rats with a rat lung epithelial cell line also resulted in elevation ofhprt
9 mutational response. This response was effectively eliminated when catalase was included in the
10 incubation mixture, providing evidence for cell-derived oxidative damage.
11 Specific-locus mutations were not induced in (C3H * 101)F, male mice exposed to diesel
12 exhaust 8 h/day, 7 days/week for either 5 or 10 weeks (Russell et al., 1980). The exhaust was a
13 1:18 dilution and the average particle concentration was 6 mg/m3. After exposure, males were
14 mated to T-stock females and matings continued for the reproductive life of the males. The
15 results were unequivocally negative; no mutants were detected in 10,635 progeny derived from
16 postspermatogonial cells or in 27,917 progeny derived from spermatogonial cells.
17 Hou et al. (1995) measured DNA adducts and hprt mutations in 47 bus maintenance
18 workers and 22 control individuals. All were nonsmoking men from garages in the Stockholm
19 area and the exposed group consisted of 16 garage workers, 25 mechanics, and 6 others. There
20 were no exposure data but the three groups were considered to be of higher to lower exposure to
21 diesel engine exhaust. Levels of DNA adducts determined by 32P-postlabeling were significantly
22 higher in workers than controls (3.2 versus 2.3 * 10"8), but hprt mutant frequencies were not
23 different (8.6 versus 8.4 * lO"6). Both adduct level and mutagenicity were highest among the 16
24 most exposed; mutant frequency was significantly correlated with adduct level. All individuals
25 were genotyped for glutathione transferase GSTM1 and aromatic amino transferase NAT2
26 polymorphism. Neither GSTM1 nulls nor NAT2 slow acetylators exhibited effects on either
27 DNA adducts or hprt mutant frequencies.
28
29 4.2. CHROMOSOME EFFECTS
30 Mitchell et al. (1981) and Brooks et al. (1984) reported increases in sister chromatid
31 exchanges (SCE) in CHO cells exposed to DPM extracts of emissions from both LD and HD
32 diesel engines. Morimoto et al. (1986) observed increased SCE from both LD and HD DPM
33 extracts in PAH-stimulated human lymphocyte cultures. Tucker et al. (1986) exposed human
34 peripheral lymphocyte cultures from four donors to direct diesel exhaust for up to 3 h. Exhaust
35 was cooled by pumping through a plastic tube about 20 feet long; airflow was 1.5 L/min.
36 Samples were taken at 16, 48, and 160 min of exposure. Cell cycle delay was observed in all
7/25/00 4-4 DRAFT—DO NOT CITE OR QUOTE
-------
cultures; significantly increased SCE levels were reported for two of the four cultures. Structural
chromosome aberrations were induced in CHO cells by DPM extracts from a Nissan diesel
3 engine (Lewtas, 1983) but not by similar extracts from an Oldsmobile diesel engine (Brooks et
4 al., 1984).
5 DPM dispersed in an aqueous mixture containing dipalmitoyl lecithin (DPL), a
6 component of pulmonary surfactant or extracted with dichloromethane (DCM), induced similar
7 responses in SCE assays in Chinese hamster V79 cells (Keane et al., 1991), micronucleus tests in
8 V79 and CHO cells (Gu et al., 1992) and unscheduled DNA synthesis (UDS) in V79 cells (Gu et
9 al., 1994). After separating the samples into supernatant and sediment fractions, mutagenic
10 activity was confined to the sediment fraction of the DPL sample and the supernatant of the
11 DCM sample. These findings suggest that the mutagenic activity of DPM inhaled into the lungs
12 could be made bioavailable through solubilization and dispersion nature of pulmonary
13 surfactants. In a later study in the same laboratory, Liu et al. (1996) found increased micronuclei
14 in V79 cells treated with crystalline quartz and a noncrystalline silica, but response was reduced
15 after pretreatment of the particles with the simulated pulmonary surfactant.
16 Pereira et al. (198la) exposed female Swiss mice to diesel exhaust 8 h/day, 5 days/week
17 for 1, 3, and 7 weeks. The incidence of micronuclei and structural aberrations was similar in
^ bone marrow cells of both control and exposed mice. Increased incidences of micronuclei, but
19 not SCE, were observed in bone marrow cells of male Chinese hamsters after 6 months of
20 exposure to diesel exhaust (Pereira et al., 1981b).
21 Guerrero et al. (1981) observed a linear concentration-related increase in SCE in lung
22 cells cultured after intratracheal instillation of DPM at doses up to 20 mg/hamster. However,
23 they did not observe any increase in SCE after 3 months of inhalation exposure to diesel exhaust
24 particles (6 mg/m3).
25 Pereira et al. (1982) measured SCE hi embryonic liver cells of Syrian hamsters. Pregnant
26 females were exposed to diesel exhaust (containing about 12 mg/m3 particles) from days 5 to 13
27 of gestation or injected intraperitoneally with diesel particles or particle extracts on gestational
28 day 13 (18 h before sacrifice). Neither the incidence of SCE nor mitotic index was affected by
29 exposure to diesel exhaust. The injection of DPM extracts but not DPM resulted in a dose-
30 related increase in SCE; however, the toxicity of the DPM was about twofold greater than the
31 DPM extract.
32 In the only studies with mammalian germ cells, Russell et al. (1980) reported no increase
33 in either dominant lethals or heritable translocations in males of T-stock mice exposed by
inhalation to diesel emissions. In the dominant lethal test, T-stock males were exposed for 7.5
weeks and immediately mated to females of different genetic backgrounds (T-stock; [C3H *
36 101]; [C3H x C57BL/6]; [SEC * C57BL/6]). There were no differences from controls in any of
7/25/00 4-5 DRAFT—DO NOT CITE OR QUOTE
-------
1 the parameters measured in this assay. For heritable translocation analysis, T-stock males were
2 exposed for 4.5 weeks and mated to (SEC * C57BL/6) females, and the F, males were tested for
3 the presence of heritable translocations. Although no translocations were detected among 358
4 progeny tested, the historical control incidence is less than 1/1,000.
5
6 4.3. OTHER GENOTOXIC EFFECTS
7 Pereira et al. (198 Ib) exposed male strain A mice to diesel exhaust emissions for 31 or 39
8 weeks using the same exposure regimen noted in the previous section. Analyses of caudal sperm
9 for sperm-head abnormalities were conducted independently in three separate laboratories.
10 Although the incidence of sperm abnormalities was not significantly above controls in any of the
11 three laboratories, there were extremely large differences in scoring among the three (control
12 values were 9.2%, 14.9%, and 27.8% in the three laboratories). Conversely, male Chinese
13 hamsters exposed for 6 months (Pereira et al., 1981c) exhibited almost a threefold increase in
14 sperm-head abnormalities. It is noted that the control incidence in the Chinese hamsters was less
15 than 0.5%. Hence, it is not clear whether the differing responses reflect true species differences
16 or experimental artifacts.
17 A number of studies measuring DNA adducts in animals exposed to DPM, carbon black,
18 or other particles have been reported and are reviewed by Shirname-More (1995). Although
19 modest increases in DNA adducts have been observed in lung tissue of rats after inhalation of
20 DPM (Wong et al., 1986; Bond et al., 1990), the increases are small in comparison with those
21 induced by chemical carcinogens present in diesel exhaust (Smith et al., 1993). While Gallagher
22 et al. (1994) found no increases in total DNA adducts in lung tissue of rats exposed to diesel
23 exhaust, carbon black or titanium dioxide, they did observe an increase in an adduct with
24 migration properties similar to nitrochrysene and nitro-benzo(a)pyrene adducts from diesel but
25 not carbon black or titanium dioxide exposures. The majority of the studies used the 32P-
26 postlabeling assay to detect adducts. Although this method is sensitive, chemical identity of
27 adducts can only be inferred if an adduct spot migrates to the same location as a known prepared
28 adduct.
?9 DNA adducts have also been measured in humans occupational!}- exposed to diesel
30 exhaust. Distinct adduct patterns were found among garage workers occupationally exposed to
31 diesel exhaust when compared with nonexposed controls (Nielsen and Autrup, 1994).
32 Furthermore, the findings were concordant with the adduct patterns observed in groups exposed
33 to low concentrations of PAHs from combustion processes. Hemminki et al. (1994) also
34 reported significantly elevated levels or DNA adducts in lymphocytes from garage workers with
35 known Jicscl exhaust exposure compared with unexposed mechanics. Hou et al. (1995) found
36 elevated adduct levels in bus maintenance workers exposed to diesel exhaust. Although no
7/25/00 4-6 DRAFT—DO NOT CITE OR QUOTE
-------
difference in mutant frequency was observed between the groups, the adduct levels were
significantly different (3.2 vs. 2.3 x lO"8). Nielsen et al. (1996) reported significantly increased
3 levels of three biomarkers (lymphocyte DNA adducts, hydroxyethylvaline adducts in
4 hemoglobin, and 1-hydroxypyrene in urine) in DE-exposed bus garage workers.
5 The role of oxidative damage in causing mutations has received increasing focus recently.
6 More than 50 different chemicals have been studied in rodents, usually measuring the formation
7 of 8-hydroxydeoxyguanosine (8-OH-dG), a highly mutagenic adduct (Loft et al., 1998).
8 Increases in the mutagenic DNA adduct 8-hydroxydeoxyguanosine were found in mouse lung
9 DNA after intratracheal instillation of diesel particles (Nagashima et al., 1995). The response
10 was dose dependent. Mice fed on a high-fat diet showed an increased response whereas the
11 responses were partially reduced when the antioxidant p-carotene was included in the diet
12 (Ichinose et al., 1997). Oxidative damage has also been measured in rat lung tissue after
13 intratracheal instillation of quartz (Nehls et al., 1997) and in rat alveolar macrophages after in
14 vitro treatment with silica dust (Zhang et al., 2000). Arimoto et al. (1999) demonstrated that
15 redissol ved methanol extracts of DPM also induced the formation of 8-OH-dG adducts in L120
16 mouse cells. The response was dependent on both DPM concentration and P450 reductase. A
17 detailed discussion of the potential role of oxidative damage in diesel exhaust carcinogenesis is
^fc presented in Chapter 7.4.
19
20 4.4. SUMMARY
21 Extensive studies with Salmonella have unequivocally demonstrated mutagenic activity
22 in both particulate and gaseous fractions of diesel exhaust. In most of the studies using
23 Salmonella, DPM extracts and individual nitropyrenes exhibited the strongest responses in strain
24 TA98 when no exogenous activation was provided. Gaseous fractions reportedly showed greater
25 response in TA100, whereas benzo(a)pyrene and other unsubstituted PAHs are mutagenic only in
26 the presence of S9 fractions. The induction of gene mutations has been reported in several in
27 vitro mammalian cell lines after exposure to extracts of DPM. Note that only the TK6 human
28 cell line did not give a positive response to DPM extracts in the absence of S9 activation.
29 Mutagenic activity was recovered in urine from animals treated with DPM by gastric intubation
30 and i.p. and s.c. implants, but not by inhalation of DPM or diluted diesel exhaust. Dilutions of
31 whole diesel exhaust did not induce sex-linked recessive lethals in Drosophila or specific-locus
32 mutations in male mouse germ cells.
33 Structural chromosome aberrations and SCE in mammalian cells have been induced by
€ particles and extracts. Whole exhaust induced micronuclei but not SCE or structural aberrations
in bone marrow of male Chinese hamsters exposed to whole diesel emissions for 6 months. In a
36 shorter exposure (7 weeks), neither micronuclei nor structural aberrations were increased in bone
7/25/00 4-7 DRAFT—DO NOT CITE OR QUOTE
-------
1 marrow of female Swiss mice. Likewise, whole diesel exhaust did not induce dominant lethals
2 or heritable translocations in male mice exposed for 7.5 and 4.5 weeks, respectively.
3 The application of mutagenicity data to the question of the potential carcinogenicity of
4 diesel emissions is based on the premise that genetic alterations are found in all cancers and that
5 several of the chemicals found in diesel emissions possess mutagenic activity in a variety of
6 genetic assays. These genetic alterations can be produce by gene mutations, deletions,
7 translocations, aneuploidy, or amplification of genes, hence no single genotoxicity assay should
8 be expected to either qualitatively or quantitatively predict rodent carcinogenicity. With diesel
9 emissions or other mixtures, additional complications arise because of the complexity of the
1 0 material being tested. Exercises that combined the Salmonella mutagenic potency with the total
1 1 concentration of mutagenic chemicals deposited in the lungs could not account for the observed
12 tumor incidence in exposed rats (Rosenkranz, 1993; Goldstein et al., 1998). However, such
1 3 calculations ignored the contribution of gaseous-phase chemicals which have been estimated to
14 contribute from less than 50% (Rannug et al., 1983) to over 90% (Matsushita et al., 1986) of the
1 5 total mutagenicity. This wide range is partly reflective of the differences in material tested:
1 6 semivolatile extracts in the former and whole gaseous emission in the latter. Of greater
1 7 importance is that these calculations are based on a reverse mutation assay hi bacteria with
18 metabolic processes strikingly different from mammals. This is at least partly reflected in the
1 9 observations that different nitro-PAHs give different responses in bacteria and in CHO cells (Li
20 and Dutcher, 1983) or in human hepatoma-derived cells (Eddy et al., 1986).
•21
22 4.5. REFERENCES
23
24 Arimoto, T; Yoshikawa, T; Takano, H; et al. (1999) Generation of reactive oxygen species and S-hydroxy-21-
25 deoxyguanosine formation from diesel exhaust particle components in L1210 cells. Jpn J Pharmacol 80:49-54.
26
27 Barfknecht, TR; Hites, RA; Cavaliers, EL; et al. (1982) Human cell mutagenicity of polycyclic aromatic
28 hydrocarbon components of diesel emissions. In: Toxicological effects of emissions from diesel engines:
29 proceedings of the Environmental Protection Agency 1981 diesel emissions symposium; October 1981; Raleigh, NC.
30 (Developments on toxicology and environmental science: v. 10.) Lewtas, J, ed. New York: Elsevier Biomedical; pp.
31 277-294.
32
33 Belisanp, MA; Buor.ocore, V; Ds Marir.is, E; ct al. (1984) Biological availaoihty of mutagenic compounds adsorbed
34 onto diesel exhaust paniculate. Mutat Res 135: 1-9.
35
36 Bond, JA; Mauderly, JL; Wolff, RK. (1990) Concentration- and time-dependent formation of DNA adducts in lungs
37 of rats exposed to diesel exhaust. Toxicology 60: 127-135.
38
39 Brooks, A; Li, AP; Dutcher, JS; et al. (1984) A comparison of genotoxicity of automobile exhaust particles from
40 laboratory and environmental sources. Environ Mutagen 6:651-668.
41
42
asiu, BC; Hatch, GG; Huang, SL; et al. (1981) Mutagenic and carcinogenic potency of extracts of diesel and
43 related environmental emissions: in vitro mutagenesis and oncogenic transformation. Environ Int 5:403-409.
7/25/00 4-8 DRAFT— DO NOT CITE OR QUOTE
-------
Chescheir, GM, III; Garrett, NE; Shelbume, JD; et al. (1981) Mutagenic effects of environmental particulates in the
CHO/HGPRT system. In: Application of short-term bioassays in the fractionation and analysis of complex .
environmental mixtures. Waters, MD; Sandhu, SS; Huisingh, JL; et al., eds. New York: Plenum Press; pp. 337-350.
4
5 Claxton, LD. (1981) Mutagenic and carcinogenic potency of diesel and related environmental emissions: Salmonella
6 bioassay. Environ Int 5:389-391.
7
8 Claxton, LD. (1983) Characterization of automotive emissions by bacterial mutagenesis bioassay: a review. Environ
9 Mutagen 5:609-631.
10
11 Claxton, L; Kohan, M. (1981) Bacterial mutagenesis and the evaluation of mobile-source emissions. In: Short-term
12 bioassays in the analysis of complex environmental mixtures II: proceedings of the second symposium on the
13 application of short-term bioassays in the fractionation and analysis of complex environmental mixtures; March
14 1980; Williamsburg, VA. (Hollaender, A; Welch, BL; Probstein, RF, eds. Environmental science research series: v.
15 22.) Waters, MD; Sandhu, SS; Huisingh, JL; et al., eds. New York: Plenum Press; pp. 299-317.
16
17 Crebelli, R; Conti, L; Crochi, B; et al. (1995) The effect of fuel composition on the mutagenicity of diesel engine
18 exhaust. Mutat Res 346:167-172.
19
20 Curren, RD; Kouri, RE; Kim, DM; et al. (1981) Mutagenic and carcinogenic potency of extracts from diesel related
21 environmental emissions: simultaneous morphological transformation and mutagenesis in BALB/c 3T3 cells.
22 Environ Int 5:411-415.
23
24 Driscoll, KE; Carter, JM; Howard, BW; et al. (1996) Pulmonary inflammatory, chemokine, and mutagenic responses
25 in rats after subchronic inhalation of carbon black. Toxicol Appl Pharmacol 136:372-380.
26
< Driscoll, KE; Deyo, KC; Carter, JM; et al. (1997) Effects of particle exposure and particle-elicited inflammatory
cells on mutation in rat alveolar epithelial cells. Carcinogenesis 18:423-430.
30 Dukovich, M; Yasbin, RE; Lestz, SS; et al. (1981) The mutagenic and SOS-inducing potential of the soluble organic
31 fraction collected from diesel paniculate emissions. Environ Mutagen 3:253-264.
32
33 Eddy, EP; McCoy, EC; Rosenkranz, HS; et al. (1986) Dichotomy in the mutagenicity and genotoxicity of
34 nitropyrenes: apparent effect of the number of electrons involved in nitroreduction. Mutat Res 161:109- 111.
35
36 Gallagher, J; Heinrich, U; George, M; et al. (1994) Formation of DNA adducts in rat lung following chronic
37 inhalation of diesel emissions, carbon black and titanium dioxide particles. Carcinogenesis 15:1291-1299.
38
39 Goldstein, LS; Weyand, EH; Safe, S; et al. (1998) Tumors and DNA adducts in mice exposed to benzo[a]pyrene and
40 coal tars: implications for risk assessment. Environ Health Perspect 106 Suppl 6:1325-1330.
41
42 Gu, ZW; Zhong, BZ; Nath, B; et al. (1992) Micronucleus induction and phagocytosis in mammalian cells treated
43 with diesel emission particles. Mutat Res 279:55-60.
44
45 Gu, ZW; Zhong, BZ; Keane, MJ; et al. (1994) Induction of unscheduled DNA synthesis in V70 cells by diesel
46 emission particles dispersed in simulated pulmonary surfactant. Ann Occup Hyg 38:345-349.
47
48 Guerrero, RR; Rounds, DE; Orthoefer, J. (1981) Sister chromatid exchange analysis of Syrian hamster lung cells
49 treated in vivo with diesel exhaust particulates. Environ Int 5:445-454.
50
51 Hemminki, K; Soderling, J; Ericson, P; et al. (1994) DNA adducts among personnel servicing and loading diesel
52 vehicles. Carcinogenesis 15:767-769.
Hinners, RG; Burkart, JK; Malanchuk, M. (1980) Facilities for diesel exhaust studies. In: Health effects of diesel
55 engine emissions: proceedings of an international symposium; December 1979. Pepelko, WE; Danner, RM; Clarke,
7/25/00 4-9 DRAFT—DO NOT CITE OR QUOTE
-------
1 NA, eds. Cincinnati, OH: U.S. Environmental Protection Agency, Health Effects Research Laboratory; pp. 681-697;
2 EPA report no. EPA-600/9-80-057b. Available from: NTIS, Springfield, VA; PB81 -173 817.
3
4 Hou, S; Lambert, B; Hemminki, K. (1995) Relationship between hprt mutant frequency, aromatic DNA adducts and
5 genotypes forGSTMI and NAT2 in bus maintenance workers. Carcinogensis 16:1913-1917.
6
7 Huisingh, J; Bradow, R; lungers, R; et al. (1978) Application of bioassay to the characterization of diesel particle
8 emissions. In: Application of short-term bioassays in the fractionation and analysis of complex environmental
9 mixtures: [proceedings of a symposium; February; Williamsburg, VA]. (Hollaender, A; Probstein, F; Welch, BL,
10 eds. Environmental science research: v. 15.) Waters, MD; Nesnow, S; Huisingh, JL; et al., eds. New York: Plenum
11 Press; pp. 383-418.
12
13 IARC (International Agency for Research on Cancer). (1989) Diesel and gasoline engine exhausts and some
14 nitroarenes. (IARC monographs on the evaluation of carcinogenic risks to humans: v. 46). Lyon, France: World
15 Health Organization; pp. 41 -185.
16
17 Ichinose, T; Yajima, Y; Nagashima, M; et al. (1997) Lung carcinogenesis and formation of 8-hydroxy-
18 deoxyguanosine in mice by diesel exhaust particles. Carcinogenesis 18:185-192.
19
20 Ishinishi, N; Koizumi, A; McClellan, RO; et al., eds. (1986) Carcinogenic and mutagenic effects of diesel engine
21 exhaust: proceedings of the international satellite symposium on toxicological effects of emissions from diesel
22 engines; July; Tsukuba Science City, Japan. (Developments in toxicology and environmental science: v. 13.)
23 Amsterdam: Elsevier Science Publishers BV.
24
25 Keane, MJ; Xing, SG; Harrison, JC; et al. (1991) Genotoxicity of diesel-exhaust particles dispersed in simulated
26 pulmonary surfactant. Mutat Res 260:233-238.
27
28 Lewtas, J. (1982) Mutagenic activity of diesel emissions. In: Toxicological effects of emissions from diesel engines:
29 proceedings of the Environmental Protection Agency 1981 diesel emissions symposium; October 1981; Raleigh, NC.
30 (Developments in toxicology and environmental science: v. 10.) Lewtas, J, ed. New York: Elsevier Biomedical; pp.
31 243-264.
32
33 Lewtas, J. (1983) Evaluation of the mutagenicity and carcinogenicity of motor vehicle emissions in short-term
34 bioassays. Environ Health Perspect 47:141-152.
35
36 Li, AP; Dutcher, JS. (1983) Mutagenicity of mono-, di-, and tri-nitropyrenes in Chinese hamster ovary cells. Mutat
37 Res 119:387-392.
38
39 Li, AP; Royer, RE. (1982) Diesel-exhaust-particle extract enhancement of chemical-induced mutagenesis in cultured
40 Chinese hamster ovary cells: possible interaction of diesel exhaust with environmental chemicals. Mutat Res
41 103:349-355.
42
43 Liber, HL; Andon, BM; Hites, RA; et al. (1981) Diesel soot: mutation measurements in bacterial and human cells.
44 Environ Int 5:281-284.
45
46 Liberti, A; Ciccioli, P; Cecinato, A; et al. (1984) Determination of nitrated-polyaromatic hydrocarbons (nitro-PAHs)
47 in environmental samples by high resolution chromatographic techniques. J High Resolut Chromatogr Commun
48 7:389-397.
49
50 Liu, X; Keane, MJ; Zong, BZ; et al. (1996) Micronucleus formation in V79 cells treated with respirable silica
51 dispersed in medium and simulated pulmonary surfactant. Mutat Res 361:89-94.
52
53 Loft, S; Deng, XS; Tuo, J; et al. (1998) Experimental study of oxidative DNA damage. Free Radical Res 29-S25-
54 3JV.
55
7/25/00 4-10 DRAFT—DO NOT CITE OR QUOTE
-------
Matsushita, H; Goto, S; Endo, O; et al. (1986) Mutagenicity of diesel exhaust and related chemicals. In:
Carcinogenic and mutagenic effects of diesel engine exhaust: proceedings of the international satellite symposium on
'3 toxicological effects of emissions from diesel engines; July; Tsukuba Science City, Japan. (Developments on
4 toxicology and environmental science: v. 13.) Ishinishi, N; Koizumi, A; McClellan, RO; et al., eds. Amsterdam:
5 Elsevier Science Publishers BV; pp. 103-118. -
6
7 Mitchell, AD; Evans, EL; Jotz, MM; et al. (1981) Mutagenic and carcinogenic potency of extracts of diesel and
8 related environmental emissions: in vitro mutagenesis and DNA damage. Environ Int 5:393-401.
9
10 Morimoto, K; Kitamura, M; Kondo, H; et al. (1986) Genotoxicity of diesel exhaust emissions in a battery of in-vitro
11 short-term bioassays. In: Carcinogenic and mutagenic effects of diesel engine exhaust: proceedings of the
12 international satellite symposium on toxicological effects of emissions from diesel engines; July; Tsukuba Science
13 City, Japan. (Developments in toxicology and environmental science: v. 13.) Ishinishi, N; Koizumi, A; McClellan,
14 RO; et al., eds. Amsterdam: Elsevier Science Publishers BV; pp. 85-102.
15
16 Nakagawa, R; Kitamori, S; Horikawa, K; et al. (1983) Identification of dinitropyrenes in diesel-exhaust particles:
17 their probable presence as the major mutagens. Mutat Res 124:201-211.
18
19 Nagashima, M; Kasai, H; Yokota, J; et al. (1995) Formation of an oxidative DNA damage, 8-
20 hydroxydeoxyguanosine, in mouse lung DNA after intratracheal instillation of diesel exhaust particles and effects of
21 high dietary fat and beta-carotene on this process. Carcinogenesis 16:1441-1445.
22
23 Nehls, P; Seiler, F; Rehn, B; et al. (1997) Formation and persistence of 8-oxoguanine in rat lung cells as an
24 important determinant in tumor formation following particle exposure. Environ Health Perspect 105(5): 1291-1296.
25
26 Nielsen, PS; Autrup, H. (1994) Diesel exhaust-related DNA adducts in garage workers. Clin Chem 40:1456-1458.
Nielsen, PS; Andreassen, A; Farmer, PB; et al. (1996) Biomonitoring of diesel-exhaust exposed workers. DNA and
29 hemoglobin adducts and urinary 1-hydroxyproline as markers of exposure. Toxicol Lett 86:27-37.
30
31 Nishioka, MG; Petersen, BA; Lewtas, J. (1982) Comparison of nitro-aromatic content and direct-acting mutagenicity
32 of diesel emissions. In: Polynuclear aromatic hydrocarbons: physical and biological chemistry. Cooke, M; Dennis,
33 AJ; Fisher, GL, eds. Columbus, OH: Battelle Press; pp. 603-613.
34
35 Pepelko, WE; Peirano, WB. (1983) Health effects of exposure to diesel engine emissions: a summary of animal
36 studies conducted by the U.S. Environmental Protection Agency's Health Effects Research Laboratories at
37 Cincinnati, Ohio. J Am Coll Toxicol 2:253-306.
38
39 Pereira, MA; Connor, TH; Meyne, J; et al. (198la) Metaphase analysis, micronucleus assay and urinary mutagenicity
40 assay of mice exposed to diesel emissions. Environ Int 5:435-438.
41
42 Pereira, MA; Sabharwal, PS; Gordon, L; et al. (1981b) The effect of diesel exhaust on sperm-shape abnormalities in
43 mice. Environ Int 5:459-460.
44
45 Pereira, MA; Sabharwal, PS; Kaur, P; et al. (1981c) In vivo detection of mutagenic effects of diesel exhaust by
46 short-term mammalian bioassays. Environ Int 5:439-443.
47
48 Pereira, MA; McMillan, L; Kaur, P; et al. (1982) Effect of diesel exhaust emissions, particulates, and extract on
49 sister chromatid exchange in transplacentally exposed fetal hamster liver. Environ Mutagen 4:215-220.
50
51 Rannug, U; Sundvall, A; Westerholm, R; et al. (1983) Some aspects of mutagenicity testing of the paniculate phase
52 and the gas phase of diluted and undiluted automobile exhaust. Environ Sci Res 27:3-16.
Rosenkranz, HS. (1993) Revisiting the role of mutagenesis in the induction of lung tumors in rats by diesel
55 emissions. Mutat Res 303:91-95.
7/25/00 4-11 DRAFT—DO NOT CITE OR QUOTE
-------
1 Russell, LB; Generoso, WM; Oakberg, EF; et al. (1980) Tests for heritable effects induced by diesel exhaust in the
2 mouse. Martin Marietta Energy Systems, Inc., Oak Ridge National Laboratory; report no. ORNL-5685.
3
4 Salmeen, I; Durisin, AM; Prater, TJ; et al. (1982) Contribution of 1-nitropyrene to direct-acting Ames assay
5 mutagenicities of diesel paniculate extracts. Mutat Res 104:17-23.
6
7 Schuetzle, D; Frazier, JA. (1986) Factors influencing the emission of vapor and paniculate phase components from
8 diesel engines. In: Carcinogenic and mutagenic effects of diesel engine exhaust: proceedings of the international
9 satellite symposium on toxicological effects of emissions from diesel engines; July; Tsukuba Science City, Japan.
10 (Developments in toxicology and environmental science: v. 13.) Ishinishi, N; Koizumi, A; McClellan, RO; et al.,
11 eds. Amsterdam: Elsevier Science Publishers BV; pp. 41-63.
12
13 Schuetzle, D; Lewtas, J. (1986) Bioassay-directed chemical analysis in environmental research. Anal Chem
14 58:1060A-1076A.
15
16 Schuetzle, D; Perez, JM. (1983) Factors influencing the emissions of nitrated-polynuclear aromatic hydrocarbons
17 (nitro-PAH) from diesel engines. J Air Pollut Control Assoc 33:751-755.
18
19 Schuler, RL; Niemeier, RW. (1981) A study of diesel emissions on Drosophila. Environ Int 5:431-434.
20
21 Shimame-More, L. (1995) Genotoxicity of diesel emissions. Part I: Mutagenicity and other genetic effects. Diesel
22 exhaust: a critical analysis of emissions, exposure, and health effects. A special report of the Institute's Diesel
23 Working Group. Cambridge, MA: Health Effects Institute, pp. 222-242.
24
25 Siak, JS; Chan, TL; Lees, PS. (1981) Diesel paniculate extracts in bacterial test systems. Environ Int 5:243-248.
26
27 SjSgren, M; Li, H; Banner, C; et al. (1996) Influence of physical and chemical characteristics of diesel fuels and
28 exhaust emissions on biological effects of panicle extracts: a multivariate statistical analysis often diesel fuels.
29 Chem Res Toxicol 9:197-207.
30
31 Smith, BA; Fullerton, NF; Aidoo, A; et al. (1993) DNA adduct formation in relation to lymphocyte mutations and
32 lung tumor induction in F344 rats treated with the environmental pollutant 1,6- dinitropyrene. Environ Health
33 Perspect 99:277-280.
34
35 Tucker, JD; Xu, J; Stewart, J; et al. (1986) Detection of sister chromatid exchanges induced by volatile
36 genotoxicants. Teratogen Carcinogen Mutagen 6:15-21.
37
38 U.S. Environmental Protection Agency (EPA). (1980) Health effects of diesel engine emissions: proceedings of an
39 international symposium. Cincinnati, OH: Office of Research and Development; EPA 600/9-80/057b.
40
41 Wong, D; Mitchell, CE; Wolff, RK; et al. (1986) Identification of DNA damage as a result of exposure of rats to
42 diesel engine exhaust. Carcinogenesis 7:1595-1597.
43
44 Zhang, Z; Shen, HM; Zhang, QF; et al. (1999) Involvement of oxidative stress in crystalline silica-induced
45 cytotoxicity and genctoxicity in rat alveolar maerophages. Environ Res 87:245-252.
7/25/00 4-12 DRAFT—DO NOT CITE OR QUOTE
-------
5. NONCANCER HEALTH EFFECTS OF DIESEL EXHAUST
1 The objective of this chapter is to review and evaluate potential health effects other than
2 cancer associated with inhalation exposure to diesel exhaust (DE). Data have been obtained
3 from diverse human, laboratory animal, and in vitro test systems. The human studies comprise
4 both occupational and human experimental exposures, the former consisting of exposure to DE
5 in the occupational environment, and the latter consisting of exposure to diluted DE or diesel
6 particulate matter (DPM) under controlled conditions. The laboratory animal studies consist of
7 both acute and chronic exposures of laboratory animals to DE or DPM. Diverse in vitro test
8 systems composed of human and laboratory animal cells treated with DPM or components of
9 DPM have also been used to investigate the effects of DPM at the cellular and molecular levels.
10 DPM mass (mg/m3) has been used as a measure of DE exposure in human and experimental
11 studies. The noncancer health effects of DPM have been reviewed previously by the Health
12 Effects Institute (HEI, 1995) and in the Air Quality for Particulate Matter Criteria Document
13 (U.S. EPA, 1996). The noncancer health effects attributable to ambient particulate matter (PM),
14 which is composed in part of DPM, as well as the potential mechanisms underlying these effects
15 have also been previously reviewed in the Air Quality for Particulate Matter Criteria Document
6 (U.S. EPA, 1996, also see chapter 6.2).
18 5.1. HEALTH EFFECTS OF WHOLE DIESEL EXHAUST
19 5.1.1. Human Studies
20 5.1.1.1. Short-Term Exposures
21 In a controlled human study, Rudell et al. (1990, 1994) exposed eight healthy subjects in
22 an exposure chamber to diluted exhaust from a diesel engine for 1 h, with intermittent exercise.
23 Dilution of the diesel exhaust was controlled to provide a median NO2 level of approximately
24 1.6 ppm. Median particle number was 4.3 * 106/cm3, and median levels of NO and CO were 3.7
25 and 27 ppm, respectively (particle size and mass concentration were not provided). There were
26 no effects on spirometry or on closing volume using nitrogen washout. Five of eight subjects
27 experienced unpleasant smell, eye irritation, and nasal irritation during exposure. Brochoalveolar
28 lavage (B AL) was preformed 18 hours after exposure and was compared with a control BAL
29 performed 3 weeks prior to exposure. There was no control air exposure. Small but statistically
30 significant reductions were seen in BAL mast cells, AM phagocytosis of opsonized yeast
31 particles, and lymphocyte CD4/CD8 ratios. A small increase in recovery of polymorphonuclear
32 cells (PMNs) was also observed. These findings suggest that diesel exhaust may induce mild
airway inflammation in the absence of spirometric changes. This study provides an intriguing
34 glimpse of the effect of diesel exhaust exposure in humans, but only one exposure level was
7/25/00 5-1 DRAFT—DO NOT CITE OR QUOTE
-------
1 used, the number of subjects was low, and a limited range of endpoints was reported, so the data
2 are inadequate to generalize about the human response. To date, no well-controlled chamber
3 study has been conducted using methodologies for assessing subtle lung inflammatory reactions.
4 Rudell et al. (1996) exposed volunteers to diesel exhaust for 1 h in an exposure chamber.
5 Light work on a bicycle ergometer was performed during exposure. Exposures included either
6 diesel exhaust or exhaust with particle numbers reduced 46% by a particle trap. The engine used
7 was a new Volvo model 1990, a six-cylinder direct-injection turbocharged diesel with an
8 intercooler, which was run at a steady speed of 900 rpm during the exposures. Comparison of
9 this study with others was difficult because neither exhaust dilution ratios nor particle
10 concentrations were reported. Carbon monoxide concentrations of 27-30 ppm and NO of
11 2.6-2.7 ppm, however, suggested DPM concentrations may have equaled several mg/m3. The
12 most prominent symptoms during exposure were irritation of the eyes and nose and an unpleasant
13 smell. Both airway resistance and specific airway resistance increased significantly during the
14 exposures. Despite the 46% reduction in particle numbers by the trap, effects on symptoms and
15 lung function were not significantly attenuated.
16 Kahn et al. (1988) reported the occurrence of 13 cases of acute overexposure to diesel
17 exhaust among Utah and Colorado coal miners. Twelve miners had symptoms of mucous
18 membrane irritation, headache, and lightheadedness. Eight individuals reported nausea; four
19 reported a sensation of unreality; four reported heartburn; three reported weakness, numbness,
20 and tingling in their extremities; three reported vomiting; two reported chest tightness; and two
21 others reported wheezing. Each miner lost time from work because of these symptoms, which
22 resolved within 24 to 48 h. No air monitoring data were presented; poor work practices were
23 described as the predisposing conditions for overexposure.
24 El Batawi and Noweir (1966) reported that among 161 workers from two garages where
25 diesel-powered buses were serviced and repaired, 42% complained of eye irritation, 37% of
26 headaches, 30% of dizziness, 19% of throat irritation, and 11% of cough and phlegm. Ranges of
27 mean concentrations of diesel exhaust components in the two diesel bus garages were as follows:
28 0.4 to 1.4 ppm NO2, 0.13 to 0.81 ppm SO23 0.6 to 44.1 ppm aldehydes, and 1.34 to 4.51 mg/in3 of
29 DPM; the highest concentrations were obtained close to the exhaust systems of the buses.
30 Eye irritation was reported by Battigelli (1965) in six subjects after 40 s of chamber
31 exposure to diluted diesel exhaust containing 4.2 ppm NO2, 1 ppm SO2, 55 ppm CO, 3.2 ppm
32 total hydrocarbons, and 1 to 2 ppm total aldehydes; after 3 min and 20 s of exposure to diluted
33 diesel exhaust containing 2.8 ppm NO2, 0.5 ppm SO2, 30 ppm CO, 2.5 ppm total hydrocarbons,
34 and <1 to 2 ppm total aldehydes; and after 6 min of exposure to diluted diesel exhaust containing
35 1.3 ppm NO,. 0.2 ppm SO,: <20 nnm CO <9 n nnm. total hydrocarbons, ar.d <1.0 ppm total
36 aldehydes. The concentration of DPM was not reported.
7/25/00 5-2 DRAFT—DO NOT CITE OR QUOTE
-------
Katz et al. (1960) described the experience of 14 chemists and their assistants monitoring
the environment of a train tunnel used by diesel-powered locomotives. Although workers
3 complained on three occasions of minor eye and throat irritation, no correlation was established
4 with concentrations of any particular component of diesel exhaust.
5 The role of antioxidant defenses in protecting against acute diesel exhaust exposure has
6 been studied. Blomberg et al. (1998) investigated changes in the antioxidant defense network
7 within the respiratory tract lining fluids of human subjects following diesel exhaust exposure.
8 Fifteen healthy, nonsmoking, asymptomatic subjects were exposed to filtered air or diesel
9 exhaust (DPM 300 mg/m3) for 1 h on two separate occasions at least 3 weeks apart. Nasal lavage
10 fluid and blood samples were collected prior to, immediately after, and 5 '/2 h post exposure.
11 Bronchoscopy was performed 6 h after the end of diesel exhaust exposure. Nasal lavage ascorbic
12 acid concentration increased tenfold during diesel exhaust exposure, but returned to basal levels
13 5.5 h postexposure. Diesel exhaust had no significant effects on nasal lavage uric acid or GSH
14 concentrations, and did not affect plasma, bronchial wash, or bronchoalveolar lavage antioxidant
15 concentrations, nor malondialdehyde or protein carbonyl concentrations. The authors concluded
16 that the physiological response to acute diesel exhaust exposure is an acute increase in the level
17 of ascorbic acid in the nasal cavity, which appears to be sufficient to prevent further oxidant
18 stress in the respiratory tract of healthy individuals.
20 5.1.1.1.1. Diesel exhaust odor. The odor of diesel exhaust is considered by most people to be
21 objectionable; at high intensities, it may produce sufficient physiological and psychological
22 effects to warrant concern for public health. The intensity of the odor of diesel exhaust is an
23 exponential function of its concentration such that a tenfold change in the concentration will alter
24 the intensity of the odor by one unit. Two human panel rating scales have been used to measure
25 diesel exhaust odor intensity. In the first (Turk, 1967), combinations of odorous materials were
26 selected to simulate diesel exhaust odor; a set of 12 mixtures, each having twice the
27 concentration of that of the previous mixture, is the basis of the diesel odor intensity scale
28 (D-scale). The second method is the TIA (total intensity of aroma) scale based on seven steps,
29 ranging from 0 to 3, with 0 being undetectable, l/2 very slight, and 1 slight and increasing in
30 one-half units up to 3, strong (Odor Panel of the CRC-APRAC Program Group on Composition
31 of Diesel Exhaust, 1979; Levins, 1981).
32 Surveys, utilizing volunteer panelists, have been taken to evaluate the general public's
33 response to the odor of diesel exhaust. Hare and Springer (1971) and Hare et al. (1974) found
34 that at a D rating of about 2 (TIA = 0.9, slight odor intensity), about 90% of the participants
perceived the odor, and almost 60% found it objectionable. At a D rating of 3.2 (TIA = 1.2,
7/25/00 5-3 DRAFT—DO NOT CITE OR QUOTE
-------
1 slight to moderate odor intensity), about 95% perceived the odor, and 75% objected to it, and, at
2 a D rating of 5 (TIA =1.8, almost moderate), about 95% objected to it.
3 Linnell and Scott (1962) reported odor threshold measurement in six subjects and found
4 that the dilution factor needed to reach the threshold ranged from 140 to 475 for this small
5 sample of people. At these dilutions, the concentrations of formaldehyde ranged from 0.012 to
6 0.088 ppm.
7
8 5.1.1.1.2. Pulmonary and respiratory effects. Battigelli (1965) exposed 13 volunteers to three
9 dilutions of diesel exhaust obtained from a one-cylinder, four-cycle, 7-hp diesel engine (fuel type
10 unspecified) and found that 15-min to 1 -h exposures had no significant effects on pulmonary
11 resistance. Pulmonary resistance was measured by plethysmography utilizing the simultaneous
12 recording of esophageal pressure and airflow determined by electrical differentiation of the
13 volume signal from a spirometer. The concentrations of the constituents in the three diluted
14 exhausts were 1.3, 2.8, and 6.2 ppm NO2; 0.2, 0.5, and 1 ppm SO2; <20, 30, and 55 ppm CO; and
15 <1.0, <1 to 2, and 1 to 2 ppm total aldehydes, respectively. DPM concentrations were not
16 reported.
17 A number of studies have evaluated changes in pulmonary function occurring over a
18 workshift hi workers occupationally exposed to diesel exhaust (specific time period not always
19 reported but assumed to be 8 h). In a study of coal miners, Reger (1979) found that both forced
20 expiratory volume in 1 s (FEV,) and forced vital capacity (FVC) decreased by 0.05 L in
21 60 diesel-exposed miners, an amount not substantially different from reductions seen in
22 non-diesel-exposed miners (0.02 and 0.04 L, respectively). Decrements in peak expiratory flow
23 rates were similar between diesel and non-diesel exhaust-exposed miners. Miners with a history
24 of smoking had an increased number of decrements over the shift than nonsmokers did.
25 Although the monitoring data were not reported, the authors stated that there was no relationship
26 between the low concentrations of measured respirable dust or NO2 (personal samplers) when
27 compared with shift changes for any lung function parameter measured for the diesel-exposed
28 miners. This study is limited because results were preliminary (abstract) and there was
29 incomplete information on the control subjects.
30 Ames ct al, (1982) compared the pulmonary function of 60 coal miners exposed to diesel
31 exhaust with that of a control group of 90 coal miners not exposed to diesel exhaust for evidence
32 of acute respiratory effects associated with exposure to diesel exhaust. Changes over the
33 workshift in FVC, FEV,, and forced expiratory flow rate at 50% FVC (FEF50) were the indices
34 for acute respiratory effects. The environmental concentrations of the primary pollutants were
35 2.0 mg/m3 respirable dust (<10 um MMAD), 0.2 ppm NO,, 12 ppm CO. and 0.3 ppm
36 formaldehyde. The investigators reported a statistically significant decline in FVC and FEV,
7/25/00 5-4 DRAFT—DO NOT CITE OR QUOTE
-------
over the workshift in both the diesel-exposed and comparison groups. Current smokers had
greater decrements in FVC, FEV,, and FEF50 than did ex-smokers and nonsmokers. There was a
3 marked disparity between the ages and the time spent underground for the two study groups.
4 Diesel-exposed miners were about 15 years younger and had worked underground for 15 fewer
5 years (4.8 versus 20.7 years) than miners not exposed to diesel exhaust. The significance to the
6 results of these differences between the populations is difficult to ascertain.
7 Except for the expected differences related to age, 120 underground iron ore miners
8 exposed to diesel exhaust had no workshift changes in FVC and FEV, when compared with
9 120 matched surface miners (Jorgensen and Svensson, 1970). Both groups had equal numbers
10 (30) of smokers and nonsmokers. The frequency of bronchitis was higher among underground
11 workers, much higher among smokers than nonsmokers, and also higher among older than
12 younger workers. The authors reported that the underground miners had exposures of 0.5 to
13 1.5 ppm NO2 and between 3 and 9 mg/m3 particulate matter, with 20% to 30% of the particles
14 <5 um MMAD. The majority of the particles were iron ore; quartz was 6% to 7% of the fraction
15 <5 um MMAD.
16 Gamble et al. (1979) measured preshift FEV, and FVC in 187 salt miners and obtained
17 peak flow forced expiratory flow rates at 25%, 50%, and 75% of FVC (FEF25, FEF50, or FEF75).
*Postshift pulmonary function values were determined from total lung capacity and flows at
preshift percentages of FVC. The miners were exposed to mean NO2 levels of 1.5 ppm and mean
20 respirable particulate levels of 0.7 mg/m3. No statistically significant changes were found
21 between changes in pulmonary function and in NO2 and respirable particles combined. Slopes of
22 the regression of NO2 and changes in FEV,, FEF25, FEF50, and FEF75 were significantly different
23 from zero. The authors concluded that these small reductions in pulmonary function were
24 attributable to variations in NO2 within each of the five salt mines that contributed to the cohort.
25 Gamble et al. (1987a) investigated the acute effects of diesel exhaust in 232 workers in
26 four diesel bus garages using an acute respiratory questionnaire and before and after workshift
27 spirometry. The prevalence of burning eyes, headaches, difficult or labored breathing, nausea,
28 and wheeze experienced at work was higher in the diesel bus garage workers than in a
29 comparison population of lead/acid battery workers who had not previously shown a statistically
30 significant association of acute symptoms with acid exposure. Comparisons between the two
31 groups were made without adjustment for age and smoking. There was no detectable association
32 of exposure to NO2 (0.23 ppm ± 0.24 S.D.) or inhalable (less than 10 um MMAD) particles
33 (0.24 mg/m3 ± 0.26 S.D.) and acute reductions in FVC, FEV,, peak flows, FEF50, and FEF7S.
34 Workers who had respiratory symptoms had slightly greater but statistically insignificant
reductions in FEV, and FEF50.
7/25/00 5-5 DRAFT—DO NOT CITE OR QUOTE
-------
1 Ulfvarson et al. (1987) evaluated workshift changes in the pulmonary function of 17 bus
2 garage workers, 25 crew members of two types of car ferries, and 37 workers on roll-on/roll-off
3 ships. The latter group was exposed primarily to diesel exhaust; the first two groups were
4 exposed to both gasoline and diesel exhaust. The diesel-only exposures that averaged 8 h
5 consisted of 0.13 to 1.0 mg/m3 particulate matter, 0.02 to 0.8 mg/m3 (0.016 to 0.65 ppm) NO,
6 0.06 to 2.3 mg/m3 (0.03 to 1.2 ppm) NO2, 1.1 to 5.1 mg/m3 (0.96 to 4.45 ppm) CO, and up to
7 0.5 mg/m3 (0.4 ppm) formaldehyde. The largest decrement in pulmonary function was observed
8 during a workshift following no exposure to diesel exhaust for 10 days. Forced vital capacity
9 and FEV, were significantly reduced over the workshift (0.44 L and 0.30 L, /K0.01 and /KO.OO 1,
10 respectively). There was no difference between smokers and nonsmokers. Maximal
11 midexpiratory flow, closing volume expressed as the percentage of expiratory vital capacity, and
12 alveolar plateau gradient (phase 3) were not affected. Similar but less pronounced effects on
13 FVC (-0.16 L) were found in a second, subsequent study of stevedores (n = 24) only following
14 5 days of no exposure to diesel truck exhaust. Pulmonary function returned to normal after
15 3 days without occupational exposure to diesel exhaust. No exposure-related correlation was
16 found between the observed pulmonary effects and concentrations of NO, NO2, CO, or
17 formaldehyde; however, it was suggested that NO2 adsorbed onto the diesel exhaust particles
18 may have contributed to the overall dose of NO2 to the lungs. In a related study, six workers (job
19 category not defined) were placed in an exposure chamber and exposed to diluted diesel exhaust
20 containing 0.6 mg/m3 DPM and 3.9 mg/m3 (2.1 ppm) NO2. The exhaust was generated by a
21 6-cylinder, 2.38-L diesel engine, operated for 3 h and 40 min at constant speed, equivalent to
22 60 km/h, and at about one-half full engine load. No effect on pulmonary function was observed.
23 The relationship between traffic density and respiratory health in children has been
24 examined in a series of studies in Holland in children attending schools located near major
25 freeways. Cough, wheeze, runny nose, and doctor-diagnosed asthma were reported more often
26 for children living within 100 m of freeways carrying between 80,000 and 150,000 vehicles per
27 day (van Vliet et al., 1997). Separate counts for truck traffic indicated a range from 8,000 to
28 17.500 trucks per day. Truck traffic intensity and the concentration of black smoke, considered
29 by the authors to be a proxy for DPM, measured in schools were found to be significantly
30 associated with chronic respiratory symptoms, with the relationships being more pronounced in
31 girls than in boys.
32 Brunekreef et ai. (1997) measured iung function in children in six areas located near
33 major motorways and assessed their exposure to traffic-related air pollution using separate traffic
34 counts for automobiles and trucks. They also measured air pollution in the children's schools.
35 While lung function was associated with track traffic density, there was a lesser association with
36 automobile traffic density. The association was stronger in those children living closest (300 m)
7/25/00 5-6 DRAFT—DO NOT CITE OR QUOTE
-------
to the roadways. Lung function was also associated with concentration of black smoke,
measured inside the schools. The associations were stronger in girls than in boys. The authors
3 conclude that exposure to vehicular pollution, hi particular DPM, may lead to reduced lung
4 function in children living near major motorways.
5 In a follow-up study of traffic-related air pollution and its effect on the respiratory health
6 of children living near roadways, Brunekreef et al. (2000) showed that the intensity of truck
7 traffic was significantly associated with the prevalence of wheeze, phlegm, bronchitis, eye
8 symptoms, and allergy to dust and pets. Associations with yearly averaged PM2 5 and "soot"
9 concentrations measured inside and outside the schools showed similar patterns. Truck traffic
10 intensity was also significantly associated with a positive skin prick test or elevated IgE for
11 outdoor allergens. There were no associations between traffic intensity or PM2 5 and "soot"
12 concentrations and lung function, bronchial responsiveness, and allergic reactions to indoor
13 allergens. Further analysis of the data showed that the associations between traffic-related air
14 pollution and symptoms were almost entirely related to children with bronchial hyperreactivity or
15 sensitization to common allergens.
16
17 5.1.1.1.3. Immunological effects. Salvi et al. (1999) exposed healthy human subjects to diluted
18 diesel exhaust (DPM 300 ng/m3) for 1 h with intermittent exercise. Although there were no
T^P changes in pulmonary function, there were significant increases in neutrophils and B
20 lymphocytes as well as histamine and fibronectin in airway lavage fluid. Bronchial biopsies
21 obtained 6 h after diesel exhaust exposure showed a significant increase in neutrophils, mast
22 cells, and CD4+ and CD8+ T lymphocytes, along with upregulation of the endothelial adhesion
23 molecules ICAM-1 and VCAM-1 and increases in the number of LFA-1+ in the bronchial tissue.
24 Significant increases in neutrophils and platelets were observed in peripheral blood following
25 exposure to diesel exhaust.
26 In a follow-up investigation of potential mechanisms underlying the DE-induced airway
27 leukocyte infiltration, Salvi et al. (2000) exposed healthy human volunteers to diluted DE, on two
28 separate occasions for 1 h each, hi an exposure chamber. Fiber-optic bronchoscopy was
29 performed 6 h after each exposure to obtain endobronchial biopsies and bronchial wash (BW)
30 cells. These workers observed that DE exposure enhanced gene transcription of IL-8 in the
31 bronchial tissue and BW cells and increased growth-regulated oncogene-a protein expression
32 and IL-8 in the bronchial epithelium; there was also a trend toward an increase in IL-5 mRNA
33 gene transcripts in the bronchial tissue.
34 In an attempt to evaluate the potential allergenic effects of DPM in humans, Diaz-
Sanchez and associates carried out a series of clinical investigations. In the first of these (Diaz-
7/25/00 5-7 DRAFT—DO NOT CITE OR QUOTE
-------
1 Sanchez et al., 1994), healthy human volunteers were challenged by spraying either saline or 0.30
2 mg DPM into their nostrils. This dose was considered equivalent to total exposure on 1-3
3 average days in Los Angeles, but could occur acutely in certain nonoccupational settings such as
4 sitting at a busy bus stop or in an express tunnel. Enhanced IgE levels were noted in nasal lavage
5 cells in as little as 24 h, with peak production observed 4 days after DPM challenge. The effects
6 seemed to be somewhat isotype-specific, because in contrast to IgE results, DPM challenge had
7 no effect on the levels of IgG, IgA, IgM, or albumin. The selective enhancement of local IgE
8 production was demonstrated by a dramatic increase in IgE-secreting cells.
9 Although direct effects of DPM on B-cells have been demonstrated by in vitro studies, it
10 was considered likely that other cells regulating the IgE response may also be affected. Cytokine
11 production was therefore measured in nasal lavage cells from healthy human volunteers
12 challenged with DPM (0 or 0.15 mg in 200 uL saline) sprayed into each nostril (Diaz-Sanchez
13 et al., 1996). Before challenge with DPM, most subjects' nasal lavage cells had detectable levels
14 of only interferon-y, IL-2, and IL-13 mRNA. After challenge with DPM, the cells produced
15 readily detectable levels of wRNA for IL-2, IL-4, IL-5, IL-6, IL-10, IL-13, and interferon-Y-
16 In addition, all levels of cytokine wRNA were increased. Although the cells in the nasal lavage
17 before and after challenge do not necessarily represent the same ones either in number or type,
18 the broad increase in cytokine production was not simply the result of an increase in T cells
19 recovered in the lavage fluid. On the basis of these findings, the authors concluded that the
20 increase in nasal cytokine expression after exposure to DPM can be predicted to contribute to
21 enhanced local IgE production and thus play a role in pollutant-induced airway disease.
22 The ability of DPM to act as an adjuvant to the ragweed allergen Amb a I was also
23 examined by nasal provocation in ragweed-allergic subjects using 0.3 mg DPM, Amb a I, or both
24 (Diaz-Sanchez et al., 1997). Although allergen and DPM each enhanced ragweed-specific IgE,
25 DPM plus allergen promoted a 16-times greater antigen-specific IgE production. Nasal challenge
26 with DPM also influenced cytokine production. Ragweed challenge resulted in a weak response,
27 DPM challenge caused a strong but nonspecific response, and allergen plus DPM caused a
28 significant increase in the expression of mRNA for THO and TH2-type cytokines (IL-4, IL-53
29 IL-6, IL-10, IL-13), with a pronounced inhibitory effect on IFN-y gene expression. The author
30 concluded that DPM can enhance B-cell differentiation and, by initiating and elevating IgE
31 production, may be a factor in the increased incidence of allergic airway disease.
32 In a further extension of these studies, Diaz-Sanchez et al. (1999) examined the potential
33 for DPM to lead to primary sensitization of humans by driving a de novo mucosal IgE response
34 to a neoantigen. keyhole limpet hemocyanin (KLH). Ten atopic subjects were given an initial
35 nasal immunization of KLH followed by two biweekly nasal challenges with KT.T-T Fifteen
36 different atopic subjects were treated identically, except that DPM was administered 24 h before
7/25/00 5-8 DRAFT—DO NOT CITE OR QUOTE
-------
each KLH exposure. Intranasal administration of KLH alone led to the generation of an anti-
KLH IgG and IgA humoral response, which was detected in nasal fluid samples. No anti-KLH
3 IgE was observed in any of these subjects. In contrast, when challenged with KLH preceded by
4 DPM, 9 of the 15 subjects produced anti-KLH-specific IgE. KLH-specific IgG and IgA at levels
5 similar to those seen with KLH alone were also detected. Subjects who received DPM and KLH
6 had significantly increased IL-4, but not IFN-gamma, levels in nasal lavage fluid, whereas these
7 levels were unchanged in subjects receiving KLH alone. These investigators concluded that
8 DPM can function as a mucosal adjuvant to a de novo IgE response and may increase allergic
9 sensitization.
10
11 5.1.1.1.4. Human cell culture studies. The potential mechanisms by which DPM may act to
12 cause allergenic effects has been examined in human cell culture studies. Takenaka et al. (1995)
13 reported that DPM extracts enhanced IgE production from purified human B cells. Interleukin-4
14 plus monoclonal antibody-stimulated IgE production was enhanced 20% to 360% by the addition
15 of DPM extracts over a period of 10-14 days. DPM extracts themselves did not induce IgE
16 production or synergize with interleukin-4 alone to induce IgE from purified B cells, suggesting
17 that the extracts were enhancing ongoing IgE production rather than inducing germline
« transcription or isotype switching. The authors concluded that enhancement of IgE production in
the human airway resulting from the organic fraction of DPM may be an important factor in the
20 increasing incidence of allergic airway disease.
21 Steerenberg et al. (1998) studied the effects of exposure to DPM on airway epithelial
22 cells, the first line of defense against inhaled pollutants. Cells from a human bronchial cell line
23 (BEAS-2B) were cultured in vitro and exposed to DPM (0.04-0.33 mg/mL) and the effects on
24 IL-6 and IL-8 production were observed. Increases in IL-6 and IL-8 production compared to the
25 nonexposed cells (11- and 4-fold, respectively) were found after 24 or 48 h exposure to DPM.
26 This increase was lower (17- and 3.3-fold) compared to silica and higher compared to titanium
27 dioxide, which showed no increase for either IL-6 or IL-8. The study was extended to observe
28 the effects of DPM on inflammation-primed cells. BEAS-2B cells were exposed to TNF-a
29 followed by DPM. Additive effects on IL-6 and IL-8 production by BEAS-2B cells were found
30 after TNF-a priming and subsequent exposure to DPM only at a low dose of DPM and TNF-a
31 (0.05-0.2 ng/mL). The investigators concluded that BEAS-2B phagocytized DPM and produced
32 an increased amount of IL-6 and IL-8, and that in TNF-oc-primed BEAS-2B cells DPM increased
33 interleukin production only at low concentrations of DPM and TNF-a.
34 Ohtoshi et al. (1998) studied the effect of suspended paniculate matter (SPM), obtained
^fc from high-volume air samplers, and DPM on the production of IL-8 and granulocyte-colony
36 stimulating factor (GM-CSF) by human airway epithelial cells in vitro. Nontoxic doses of DPMs
7/25/00 5-9 DRAFT—DO NOT CITE OR QUOTE
-------
1 stimulated production of IL-8 and GM-CSF by three kinds of human epithelial cells (nasal
2 polyp-derived upper airway, normal bronchial, and transformed bronchial epithelial cells) in a
3 dose- and time-dependent fashion. SPM had a stimulatory effect on GM-CSF, but not on IL-8
4 production. The effects could be blocked with a protein synthesis inhibitor, suggesting that the
5 process required de novo protein synthesis, and appeared to be due to an extractable component
6 because neither charcoal nor graphite showed such stimulatory effects. The authors concluded
7 that SPM and DPM, a component of SPM, may be important air pollutants in the activation of
8 airway cells for the release of cytokines relevant to allergic airway inflammation.
9 The mechanisms underlying DPM-induced injury to airway cells were investigated in
10 human bronchial epithelial cells (HBECs) in culture (Bayram et al., 1998a). HBECs from
11 bronchial explants obtained at surgery were cultured and exposed to DPM (10-100 ng/mL)
12 suspended in a serum-free supplemented medium (SF-medium) or to a SF-medium filtrate of
13 DPM. The filtrate was obtained by incubating DPM (50 ng/mL) in SF-medium for 24 h. The
14 effects of DPM and DPM filtrate on permeability, ciliary beat frequency (CBF), and release of
15 inflammatory mediators were observed. DPM and filtered solution of DPM significantly
16 increased the electrical resistance of the cultures but did not affect movement of bovine serum
17 albumin across cell cultures. DPM and filtered DPM solution significantly attenuated the CBF of
18 these cultures and significantly increased the release of IL-8. DPM also increased the release by
19 these cultures of GM-CSF and soluble intercellular adhesion molecule-1 (sICAM-1). These
20 authors also observed that activated charcoal was not able to induce changes in electrical
21 resistance, attenuate CBF, and increase the release of inflammatory mediators from HBEC, and
22 proposed that these effects were due most likely to the compounds adsorbed onto the DPM rather
23 than the size of DPM. The authors concluded that exposure of airway cells to DPM may lead to
24 functional changes and release of proinflammatory mediators and that these effects may influence
25 the development of airway disease.
26 Bayram et al. (1998b) investigated the sensitivity of cultured airway cells from asthmatic
27 patients to DPM. Incubation with DPM significantly attenuated the CBF in both the asthmatic
28 and nonasthmatic bronchial epithelial cell cultures. Cultured airway cells from asthmatic patients
29 constitutively released significantly greater amounts of IL-8, GM-CSF, and sICAM-1 than cell
30 cultures from nonasthmatic subjects. Only cultures from asthmatic patients additionally released
31 RANTES. The authors concluded that cultured airway cells from asthmatic subjects differ with
32 regard to the amounts and types of proinflammatory mediators they can release and that the
33 increased sensitivity of bronchial epithelial cells of asthmatic subjects to DPM may result in
34 exacerbation of their disease symptoms.
35 Devalia et al. (1999*) investigated the potential sensitivity of HBECs biopsied from atopic
36 mild asthmatic patients and non-atopic nonasthmatic subjects to DPM. HBECs from asthmatic
7/25/00 5-10 DRAFT—DO NOT CITE OR QUOTE
-------
patients constitutively released significantly greater amounts of IL-8, GM-CSF, and sICAM-1
than HBECs from nonasthmatic subjects. RANTES was only released by HBECs of asthmatic
3 patients. Incubation of the asthmatic cultures with 10 p,g/mL DPM significantly increased the
4 release of IL-8, GM-CSF, and sICAM-1 after 24 h. In contrast, only higher concentrations (50-
5 100 jig/mL DPM) significantly increased the release of IL-8 and GM-CSF from HBECs of
6 nonasthmatics. The authors conclude that the increased sensitivity of the airways of asthmatics
7 to DPM may be, at least in part, a consequence of greater constitutive and DPM-induced release
8 of specific pro-inflammatory mediators from bronchial epithelial cells.
9 To elucidate the intracellular signal transduction pathway regulating IL-8 and RANTES
10 production, Hashimoto et al. (2000) examined the role of p38 mitogen-activated protein (MAP)
11 kinase in DPM-induced IL-8 and RANTES production by HBECs. They also examined the
12 effect of a thiol-reducing agent, N-acetylcysteine (NAC), on DPM-induced p38 MAP kinase
13 activation and cytokine production. The authors conclude that p38 MAP kinase plays an
14 important role in the DPM-activated signaling pathway that regulates IL-8 and RANTES
15 production by HBECs and that the cellular redox state is critical for DPM-induced p38 MAP
16 kinase activation leading to IL-8 and RANTES production.
17 Boland et al. (1999) compared the biological effects of carbon black and DPM collected
«from catalyst- and noncatalyst-equipped diesel vehicles in cultures of both human bronchial
epithelial cells and human nasal epithelial cells. Transmission electron microscopy indicated that
20 DPM was phagocytosed by epithelial cells and translocated through the epithelial cell sheet. The
21 time and dose dependency of phagocytosis and its nonspecificity for different particles (DPM,
22 carbon black, and latex particles) were established by flow cytometry. DPM also induced a
23 time-dependent increase in interleukin-8, GM-CSF, and interleukin-lp release. The
24 inflammatory response occurred later than phagocytosis and, because carbon black had no effect
25 on cytokine release, its extent appeared to depend on the content of adsorbed organic compounds.
26 Furthermore, treatment of the exhaust gas to decrease the adsorbed organic fraction reduced the
27 DPM-induced increase in GM-CSF factor release. These results indicate that DPM can be
28 phagocytosed by and induce a specific inflammatory response hi airway epithelial cells.
29
30 5.1.1.1.5. Summary. In the available exposure studies, considerable variability is reported in
31 diesel exhaust detection threshold. The odor scales described in some of these studies have no
32 general use at present because they are not objectively defined; however, the studies do clearly
33 indicate substantial interindividual variability in the ability to detect odor and the level at which it
34 becomes objectionable. Much of what is known about the acute effects of diesel exhaust comes
^^ from case reports that lack clear measurements of exposure concentrations. The studies of
36 pulmonary function changes in exposed humans have looked for changes occurring over a
7/25/00 5-11 DRAFT—DO NOT CITE OR QUOTE
-------
1 workshift or after a short-term exposure. The overall conclusion of these studies is that
2 reversible changes in pulmonary function in humans can occur in relation to diesel exhaust
3 exposure, although it is not possible to relate these changes to specific exposure levels. Exposure
4 studies in humans and in isolated cell systems derived from humans reveal that DPM has the
5 potential to elicit inflammatory and immunological responses and responses typical of asthma;
6 DPM may be a likely factor in the increasing incidence of allergic hypersensitivity. These
7 studies have also shown that effects are due primarily to the organic fraction and that DPM
8 synergizes with known allergens to increase their effectiveness. Results from human cell culture
9 studies indicate that DPM has the potential to influence the development of airway inflammation
1 0 and disease through its adjuvant properties and by causing the release of proinflammatory
1 1 mediators.
12
13 5.1.1.2. Long-Term Exposures
1 4 Several epidemiologic studies have evaluated the effects of chronic exposure to diesel
1 5 exhaust on occupationally exposed workers.
1 6 Battigelli et al. (1964) measured several indices of pulmonary function, including vital
1 7 capacity, FEV,, peak flow, nitrogen washout, and diffusion capacity in 210 locomotive repairmen
1 8 exposed to diesel exhaust in 3 engine houses. The average exposure of these locomotive
1 9 repairmen to diesel exhaust was 9.6 years. When compared with a control group matched for
20 age, body size, "past extrapulmonary medical history" (no explanation given), and job status
21 (1 54 railroad yard workers), no significant clinical differences were found in pulmonary function
22 or in the prevalence of dyspnea, cough, or sputum between the diesel exhaust-exposed and
23 nonexposed groups. Exposure to diesel exhaust showed marked seasonal variations because the
24 doors of the engine house were open in the summer and closed in the winter. For the exposed
25 group, the maximum daily workplace concentrations of air pollutants measured were 1.8 ppm
26 NO2, 1 .7 ppm total aldehydes, 0. 1 5 ppm acrolein, 4.0 ppm SO2, and 5.0 ppm total hydrocarbons.
27 The concentration of airborne particles was not reported.
28 Gamble et al. (1987b) examined 283 diesel bus garage workers from four garages in two
29 cities to determine if there was excess chronic respiratory morbidity associated with exposure to
30 diesel exhaust. Tenure of employment was used as a surrogate of exposure; mean tenure of the
31 study population was 9 years ±10 years S.D. Exposure-effect relationships within the study
32 population showed no detectable associations of symptoms with tenure. Reductions in FVC,
33 FEV,, peak flow, and FEF50 (but not FEF75) were associated with increasing tenure. Compared
r»nTMilatir»n f7\f\ nnn*»vnr»c«»H Wiip-rollar wnrlrgrc^ atiH after inHirpot urliiictmi»nt fnr
f^vf «*_•» . W-- y . - — ..*, .-V~J~ .^-J— ~ . ~ • .._* ~~/ — — -— - «-— ~ - — ~ " — — -J —."-..— »-- - v_
36 wheezing; however, there was no correlation between symptoms and length of employment.
7/25/00 5- 1 2 DRAFT— DO NOT CITE OR QUOTE
-------
Dyspnea showed an exposure-response trend but no apparent increase in prevalence. Mean
FEV,, FVC, FEF50, and peak flow were not reduced in the total cohort compared with the
3 reference population, but were reduced in workers with 10 years or more tenure.
4 Purdham et al. (1987) evaluated respiratory symptoms and pulmonary function in
5 17 stevedores employed in car ferry operations who were exposed to both diesel and gasoline
6 exhausts and in a control group of 11 on-site office workers. Twenty-four percent of the exposed
7 group and 36% of the controls were smokers. If a particular symptom was considered to be
8 influenced by smoking, smoking status was used as a covariate in the logistic regression analysis;
9 pack-years smoked was a covariate for lung function indices. The frequency of respiratory
10 symptoms was not significantly different between the two groups; however, baseline pulmonary
11 function measurements were significantly different. The latter comparisons were measured by
12 multiple regression analysis using the actual (not percentage predicted) results and correcting for
13 age, height, and pack-years smoked. The stevedores had significantly lower FEV,, FEV,/FVC,
14 FEFSO, and FEF75 (p<0.021,;7<0.023,/7<0.001, and/KO.008, respectively), but not FVC. The
15 results from the stevedores were also compared with those obtained from a study of the
16 respiratory health status of Sydney, Nova Scotia, residents. These comparisons showed that the
17 dock workers had higher FVC, similar FEV,, but lower FEV,/FVC and flow rates than the
18 residents of Sydney. Based on these consistent findings, the authors concluded that the lower
^P baseline function measurements in the stevedores provided evidence of an obstructive ventilatory
20 defect, but caution in interpretation was warranted because of the small sample size. There were
21 no significant changes in lung function over the workshift, nor was there a difference between the
22 two groups. The stevedores were exposed to significantly (p<0.04) higher concentrations of
23 particulate matter (0.06 to 1.72 mg/m3, mean 0.50 mg/m3) than the controls (0.13 to 0.58 mg/m3,
24 mean not reported). Exposures of stevedores to SO2, NO2, aldehydes, and PAHs were very low;
25 occasional CO concentrations in the 20 to 100 ppm range could be detected for periods up to 1 h
26 in areas where blockers were chaining gasoline-powered vehicles.
27 Additional epidemiological studies on the health hazards posed by exposure to diesel
28 exhaust have been conducted for mining operations. Reger et al. (1982) evaluated the respiratory
29 health status of 823 male coal miners from six diesel-equipped mines compared with
30 823 matched coal miners not exposed to diesel exhaust. The average tenure of underground
31 work for the underground miners and their controls was only about 5 years; on average, the
32 underground workers in diesel mines spent only 3 of those 5 years underground in diesel-use
33 mines. Underground miners exposed to diesel exhaust reported a higher incidence of symptoms
34 of cough and phlegm but proportionally fewer symptoms of moderate to severe dyspnea than
^^ their matched counterparts. These differences in prevalence of symptoms were not statistically
36 significant. The diesel-exposed underground miners, on the average, had lower FVC, FEV,,
7/25/00 5-13 DRAFT—DO NOT CITE OR QUOTE
-------
1 FEF50, FEF75, and FEF^ but higher peak flow and FEF2J than their matched controls. These
2 differences, however, were not statistically significant. Health indicators for surface workers and
3 their matched controls were directionally the same as for matched underground workers. There
4 were no consistent relationships between the findings of increased respiratory symptoms,
5 decreased pulmonary function, smoking history, years of exposure, or monitored atmosphere
6 pollutants (NOX, CO, particles, and aldehydes). Mean concentrations of NOX at the six mines
7 ranged from 0 to 0.6 ppm for short-term area samples, 0.13 to 0.28 ppm for full-shift personal
8 samples, and 0.03 to 0.80 for full-shift area samples. Inhalable particles (less than 10 urn
9 MMAD) averaged 0.93 to 2.73 mg/m3 for personal samples and 0 to 16.1 mg/m3 for full-shift
10 area samples. Ames et al. (1984), using a portion of the miners studied by Reger, examined
11 280 diesel-exposed underground miners in 1977 and again in 1982. Each miner in this group had
12 at least 1 year of underground mining work history in 1977. The control group was 838 miners
13 with no exposure to diesel exhaust. The miners were evaluated for prevalence of respiratory
14 symptoms, chronic cough, phlegm, dyspnea, and changes in FVC, FEV,, and FEF50. No air
15 monitoring data were reported; exposure to diesel exhaust gases and mine dust particles were
16 described as very low. These authors found no decrements in pulmonary function or increased
17 prevalence of respiratory symptoms attributable to exposure to diesel exhaust. In fact, the 5-year
18 incidences of cough, phlegm, and dyspnea were greater in miners without exposure to diesel
19 exhaust.
20 Attfield (1978) studied 2,659 miners from 21 mines (8 metal, 6 potash, 5 salt, and
21 2 trona). Diesels were employed in only 18 of the mines, but the 3 mines not using diesels were
22 not identified. The years of diesel usage, ranging from 8 in trona mines to 16 in potash mines,
23 were used as a surrogate for exposure to diesel exhaust. Based on a questionnaire, an increased
24 prevalence of persistent cough was associated with exposure to aldehydes; this finding, however,
25 was not supported by the pulmonary function data. No adverse respiratory symptoms or
26 pulmonary function impairments were related to CO2, CO, NO2, inhalable dust, or inhalable
27 quartz. The author failed to comment on whether the prevalence of cough was related to the high
28 incidence (70%) of smokers in the cohort.
29 Questionnaire, chest radiograph, and spirometric data were collected by Attfield et al.
30 (1982) on 630 potash miners from six potash mines. These miners were exposed for an average
31 of 10 years (range 5 to 14 years) to 0.1 to 3.3 ppm NO2, 0.1 to 4.0 ppm aldehyde, 5 to 9 ppm CO,
32 and total dust concentrations of 9 to 23 mg/m3. No attempt was made to measure diesei-derived
33 particles separately from other dusts. The ratio of total to inhalable (<10 (im MMAD) dust
34 ranged from 2 to 11. An increased prevalence of respiratory symptoms was related solely to
35 smoking. No association was found between symptoms and tenure nf employment, dust
36 exposure, NO2, CO, or aldehydes. A higher prevalence of symptoms of cough and phlegm was
7/25/00 5-14 DRAFT—DO NOT CITE OR QUOTE
-------
1 found, but no differences in pulmonary function (FVC and FEV,) were found in these
(jm) diesel-exposed potash miners when compared with the predicted values derived from a logistics
3 model based on blue-collar workers working in nondusty jobs.
4 Gamble et al. (1983) investigated respiratory morbidity in 259 miners from 5 salt mines
5 in terms of increased respiratory symptoms, radiographic findings, and reduced pulmonary
6 function associated with exposure to NO2, inhalable particles (<10 um MMAB), or years worked
7 underground. Two of the mines used diesel extensively; no diesels were used in one salt mine.
8 Diesels were introduced into each mine in 1956,1957, 1963, or 1963 through 1967. Several
9 working populations were compared with the salt miner cohort. After adjustment for age and
10 smoking, the salt miners showed no increased prevalence of cough, phlegm, dyspnea, or airway
11 obstruction (FEV,/FVC) compared with aboveground coal miners, potash miners, or blue-collar
12 workers. The underground coal miners consistently had an elevated level of symptoms. Forced
13 expiratory volume at 1 s, FVC, FEF50, and FEF75 were uniformly lower for salt miners in relation
14 to all the comparison populations. There was, however, no association between changes in
15 pulmonary function and years worked, estimated cumulative inhalable particles, or estimated
16 NO2 exposure. The highest average exposure to participate matter was 1.4 mg/m3 (particle size
17 not reported, measurement includes NaCl). Mean NO2 exposure was 1.3 ppm, with a range of
0.17 ppm to 2.5 ppm. In a continuation of these studies, Gamble and Jones (1983) grouped the
salt miners into low-, intermediate-, and high-exposure categories based on tenure in jobs with
20 diesel exhaust exposure. Average concentrations of inhalable particles and NO2 were 0.40, 0.60,
21 and 0.82 mg/m3 and 0.64,1.77, and 2.21 ppm for the three diesel exposure categories,
22 respectively. A statistically significant concentration-response association was found between
23 the prevalence of phlegm in the salt miners and exposure to diesel exhaust (pO.OOOl) and a
24 similar, but nonsignificant, trend for cough and dyspnea. Changes in pulmonary function showed
25 no association with diesel tenure. In a comparison with the control group of nonexposed,
26 blue-collar workers, adjusted for age and smoking, the overall prevalence of cough and phlegm
27 (but not dyspnea) was elevated in the diesel-exposed workers. Forced expiratory volumes at 1 s
28 and FVC were within 4% of expected, which was considered to be within the normal range of
29 variation for a nonexposed population.
30 In a preliminary study of three subcohorts from bus company personnel (clerks [lowest
31 exposure], bus drivers [intermediate exposure], and bus garage workers [highest exposure])
32 representing different levels of exposure to diesel exhaust, Edling and Axelson (1984) found a
33 fourfold higher risk ratio for cardiovascular mortality in bus garage workers, even after adjusting
34 for smoking history and allowing for at least 10 years of exposure and 15 years or more of
3*j induction latency. Carbon monoxide was hypothesized as the etiologic agent for the increased
36 cardiovascular disease but was not measured. However, in a more comprehensive
7/25/00 5-15 DRAFT—DO NOT CITE OR QUOTE
-------
1 epidemiological study, Edling et al. (1987) evaluated mortality data covering a 32-year period for
2 a cohort of 694 bus garage employees and found no significant differences between the observed
3 and expected number of deaths from cardiovascular disease. Information on exposure
4 components and their concentrations was not reported.
5 The absence of reported noncancerous human health effects, other than infrequently
6 occurring effects related to respiratory symptoms and pulmonary function changes, is notable.
7 Unlike studies in laboratory animals, to be described later in this chapter, studies of the impact of
8 diesel exhaust on the defense mechanisms of the human lung have not been performed.
9 No direct evidence is available in humans regarding doses of diesel exhaust, gas phase,
10 participate phase, or total exhaust that lead to impaired particle clearance or enhanced
11 susceptibility to infection. A summary of epidemiology studies is presented in Table 5-1.
12 To date, no large-scale epidemiological study has looked for effects of chronic exposure
13 to diesel exhaust on pulmonary function. In the long-term longitudinal and cross-sectional
14 studies, a relationship was generally observed between work in a job with diesel exposure and
15 respiratory symptoms (such as cough and phlegm), but there was no consistent effect on
16 pulmonary function. The interpretation of these results is hampered by lack of measured diesel
17 exhaust exposure levels and the short duration of exposure in these cohorts. The studies are
18 further limited in that only active workers were included, and it is possible that workers who
19 have developed symptoms or severe respiratory disease are likely to have moved away from
20 these jobs. The relationship between work in a job with diesel exposure and respiratory
21 symptoms may be due to short-term exposure.
22
23 5.1.2. Laboratory Animal Studies
24 Because humans and laboratory animals show similar nonneoplastic responses to inhaled
25 particles (ILSI, 2000), animal studies have been conducted to assess the pathophysiologic effects
26 ofDPM. Because of the large number of statistical comparisons made in the laboratory animal
27 studies, and to permit uniform, objective evaluations within and among studies, data will be
28 reported as significantly different (i.e../?<0.05) unless otherwise specified. The exposure
29 regimens used and the resultant exposure conditions employed in the laboratory animal
30 inhalation studies are summarized in Tables 5-2 through 5-16. Other than the pulmonary
31 function studies performed by Wiester et al. (1980) on guinea pigs during their exposure in
32 inhalation chambers, the pulmonary function studies performed by other investigators, although
33 sometimes unreported, were interpreted as being conducted on the following day or thereafter
34 and not immediately following exposure.
35
7/25/00 5-16 DRAFT—DO NOT CITE OR QUOTE
-------
5.1.2.1. Acute Exposures
The acute toxicity of undiluted diesel exhaust to rabbits, guinea pigs, and mice was
3 assessed by Pattle et al. (1957). Four engine operating conditions were used, and 4 rabbits,
4 10 guinea pigs, and 40 mice were tested under each exposure condition for 5 h (no controls were
5 used). Mortality was assessed up to 7 days after exposure. With the engine operating under light
6 load, the exhaust was highly irritating but not lethal to the test species, and only mild tracheal and
7 lung damage was observed in the exposed animals. The exhaust contained 74 mg/m3 DPM
8 (particle size not reported), 560 ppm CO, 23 ppm NO2, and 16 ppm aldehydes. Exhaust
9 containing 5 mg/m3 DPM, 380 ppm CO, 43 ppm NO2, and 6.4 ppm aldehydes resulted in low
10 mortality rates (mostly below 10%) and moderate lung damage. Exhaust containing 122 mg/m3
11 DPM, 418 ppm CO, 51 ppm NO2, and 6.0 ppm aldehydes produced high mortality rates (mostly
12 above 50%) and severe lung damage. Exhaust containing 1,070 mg/m3 DPM, 1,700 ppm CO,
13 12 ppm NO2, and 154 ppm aldehydes resulted in 100% mortality in all three species. High CO
14 levels, which resulted in a carboxyhemoglobin value of 60% in mice and 50% in rabbits and
15 guinea pigs, were considered to be the main cause of death in the latter case. High NO2 levels
16 were considered to be the main cause of lung damage and mortality seen in the other three tests.
17 Aldehydes and NO2 were considered to be the main irritants in the light load test.
Kobayashi and Ito (1995) administered 1, 10, or 20 mg/kg DPM in phosphate-buffered
saline to the nasal mucosa of guinea pigs. The administration increased nasal airway resistance,
20 augmented increased airway resistance and nasal secretion induced by a histamine aerosol,
21 increased vascular permeability in dorsal skin, and augmented vascular permeability induced by
22 histamine. The increases in nasal airway resistance and secretion are considered typical
23 responses of nasal mucosa against allergic stimulation. Similar results were reported for guinea
24 pigs exposed via inhalation for 3 h to diesel exhaust diluted to DPM concentrations of either 1 or
25 3.2 mg/m3 (Kobayashi et al., 1997). These studies show that short-term exposure to DPM
26 augments nasal mucosal hyperresponsiveness induced by histamine in guinea pigs.
27 The effects of DPM and its components (extracted particles and particle extracts) on the
28 release of proinflammatory cytokines, interleukin-1 (IL-1), and tumor necrosis factor-OC (TNF-a)
29 by alveolar macrophages (AMs) were investigated by Yang et al. (1997). Rat AMs were
30 incubated with 0, 5, 10, 20, 50, or 100 ug/106 AM/mL of DPM, methanol-extracted DPM, or
31 equivalent concentrations of DPM at 37 °C for 24 h. At high concentrations, both DPM and
32 DPM extracts were shown to increase IL-1-like activity secreted by AMs, whereas extracted
33 particles had no effect. Neither particles, particle extracts, or extracted particles stimulated
34 secretion of TNF-CC. DPM inhibited lipid polysaccharide (LPS)-stimulated production of IL-1
^fc and TNF-a. In contrast, interferon (IFN)-v stimulated production of TNF-a was not affected by
36 DPM. Results of this study indicate that the organic fraction of exhaust particles is responsible
7/25/00 5-17 DRAFT—DO NOT CITE OR QUOTE
-------
1 for the effects noted. Stimulation of IL- 1 but not TNF-CC suggests that IL- 1 , but not TNF-a, may
2 play an important role in the development of DPM-induced inflammatory and immune responses.
3 The cellular mechanism involved in inhibiting increased release of IL-1 and TNF-a by LPS is
4 unknown, but may be a contributing factor to the decreased AM phagocytic activity and
5 increased susceptibility to pulmonary infection after prolonged exposure to DPM.
6 Takano et al. (1997) designed a study to evaluate the effects of DPM on the
7 manifestations of allergic asthma in mice, with emphasis on antigen-induced airway
8 inflammation, the local expression of IL-5, GM-CSF, IL-2 and IFN-y. and the production of
9 antigen-specific IgE and IgG. Male ICR mice were intratracheally instilled with ovalbumin
1 0 (OVA), DPM, and DPM+OVA. DPM was obtained from a 4JBl-type, light-duty 2.74 L, four-
1 1 cylinder Izuzu diesel engine operated at a steady speed of 1 ,500 rpm under a load of 1 0 torque
1 2 (kg/m). The OVA-group mice were instilled with 1 ug OVA at 3 and 6 weeks. The mice
1 3 receiving DPM alone were instilled with 1 00 ug DPM weekly for 6 weeks. The OVA + DPM
1 4 group received the combined treatment in the same protocol as the OVA and the DPM groups,
1 5 respectively. Additional groups were exposed for 9 weeks. DPM aggravated OVA-induced
1 6 airway inflammation, characterized by infiltration of eosinophils and lymphocytes and an
1 7 increase in goblet cells in the bronchial epithelium. DPM in combination with antigen markedly
1 8 increased IL-5 protein levels in lung tissue and bronchoalveoiar lavage supernatants compared
19 with either antigen or DPM alone. The combination of DPM and antigen induced significant
20 increases in local expression of IL-4, GM-CSF, and IL-2, whereas expression of IFN-y was not
2 1 affected. In addition, DPM exhibited adjuvant activity for the antigen-specific production of IgG
22 and IgE.
23
24 5.1.2.2. Short-Term and Subchronic Exposures
25 A number of inhalation studies have employed a regimen of 20 h/day, 7 days/week for
26 varying exposure periods up to 20 weeks to differing concentrations of airborne paniculate
27 matter, vapor, and gas concentrations of diluted diesel exhaust. Exposure regimens and
28 characterization of gas-phase components for these studies are summarized in Table 5-2.
29 Pepelko et al. (1980a) evaluated the pulmonary function of cats exposed under these conditions
30 for 28 days to 6.4 mg/m3 DPM. The only significant functional change observed was a decrease
31 in maximum expiratory flow rate at 10% vital capacity. The excised lungs of the exposed cats
32 appeared charcoal gray, with focal black spots visible on the pleura! surface. Pathologic changes
33 included a predominantly peribronchial localization of black-pigmented macrophages within the
34 alveoli characteristic of focal pneumonitis or alveolitis.
35 The effects of a short-term diesel exhaust exposure on arterial blood gases, pH, blood
36 buffering, body weight changes, lung volumes, and deflation pressure- volume (PV) curves of
-7 1^ C 1f\f\
T-\l~» A T7T1 T^/~\ X T/~\-TT /"lTT»T-« S~\T\
uru\r i — JL/O INV-H \^L in vjjtv
-------
young adult rats were evaluated by Pepelko (1982a). Exposures were 20 h/day, 7 days/week for
8 days to a concentration of 6.4 mg/m3 DPM in the nonirradiated exhaust (RE) and 6.75 mg/m3 in
3 the irradiated exhaust (IE). In spite of the irradiation, levels of gaseous compounds were not
4 substantially different between the two groups (Table 5-2). Body weight gains were significantly
5 reduced in the RE-exposed rats and to an even greater degree in rats exposed to IE. Arterial
6 blood gases and standard bicarbonate were unaffected, but arterial blood pH was significantly
7 reduced in rats exposed to IE. Residual volume and wet lung weight were not affected by either
8 exposure, but vital capacity and total lung capacity were increased significantly following
9 exposure to RE. The shape of the deflation PV curves were nearly identical for the control, RE,
10 and IE groups.
11 In related studies, Wiester et al. (1980) evaluated pulmonary function in 4-day-old guinea
12 pigs exposed for 20 h/day, 7 days/week for 28 days to IE having a concentration of 6.3 mg/m3
13 DPM. When housed in the exposure chamber, pulmonary flow resistance increased 35%, and a
14 small but significant sinus bradycardia occurred as compared with controls housed and measured
15 in control air chambers (/K0.002). Respiratory rate, tidal volume, minute volume, and dynamic
16 compliance were unaffected, as were lead-1 electrocardiograms.
17 A separate group of adult guinea pigs was necropsied after 56 days of exposure to IE, to
diluted RE, or to clean air (Wiester et al., 1980). Exposure resulted in a significant increase in
the ratio of lung weight to body weight (0.68% for controls, 0.78% for IE, and 0.82% for RE).
20 Heart/body weight ratios were not affected by exposure. Microscopically, there was a marked
21 accumulation of black pigment-laden AMs throughout the lung, with a slight to moderate
22 accumulation in bronchial and carinal lymph nodes. Hypertrophy of goblet cells in the
23 tracheobronchial tree was frequently observed, and focal hyperplasia of alveolar lining cells was
24 occasionally observed. No evidence of squamous metaplasia of the tracheobronchial tree,
25 emphysema, peribronchitis, or peribronchiolitis was noted.
26 White and Garg (1981) studied pathologic alterations in the lungs of rats (16 exposed and
27 8 controls) after exposure to diesel exhaust containing 6 mg/m3 DPM. Two rats from the
28 exposed group and one rat from the control group (filtered room air) were sacrificed after each
29 exposure interval of 6 h and 1, 3, 7, 14,28,42, and 63 days; daily exposures were for 20 h and
30 were 5.5 days/week. Evidence of AM recruitment and phagocytosis of diesel particles was found
31 at the 6-h sacrifice; after 24 h of exposure there was a focal, scattered increase in the number of
32 Type II cells. After 4 weeks of exposure, there were morphologic changes in size, content, and
33 shape of AM, septal thickening adjacent to clusters of AMs, and an appearance of inflammatory
34 cells, primarily within the septa. At 9 weeks of exposure, focal aggregations of particle-laden
*H macrophages developed near the terminal bronchi, along with an influx of PMNs, Type II cell
36 proliferation, and thickening of alveolar walls. The affected alveoli occurred in clusters that, for
7/25/00 5-19 DRAFT—DO NOT CITE OR QUOTE
-------
1 the most part, were located near the terminal bronchioles, but occasionally were focally located
2 in the lung parenchyma. Hypertrophy of goblet cells in the tracheobronchial tree was frequently
3 observed, and focal hyperplasia of alveolar lining cells was occasionally observed. No evidence
4 of squamous metaplasia of the tracheobronchial tree, emphysema, peribronchitis, or
5 peribronchiolitis was noted.
6 Mauderly et al. (1981) exposed rats and mice by inhalation to diluted diesel exhaust for
7 545 h over a 19-week period on a regimen of 7 h/day, 5 days/week at concentrations of 0, 0.21,
8 1.02, or 4.38 mg/m3 DPM. Indices of health effects were minimal following 19 weeks of
9 exposure. There were no significant exposure-related differences in mortality or body weights of
10 the rats or mice. There also were no significant differences in respiratory function (breathing
11 patterns, dynamic lung mechanics, lung volumes, quasi-static PV relationships, forced
12 expirograms, and CO-diffusing capacity) in rats; pulmonary function was not measured in mice.
13 No effect on trachea! mucociliary or deep lung clearances were observed in the exposed groups.
14 Rats, but not mice, had elevated immune responses in lung-associated lymph nodes at the two
15 higher exposure levels. Inflammation in the lungs of rats exposed to 4.38 mg/m3 DPM was
16 indicated by increases in PMNs and lung tissue proteases. Histopathologic findings included
17 AMs that contained DPM, an increase in Type II cells, and the presence of particles hi the
18 interstitium and tracheobronchial lymph nodes.
19 Kaplan et al. (1982) evaluated the effects of subchronic exposure to diesel exhaust on
20 rats, hamsters, and mice. The exhaust was diluted to a concentration of 1.5 mg/m3 DPM;
21 exposures were 20 h/day, 7 days/week. Hamsters were exposed for 86 days, rats and mice for
22 90 days. There were no significant differences hi mortality or growth rates between exposed and
23 control animals. Lung weight relative to body weight of rats exposed for 90 days was
24 significantly higher than the mean for the control group. Histological examination of tissues of
25 all three species indicated particle accumulation in the lungs and mediastinal lymph nodes.
26 Associated with the larger accumulations, there was a minimal increase in the thickness of the
27 alveolar walls, but the vast majority of the particles elicited no response. After 6 mo of recovery,
28 considerable clearance of the DPM from the lungs occurred in all three species, as evaluated by
29 gross pathology and histopathology. However, no quantitative estimate of clearance was
30 provided.
31 Toxic effects hi animals from acute exposure to diesel exhaust appear to be primarily
32 attributable to the gaseous components (i.e., mortality from CO intoxication and lung injury
33 caused by cellular damage resulting from NO2 exposure). The results from short-term exposures
34 indicate that rats experience minimal lung function impairment even at diesel exhaust levels
3R sufficiently high to cause histological and cytological changes in the lung. In subchronic studies
36 of durations of 4 weeks or more, frank adverse health effects are not readily apparent and, when
7/25/00 5-20 DRAFT—DO NOT CITE OR QUOTE
-------
found, are mild and result from exposure to concentrations of about 6 mg/m3 DPM and durations
of exposures of 20 h/day. There is ample evidence that subchronic exposure to lower levels of
3 diesel exhaust affects the lung, as indicated by accumulation of particles, evidence of
4 inflammatory response, AM aggregation and accumulation near the terminal bronchioles, Type II
5 cell proliferation, and thickening of alveolar walls adjacent to AM aggregates. Little evidence
6 exists, however, that subchronic exposure to diesel exhaust impairs lung function. Recent
7 studies have implicated the organic fraction of DPM in the induction of respiratory allergic
8 disease.
9
10 5.1.23. Chromic Exgwswes
11 5.1.2.3.1. Effects ®m gmwth and longevity. Changes in growth, body weight, absolute or
12 relative organ weights, and longevity can be measurable indicators of chronic toxic effects. Such
13 effects have been observed in some, but not all, of the long-term studies conducted on laboratory
14 animals exposed to diesel exhaust. There was limited evidence for an effect on survival in the
15 published chronic animal studies; deaths occurred intermittently early in one study in female rats
16 exposed to 3.7 mg/m3 DPM; however, the death rate began to decrease after 15 mo, and the
17 survival rate after 30 mo was slightly higher than that of the control group (Research Committee
for HERP Studies, 1988). Studies of the effects of chronic exposure to diesel exhaust on survival
and body weight or growth are detailed in Table 5-3.
20 Increased lung weights and lung-to-body weight ratios have been reported in rats, mice,
21 and hamsters. These data are summarized in Table 5-4. In rats exposed for up to 36 weeks to
22 0.25 or 1.5 mg/m3 DPM, lung wet weights (normalized to body weight) were significantly higher
23 in the 1.5 mg/m3 exposure group than control values after 12 weeks of exposure (Misiorowski
24 et al., 1980). Rats and Syrian hamsters were exposed for 2 years (five 16-h periods per week) to
25 diesel exhaust diluted to achieve concentrations of 0.7,2.2, and 6.6 mg/m3 DPM (Brightwell
26 et al., 1986). At necropsy, a significant increase in lung weight was seen in both rats and
27 hamsters exposed to diesel exhaust compared with controls. This finding was more pronounced
28 in the rats in which the increase was progressive with both duration of exposure and particulate
29 matter level. The increase was greatest at 30 mo (after the end of a 6-mo observation period in
30 the high-concentration male group where the lung weight was 2.7 times the control and at 24 mo
31 in the high-concentration female group [3.9 times control]). Heinrich et al. (1986a,b; see also
32 Stober, 1986) found a significant increase in wet and dry weights of the lungs of rats and mice
33 exposed at 4.24 mg/m3 DPM for 1 year in comparison with controls. After 2 years, the difference
34 was a factor of 2 (mice) or 3 (rats). After the same exposure periods, the hamsters showed
(fej increases of 50% to 75%, respectively. Exposure to equivalent filtered diesel exhaust caused no
36 significant effects in any of the species. Vinegar et al. (1980, 1981a,b) exposed hamsters to two
7/25/00 5-21 DRAFT—DO NOT CITE OR QUOTE
-------
1 levels of diesel exhaust with resultant concentrations of about 6 and 12 mg/m3 DPM for 8 h/day,
2 7 days/week for 6 mo. Both exposures significantly increased lung weight and lung-weight to
3 body-weight ratios. The difference between lung weights of exposed and control hamsters
4 exposed to 12 mg/m3 DPM was approximately twice that of those exposed to 6 mg/m3.
5 Heinrich et al. (1995) reported that rats exposed to 2.5 and 7 mg/m3 DPM for 18 h/day,
6 5 days/week for 24 mo showed significantly lower body weights than controls starting at day
7 200 in the high-concentration group and at day 440 in the low-concentration group. Body weight
8 in the low-concentration group was unaffected, as was mortality in any group. Lung weight was
9 increased in the 7 mg/m3 group starting at 3 mo and persisting throughout the study, while the
10 2.5 mg/m3 group showed increased lung weight only at 22 and 24 mo of exposure. Mice (NMRI
11 strain) exposed to 7 mg/m3 in this study for 13.5 mo had no increase in mortality and
12 insignificant decreases in body weight. Lung weights were dramatically affected, with increases
13 progressing throughout the study from 1.5-fold at 3 mo to 3-fold at 12 mo. Mice (NMRI and
14 C57BL/6N strains) were also exposed to 4.5 mg/m3 for 23 mo. In NMRI mice, the body weights
15 were reported to be significantly lower than controls, but the magnitude of the change is not
16 reported, so biological significance cannot be assessed. Mortality was slightly increased, but
17 statistical significance is not reported. The C57BL/6N mice showed minimal effects on body
18 weight and mortality, which were not statistically significant. Lung weights were dramatically
19 affected in both strains.
20 Nikula et al. (1995) exposed male and female F344 rats to DPM concentrations of 2.4 and
21 6.3 mg/m3 for 16 h/day, 5 days/week for 23 mo in a study designed to compare the effects of
22 DPM with those of carbon black. Significantly reduced survival was observed in males exposed
23 to 6.3 mg/m3 but not in females or at the lower concentration. Body weights were decreased by
24 exposure to 6.3 mg/m3 DPM in both male and female rats throughout the exposure period.
25 Significant increases in lung weight were first seen at 6 mo in the high-exposure group and at
26 12 to 18 mo in the low-exposure group.
27 No evidence was found in the published literature that chronic exposure to diesel exhaust
28 affected the weight of body organs other than the lung and heart (e.g.s liver, kidney, spleen, or
29 testes) (Table 5-4). Morphometric analysis of hearts from rats and guinea pigs exposed to 0.25,
30 0.75, or 1.5 mg/m3 DPM 20 h/day, 5.5 days/week for 78 weeks revealed no significant alteration
31 in mass at any exposure level or duration of exposure (Penney et al., 1981). The analysis
32 included relative wet weights of the right ventricle, left ventricle, combined atria, and ratio of
33 right to left ventricle. Vallyathan et al. (1986) found no significant differences in heart weights
34 and the ratio of heart weight to body weight between rats exposed to 2 mg/m3 DPM for 7 h/day,
35 5 davs/week for 24 mo and their resneetive clean-air rhamhpr rnnrrnlc >Jr> significant
36 differences were found in the lungs, heart, liver, spleen, kidney, and testes of rats exposed for
7/25/00 5-22 DRAFT—DO NOT CITE OR QUOTE
-------
52 weeks, 7 h/day, 5 days/week to diluted diesel exhaust containing 2 mg/m3 DPM compared
with their respective controls (Green et al., 1983).
3
4 5.1.23.2. Effects turn pwlmtammvy function. The effect of long-term exposure to diesel exhaust
5 on pulmonary function has been evaluated in laboratory studies of rats, hamsters, cats, and
6 monkeys. These studies are summarized in Table 5-5, along with more details on the exposure
7 characteristics, in general order of increasing dose (C * T) of DPM. The text will be presented
8 using the same approach.
Q Lewis et al. (1989) evaluated functional residual capacity and airway resistance and
10 conductance in 10 control and 10 diesel-exposed rats (2 mg/m3 DPM, 7 h/day, 5 days/week for
11 52 or 104 weeks). At the 104-week evaluation, the rats were also examined for maximum flow
12 volume impairments. No evidence of impaired pulmonary function as a result of the exposure to
13 diesel exhaust was found in rats. Lewis et al. (1989) exposed male cynomolgus monkeys to
14 diesel exhaust for 7 h/day, 5 days/week for 24 mo. Groups of 15 monkeys were exposed to air,
15 diesel exhaust (2 mg/m3), coal dust, or combined coal dust and diesel exhaust. Pulmonary
16 function was evaluated prior to exposure and at 6-mo intervals during the 2-year exposure,
17 including compliance and resistance, static and dynamic lung volumes, distribution of
ventilation, diffusing capacity, and maximum ventilatory performance. There were no effects on
lung volumes, diffusing capacity, or ventilation distribution, so there was no evidence of
20 restrictive disease. There was, however, evidence of obstructive airway disease as measured by
21 low maximal flows in diesel-exposed monkeys. At 18 mo of exposure, forced expiratory flow at
22 25% of vital capacity and forced expiratory flow normalized to FVC were decreased. The
23 measurement of forced expiratory flow at 40% of total lung capacity was significantly decreased
24 at 12, 18, and 24 mo of exposure. The finding of an obstructive effect in monkeys contrasts with
25 the finding of restrictive type effects in other laboratory animal species (Vinegar et al., 1980,
26 1981a; Mauderly et al., 1988; Pepelko et al., 1980b, 1981) and suggests a possible difference in
27 effect between primate and small animal respiratory tracts. In these monkeys there were no
28 specific histopathological effects reported (see next section), although particle aggregates were
29 reported in the distal airways, suggesting more small airway deposition.
30 Gross (1981) exposed rats for 20 h/day, 5.5 days/week for 87 weeks to diesel exhaust
31 containing 1.5 mg/m3 DPM. When the data were normalized (e.g., indices expressed in units of
32 airflow or volume for each animal by its own forced expiratory volume), there were no apparent
33 functionally significant changes occurring in the lungs at 38 weeks of exposure that might be
34 attributable to the inhalation of diesel exhaust. After 87 weeks of exposure, functional residual
(fy capacity (FRC) and its component volumes (expiratory reserve [ER] and residual volume [RV]),
36 maximum expiratory flow (MEF) at 40% FVC, MEF at 20% FVC, and FEV0, were significantly
7/25/00 5-23 DRAFT—DO NOT CITE OR QUOTE
-------
1 greater in the diesel-exposed rats. An observed increase in airflow at the end of the forced
2 expiratory maneuver when a decreased airflow would be expected from the increased FRC, ER,
3 and RV data (the typical scenario of human pulmonary disease) showed these data to be
4 inconsistent with known clinically significant health effects. Furthermore, although the lung
5 volume changes in the diesel-exposed rats could have been indicative of emphysema or chronic
6 obstructive lung disease, this interpretation was contradicted by the airflow data, which suggest
7 simultaneous lowering of the resistance of the distal airways.
8 Heinrich et al. (1982) evaluated the pulmonary function of rats exposed to a concentration
9 of 3.9 mg/m3 DPM for 7 to 8 h/day, 5 days/week for 2 years. When compared with a control
10 group, no significant changes in respiratory rate, minute volume, compliance, or resistance
11 occurred in the exposed group (number of rats per group was not stated).
12 Chinese hamsters (eight or nine per group) were exposed 8 h/day, 7 days/week, for 6 mo
13 to concentrations of either about 6 mg/m3 or about 12 mg/m3 DPM (Vinegar et al., 1980,
14 1981a,b). Vital capacity, vital capacity/lung weight ratio, residual lung volume by water
15 displacement, and CO2 diffusing capacity decreased significantly in hamsters exposed to 6 mg/m3
16 DPM. Static deflation volume-pressure curves showed depressed deflation volumes for
17 diesel-exposed hamsters when volumes were corrected for body weight and even greater
18 depressed volumes when volumes were corrected for lung weight. However, when volumes
19 were expressed as percentage of vital capacity, the diesel-exposed hamsters had higher lung
20 volumes at 0 and 5 cm H2O. In the absence of confirmatory histopathology, the authors
21 tentatively concluded that these elevated lung volumes and the significantly reduced diffusing
22 capacity in the same hamsters were indicative of possible emphysematous changes in the lung.
23 Similar lung function changes were reported in hamsters exposed at 12 mg/m3 DPM, but detailed
24 information was not reported. It was stated, however, that the decrease in vital capacity was
25 176% greater in the second experiment than in the first.
26 Mauderly et al. (1988; see also McClellan et al., 1986) examined the impairment of
27 respiratory function in rats exposed for 7 h/day, 5 days/week for 24 mo to diluted diesel exhaust
28 with 0.35, 3.5, or 7.1 mg/m3 DPM. After 12 mo of exposure to the highest concentration of
29 diesel exhaust, the exposed rats (n = 22) had lower total lung capacity (TLC). dynamic lung
30 compliance (C^, FVC, and CO diffusing capacity than controls (n = 23). After 24 mo of
31 exposure to 7.1 mg/m3 DPM, mean TLC, Cdyn, quasi-static chord compliance, and CO diffusing
32 capacity were significantly lower than control values. Nitrogen washout and percentage of FVC
33 expired in 0.1 s were significantly greater than control values. There was no evidence of airflow
34 obstruction The functional alterations were attributed to focal fibrotic and emphyssinatcus
3B legion? and thickened alveolar membranes observed by histologies! examination. Similar
36 functional alterations and histopathologic lesions were observed in the rats exposed to 3.5 mg/m3
7/25/00 5-24 DRAFT—DO NOT CITE OR QUOTE
-------
DPM, but such changes usually occurred later in the exposure period and were generally less
pronounced. There were no significant decrements in pulmonary function for the 0.35 mg/m3
3 group at any time during the study nor were there reported histopathologic changes in this group.
4 Additional studies were conducted by Heinrich et al. (1986a,b; see also Stober, 1986) on
5 the effects of long-term exposure to diesel exhaust on the pulmonary function of hamsters and
6 rats. The exhaust was diluted to achieve a concentration of 4.24 mg/m3 DPM; exposures were
7 for 19 h/day, 5 days/week for a maximum of 120 weeks (hamsters) or 140 weeks (rats). After
8 1 year of exposure to the diesel exhaust, the hamsters exhibited a significant increase in airway
9 resistance and a nonsignificant reduction in lung compliance. For the same time period, rats
10 showed increased lung weights, a significant decrease hi C^,,, and a significant increase in airway
11 resistance. These indices did not change during the second year of exposure.
12 Syrian hamsters and rats were exposed to 0.7,2.2, or 6.6 mg/m3 DPM for five 16-h
13 periods per week for 2 years (Brightwell et al., 1986). There were no treatment-related changes
14 in pulmonary function hi the hamster. Rats exposed to the highest concentration of diesel
15 exhaust exhibited changes in pulmonary function (data not presented) that were reported to be
16 consistent with a concentration-related obstructive and restrictive disease.
17 Pepelko et al. (1980b; 1981; see also Pepelko, 1982b) and Moorman et al. (1985) .
« measured the lung function of adult cats chronically exposed to diesel exhaust. The cats were
exposed for 8 h/day and 7 days/week for 124 weeks. Exposures were at 6 mg/m3 for the first
20 61 weeks and 12 mg/m3 from weeks 62 to 124. No definitive pattern of pulmonary function
21 changes was observed following 61 weeks of exposure; however, a classic pattern of restrictive
22 lung disease was found at 124 weeks. The significantly reduced lung volumes (TLC, FVC, FRC,
23 and inspiratory capacity [1C]) and the significantly lower single-breath diffusing capacity,
24 coupled with normal values for dynamic ventilatory function (mechanics of breathing), indicate
25 the presence of a lesion that restricts inspiration but does not cause airway obstruction or loss of
26 elasticity. This pulmonary physiological syndrome is consistent with an interstitial fibrotic
27 response that was later verified by histopathology (Plopper et al., 1983).
28 Pulmonary function impairment has been reported in rats, hamsters, cats, and monkeys
29 chronically exposed to diesel exhaust. In all species but the monkey, the pulmonary function
30 testing results have been consistent with restrictive lung disease. The monkeys demonstrated
31 evidence of small airway obstructive responses. The disparity between the findings in monkeys
32 and those in rats, hamsters, and cats could be in part the result of increased particle retention in
33 the smaller species resulting from (1) exposure to diesel exhaust that has higher airborne
34 concentrations of gases, vapors, and particles and/or (2) longer duration of exposure. The nature
^P of the pulmonary impairment is also dependent on the site of deposition and routes of clearance,
36 which are determined by the anatomy and physiology of the test laboratory species and the
7/25/00 5-25 DRAFT—DO NOT CITE OR QUOTE
-------
1 exposure regimen. The data on pulmonary function effects raise the possibility that diesel
2 exhaust produces small airway disease in primates compared with primarily alveolar effects in
3 small animals and that similar changes might be expected in humans and monkeys.
4 Unfortunately, the available data hi primates are too limited to draw clear conclusions.
5
6 5.1.23.3. Lung morphology, biochemistry, and lung lavage analysis. Several studies have
7 examined the morphological, histological, and histochemical changes occurring in the lungs of
8 laboratory animals chronically exposed to diesel exhaust. The histopathological effects of diesel
9 exposure hi the lungs of laboratory animals are summarized in Table 5-6, ranked hi order of
10 C x T. Table 5-6 also contains an expanded description of exposures.
1 1 Kaplan et al. (1 982) performed macroscopic and microscopic examinations of the lungs
1 2 of rats, mice, and hamsters exposed for 20 h/day, 7 days/week for 3 mo to diesei exhaust
1 3 containing 1 .5 mg/m3 DPM. Gross examination revealed diffuse and focal deposition of the
1 4 diesel particles that produced a grayish overall appearance of the lungs with scattered, denser
1 5 black areas. There was clearance of particles via the lymphatics to regional lymph nodes.
1 6 Microscopic examination revealed no anatomic changes in the upper respiratory tract; the
1 7 mucociliary border was normal in appearance. Most of the particles were in macrophages, but
1 8 some were free as small aggregates on alveolar and bronchiolar surfaces. The particle-laden
1 9 macrophages were often hi masses near the entrances of the lymphatic drainage and respiratory
20 ducts. Associated with these masses was a minimal increase in the thickness of the alveolar
21 walls; however, the vast majority of the particles elicited no response. After 6 mo of recovery,
22 the lungs of all three species contained considerably less pigment, as assessed by gross
23 pathological and histopathological examinations.
24 Lewis et al. (1989; see also Green et al., 1983) performed serial histological examinations
25 of rat lung tissue exposed to diesel exhaust containing 2 mg/m3 DPM for 7 h/day, 7 days/week
26 for 2 years. Accumulations of black-pigmented AMs were seen in the alveolar ducts adjacent to
27 terminal bronchioles as early as 3 mo of exposure, and particles were seen within the interstitium
28 of the alveolar ducts. These macular lesions increased hi size up to 12 mo of exposure. Collagen
29 or reticulum fibers were seen only rarely in association with deposited particles; the vast majority
30 of lesions showed no evidence of fibrosis. There was no evidence of focal emphysema with the
3 1 macules. Multifocal histiocytosis (24% of exposed rats) was observed only after 24 mo of
32 exposure. These lesions were most commonly observed subpleurally and were composed of
33 collections of degenerating macrophages and amorphous granular material within alveoli,
35 adjacent to collections of pigmented macrophages showed a narked Type II cell hyperplasia;
36 degenerative changes were not observed in Type I cells. Histological examination of lung tissue
7/25/00 5-26 DRAFT— DO NOT CITE OR QUOTE
-------
from monkeys exposed for 24 mo in the same regimen as used for rats revealed aggregates of
black particles, principally in the distal airways of the lung. Particles were present within the
3 cytoplasm of macrophages in the alveolar spaces as well as the interstitium. Fibrosis, focal
4 emphysema, or inflammation was not observed. No specific histopathological lesions were
5 reported for the monkey.
6 Nikula et al. (1997) reevaluated the lung tissue from this study. They concluded that
7 there were no significant differences in the amount of retained particulate matter between
8 monkeys and rats exposed under the same conditions. The rats, however, retained a greater
9 portion of the particulate matter in lumens of the alveolar ducts and alveoli than did the monkeys.
10 Conversely, monkeys retained a greater portion of the particulate material in the interstitium than
11 did rats. Aggregations of particle-laden macrophages in the alveoli were rare, and there were few
12 signs of particle-associated inflammation in the monkeys. Minimal histopathologic lesions were
13 detected in the interstitium.
14 Histopathological effects of diesel exhaust on the lungs of rats have been investigated by
15 the Health Effects Research Program on Diesel Exhaust (HERP) in Japan. Both light-duty (LD)
16 and heavy-duty (HD) diesel engines were used. The exhaust was diluted to achieve nominal
17 concentrations of 0.1 (LD only), 0.4 (LD and HD), 1 (LD and HD), 2 (LD and HD), and 4 (HD
only) mg/m3 DPM. Rats were exposed for 16 h/day, 6 days/week for 30 mo. No
histopathological changes were observed in the lungs of rats exposed to 0.4 mg/m3 DPM or less.
20 At concentrations above 0.4 mg/m3 DPM, severe morphological changes were observed. These
21 changes consisted of shortened and absent cilia in the trachea! and bronchial epithelium, marked
22 hyperplasia of the bronchiolar epithelium, and swelling of the Type II cellular epithelium. These
23 lesions appeared to increase in severity with increases hi exhaust concentration and duration of
24 exposure. There was no difference in the degree of changes in pulmonary pathology at the same
25 concentrations between the LD and the HD series.
26 Heinrich et al. (1982) investigated histological changes occurring in the respiratory tract
27 of hamsters exposed to diesel exhaust. Exposures were for 7 to 8 h/day, 5 days/week for
28 104 weeks to diesel exhaust diluted to achieve a concentration of 3.9 mg/m3 DPM. Significantly
29 higher numbers of hamsters in the group exposed to diesel exhaust exhibited definite
30 proliferative changes in the lungs compared with the groups exposed to particle-free diesel
31 exhaust or clean air. Sixty percent of these changes were described as adenomatous
32 proliferations.
33 Heinrich et al. (1995) reported increased incidence and severity of bronchioloalveolar
34 hyperplasia in rats exposed to 0.8, 2.5, and 7 mg/m3. The lesion in the lowest concentration
^fc group was described as very slight to moderate. Slight to moderate interstitial fibrosis also
36 increased in incidence and severity in all exposed groups, but incidences were not reported. This
7/25/00 5-27 DRAFT—DO NOT CITE OR QUOTE
-------
1 chronic study also exposed NMRI mice to 7 mg/m3 for 13.5 mo and both NMRI and C56BL/6N
2 mice to 4.5 mg/m3 for 24 mo. Noncancer histological endpoints are not discussed in any detail in
3 the report, which is focused on the carcinogenicity of diesel as compared with titanium dioxide
4 and carbon black.
5 Iwai et al. (1986) performed serial histopathology on the lungs of rats at 1, 3, 6, 12, and
6 24 mo of exposure to diesel exhaust. Exposures were for 8 h/day, 7 days/week for 24 mo; the
7 exposure atmosphere contained 4.9 mg/m3 DPM. At 1 and 3 mo of exposure, there were
8 minimal histological changes in the lungs of the exposed rats. After 6 mo of exposure, there
9 were particle-laden macrophages distributed irregularly throughout the lung and a proliferation of
1 0 Type II cells with adenomatous metaplasia in areas where the macrophages had accumulated.
1 1 After 1 year of exposure, foci of heterotrophic hyperplasia of ciliated or nonciliated bronchiolar
1 2 epithelium on the adjacent alveolar walls were more common, the quantity of deposited
1 3 particulate matter increased, and the number of degenerative AMs and proliferative lesions of
1 4 Type II or bronchiolar epithelial cells increased. After 2 years of exposure, there was a fibrous
1 5 thickening of the alveolar walls, mast cell infiltration with epithelial hyperplasia in areas where
1 6 the macrophages had accumulated, and neoplasms.
1 7 Heinrich et al. (1986a; see also Stober, 1986) performed histopathologic examinations of
1 8 the respiratory tract of hamsters, mice, and rats exposed to diesel exhaust that had 4 mg/m3 DPM.
1 9 Exposures were for 19 h/day, 5 days/week; the maximum exposure period was 120 weeks for
20 hamsters and mice and 140 weeks for rats. Histological examination revealed different levels of
21 response among the three species. In hamsters, the exhaust produced thickened alveolar septa,
22 bronchioloalveolar hyperplasia, and what were termed emphysematous lesions (diagnostic
23 methodology not described). In mice, bronchoalveolar hyperplasia occurred in 64% of the mice
24 exposed to the exhaust and in 5% of the controls. Multifocal alveolar lipoproteinosis occurred hi
25 71% and multifocal interstitial fibrosis occurred in 43% of the mice exposed to exhaust but hi
26 only 4% of the controls. In exposed rats, there were severe inflammatory changes in the lungs, as
27 well as thickened septa, foci of macrophages, and hyperplastic and metaplastic lesions.
28 Nikula et al. (1995) reported in detail the nonneoplastic effects in male and female
29 F344 rats exposed to 24 or 6.3 mg/m3 of DPM. At 3 rnc in the low-concentration group,
30 enlarged particle-containing macrophages were found with minimal aggregation. With higher
3 1 concentration and longer duration of exposure, the number and size of macrophages and
32 aggregates increased. Alveolar epithelial hyperplasia was found starting at 3 mo and in all rats at
33 6 mo. These lesions progressed to chronic active inflammation, alveolar proteinosis, and septal
34 fibrosis at 12 mo. Other lesions observed late in the study included bronchioiar-aiveoiar
35 metaplasia, squamcus nietaplasia, and squarnous Cysii. Tliis study reports in deiaii ihe
36 progression of lesions in diesel exhaust exposure and finds relatively little difference between the
7/25/00 5-28 DRAFT— DO NOT CITE OR QUOTE
-------
lesions caused by diesel exhaust exposure and exposure to similar levels of carbon black
particles.
3 The effects of diesel exhaust on the lungs of rats exposed to 8.3 ± 2.0 mg/m3 DPM were
4 investigated by Karagianes et al. (1981). Exposures were for 6 h/day, 5 days/week, for 4, 8, 16,
5 or 20 mo. Histological examinations of lung tissue noted focal aggregation of particle-laden
6 AMs, alveolar histiocytosis, interstitial fibrosis, and alveolar emphysema (diagnostic
7 methodology not described). Lesion severity was related to length of exposure. No significant
8 differences were noted in lesion severity among the diesel exhaust, the diesel exhaust plus coal
9 dust (5.8 ± 3.5 mg/m3), or the high-concentration (14.9 ± 6.2 mg/m3) coal dust exposure groups
10 following 20 mo of exposure.
11 Histological changes in the lungs of guinea pigs exposed to diluted diesel exhaust
12 containing either 0.25, 0.75, 1.5, or 6.0 mg/m3 DPM were reported by Barnhart et al. (1981;
13 1982). Exposures at 0.75 and 1.5 mg/m3 for 2 weeks to 6 mo resulted in an uptake of exhaust
14 particles by three alveolar cell types (AMs, Type I cells, and interstitial macrophages) and also by
15 granulocytic leukocytes (eosinophils). The alveolar-capillary membrane increased hi thickness
16 as a result of an increase in the absolute tissue volume of interstitium and Type II cells. In a
17 continuation of these studies, guinea pigs were exposed to diesel exhaust (up to 6.0 mg/m3 DPM)
«for 2 years (Barnhart et al., 1982). A minimal tissue response occurred at a concentration of 0.25
mg/m3 After 9 mo of exposure, there was a significant increase, about 30%, in Type I and II
20 cells, endothelial cells, and interstitial cells over concurrent age-matched controls; by 24 mo only
21 macrophages and Type II cells were significantly increased. As in the earlier study,
22 ultrastructural evaluation showed that Type I cells, AMs, and eosinophils phagocytized the diesel
23 particles. Exposure to 0.75 mg/m3 for 6 mo resulted in fibrosis in regions of macrophage clusters
24 and in focal Type II cell proliferation. No additional information was provided regarding the
25 fibrotic changes with increasing concentration or duration of exposure. With increasing
26 concentration/duration of diesel exhaust exposure, Type II cell clusters occurred hi some alveoli.
27 Intraalveolar debris was particularly prominent after exposures at 1.5 and 6.0 mg/m3 and
28 consisted of secretory products from Type II cells.
29 In studies conducted on hamsters, Pepelko (1982b) found that the lungs of hamsters
30 exposed for 8 h/day, 7 days/week for 6 mo to 6 or 12 mg/m3 DPM were characterized by large
31 numbers of black AMs in the alveolar spaces, thickening of the alveolar epithelium, hyperplasia
32 of Type II cells, and edema.
33 Lungs from rats and mice exposed to 0.35, 3.5, or 7.1 mg/m3 (0.23 to 0.26 um mass
34 median diameter [MMD]) for 7 h/day and 5 days/week showed pathologic lesions (Mauderly
^P et al., 1987a; Henderson et al., 1988). After 1 year of exposure at 7.1 mg/m3, the lungs of the rats
36 exhibited focal areas of fibrosis; fibrosis increased with increasing duration of exposure and was
7/25/00 5-29 DRAFT—DO NOT CITE OR QUOTE
-------
1 observable in the 3.5-mg/m3 group of rats at 18 mo. The severity of inflammatory responses and
2 fibrosis was directly related to the exposure level. In the 0.35 mg/m3 group of rats, there was no
3 inflammation or fibrosis. Although the mouse lungs contained higher burdens of diesel particles
4 per gram of lung weight at each equivalent exposure concentration, there was substantially less
5 inflammatory reaction and fibrosis than was the case in rats. Fibrosis was observed only in the
6 lungs of mice exposed at 7.1 mg/m3 and consisted of fine fibrillar thickening of occasional
7 alveolar septa.
8 Histological examinations were performed on the lungs of cats initially exposed to
9 6 mg/m3 DPM for 61 weeks and subsequently increased to 12 mg/m3 for Weeks 62 to 124 of
10 exposure. Plopper et al. (1983; see also Hyde et al., 1985) concluded from the results of this
11 study that exposure to diesel exhaust produced changes in both epithelial and interstitial tissue
12 compartments and that the focus of these lesions in the peripheral lung was the centriacinar
13 region where the alveolar ducts join the terminal conducting airways. This conclusion was based
14 on the following evidence. The epithelium of the terminal and respiratory bronchioles in
15 exposed cats consisted of three cell types (ciliated, basal, and Clara cells) compared with only
16 one type (Clara cells) in the controls. The proximal acinar region showed evidence of
17 peribronchial fibrosis and bronchiolar epithelial metaplasia. Type II cell hyperplasia was present
18 in the proximal interalveolar septa. The more distal alveolar ducts and the majority of the rest of
19 the parenchyma were unchanged from controls. Peribronchial fibrosis was greater at the end of
20 6 mo in clean air following exposure, whereas the bronchiolar epithelial metaplasia was most
21 severe at the end of exposure. Following an additional 6 mo in clean air, the bronchiolar
22 epithelium more closely resembled the control epithelial cell population.
23 Wallace et al. (1987) used transmission electron microscopy (TEM) to determine the
24 effect of diesel exhaust on the intravascular and interstitial cellular populations of the lungs of
25 exposed rats and guinea pigs. Exposed animals and matched controls were exposed to 0.25,
26 0.75, 1.5, or 6.0 mg/m3 DPM for 2, 6, or 10 weeks or 18 mo. The results inferred the following:
27 (1) exposure to 6.0 mg/m3 for 2 weeks was insufficient to elicit any cellular response, (2) both
28 species demonstrated an adaptive multicellular response to diesel exhaust, (3) increased numbers
29 of fibroblasts were found in the interstitium from week 6 of exposure through month 18, and
30 (4) there was no significant difference in either cell type or number in alveolar capillaries, but
31 there was a significant increase at 18 mo in the mononuclear population in the interstitium of
32 both species.
33 Additional means for assessing the adverse effects of diesel exhaust on the lung are to
34 examine biochemical and cytological changes in brcnchoalveclar iavage fluid (BALF) and in
35 lung tissue. Fedan et a! (1Q85) performed studies tc determine v/hether chronic exposure of raU>
36 affected the pharmacologic characteristics of rat airway smooth muscle. Concentration-response
7/25/00 5-30 DRAFT—DO NOT CITE OR QUOTE
-------
relationships for tension changes induced with acetylcholine, 5-hydroxytryptamine, potassium
chloride, and isoproterenol were assessed in vitro on isolated preparations of airway smooth
3 muscle (trachealis). Chronic exposure to diesel exhaust significantly increased the maximal
4 contractile responses to acetylcholine compared with control values; exposure did not alter the
5 sensitivity (EC50 values) of the muscles to the agonists. Exposures were to diesel exhaust
6 containing 2 mg/m3 DPM for 7 h/day, 5 days/week for 2 years.
7 Biochemical studies of HALF obtained from hamsters and rats revealed that exposures to
8 diesel exhaust caused significant increases in lactic dehydrogenase, alkaline phosphatase,
9 glucose-6-phosphate dehydrogenase (G6P-DH), total protein, collagen, and protease (pH 5.1)
10 after approximately 1 year and 2 years of exposure (Heinrich et al., 1986a). These responses
11 were generally much greater in rats than in hamsters. Exposures were to diesel exhaust
12 containing 4.24 mg/m3 DPM for 19 h/day, 5 days/week for 120 (hamsters) to 140 (rats) weeks.
13 Protein, P-glucuronidase activity, and acid phosphatase activity were significantly
14 elevated in HALF obtained from rats exposed to diesel exhaust containing 0.75 or 1.5 mg/m3
15 DPM for 12 mo (Strom, 1984). Exposure for 6 mo resulted in significant increases hi acid
16 phosphatase activity at 0.75 mg/m3 and hi protein, P-glucuronidase, and acid phosphatase activity
17 at the 1.5 mg/m3 concentration. Exposure at 0.25 mg/m3 DPM did not affect the three indices
measured at either time period. The exposures were for 20 h/day, 5.5 days/week for 52 weeks.
Additional biochemical studies (Misiorowski et al., 1980) were conducted on laboratory
20 animals exposed under the same conditions and at the same site as reported on by Strom (1984).
21 In most cases, exposures at 0.25 mg/m3 did not cause any significant changes. The DNA content
22 hi lung tissue and the rate of collagen synthesis were significantly increased at 1.5 mg/m3 DPM
23 after 6 mo. Collagen deposition was not affected. Total lung collagen content increased in
24 proportion to the increase in lung weight. The activity of prolyl hydroxylase was significantly
25 increased at 12 weeks at 0.25 and 1.5 mg/m3; it then decreased with age. Lysal oxidase activity
26 did not change. After 9 mo of exposure, there were significant increases in lung phospholipids in
27 rats and guinea pigs exposed to 0.75 mg/m3 and in lung cholesterol hi rats and guinea pigs
28 exposed to 1.5 mg/m3. Pulmonary prostaglandin dehydrogenase activity was stimulated by an
29 exposure at 0.25 mg/m3 but was not affected by exposure at 1.5 mg/m3 (Chaudhari et al., 1980,
30 1981). Exposures for 12 or 24 weeks resulted in a concentration-dependent lowering of this
31 enzyme activity. Exposure of male rats and guinea pigs at 0.75 mg/m3 for 12 weeks did not
32 cause any changes in glutathione levels of the lung, heart, or liver. Rats exposed for 2 mo at
33 6 mg/m3 showed a significant depletion of hepatic glutathione, whereas the lung showed an
34 increase of glutathione (Chaudhari and Dutta, 1982). Schneider and Felt (1981) reported that
^ similar exposures did not substantially change adenylate cyclase and guanylate cyclase activities
36 in lung or liver tissue of exposed rats and guinea pigs.
7/25/00 5-31 DRAFT—DO NOT CITE OR QUOTE
-------
1 Bhatnagar et al. (1980; see also Pepelko, 1982a) evaluated changes in the biochemistry of
2 lung connective tissue of diesel-exposed rats and mice. The mice were exposed for 8 h/day and
3 7 days/week for up to 9 mo to exhaust containing 6 mg/m3 DPM. Total lung protein content was
4 measured, as was labeled proline and labeled leucine. Leucine incorporation is an index of total
5 protein synthesis, although collagen is very low in leucine. Proline incorporation reflects
6 collagen synthesis. Amino acid incorporation was measured in vivo in the rat and in short-term
7 organ culture in mice. Both rats and mice showed a large increase in total protein (41% to 47%
8 in rats), while leucine incorporation declined and proline incorporation was unchanged. These
9 data are consistent with an overall depression of protein synthesis in diesel-exposed animals and
10 also with a relative increase hi collagen synthesis compared to other proteins. The increase in
11 collagen synthesis suggested proliferation of connective tissue and possible fibrosis (Pepelko,
12 1982a).
13 A number of reports (McClellan et al., 1986; Mauderly et al., 1987a, 1990a; Henderson
14 et al., 1988) have addressed biochemical and cytological changes in lung tissue and BALF of
15 rodents exposed for 7 h/day, 5 days/week for up to 30 mo at concentrations of 0, 0.35, 3.5, or
16 7.1 mg/m3 DPM. At the lowest exposure level (0.35 mg/m3), no biochemical or cytological
17 changes occurred in the BALF or in lung tissue in either Fischer 344 rats or CD-I mice.
18 Henderson et al. (1988) provide considerable time-course information on inflammatory events
19 taking place throughout a chronic exposure. A chronic inflammatory response was seen at the
20 two higher exposure levels in both species, as evidenced by increases in inflammatory cells
21 (macrophages and neutrophils), cytoplasmic and lysosomal enzymes (lactate dehydrogenase,
22 glutathione reductase, and p-glucuronidase), and protein (hydroxyproline) in BALF. Analysis of
23 lung tissue indicated similar changes in enzyme levels as well as an increase in total lung
24 collagen content. After 18 mo of exposure, lung tissue glutathione was depleted in a
25 concentration-dependent fashion hi rats but was slightly increased in mice. Lavage fluid levels of
26 glutathione and glutathione reductase activity increased in a concentration-dependent manner and
27 were higher in mice than in rats.
28 Rats exposed for up to 17 days to diluted diesel exhaust (3.5 mg/m3 DPM) had a fivefold
29 increase in the bronchoconstrictive prostaglandin PGF2o and a twofold increase in the
30 inflammatory leukotriene LTB4. In similarly exposed mice, there was a twofold increase in both
31 parameters. These investigators (Henderson et al., 1988) concluded that the release of larger
3 2 amounts of such mediators of inflammation from the alveolar phagocytic cells of rats accounted
33 for the greater fibrogenic response seen in that species.
34 Biochemical analysis of lung tissue from cats exposed for 124 weeks and held in clean air
35 for an additional 26 weeks indicated increases of lung collagen: this finding was confirmed by an
36 observed increase in total lung wet weight and in connective tissue fibers estimated
7/25/00 5-32 DRAFT—DO NOT CITE OR QUOTE
-------
morphometrically (Hyde et al., 1985). Exposures were for 7 h/day, 5 days/week at 6 mg/m3
DPM for 61 weeks and at 12 mg/m3 for weeks 62 to 124.
3 Heinrich et al. (1995) reported on bronchoalveolar lavage in animals exposed for 24 mo
4 and found exposure-related increases in lactate dehydrogenase, p-glucuronidase, protein, and
5 hydroxyproline in groups exposed to 2.5 or 7 mg/m3, although detailed data are not presented.
6 Lavage analyses were not carried out in concurrent studies in mice.
7 The pathogenic sequence following the inhalation of diesel exhaust as determined
8 histopathologically and biochemically begins with the interaction of diesel particles with airway
9 epithelial cells and phagocytosis by AMs. The airway epithelial cells and activated macrophages
10 release chemotactic factors that attract neutrophils and additional AMs. As the lung burden of
11 DPM increases, there is an aggregation of particle-laden AMs in alveoli adjacent to terminal
12 bronchioles, increases in the number of Type II cells lining particle-laden alveoli, and the
13 presence of particles within alveolar and peribronchial interstitial tissues and associated lymph
14 nodes. The neutrophils and macrophages release mediators of inflammation and oxygen radicals
15 that deplete a biochemical defense mechanism of the lung (i.e., glutathione). As will be
16 described later in more detail, other defense mechanisms are affected, particularly the decreased
17 viability of AMs, which leads to decreased phagocytic activity and death of the macrophage. The
«latter series of events may result in the presence of pulmonary inflammatory, fibrotic, or
emphysematous lesions. The data suggest that there may be a threshold of exposure to diesel
20 exhaust below which adverse structural and biochemical effects may not occur in the lung;
21 however, differences in the anatomy and pathological responses of laboratory animals coupled
22 with their lifespans compared with humans make a determination of human levels of exposure to
23 diesel exhaust without resultant pulmonary injury a difficult and challenging endeavor.
24
25 5.1.2.3.4. Effects on pulmonary defense mechanisms. The respiratory system has a number of
26 defense mechanisms that negate or compensate for the effects produced by the injurious
27 substances that repeatedly insult the upper respiratory tract, the tracheobronchial airways, and the
28 alveoli. The effects of exposure to diesel exhaust on the pulmonary defense mechanisms of
29 laboratory animals as well as more details on exposure atmosphere are summarized in Table 5-7
30 and ranked by cumulative exposure (C * T).
31 Several studies have been conducted investigating the effect of inhaled diesel exhaust on
32 the deposition and fate of inert tracer particles or diesel particles themselves. Lung clearance of
33 deposited particles occurs in two distinct phases: a rapid phase (hours to days) from the
34 tracheobronchial region via the mucociliary escalator and a much slower phase (weeks to
^fc months) from the nonciliated pulmonary region via, primarily but not solely, AMs. Battigelli et
36 al. (1966) reported impaired tracheal mucociliary clearance in vitro in excised trachea from rats
7/25/00 5-33 DRAFT—DO NOT CITE OR QUOTE
-------
1 exposed for single or repeated exposures of 4 to 6 h at two dilutions of diesel exhaust that
2 resulted in exposures of approximately 8 and 17 mg/m3 DPM. The exposure to 17 mg/m3
3 resulted in decreased clearance after a single exposure as well as after a cumulative exposure of
4 34 or 100 h. Clearance was reduced to a lesser extent and in fewer tracheas from animals
5 exposed to 8 mg/m3 for a cumulative exposure of 40 h. Lewis et al. (1989) found no difference
6 in the clearance of 59Fe3O4 particles (1.5 jam MMAD, Og 1.8) 1 day after dosing control and
7 diesel exhaust-exposed rats (2 mg/m3, 7 h/day, 5 days/week for 8 weeks).
8 Wolff et al. (1987) and Wolff and Gray (1980) studied the effects of both subchronic and
9 chronic diesel exhaust exposure on the trachea! clearance of particles. Trachea! clearance
10 assessments were made by measuring the retention of radiolabeled technetium
11 macroaggregated-albumin remaining 1 h after instillation in the distal trachea of rats. In the
12 subchronic studies, rats were exposed to 4.5,1.0, or 0.2 mg/m3 DPM on a 7 h/day, 5 days/week
13 schedule for up to 12 weeks. After 1 week there was an apparent speeding of trachea! clearance
14 at the 4.5 mg/m3 exposure level (p=0.10), which returned toward baseline after 6 weeks and was
15 slightly below the baseline rate at 12 weeks. In the 1.0 mg/m3 group, there was a progressive
16 significant reduction in the clearance rate at 6 and 12 weeks of exposure. There was a trend
17 toward reduced clearance in the 0.2 mg/m3 group. Scanning electron micrographs indicated
18 minimal changes in ciliary morphology; however, there was an indication of a lower percentage
19 of ciliated cells at the 1.0 and 4.5 mg/m3 levels. In the chronic studies, rats were exposed to 0,
20 0.35, 3.5, or 7.1 mg/m3 for 7 h/day, 5 days/week for 30 mo. There were no significant
21 differences in trachea! clearance rates between the control group and any of the exposure groups
22 after 6, 12,18, 24, or 30 mo of exposure. The preexposure measurements for all groups,
23 however, were significantly lower than those during the exposure period, suggesting a possible
24 age effect. The preexposure value for the 3.5-mg/m3 group was also significantly lower than the
25 control group.
26 There is a substantial body of evidence for an impairment of particle clearance from the
27 bronchiole-alveolar region of rats following exposure to diesel exhaust. Griffis et al. (1983)
28 exposed rats 7 h/day. 5 days/week for 18 weeks to diesel exhaust at 0.15. 0.94. or 4.1 mg/m3
29 DPM. Lung burdens of the 0.15, 0.94, and 4.1 mg/m3 levels were 35, 220, and 1,890 ng/g lung,
30 respectively, 1 day after the 18-week exposure. The clearance half-time of the DPM was
31 significantly greater, almost double, for the 4.1 mg/m3 exposure group than for those of the lower
32 exposure groups, 165 ± 8 days versus 99 ± 8 days (0.94 mg/m3) and 87 ± 28 days (0.15 mg/m3),
33 respectively.
34 Chan et al. (1981) showed a dose-related slowing of 14C-diese! particle clearance in rats
35 preexposed to diesel exhaust at 0.25 or 6 mg/m3 particulate matter for 20 h/day. 7 days/week for
7/25/00 5-34 DRAFT—DO NOT CITE OR QUOTE
-------
7 to 112 days. Clearance was inhibited in the 6 mg/m3 group when compared by length of
exposure or compared with the 0.25 mg/m3 or control rats at the same time periods.
3 Heinrich et al. (1982) evaluated lung clearance in rats exposed for approximately 18 mo
4 at 3.9 mg/m3 DPM for 7 to 8 h/day, 5 days/week. Following exposure to 59Fe2O3-aerosol, the rats
5 were returned to the diesel exhaust exposure and the radioactivity was measured over the thoracic
6 area at subsequent times. The biological half-life of the iron oxide deposited in the rats' lungs
7 was nearly twice that of controls.
8 Heinrich also used labeled iron oxide aerosols to study clearance in rats exposed to 0.8,
9 2.5, or 7 mg/m3 diesel DPM for 24 mo (Heinrich et al., 1995). Clearance measurements were
10 carried out at 3, 12, and 18 mo of exposure. Half-times of clearance were increased in a
11 concentration- and duration-related way in all exposed groups, with a range of a 50% increase in
12 the 0.8 mg/m3 group at 3 mo to an 11-fold increase in the 7 mg/m3 group at 19 mo. The
13 differential cell counts in these animals were stated to have shown clear effects in the 2.5 and
14 7 mg/m3 groups, but specific information about the changes is not reported.
15 Wolff et al. (1987) investigated alterations in DPM clearance from the lungs of rats
16 chronically exposed to diesel exhaust at 0, 0.35, 3.5, or 7.1 mg/m3 DPM for 7 h/day, 5 days/week
17 for up to 24 mo. Progressive increases in lung burdens were observed over time in all groups;
f levels of DPM in terms of milligrams per lung were 0.60, 11.5, and 20.5 after 24 mo of exposure
at the 0.35, 3.5, or 7.1 mg/m3 exposure levels, respectively. There were significant increases in
20 16-day clearance half-times of inhaled radiolabeied particles of ^Ga^ (0.1 urn MMD) as early
21 as 6 mo at the 7.1 mg/m3 level and 18 mo at the 3.5 mg/m3 level; no significant changes were
22 seen at the 0.35 mg/m3 level. Rats inhaled fused aluminosilicate particles (2 urn MMAD)
23 labeled with 134Cs after 24 mo of diesel exhaust exposure; long-term clearance half-times were
24 79, 81, 264, and 240 days for the 0,0.35, 3.5, and 7.1 mg/m3 groups, respectively. Differences
25 were significant between the control and the 3.5 and 7.1 mg/m3 groups (p < 0.01).
26 Mauderly et al. (1987b) compared the effects of diesel exhaust in the developing lung to
27 the adult lung by exposing groups of male F344 rats to 3.5 mg/m3 for 7 h/day, 5 days/week for
28 6 mo. One group (adult) was exposed between 6 and 12 mo of age, and the other was exposed
29 beginning in utero and until 6 mo of age. Clearance of an inhaled monodisperse 2 urn
30 aluminosilicate particle was measured after exposure for 6 mo. The clearance half-time of the
31 slow phase was found to be doubled in adult rats compared with age-matched controls and was
32 not significantly affected in developing rat lungs.
33 Mauderly et al. compared the effects of diesel exhaust in normal lungs with rats in which
34 emphysema had been induced experimentally by instillation of elastase 6 weeks before diesel
^P exhaust exposures. The rats were exposed to 3.5 mg/m3 DPM for 7 h/day, 5 days/week for
36 24 mo. Measurements included histopathology, clearance, pulmonary function, lung lavage, and
7/25/00 5-35 DRAFT—DO NOT CITE OR QUOTE
-------
1 immune response. In the rats that were not pretreated with elastase, there was a significant
2 reduction in the number of macrophages recovered by pulmonary lavage in contrast to the
3 increases in macrophages reported by Strom (1984) and Henderson et al. (1988). The half-time
4 of the slow phase of clearance of inhaled, 1 um, monodisperse particles was doubled in the
5 exposure animals without elastase pretreatment. The elastase pretreatment did not affect
6 clearance in unexposed animals but significantly reduced the effect of diesel. The clearance
7 half-time was significantly less in elastase-pretreated, diesel-exposed animals than hi
8 diesel-exposed normal animals. Many other effects measured in this study were also less
9 affected by diesel exposure in elastase-treated animals. Measurements of lung burden of DPM
10 showed that elastase-pretreated animals accumulated less than half as much DPM mass as
11 normal animals exposed at the same time, suggesting that the difference in effect could be
12 explained by differences in dose to the lung.
13 Lewis et al. (1989) conducted lung burden and 59Fe304 tracer studies in rats exposed for
14 12 and 24 mo to 2 mg/m3 DPM (7 h/day, 5 days/week). The slope of the Fe3O4 clearance curve
15 was significantly steeper than that of the controls, indicating a more rapid alveolar clearance of
16 the deposited 59Fe3O4. After 120 days from the inhalation of the tracer particle, 19% and 8% of
17 the initially deposited 59Fe3O4 were present in the lungs of control and diesel exhaust-exposed
18 rats, respectively. The lung burden of DPM, however, increased significantly between 12 and
19 24 mo of exposure (0.52 to 0.97% lung dry weight), indicating a later dose-dependent inhibition
20 of clearance.
21 Alveolar macrophages, because of their phagocytic and digestive capabilities, are one of
22 the prime defense mechanisms of the alveolar region of the lung against inhaled particles. Thus,
23 characterization of the effects of diesel exhaust on various properties of AMs provides
24 information on the integrity or compromise of a key pulmonary defense mechanism. The
25 physiological viability of AMs from diesel-exposed rats was assessed after 2 years of exposure
26 by Castranova et al. (1985). The 7 h/day, 5 days/week exposure at 2 mg/m3 DPM had little effect
27 on the following: viability, cell number, oxygen consumption, membrane integrity, lysosomal
28 enzyme activity, or protein content of the AMs. A slight decrease in cell volume, a decrease in
29 chemiluminescence indicative of a decreased secretion of reactive oxygen species, and a decrease
30 in ruffling of the cell membrane were observed. These findings could be reflective of an overall
31 reduction in phagocytic activity.
32 Exposure to diesei exhaust has been reported both to increase the number of recoverable
33 AMs from the lung (Strom, 1984; Vostal et al., 1982; Henderson et al., 1988) or to produce no
34 change in numbers (Chen et al., 1980; Castranova et al., 1985). Strom (] 984) found that in rats
3R exposed to 0.25 rnp/m3 DPM for 20 h/dav. 5.5 davs/week for f> rno or 1 vear as 'well as in the
36 controls, BAL cells consisted entirely of AMs, with no differences in the cell counts in the lavage
7/25/00 5-36 DRAFT—DO NOT CITE OR QUOTE
-------
fluid. At the higher concentrations, 0.75 or 1.5 mgDPM/m3, the count of AM increased
proportionally with the exposure concentration; the results were identical for AMs at both 6 and
3 11 or 12 mo of exposure. The increase in AM counts was much larger after exposure to
4 1.5 mg/m3 DPM for 6 mo than after exposure to 0.75 mg/m* for 1 year, although the total mass
5 (calculated as C * T) of deposited particulate burden was the same. These data suggested to the
6 authors that the number of lavaged AMs was proportional to the mass influx of particles rather
7 than to the actual DPM burden in the lung. These results further implied that there may be a
8 threshold for the rate of mass influx of DPM into the lungs of rats above which there was an
9 increased recruitment of AMs. Henderson et al. (1988) reported similar findings of significant
10 increases of AMs in rats and mice exposed to 7.1 mg/m3 DPM for 18 and 24 mo, respectively,
11 for 7 h/day, 5 days/week, but not at concentrations of 3.5 or 0.35 mg/m3 for the same exposure
12 durations. Chen et al. (1980), using an exposure regimen of 0.25 and 1.5 mg/m3 DPM for 2 mo
13 and 20 h/day and 5.5 days/week, found no significant changes in absolute numbers of AMs from
14 guinea pig BALF, nor did Castranova et al. (1985) in rat BALF following exposure to 2 mg/m3
15 DPM for 7 h/day, 5 days/week for 2 years.
16 A similar inflammatory response was noted by Henderson et al. (1988) and Strom (1984),
17 as evidenced by an increased number of PMNs present in BALF from rodents exposed to diesel
exhaust. Henderson et al. (1988) found these changes in rats and mice exposed to 7.1 and
3.5 mg/m3 DPM for 7 h/day, 5 days/week. Significant increases in BALF PMNs were observed
20 in mice at 6 mo of exposure and thereafter at the 7.1 and 3.5 mg/m3 exposure levels, but in rats
21 only the 7.1 mg/m3 exposure level showed an increase in BALF PMNs at 6 mo of exposure and
22 thereafter. Significant increases in BALF PMNs occurred in rats at 12,18, and 24 mo of
23 exposure to 3.5 mg/m3 DPM. Although increases in PMNs were usually greater in mice in terms
24 of absolute numbers, the PMN response in terms of increase relative to controls was only about
25 one-third that of rats. Strom (1984) reported that the increased numbers of PMNs in BALF were
26 proportional to the inhaled concentrations and/or duration of exposure. The PMNs also appeared
27 to be affiliated with clusters of aggregated AMs rather than to the diesel particles per se.
28 Proliferation of Type II cells likewise occurred in response to the formed aggregates of AMs
29 (White and Garg, 1981).
30 The integrity of pulmonary defense mechanisms can also be ascertained by assessing if
31 exposure to diesel exhaust affects colonization and clearance of pathogens and alters the response
32 of the challenged animals to respiratory tract infections. Campbell et al. (1980, 1981) exposed
33 mice to diesel exhaust followed by infectious challenge with Salmonella typhimurium,
34 Streptococcus pyogenes, or A/PR8-3 influenza virus and measured microbial-induced mortality.
?M> Exposures to diesel exhaust were to 6 mg/m3 DPM for 8 h/day, 7 days/week for up to 321 days.
36 Exposure to diesel exhaust resulted in enhanced susceptibility to the lethal effects of S. pyogenes
7/25/00 5-37 DRAFT—DO NOT CITE OR QUOTE
-------
1 infection at all exposure durations (2 h, 6 h; 8, 15, 16,307, and 321 days). Tests with S.
2 typhimuriwn were inconclusive because of high mortality rates in the controls. Mice exposed to
3 diesel exhaust did not exhibit an enhanced mortality when challenged with the influenza virus.
4 Hatch et al. (1985) found no changes in the susceptibility of mice to Group C Streptococcus sp.
5 infection following intratracheal injection of 100 ng of DPM suspended in unbuffered saline.
6 Hahon et al. (1985) assessed virus-induced mortality, virus multiplication with
7 concomitant IFN levels (lungs and sera), antibody response, and lung histopathology in mice
8 exposed to diesel exhaust prior to infectious challenge with Ao/PR/8/34 influenza virus.
9 Weanling mice were exposed to diesel exhaust containing 2 mg/m3 DPM for 7 h/day,
10 5 days/week. In mice exposed for 1, 3, and 6 mo, mortality was similar between the exposed and
11 control mice. In mice exposed for 3 and 6 mo, however, there were significant increases in the
12 percentage of mice having lung consolidation, higher virus growth, depressed IFN levels, and a
13 fourfold reduction in hemagglutinin antibody levels; these effects were not seen after the 1-mo
14 exposure.
15 The effects of diesel exhaust on the pulmonary defense mechanisms are determined by
16 three critical factors related to exposure: the concentrations of the pollutants, the exposure
17 duration, and the exposure pattern. Higher doses of diesel exhaust as determined by an increase
18 in one or more of these three variables have been reported to increase the numbers of AMs,
19 PMNs, and Type II cells in the lung, whereas lower doses fail to produce such changes. The
20 single most significant contributor to the impairment of the pulmonary defense mechanisms
21 appears to be an excessive accumulation of DPM, particularly as particle-laden aggregates of
22 AMs. Such an accumulation would result from an increase in deposition and/or a reduction in
23 clearance. The deposition of particles does not appear to change significantly following exposure
24 to equivalent diesel exhaust doses over time. Because of the significant nonlinearity in particle
25 accumulation between low and high doses of diesel exhaust exposure, coupled with no evidence
26 of increased particle deposition, an impairment in one or more of the mechanisms of pulmonary
27 defense appears to be responsible for the DPM accumulation and subsequent pathological
28 sequelae. The time of onset of pulmonary clearance impairment was dependent both on the
29 magnitude and on the duration of exposures. For example, for rats exposed for 7 h/day,
30 5 days/week for 104 weeks, the concentration needed to induce pulmonary clearance impairment
31 appears to lie between 0.35 and 2.0 mg/m3 DPM.
32
33 5.1.2.3.5. Effects on the immune system—inhalation studies. The effects of diesel exhaust on
34 the immune system of guinea pigs were investigated by Dziedzic (1981). Exposures were to
35 1.5 rng/rn3 DPM for 20 h/day, 5.5 days/week for up tn 8 weeks- There was no effect of diesel
36 exposure when compared with matched controls for the number of B and T lymphocytes and null
7/25/00 5-38 DRAFT—DO NOT CITE OR QUOTE
-------
cells isolated from the tracheobronchial lymph nodes, spleen, and blood. Cell viability as
measured by trypan blue exclusion was comparable between the exposed and control groups.
3 The results of this study and others on the effects of exposure to diesel exhaust on the immune
4 system are summarized in Table 5-8.
5 Mentnech et al. (1984) examined the effect of diesel exhaust on the immune system of
6 rats. Exposures were to 2 mg/m3 DPM for 7 h/day, 5 days/week for up to 2 years. Rats exposed
7 for 12 and 24 mo were tested for immunocompetency by determining antibody-producing cells in
8 the spleen 4 days after immunization with sheep erythrocytes. The proliferative response of
9 splenic T-lymphocytes to the mitogens concanavalin A and phytohemagglutinin was assessed in
10 rats exposed for 24 mo. There were no significant differences between the exposed and control
11 animals. Results obtained from these two assays indicate that neither humoral immunity
12 (assessed by enumerating antibody-producing cells) nor cellular immunity (assessed by the
13 lymphocyte blast transformation assay) were markedly affected by the exposures.
14 Bice et al. (1985) evaluated whether or not exposure to diesel exhaust would alter
15 antibody immune responses induced after lung immunization of rats and mice. Exposures were
16 to 0.35, 3.5, or 7.1 mg/m3 DPM for 7 h/day, 5 days/week for 24 mo. Chamber controls and
17 exposed animals were immunized by intratracheal instillation of SRBCs after 6,12,18, or 24 mo
of exposure. No suppression in the immune response occurred in either species. After 12, 18,
and 24 mo of exposure, the total number of anti-SRBC IgM antibody forming cells (AFCs) was
20 elevated in rats, but not in mice, exposed to 3.5 or 7.1 mg/m3 DPM; after 6 mo of exposure, only
21 the 7.1 mg/m3 level was found to have caused this response in rats. The number of AFCs per 106
22 lymphoid cells in lung-associated lymph nodes and the levels of specific IgM, IgG, or IgA in rat
23 sera were not significantly altered. The investigators concluded that the increased cellularity and
24 the presence of DPM in the lung-associated lymph nodes had only a minimal effect on the
25 immune and antigen filtration function of these tissues.
26 The effects of inhaled diesel exhaust and DPM have been studied in a murine model of
27 allergic asthma (Takano et al., 1998a,b). ICR mice were exposed for 12 h/day, 7 days/week for
28 40 weeks to diesel exhaust (0.3, 1.0, or 3.0 mg/m3). The mice were sensitized with ovalbumin
29 (OA) after 16 weeks exposure and subsequently challenged with aerosol allergen (1% OA in
30 isotonic saline for 6 min) at 3-week intervals during the last 24 weeks of exposure. Exposure to
31 diesel exhaust enhanced allergen-related eosinophil recruitment to the submucosal layers of the
32 airways and to the bronchoalveolar space, and increased protein levels of GM-CSF and IL-5 in
33 the lung in a dose-dependent manner. In the diesel exhaust-exposed mice, increases in
34 eosinophil recruitment and local cytokine expression were accompanied by goblet cell
proliferation in the bronchial epithelium and airway hyperresponsiveness to inhaled
36 acetylcholine. In contrast, mice exposed to clean air or diesel exhaust without allergen
7/25/00 5-39 DRAFT—DO NOT CITE OR QUOTE
-------
1 provocation showed no eosinophil recruitment to the submucosal layers of the airways or to the
2 bronchoalveolar space, and few goblet cells in the bronchial epithelium. The authors concluded
3 that daily inhalation of diesel exhaust can enhance allergen-related respiratory diseases such as
4 allergic asthma, and that this effect may be mediated by the enhanced local expression of IL-5
5 andGM-CSF. The effect of DPM on a second characteristic of allergic asthma, airway
6 hyperresponsiveness, was examined by Takano et al. (1998b). Laboratory mice were
7 administered OA, DPM, or OA and DPM combined by intratracheal instillation for 6 wk.
8 Respiratory resistance (Rrs) after acetylcholine challenge was measured 24 h after the final
9 instillation. Rrs was significantly greater in the mice treated with OA and DPM than in the other
10 treatments. The authors concluded that DPM can enhance airway responsiveness associated with
11 allergen exposure.
12 In a series of inhalation studies following earlier instillation studies, Miyabara and
13 co-workers investigated whether inhalation of diesel exhaust could enhance allergic reactions in
14 laboratory mice. C3H/Hen mice were exposed to diesel exhaust (3 mg DPM/m3) by inhalation
15 for 5 weeks (Miyabara et al., 1998b) and, after 7 days of exposure, were sensitized to OA
16 injected intraperjtoneally. At the end of the diesel exhaust exposure, the mice were challenged
17 with an OA aerosol for 15 min. Diesel exhaust caused an increase in the numbers of neutrophils
18 and macrophages in bronchoalveolar lavage fluid independent of OA sensitization, whereas a
19 significant increase in eosinophil numbers occurred only after diesel exhaust exposure was
20 combined with antigen challenge. Even though OA alone caused an increase hi eosinophil
21 numbers in lung tissue, this response was enhanced by diesel exhaust. Diesel exhaust exposure
22 combined with OA sensitization enhanced the number of goblet cells in lung tissue, respiratory
23 resistance, production of OA-specific IgE and Igl in the serum, and overexpression of IL-5 in
24 lung tissue. In a second study, C3H/Hen mice were sensitized with OA injected intraperitoneally
25 and then exposed to diesel exhaust by inhalation for 12 h/day for 3 mo (Miyabara et al., 1998a).
26 After 3 weeks of diesel exhaust exposure, and every 3 weeks thereafter, the mice were challenged
27 with an OA aerosol. Exposure to diesel exhaust with antigen challenge induced airway
28 hyperresponsiveness and airway inflammation, which was characterized by increased numbers of
29 eosinophils and mast cells in lung tissue. The increase in inflammatory cells was accompanied
30 by an increase in gobiet ceils in the bronchial epithelium. Airway hyperresponsiveness, but not
31 eosinophilic infiltration or increased goblet cells, was increased by diesel exhaust exposure
32 alone. These workers concluded that inhalation of diesel exhaust can enhance airway
33 hyperresponsiveness and airway inflammation caused by OA sensitization in mice.
34 The effects of diesel exhaust on IgE antibody production were investigated in B ALB/c
35 mice sensitized with OA and exposed by inhalation to diesel exhaust (3.0 and 6.0 mg/m3) for
36 3 weeks (Fujimaki et al., 1997). The mice were sensitized by intranasal administration of OA
7/25/00 5-40 DRAFT—DO NOT CITE OR QUOTE
-------
alone before, immediately after, and 3 weeks after diesel exhaust inhalation. While body and
thymus weights were unchanged in the diesel exhaust-exposed and control mice, spleen weights
3 in mice exposed to 6 mg/m3 diesel exhaust increased significantly. Anti-OA IgE antibody titers
4 in the sera of mice exposed to 6 mg/m3 diesel exhaust were significantly higher than control.
5 Total IgE and anti-OA IgG in sera from diesel exhaust-exposed and control mice remained
6 unchanged. Cytokine production was measured in vitro stimulated with OA in spleen cells from
7 mice exposed to diesel exhaust (6 mg/m3). Antigen-stimulated interleukin-4 (IL-4) and
8 -10 (IL-10) production increased significantly in vitro in spleen cells from diesel exhaust-
9 exposed mice compared with controls, while IFN-y production decreased markedly. The authors
10 concluded that diesel exhaust inhalation in mice may affect antigen-specific IgE antibody
11 production through alteration of the cytokine network.
12
13 5.1.2.3.6. Effects on the immune system—noninhalation studies. The immune response of
14 laboratory animals to DPM has been studied in various noninhalation models, and the results of
15 these studies are presented in Table 5-9. Takafuji et al. (1987) evaluated the IgE antibody
16 response of mice inoculated intranasally at intervals of 3 weeks with varying doses of a
17 suspension of DPM in ovalbumin. Antiovalbumin IgE antibody titers, assayed by passive
« cutaneous anaphylaxis, were enhanced by doses as low as 1 ug of particles compared with
immunization with ovalbumin alone.
20 Muranaka et al. (1986) studied the effects of DPM on IgE antibody production in
21 immunized mice. A greater IgE antibody response was noted in mice immunized by ip injection
22 of ovalbumin (OA) mixed with DPM than in animals immunized with OA alone. This effect of
23 DPM on IgE antibody production in mice was also demonstrated in mice immunized with
24 repeated injections of dinitrophenylated-OA. Moreover, a persistent IgE-antibody response to
25 Japanese cedar pollen (JCPA), a common pollen allergen causing allergic rhinitis in Japan, was
26 observed in mice immunized with JCPA mixed with DPM but not in animals immunized with
27 JCPA alone. The results suggest an association between the adjuvant activity of DPM and
28 allergic rhinitis caused by JCPA.
29 The potential role of oxygen radicals in injury caused by DPM was investigated by Sagai
30 et al. (1996). These workers reported that repeated intratracheal instillation of DPM (once/week
31 for 16 weeks) in mice caused marked infiltration of inflammatory cells, proliferation of goblet
32 ceils, increased mucus secretion, respiratory resistance, and airway constriction. Eosinophils in
33 the submucosa of the proximal bronchi and medium bronchioles increased eightfold following
34 instillation. Eosinophil infiltration was significantly suppressed by pretreatment with
^P polyethyleneglycol-conjugated superoxide dismutase (PEG-SOD). Bound sialic acid
36 concentrations in bronchial alveolar lavage fluids, an index of mucus secretion, increased with
7/25/00 5-41 DRAFT—DO NOT CITE OR QUOTE
-------
1 DPM, but were also suppressed by pretreatment with PEG-SOD. Goblet cell hyperplasia, airway
2 narrowing, and airway constriction also were observed with DPM.
3 Respiratory resistance to acetylcholine in the DPM group was 11 times higher than in
4 controls, and the increased resistance was significantly suppressed by PEG-SOD pretreatment.
5 These findings indicate that oxygen radicals caused by intratracheally instilled DPM elicit
6 responses characteristic of bronchial asthma.
7 Potential adjuvant effects of DPM on the response to the model allergen OA were
8 investigated in BALB/c mice using the popliteal lymph node (PLN) assay (L0vik et al., 1997).
9 DPM inoculated together with OA into one hind footpad gave a significantly augmented
10 response (increase in weight, cell numbers, and cell proliferation) in the draining popliteal lymph
11 node as compared to DPM or OA alone. The duration of the local lymph node response was also
12 longer when DPM was given with the allergen. The lymph node response appeared to be of a
13 specific immunologic character and not an unspecific inflammatory reaction. The OA-specific
14 response IgE was increased in mice receiving OA together with DPM as compared with the
15 response in mice receiving OA alone. Further studies using carbon black (CB) as a surrogate for
16 the nonextractable core of DPM found that while CB resembled DPM in its capacity to increase
17 the local lymph node response and serum-specific IgE response to OA, CB appeared to be
18 slightly less potent than DPM. The results indicate that the nonextractable particle core
19 contributes substantially to the adjuvant activity of DPM.
20 Nilsen et al. (1997) investigated which part of the particle was responsible, the carbon
21 core and/or the adsorbed organic substances, for the adjuvant activity of DPM. Female
22 BALB/cA mice were immunized with OA alone or in combination with DPM or CB particles by
23 intranasal administration. There was an increased response to the antigen in animals receiving
24 OA together with DPM or CB, compared with animals receiving OA alone. The response was
25 seen as both an increased number of responding animals and increased serum anti OA IgE
26 response. The workers concluded that both DPM and CB have an adjuvant activity for specific
27 IgE production, but that the activity of DPM may be more pronounced than that of CB. The
28 results suggest that both the organic matter adsorbed to DPM and the nonextractable carbon are
29 responsible for the observed adjuvant effect.
30 Fujimaki et al. (1994) investigated the relationship between DPM and IgE antibody
31 production, interleukin 4 (IL-4) production hi BALB/c mice treated with DPM mixed with
32 antigen OA or JCPA by intratracheal instillation. BALB/c mice were injected with DPM plus OA
33 or OA alone and, after the last instillation, the proliferative response and lymphokine production
34 by mediastinal lymph node cells (LNC) were examined in vitro. The proliferative response to OA
35 in mediastinal LNC from mice injected with DPM plus OA was enhanced to 4-17 times that of
36 control mice. IL-4 production by OA stimulation was also enhanced in mediastinal LNC from
•7 /*> C /,
/GO 5-42 DRAFT—DO NOT CITE OR QUOTE
-------
mice injected with DPM plus OA. A significantly larger amount of anti-OA IgE antibody was
detected in sera from DPM- and OA-injected mice compared with those from control mice. The
3 levels of IL-4, estimated by JCP antigen in mediastinal LNC, from mice injected with DPM plus
4 JCPA were twofold higher than those from mice injected with JCP alone. These results suggest
5 that intratracheal instillation of DPM affects antigen-specific IgE antibody responses via local T-
6 cell activation, especially enhanced IL-4 production.
7 Suzuki et al. (1993) investigated the adjuvant activity of pyrene, a compound contained in
8 DPM, on IgE antibody production in mice. In the first experiment, mice were immunized with
9 1 jig of OA alone, 1 jig of OA plus 1 mg of pyrene, or 1 ng of OA plus 1 mg of DPM,
10 respectively. The IgE antibody responses to OA in mice immunized with OA plus pyrene or OA
11 plus DPM were enhanced as compared to those in mice immunized with OA alone; the highest
12 responses were observed in mice immunized with OA plus DPM. In the second experiment, mice
13 were immunized with 10 ug of JCPA alone or 10 |ig of JCPA plus 5 mg of pyrene. The IgE
14 antibody responses to JCPA in mice immunized with JCPA plus pyrene were higher than those
15 in mice immunized with JCPA alone. The results indicate that pyrene contained in DPM acts as
16 an adjuvant in IgE antibody production in immunized mice.
17 Suzuki et al. (1996) investigated the effect of pyrene on IgE and IgGl antibody
production in mice to clarify the relation between mite allergy and adjuvancy of the chemical
compounds in DPM. The mite allergen was Der f II, one of the major allergens of house dust
20 mite (Dermatophagoides farinae). Allergen mice were grouped and immunized with Der f II
21 (5 ug), Der f II (5 ug) plus pyrene (200 ug), and Der f II (5 ug) plus DPM (100 ug) intranasally
22 seven times at 2-week intervals. The separate groups of mice were also immunized with Der f II
23 (10 ug) plus the same dose of adjuvants in the same way. The IgE antibody responses to Der f II
24 in mice immunized with Der f II plus pyrene or Der f II plus DPM were markedly enhanced
25 compared with those immunized with Der f II alone. The anti-Der f II IgE antibody production
26 increased with increasing the dose of Der f II from 5 ug to 10 |ig hi mice immunized with
27 Der f II plus the same dose of adjuvants. The IgGl antibody responses to Der f II in mice
2 8 immunized with Der f II (10 ug) plus pyrene (200 ug) or Der f II (10 ug) plus DPM (100 ug)
29 were greater than those immunized with 10 ug of Der f II alone. In addition, when peritoneal
30 macrophages obtained from normal mice were incubated with pyrene or DPM in vitro, an
31 enhanced IL-la production by the macrophages was observed. When spleen lymphocytes
32 obtained from the mice immunized with Der f II (10 ug) plus DPM (100 ug) or Der f II (10 u.g)
33 plus pyrene (200 ug) were stimulated with 10 jig of Der f II in vitro, an enhanced IL-4 production
34 of the lymphocytes was also observed compared with those immunized with Der f II alone. This
7/25/00 5-43 DRAFT—DO NOT CITE OR QUOTE
-------
1 study indicates that DPM and pyrene have an adjuvant activity on IgE and IgGl antibody
2 production in mice immunized intranasally with a house dust mite allergen.
3 Maejima et al. (1997) examined the potential adjuvant activity of several different fine
4 particles. These workers administered 25 ug of each of 5 particles (Kanto loam dust, fly ash, CB,
5 DPM, and aluminum hydroxide [alum]) intranasally in mice and exposed them to aerosolized
6 JCPA for intervals up to 18 weeks. Measurements were made of JCPA-specific IgE and IgG
7 antibody titers, the protein-adsorbing capacity of each type of particle, and nasal rubbing
8 movements (a parameter of allergic rhinitis in mice). The increases in anti-JPCA IgE and IgG
9 antibody titers were significantly greater in mice treated with particles and aerosolized JCPA
10 than hi mice treated with aerosolized JCPA alone. In a subsequent experiment, the mice received
11 the particles as before, but about 160,000 grains of JCP were dropped onto the tip of the nose of
12 each mouse twice a week for 16 weeks. After 18 weeks there were no significant differences in
13 the anti-JCPA IgE and IgG production, nasal rubbing, or histopathological changes. The workers
14 concluded that the nature of the particle, the ability of the particle to absorb antigens, and particle
15 size are not related to the enhancement of IgE antibody production or symptoms of allergic
16 rhinitis. However, IgE antibody production did appear to occur earlier in mice treated with
17 particles than in mice immunized with allergens alone.
18 Eosinophils are major components of allergic inflammatory disorders including asthma
19 and nasal allergy. Terada et al. (1997) examined the effects of DPM and DPM extract on
20 eosinophil adhesion, survival rate, and degranulation. Eosinophils, human mucosal
21 microvascular endothelial cells (HMMECs), and human nasal epithelial cells (HNECs) were
22 preincubated in the presence of DPM and DPM extract. 35S-labeled eosinophils were allowed to
23 adhere to monolayers of HMMECs and HNECs. Although neither DPM nor DPM extract
24 affected the adhesiveness of HMMECs and HNECs to eosinophils, DPM and DPM extract each
25 significantly increased eosinophil adhesiveness to HNECs; neither affected eosinophil
26 adhesiveness to HMMECs. DPM extract also induced eosinophil degranulation without changing
27 the eosinophil survival rate. These results indicate that DPM may play an important role in
28 promoting the nasal hypersensitivity induced by enhanced eosinophil infiltration of epithelium
29 and eosinophil degranulation.
30 Histamine is the most important chemical mediator in the pathogenesis of nasal allergy.
31 Terada et al. (1999) examined the effects of DPM extract on the expression of histamine H1
3 2 receptor (H 1R) mRNA in HNECs and HMMECs, and on the production of IL-8 and GM-CSF
33 induced by histamine. HNECs and HMMECs, isolated from human nasal mucosa specimens,
34 were cultured with DPM extract. DPM extract increased the expression of H1R mRNA in both
35 HNECs and HMMECs. The amount of IL-8 and GM-CSF, induced by histamine. was also
36 significantly higher in HNECs and HMMECs treated with DPM extract. These results strongly
7/25/00 5-44 DRAFT—DO NOT CITE OR QUOTE
-------
1 suggest that DPM accelerates the inflammatory change by not only directly upregulating H1R
2 expression but also by increasing histamine-induced IL-8 and GM-CSF production.
3 The potential for DPM to modulate cytokine production has been demonstrated in
4 cultured mouse bone marrow-derived mast cells (BMMC). Saneyoshi et al. (1997) examined the
5 production of cytokines in BMMC treated with DPM (0.8,2 and 4 ng/mL). Production of
6 interleukin-4 (IL-4) and IL-6 was higher in BMMC stimulated with A23187 and treated with low
7 concentrations of DPM than in controls, but no increase was seen in BMMC treated with high
8 DPM. After pretreatment with low DPM for 24 h, IL-4 production in BMMC stimulated with
9 A23187 was lower than in controls. Antigen-induced IL-4 production increased significantly in
10 BMMC treated with 0.4 or 0.8 ng/mL DPM, but did not increase with low DPM. Although the
11 enhancement of IL-4 production of BMMC stimulated with A23187 plus DPM was not
12 completely inhibited by 2-mercaptoethanol, treatment with dexamethasone inhibited further IL-4
13 production. Thus, DPM may affect the immune response via the modulation of cytokine
14 production in mast cells.
15 Ormstad et al. (1998) investigated the potential for DPM as well as other suspended
16 particulate matter (SPM) to act as a carrier for allergens into the airways. These investigators
17 found both Can f 1 (dog) and Bet v 1 (birch pollen) on the surface of SPM collected in air from
18 different homes. In an extension of the study, they found that DPM had the potential of binding,
19 in vitro, both of these allergens as well as Pel d 1 (cat) and Der p 1 (house mite). The authors
20 conclude that soot particles in indoor air house dust may act as carrier of several allergens in
21 indoor air.
22 Knox et al. (1997) investigated whether free grass pollen allergen molecules, released
23 from pollen grains by osmotic shock (Suphioglu et al., 1992) and dispersed in microdroplets of
24 water in aerosols, can bind to DPM in air. Using natural highly purified Lol p 1, immunogold
25 labeling with specific monoclonal antibodies, and a high-voltage transmission electron-
26 microscopic imaging technique, these workers demonstrated binding of the major grass pollen
27 allergen, Lol p 1, to DPM in vitro. These workers conclude that binding of DPM with Lol p 1
28 might be a mechanism by which allergens can become concentrated in air and trigger attacks of
29 asthma.
30 The inhalation of diesel exhaust appeared to have minimal effects on the immune status
31 of rats and guinea pigs. Conversely, intranasally delivered doses as low as 1 ug of DPM exerted
32 an adjuvant activity for IgE antibody production in mice. Further studies of the effects of diesel
33 exhaust on the immune system are needed to clarify the impact of such variables as route of
34 exposure, species, dose, and atopy.
7/25/00 5-45 DRAFT—DO NOT CITE OR QUOTE
-------
1 Murphy et al. (1999) examined the comparative toxicities to the lung of four CB particles
2 and DPM, in primary cultures of mouse Clara and rat type II epithelial cells. Particle toxicity
3 was assessed by cell attachment to an extracellular matrix substratum. The CB particles varied in
4 toxicity to Clara and type II cells. DPM stored for 2 weeks was equally toxic to both cell types.
5 DPM became progressively less toxic to type II cells with time of storage. Both primary
6 epithelial cell types internalized the particles in culture. These workers concluded that
7 bioreactivity was related to CB particle size and surface area, with the smaller particles having
8 the larger surface area being the more toxic. Although freshly prepared DPM was equally toxic
9 to type II and Clara cells, DPM became progressively less toxic to the type II cells with time.
10 Exposure studies in laboratory animals and isolated cell systems derived from animals
11 also indicate that DPM can elicit both inflammatory and immunological changes. Moreover, the
12 effects appear to be due to both the nonextractable carbon core and the adsorbed organic fraction
13 of the diesel particle. The data further indicate a role for oxygen radicals in DPM injury because
14 the extent of the injury can be reduced by treatment with antioxidants. DPM also has the
15 capacity to bind and transport airborne allergens.
16
17 5.1.23.7. Effects on the liver. Meiss et al. (1981) examined alterations in the hepatic
18 parenchyma of hamsters by using thin-section and freeze-fracture histological techniques.
19 Exposures to diesel exhaust were for 7 to 8 h/day, 5 days/week, for 5 mo at about 4 or 11 mg/m3
20 DPM. The livers of the hamsters exposed to both concentrations of diesel exhaust exhibited
21 moderate dilatation of the sinusoids, with activation of the Kupffer cells and slight changes in the
22 cell nuclei. Fatty deposits were observed in the sinusoids, and small fat droplets were
23 occasionally observed in the peripheral hepatocytes. Mitochondria often had a loss of cristae and
24. exhibited a pleomorphic character. Giant microbodies were seen in the hepatocytes, which were
25 moderately enlarged, and gap junctions between hepatocytes exhibited a wide range in structural
26 diversity. The results of this study and others on the effect of exposure of diesel exhaust on the
27 liver of laboratory animals are summarized in Table 5-10.
28 Green et al. (1983) and Plopper et al (1983) reported no changes in liver weights of rats
29 exposed to 2 mg/m3 DPM for 7 h/day, 5 days/week for 52 weeks or of cats exposed to 6 to
30 12 mg/m3, 8 h/day, 7 days/week for 124 weeks. The use of light and electron microscopy
31 revealed that long-term inhalation of varying high concentrations of diesel exhaust caused
32 numerous alterations to the hepatic parenchyma of guinea pigs. A less sensitive index of liver
33 toxicity, increased liver weight, failed to detect an effect of diesel exhaust on the liver of the rat
34 and cat following long-term exposure to diesel exhaust. These results are ton lirnjterl to
35 understand potential impacts on the liver,
36
7/25/00 5-46 DRAFT—DO NOT CITE OR QUOTE
-------
5.1.23.8. Blood and cardiovascular systems. Several studies have evaluated the effects of
diesel exhaust exposure on hematological and cardiovascular parameters of laboratory animals.
3 These studies are summarized in Table 5-11. Standard hematological indices of toxicological
4 effects on red and white blood cells failed to detect dramatic and consistent responses.
5 Erythrocyte (RBC) counts were reported as being unaffected in cats (Pepelko and Peirano, 1983),
6 rats and monkeys (Lewis et al., 1989), guinea pigs and rats (Penney et al., 1981), and rats
7 (Karagianes et al., 1981); lowered in rats (Heinrich et al., 1982); and elevated in rats (Research
8 Committee for HERP Studies, 1988; Brightwell et al., 1986). Mean corpuscular volume was
9 significantly increased hi monkeys, 69 versus 64 (Lewis et al., 1989), and hamsters (Heinrich et
10 al., 1982), and lowered in rats (Research Committee for HERP Studies, 1988). The only other
11 parameters of erythrocyte status and related events were lowered mean corpuscular hemoglobin
12 and mean corpuscular hemoglobin concentration in rats (Research Committee for HERP Studies,
13 1988), a 3% to 5% increase in carboxyhemoglobin saturation in rats (Karagianes et al., 1981),
14 and a suggestion of an increase in prothrombin time (Brightwell et al., 1986). The biological
15 significance of these findings regarding adverse health effects is deemed to be inconsequential.
16 Three investigators (Pepelko and Peirano, 1983; Lewis et al., 1989; Brightwell et al.,
17 1986) reported an increase in the percentage of banded neutrophils hi cats and rats. This effect
was not observed in monkeys (Lewis et al., 1989). The health implications of an increase in
9 abnormal maturation of circulating neutrophils are uncertain but indicate a toxic response of
20 leukocytes following exposures to diesel exhaust. Leukocyte counts were reported to be reduced
21 in hamsters (Heinrich et al., 1982); increased hi rats (Brightwell et al., 1986); and unaffected in
22 cats, rats, and monkeys (Pepelko and Peirano, 1983; Research Committee for HERP Studies,
23 1988; Lewis et al., 1989). These inconsistent findings indicate that the leukocyte counts are more
24 indicative of the clinical status of the laboratory animals than any direct effect of exposure to
25 diesel exhaust.
26 No significant changes in heart mass were found hi guinea pigs or rats exposed to diesel
27 exhaust (Wiester et al., 1980; Penney et al., 1981; Lewis et al., 1989). Rats exposed to diesel
28 exhaust showed a greater increase in the medial wall thickness of pulmonary arteries of differing
29 diameters and right ventricular wall thickness; these increases, however, did not achieve
30 statistically significant levels (Vallyathan et al., 1986). Brightwell et al. (1986) reported
31 increased heart/body weight and right ventricular/heart weight ratios and decreased left
32 ventricular contractility in rats exposed to 6.6 mg/m3 DPM for 16 h/day, 5 days/week for
33 104 weeks.
34 The effects of DPM on the endothelium-dependent relaxation (EDR) of vascular smooth
muscle cells have been investigated (Ikeda et al., 1995, 1998). Incubation of rat thoracic aortae
36 with suspensions of DPM (10-100 ng/mL) markedly attenuated acetylcholine-induced EDR. The
7/25/00 5-47 DRAFT—DO NOT CITE OR QUOTE
-------
1 mechanism of this effect was studied further in cultured porcine endothelial cells (CPE).
2 A 1 0-min incubation of PEC with DPM (0. 1 - 1 00 ug/mL) inhibited endothelium-dependent
3 relaxing factor (EDRF) or nitric oxide (NO) release. A 1 0-min incubation of DPM with NO
4 synthase inhibited formation of NO2", a product of NO metabolism. The authors concluded that
5 DPM, at the concentrations tested, neither induced cell damage nor inhibited EDRF release from
6 PEC, but scavenged and thereby blocked the physiological action of NO.
7
8 5.1.23.9. Serum chemistry. A number of investigators have studied the effects of exposure to
9 diesel exhaust on serum biochemistry, and no consistent effects have been found. Such studies
1 0 are summarized in Table 5-12.
1 1 The biological significance of changes in serum chemistry in female but not male rats
1 2 exposed at 2 mg/m3 DPM for 7 h/day, 5 days/week for 104 weeks (Lewis et al., 1989) is difficult
13 to interpret. Not only were the effects noted in one sex (females) only, but the serum enzymes,
1 4 lactate dehydrogenase (LDH), serum glutamic-oxaloacetic transaminase (SGOT), and serum
1 5 glutamic-pyruvic transaminase (SGPT), were elevated in the control group, a circumstance
1 6 contrary to denoting organ damage in the exposed female rats. The elevations of liver-related
1 7 serum enzymes in the control versus the exposed female rats appear to be a random event among
1 8 these aged subjects. The incidence of age-related disease, such as mononuclear cell leukemia,
1 9 can markedly affect such enzyme levels, seriously compromising the usefulness of a comparison
20 to historical controls. The serum sodium values of 144 versus 148 mmol/L in control and
2 1 exposed rats, respectively, although statistically different, would have no biological import.
22 The increased serum enzyme activities, alkaline phosphatase, SGOT, SGPT,
23 gamma-glutamyl transpeptidase, and decreased cholinesterase activity suggest an impaired liver;
24 however, such an impairment was not established histopathologically (Heinrich et al., 1982;
25 Research Committee for HERP Studies, 1988; Brightwell et al., 1986). The increased urea
26 nitrogen, electrolyte levels, and gamma globulin concentration and reduction in total blood
27 proteins are indicative of impaired kidney function. Again, there was no histopathological
28 confirmation of impaired kidneys in these studies.
29 Clinical chemistry studies suggest impairment of both liver and kidney functions in rats
30 and hamsters chronically exposed to high concentrations of diesel exhaust. The absence of
3 1 histopathological confirmation, the appearance of such effects near the end of the lifespan of the
32 laboratory animal, and the failure to find such biochemical changes in cats exposed to a higher
33 dose, however, tend to discredit the probability of hepatic and renal hazards to humans exposed
34 at atmosheric levels of diesel exhaust.
7/25/00 5-48 DRAFT— DO NOT CITE OR QUOTE
-------
5.1.23.10. Effects on microsomal enzymes. Several studies have examined the effects of diesel
exhaust exposure on microsomal enzymes associated with the metabolism and possible
3 activation of xenobiotics, especially polynuclear aromatic hydrocarbons. These studies are
4 summarized in Table 5-13. Lee etal. (1980) measured the activities of aryl hydrocarbon
5 hydroxylase (AHH) and epoxide hydrase (EH) in liver, lung, testis, and prostate gland of adult
6 male rats exposed to 6.32 mg/m3 DPM 20 h/day for 42 days. Maximal significant AHH activities
7 (pmol/min/mg microsomal protein) occurred at different times during the exposure period, and
8 differences between controls and exposed rats, respectively, were as follows: prostate
9 0.29 versus 1.31, lung 3.67 versus 5.11, and liver 113.9 versus 164.0. There was no difference in
10 AHH activity in the testis between exposed and control rats. Epoxide hydrase activity was not
11 significantly different from control values for any of the organs tested.
12 Pepelko and Peirano (1983) found no statistically significant differences in liver
13 microsomal cytochrome P448-450 levels and liver microsomal AHH between control and diesel-
14 exposed mice at either 6 or 8 mo of exposure. Small differences were noted in the lung
15 microsomal AHH activities, but these were believed to be artifactual differences, due to increases
16 in nonmicrosomal lung protein present in the microsomal preparations. Exposures to 6 mg/m3
17 DPM were for 8 h/day, 7 days/week.
8 Rabovsky et al. (1984) investigated the effect of chronic exposure to diesel exhaust on
\ 9 microsomal cytochrome P450-associated benzo[a]pyrene hydroxylase and 7-ethoxycoumarin
20 deethylase activities in rat lung and liver. Male rats were exposed for 7 h/day, 5 days/week for
21 104 weeks to 2 mg/m3 DPM. The exposure had no effect on B[a]P hydroxylase or
22 7-ethoxycoumarin deethylase activities in lung or liver. In related studies, Rabovsky et al. (1986)
23 examined the effects of diesel exhaust on vitally induced enzyme activity and interferon
24 production in female mice. The mice were exposed for 7 h/day, 5 days/week for 1 mo to diesel
25 exhaust diluted to achieve a concentration of 2 mg/m3 DPM. After the exposure, the mice were
26 inoculated intranasally with influenza virus. Changes in serum levels of interferon and liver
27 microsomal activities of 7-ethoxycoumarin, ethylmorphine demethylase, and nicotinamide
28 adenine dinucleotide phosphate (NADPH)-dependent cytochrome c reductase were measured.
29 In the absence of viral inoculation, exposure to diesel exhaust had no significant effects on the
30 activity levels of the two liver microsomal monooxygenases and NADPH-dependent cytochrome
31 c reductase. Exposure to diesel exhaust produced smaller increases in ethylmorphine
32 demethylase activity on days 2 to 4 postvirus infection and also abolished the day 4 postinfection
33 increase in NADPH-dependent cytochrome c reductase when compared with nonexposed mice.
34 These data suggested to the authors that the relationship that exists between metabolic
35 detoxification and resistance to infection in unexposed mice was altered during a short-term
36 exposure to diesel exhaust.
7/25/00 5-49 DRAFT—DO NOT CITE OR QUOTE
-------
1 Chen and Vostal (1981) measured the activity of AHH and the content of cytochrome
2 P450 in the lungs and livers of rats exposed by inhalation or intraperitoneal (i.p.) injection of a
3 dichloromethane extract of DPM. In the inhalation exposures, the exhaust was diluted to achieve
4 concentrations of 0.75 or 1.5 mg/m3 DPM, and the exposure regimen was 20 h/day,
5 5.5 days/week for up to 9 mo. The concentration of total hydrocarbons and particle-phase
6 hydrocarbons was not reported. Parenteral administration involved repeated i.p. injections at
7 several dose levels for 4 days. Inhalation exposure had no significant effect on liver microsomal
8 AHH activity; however, lung AHH activity was slightly reduced after 6 mo exposure to
9 1.5 mg/m3. An i.p. dose of DPM extract, estimated to be equivalent to the inhalation exposure,
10 had no effect on AHH activity in liver or lungs. No changes were observed in cytochrome
11 P450 contents in lungs or liver following inhalation exposure or i.p. treatment. Direct
12 intratracheal administration of a dichloromethane DPM extract required doses greater than
13 6 mg/kg body weight before the activity of induced AHH in the lung was barely doubled; liver
14 AHH activity remained unchanged (Chen, 1986).
15 In related studies, Navarro et al. (1981) evaluated the effect of exposure to diesel exhaust
16 on rat hepatic and pulmonary microsomal enzyme activities. The same exposure regimen was
17 employed (20 h/day, 5.5 days/week, for up to 1 year), and the exhaust was diluted to achieve
18 concentrations of 0.25 and 1.5 mg/m3 DPM (a few studies were also conducted at 0.75 mg/m3).
19 After 8 weeks of exposure, there was no evidence for the induction of cytochrome P450,
20 cytochrome P448, or NADPH-dependent cytochrome c reductase in rat liver microsomes. One
21 year of exposure had little, if any, effect on the hepatic metabolism of B[a]P. However, 1 year
22 of exposure to 0.25 and 1.5 mg/m3 significantly impaired the ability of lung microsomes to
23 metabolize B[ot]P (0.15 and 0.02 nmole/30 min/mg protein, respectively, versus
24 0.32 nmole/30 min/mg protein for the controls).
25 There are conflicting results regarding the induction of microsomal AHH activities in the
26 lungs and liver of rodents exposed to diesel exhaust. One study reported induced AHH activity
27 in the lungs, liver, and prostate of rats exposed to diesel exhaust containing 6.32 mg/m3 DPM for
28 20 h/day for 42 days; however, no induction of AHH was observed in the lungs of rats and mice
29 exposed to 6 mg/m3 DPM for 8 h/day, 7 days/week for up to 8 mo or to 0.25 to 2 mg/m3 for
30 periods up to 2 years. Exposure to diesel exhaust has not been shown to produce adverse effects
31 on microsomal cytochrome P450 in the lungs or liver of rats or mice. The weight of evidence
32 suggests that the absence of enzyme induction in the rodent lung exposed to diesel exhaust is
33 caused either by the unavailability of the adsorbed hydrocarbons or by their presence in
34 insufficient quantities for enzyme induction.
35
7/25/00 5-50 DRAFT—DO NOT CITE OR QUOTE
-------
5.1 JJ3.11. Effects on behavior and neurophysiology. Studies on the effects of exposure to
diesel exhaust on the behavior and neurophysiology of laboratory animals are summarized in
3 Table 5-14. Laurie et al. (1978) and Laurie et al. (1980) examined behavioral alterations in adult
4 and neonatal rats exposed to diesel exhaust. Exposure for 20 h/day, 7 days/week, for 6 weeks to
5 exhaust containing 6 mg/m3 DPM produced a significant reduction in adult spontaneous
6 locomotor activity (SLA) and in neonatal pivoting (Laurie et al., 1978). In a follow-up study,
7 Laurie et al. (1980) found that shorter exposure (8 h/day) to 6 mg/m3 DPM also resulted hi a
8 reduction of SLA in adult rats. Laurie et al. (1980) conducted additional behavioral tests on adult
9 rats exposed during their neonatal period. For two of three exposure situations (20 h/day for
10 17 days postparturition, or 8 h/day for the first 28 or 42 days postparturition), significantly lower
11 SLA was observed hi the majority of the tests conducted on the adults after week 5 of
12 measurement. When compared with control rats, adult 15-month-old rats that had been exposed
13 as neonates (20 h/day for 17 days) also exhibited a significantly slower rate of acquisition of a
14 bar-pressing task to obtain food. The investigators noted that the evidence was insufficient to
15 determine whether the differences were the result of a learning deficit or due to some other cause
16 (e-g-, motivational or arousal differences).
17 These data are difficult to interpret hi terms of health hazards to humans under ambient
8 environmental conditions because of the high concentration of diesel exhaust to which the
9 laboratory rats were exposed. Additionally, there are no further concentration-response studies to
20 assess at what exposure levels these observed results persist or abate. A permanent alteration in
21 both learning ability and activity resulting from exposures early in life is a health hazard whose
22 significance to humans should be pursued further.
23 Neurophysiological effects from exposure to diesel exhaust were investigated in rats by
24 Laurie and Boyes (1980, 1981). Rats were exposed to diluted diesel exhaust containing 6 mg/m3
25 DPM for 8 h/day, 7 days/week from birth up until 28 days of age. Somatosensory evoked
26 potential, as elicited by a 1 mA electrical pulse to the tibial nerve hi the left hind limb, and visual
27 evoked potential, as elicited by a flash of light, were the endpoints tested. An increased pulse
28 latency was reported for the rats exposed to diesel exhaust, and this was thought to be caused by
29 a reduction in the degree of nerve myelinization. There was no neuropathological examination,
30 however, to corifirm this supposition.
31 Based on the data presented, it is not possible to specify the particular neurological
32 impairment(s) induced by the exposure to diesel exhaust. Again, these results occurred following
33 exposure to a high level of diesel exhaust and no additional concentration-response studies were
34 performed.
7/25/00 5-51 DRAFT—DO NOT CITE OR QUOTE
-------
1 5.1.23.12. Effects on reproduction and development. Studies of the effects of exposure to
2 diesel exhaust on reproduction and development are summarized in Table 5-15. Twenty rats
3 were exposed 8 h/day on days 6 through 15 of gestation to diluted diesel exhaust containing
4 6 mg/m3 DPM (Werchowski et ah, 1980a,b; Pepelko and Peirano, 1983). There were no signs of
5 maternal toxicity or decreased fertility. No skeletal or visceral teratogenic effects were observed
6 in 20-day-old fetuses (Werchowski et al., 1980a). In a second study, 42 rabbits were exposed to
7 6 mg/m3 DPM for 8 h/day on gestation days 6 through 18. No adverse effects on body weight
8 gain or fertility were seen in the does exposed to diesel exhaust. No visceral or skeletal
9 developmental abnormalities were observed in the fetuses (Werchowski et al., 1980b).
10 Pepelko and Peirano (1983) evaluated the potential for diesel exhaust to affect
11 reproductive performance in mice exposed from 100 days prior to exposure throughout maturity
12 of the F2 generation. The mice were exposed for 8 h/day, 7 days/week to 12 mg/m3 DPM.
13 In general, treatment-related effects were minimal. Some differences in organ and body weights
14 were noted, but overall fertility and survival rates were not altered by exposure to diesel exhaust.
15 The only consistent change, an increase in lung weights, was accompanied by a gross
16 pathological diagnosis of anthracosis. These data denoted that exposure to diesel exhaust at a
17 concentration of 12 mg/m3 did not affect reproduction. See Section 5.3, which reports a lack of
18 effects of exposure to diesel exhaust on rat lung development (Mauderly et al., 1987b).
19 Several studies have evaluated the effect of exposure to diesel exhaust on sperm. Lewis
20 et al. (1989) found no adverse sperm effects (sperm motility, velocity, densities, morphology, or
21 incidence of abnormal sperm) in monkeys exposed for 7 h/day, 5 days/week for 104 weeks to 2
22 mg/m3 DPM. In another study in which A/Strong mice were exposed to diesel exhaust
23 containing 6 mg/m3 DPM for 8 h/day for 31 or 38 weeks, no significant differences were
24 observed in sperm morphology between exposed and control mice (Pereira et al., 1981). It was
25 noted, however, that there was a high rate of spontaneous sperm abnormalities in this strain of
26 mice, and this may have masked any small positive effect. Quinto and De Marinis (1984)
27 reported a statistically significant and dose-related increase in sperm abnormalities in mice
28 injected intraperitoneally for 5 days with 50,100, or 200 mg/kg of DPM suspended in corn oil.
29 A significant decrease in sperm number was seen at the highest dose, but testicular weight was
30 unaffected by the treatment.
31 Watanabe and Oonuki (1999) investigated the effects of diesel engine exhaust on
32 reproductive endocrine function in growing rats. The rats were exposed to whole diesel engine
33 exhaust (5.63 mg/m3 DPM, 4.10 ppm NO2, and 8.10 ppm NOX); a group was exposed to filtered
34 exhaust without DPM, and a group was exposed to clean air. Exposures were for 3 mo
35 beginning at birth (6 h/day for 5 days/week).
7/25/00 5-52 DRAFT—DO NOT CITE OR QUOTE
-------
Serum levels of testosterone and estradiol were significantly higher and follicle-
stimulating hormone significantly lower in animals exposed to whole diesel exhaust and filtered
3 exhaust compared to controls. Luteinizing hormone was significantly decreased in the whole-
4 exhaust-exposed group as compared to the control and filtered groups. Sperm production and
5 activity of testicular hyaluronidase were significantly reduced in both exhaust-exposed groups as
6 compared to the control group. This study suggests that diesel exhaust stimulates hormonal
7 secretion of the adrenal cortex, depresses gonadotropin-releasing hormone, and inhibits
8 spermatogenesis in rats. Because these effects were not inhibited by filtration, the gaseous phase
9 of the exhaust appears more responsible than particulate matter for disrupting the endocrine
10 system.
11 No teratogenic, embryotoxic, fetotoxic, or female reproductive effects were observed in
12 mice, rats, or rabbits at exposure levels up to 12 mg/m3 DPM. Effects on sperm morphology and
13 number were reported in hamsters and mice exposed to high doses of DPM; however, no adverse
14 effects were observed in sperm obtained from monkeys exposed at 2 mg/m3 for 7 h/day,
15 5 days/week for 104 weeks. Concentrations of 12 mg/m3 DPM did not affect male rat
16 reproductive fertility hi the F0 and F, generation breeders. Thus, exposure to diesel exhaust
17 would not appear to be a reproductive or developmental hazard.
8
5.2. MODE OF ACTION OF DIESEL EMISSIONS-INDUCED NONCANCER
20 EFFECTS
21 5.2.1. Comparison of Health Effects of Filtered and Unfiltered Diesel Exhaust
22 In four chronic toxicity studies of diesel exhaust, the experimental protocol included
23 exposing test animals to exhaust containing no particles. Comparisons were then made between
24 the effects caused by whole, unfiltered exhaust and those caused by the gaseous components of
25 the exhaust. Concentrations of components of the exposure atmospheres in these four studies are
26 given in Table 5-16.
27 Heinrich et al. (1982) compared the toxic effects of whole and filtered diesel exhaust on
28 hamsters and rats. Exposures were for 7 to 8 h/day and 5 days/week. Rats exposed for 24 mo to
29 either whole or filtered exhaust exhibited no significant changes in respiratory frequency,
30 respiratory minute volume, compliance or resistance as measured by a whole-body
31 plethysmography, or heart rate. In the hamsters, histological changes (adenomatous
32 proliferations) were seen in the lungs of animals exposed to either whole or filtered exhaust;
33 however, in all groups exposed to the whole exhaust the number of hamsters exhibiting such
34 lesions was significantly higher than for the corresponding groups exposed to filtered exhaust or
clean air. Severity of the lesions was, however, not reported.
7/25/00 5-53 DRAFT—DO NOT CITE OR QUOTE
-------
1 In a second study, Heinrich et al. (1986a, see also Stober, 1986) compared the toxic
2 effects of whole and filtered diesel exhaust on hamsters, rats, and mice. The test animals (96 per
3 test group) were exposed for 19 h/day, 5 days/week for 120 (hamsters and mice) or 140 (rats)
4 weeks. Body weights of hamsters were unaffected by either exposure. Body weights of rats and
5 mice were reduced by the whole exhaust but not by the filtered exhaust. Exposure-related higher
6 mortality rates occurred in mice after 2 years of exposure to whole exhaust. After 1 year of
7 exposure to the whole exhaust, hamsters exhibited increased lung weights, a significant increase
8 in airway resistance, and a nonsignificant reduction in lung compliance. For the same time
9 period, rats exhibited increased lung weights, a significant decrease in dynamic lung compliance,
10 and a significant increase in airway resistance. Test animals exposed to filtered exhaust did not
11 exhibit such effects. Histopathological examination indicated that different levels of response
12 occurred in the three species. In hamsters, filtered exhaust caused no significant
13 histopathological effects in the lung; whole exhaust caused thickened alveolar septa,
14 bronchioloalveolar hyperplasia, and emphysematous lesions. In mice, whole exhaust, but not
15 filtered exhaust, caused multifocal bronchioloalveolar hyperplasia, multifocal alveolar
16 lipoproteinosis, and multifocal interstitial fibrosis. In rats, there were no significant
17 morphological changes in the lungs following exposure to filtered exhaust. In rats exposed to
18 whole exhaust, there were severe inflammatory changes in the lungs, thickened alveolar septa,
19 foci of macrophages, crystals of cholesterol, and hyperplastic and metaplastic lesions.
20 Biochemical studies of lung lavage fluids of hamsters and mice indicated that exposure to filtered
21 exhaust caused fewer changes than did exposure to whole exhaust. The latter produced
22 significant increases in lactate dehydrogenase, alkaline phosphatase, glucose-6-phosphate
23 dehydrogenase, total protein, protease (pH 5.1), and collagen. The filtered exhaust had a slight
24 but nonsignificant effect on G6P-DH, total protein, and collagen. Similarly, cytological studies
25 showed that while the filtered exhaust had no effect on differential cell counts, the whole exhaust
26 resulted in an increase in leukocytes (161 ± 43.3/uL versus 55.7 ± 12.8/uL in the controls),
27 a decrease in AMs (30.0 ± 12.5 versus 51.3 ± 12.5/uL in the controls), and an increase in
28 granulocytes (125 ± 39.7 versus 1.23 ± 1.14/uL in the controls). All values presented for this
29 study are the mean with its standard deviation. The differences were significant for each cell
30 type. There was also a small increase in lymphocytes (5.81 ± 4.72 versus 3.01 ± 1.23/uL in the
31 controls).
32 Iwai et ai. (1986) exposed rats (24 per group) to whole or filtered diesel exhaust 8 h/day,
33 7 days/week for 24 mo. The whole exhaust was diluted to achieve a concentration of
34 4.9 ± 1.6 mg/m3 DPM. Body weights in the whole exhaust group began to decrease after 6 mo
3R and in both exposed groups began to decrease after 18 mo when compared with controls.
36 Lung-to-body weight ratios of the rats exposed to the whole exhaust showed a significant
7/25/00 5-54 DRAFT—DO NOT CITE OR QUOTE
-------
increase (/?<0.01) after 12 mo in comparison with control values. Spleen-to-body weight ratios
of both exposed groups were higher than control values after 24 mo. After 6 rno of exposure to
3 whole exhaust, DPM accumulated in AMs, and Type II cell hyperplasia was observed. After
4 2 years of exposure, the alveolar walls had become fibrotic with mast cell infiltration and
5 epithelial hyperplasia. In rats exposed to filtered exhaust, after 2 years there were only minimal
6 histologic changes in the lungs, with slight hyperplasia and stratification of bronchiolar
7 epithelium and infiltration of atypical lymphocytic cells in the spleen.
8 Brightwell et al. (1986) evaluated the toxic effects of whole and filtered diesel exhaust on
9 rats and hamsters. Three exhaust dilutions were tested, producing concentrations of 0.7,2.2, and
10 6.6 mg/m3 DPM. The test animals (144 rats and 312 hamsters per exposure group) were exposed
11 for five 16-h periods per week for 2 years. The four exposure types were gasoline, gasoline
12 catalyst, diesel, and filtered diesel. The results presented were limited to statistically significant
13 differences between exhaust-exposed and control animals. The inference from the discussion
14 section of the paper was that there was a minimum of toxicity in the animals exposed to filtered
15 diesel exhaust: "It is clear from the results presented that statistically significant differences
16 between exhaust-exposed and control animals are almost exclusively limited to animals exposed
17 to either gasoline or unfiltered diesel exhaust." Additional results are described in
18 Section 5.1.2.3.
/9 Heinrich et al. (1995) exposed female NMRI and C57BL/6N mice to a diesel exhaust
20 dilution that resulted in a DPM concentration of 4.5 mg/m3 and to the same dilution after filtering
21 to remove the particles. This study is focused on the carcinogenic effects of DPM exposure, and
22 inadequate information was presented to compare noncancer effects in filtered versus unfiltered
23 exhaust.
24 A comparison of the toxic responses in laboratory animals exposed to whole exhaust or
25 filtered exhaust containing no particles demonstrates across studies that when the exhaust is
26 sufficiently diluted to limit the concentrations of gaseous irritants (NO2 and SO^, irritant vapors
27 (aldehydes), CO, or other systemic toxicants, the diesel particles are the prime etiologic agents of
28 noncancer health effects, although additivity or synergism with the gases cannot be ruled out.
29 These toxic responses are both functional and pathological and represent cascading sequelae of
30 lung pathology based on concentration and species. The diesel particles plus gas exposures
31 produced biochemical and cytological changes in the lung that are much more prominent than
32 those evoked by the gas phase alone. Such marked differences between whole and filtered diesel
33 exhaust are also evident from general toxicological indices, such as decreases in body weight and
34 increases in lung weights, pulmonary function measurements, and pulmonary histopathology
(e.g., proliferative changes in Type II cells and respiratory bronchiolar epithelium, fibrosis).
36 Hamsters, under equivalent exposure regimens, have lower levels of retained DPM in their lungs
7/25/00 5-55 DRAFT—DO NOT CITE OR QUOTE
-------
1 than rats and mice do and, consequently, less pulmonary function impairment and pulmonary
2 pathology. These differences may result from lower DPM inspiration and deposition during
3 exposure, greater DPM clearance, or lung tissue less susceptible to the cytotoxicity of deposited
4 DPM.
5
6 5.2.2. Mode of Action for the Noncarcinogenic Effects of DPM
7 As noted in Chapter 2, diesel emissions are a complex mixture that includes both a vapor
8 phase and a particle phase. The particle phase consists of poorly soluble carbon particles on the
9 surfaces of which are adsorbed a large number of organic and inorganic compounds. Although
10 the effects to be discussed are considered attributable to the particle phase (termed diesel
11 particulate matter or DPM), additive or synergistic effects due to the vapor phase cannot be
12 totally discounted. This may be especially so in the human studies and the animal toxicology
13 studies where exposure is to various dilutions of diesel emissions, or in the in vitro studies in
14 which the test material was captured by filtration.
15 The mechanisms by which DPM is inhaled, deposited, and cleared from the respiratory
16 tract are discussed in Chapter 3. DPM deposited upon airway surfaces may be cleared from the
17 respiratory tract completely, or may be translocated to other sites within the respiratory system.
18 The pathogenic sequence following the deposition of inhaled DPM begins with the interaction of
19 DPM with airway epithelial cells and phagocytosis by AMs. The airway epithelial cells and
20 activated AMs release chemotactic factors that attract neutrophils and additional AMs. As the
21 lung burden of DPM increases, there is an aggregation of particle-laden AMs in alveoli adjacent
22 to terminal bronchioles, increases in the number of Type II cells lining particle-laden alveoli, and
23 the presence of particles within alveolar and peribronchial interstitial tissues and associated
24 lymph nodes.
25 The macrophages engulfing the DPM may release cytokines, growth factors, and
26 proteases, which may cause inflammation, cell injury, cell proliferation, hyperplasia, and fibrosis.
27 This is especially true under lung overload conditions occurring in laboratory rats when the rate
28 of deposition exceeds the rate of alveolar clearance. This phenomenon is described in Chapter 3.
29 The mechanisms leading to the generation of oxygen radicals and subsequent lung injury are
30 described in Chapter 7.
31 DPM is a poorly soluble particle whose rate of clearance by dissolution is insignificant
32 compared to its rate of clearance as an intact particle. The organic material adsorbed to the
33 surface is desorbed from the DPM and may enter into metabolic reactions and be activated and
34 enter into reactions with other macromolecules or be detoxified and excreted (Figure 7-1). The
35 diesel particle may be cleared directly by the clearance mechanisms described in Chapter 3.
7/25/00 5-56 DRAFT—DO NOT CITE OR QUOTE
-------
The organic material desorbed from the particle (described in Chapter 7) appears to be
associated with the immunological changes described in Section 5.1.1.1.4. The potential
3 adjuvant effects of DPM have also been studied. The results, described in Section 5.1.2.3.6,
4 indicate that while the nonextractable particle core contributes substantially to the adjuvant
5 activity of DPM, the organic matter adsorbed to DPM, notably pyrene, also augments the
6 adjuvant effect.
7 Thus, the available evidence indicates that DPM has the potential to produce pathological
8 and immunological changes in the respiratory tract. Moreover, the magnitude of these responses
9 is determined by the dose delivered to the respiratory tract and is attributable to both the carbon
10 core and the adsorbed organic materials.
11
12 5.3. INTERACTIVE EFFECTS OF DIESEL EXHAUST
13 A multitude of factors may influence the susceptibility to exposure to diesel exhaust as
14 well as the resulting response. Some of these have already been discussed in detail (e.g., the
15 composition of diesel exhaust and concentration-response data); others will be addressed in this
16 section (e.g., the interaction of diesel exhaust with factors particular to the exposed individual
17 and the interaction of diesel exhaust components with other airborne contaminants).
Mauderly et al. (1990a) compared the susceptibility of normal rats and rats with
T9 preexisting laboratory-induced pulmonary emphysema exposed for 7 h/day, 5 days/week for
20 24 mo to diesel exhaust containing 3.5 mg/m3 DPM or to clean air (controls). Emphysema was
21 induced in one-half of the rats by intratracheal instillation of elastase 6 weeks before exhaust
22 exposure. Measurements included lung burdens of DPM, respiratory function, bronchoalveolar
23 lavage, clearance of radiolabeled particles, pulmonary immune responses, lung collagen, excised
24 lung weight and volume, histopathology, and mean linear intercept of terminal air spaces. None
25 of the data for the 63 parameters measured suggest that rats with emphysematous lungs were
26 more susceptible than rats with normal lungs to the effects of diesel exhaust exposure. In fact,
27 each of the 14 emphysema-exhaust interactions detected by statistical analysis of variance
28 indicated that emphysema acted to reduce the effects of diesel exhaust exposure. DPM
29 accumulated much less rapidly in the lungs of emphysematous rats than in those of normal rats.
30 The mean lung burdens of DPM in the emphysematous rats were 39%, 36%, and 37% of the lung
31 burdens of normal rats at 12,18, and 24 mo, respectively. No significant interactions were
32 observed among lung morphometric parameters. Emphysema prevented the exhaust-induced
33 increase for three respiratory indices of expiratory flow rate at low lung volumes, reduced the
_34 exhaust-induced increase in nine lavage fluid indicators of lung damage, prevented the
15 expression of an exhaust-induced increase in lung collagen, and reduced the exhaust-induced
36 delay in DPM clearance.
7/25/00 5-57 DRAFT—DO NOT CITE OR QUOTE
-------
1 Mauderly et al. (1987b) evaluated the relative susceptibility of developing and adult rat
2 lungs to damage by exposure to diesel exhaust. Rats (48 per test group) were exposed to diesel
3 exhaust containing 3.5 mg/m3 DPM and about 0.8 ppm NO2. Exposures were for 7 h/day,
4 5 days/week through gestation to the age of 6 mo, or from the age of 6 to 12 mo. Comparative
5 studies were conducted on respiratory function, immune response, lung clearance, airway fluid
6 enzymes, protein and cytology, lung tissue collagen, and proteinases in both age groups. After
7 the 6-mo exposure, adult rats, compared with controls, exhibited (1) more focal aggregates of
8 particle-containing AMs in the alveolar ducts near the terminal bronchioles, (2) a sixfold increase
9 in the neutrophils (as a percentage of total leukocytes) hi the airway fluids, (3) a significantly
10 higher number of total lymphoid cells in the pulmonary lymph nodes, (4) delayed clearance of
11 DPM and radiolabeled particles (t1/2 = 90 days versus 47 days for controls), and (5) increased
12 lung weights. These effects were not seen in the developing rats. On a weight-for-weight
13 (milligrams of DPM per gram of lung) basis, DPM accumulation in the lungs was similar in
14 developing and adult rats immediately after the exposure. During the 6-mo postexposure period,
15 DPM clearance was much more rapid in the developing rats, approximately 2.5-fold. During
16 postexposure, diesel particle-laden macrophages became aggregated in the developing rats, but
17 these aggregations were located primarily in a subpleural position. The authors concluded that
18 exposure to diesel exhaust, using pulmonary function, structural (qualitative or quantitative)
19 biochemistry as the indices, did not affect the developing rat lung more severely than the adult rat
20 lung.
21 As a result of the increasing trend of using diesel-powered equipment in coal mining
22 operations and the concern for adverse health effects in coal miners exposed to both coal dust or
23 coal mine dust and diesel exhaust, Lewis et al. (1989) and Karagianes et al. (1981) investigated
24 the interaction of coal dust and diesel exhaust. Lewis et al. (1989) exposed rats, mice, and
25 cynomolgus monkeys to (1) filtered ambient air, (2) 2 mg/m3 DPM, (3) 2 mg/m3 respirable coal
26 dust, and (4) 1 mg/m3 of both DPM and respirable coal dust. Gaseous and vapor concentrations
27 were identical in both diesel exhaust exposures. Exposures were for 7 h/day, 5 days/week for up
28 to 24 mo. Synergistic effects between diesel exhaust and coal dust were not demonstrated;
29 additive toxic effects were the predominant effects noted.
3O Karagianes et al. (1981) exposed rats (24 per group) to diesel exhaust containing
31 8.3 mg/m3 of DPM alone or in combination with about 6 mg/m3 of coal dust. No synergistic
32 effects were found between diesel exhaust and coal dust; additive effects in terms of visual dust
33 burdens in necropsied lungs were related to dose (i.e., length of exposure and airborne particulate
34 concentrations).
35 The health effects of airborne contaminants from sources other than diesel engines may
36 be altered in the presence of DPM by their adsorption onto the diesel particles. When adsorbed
•7 /•"» C /f\f\
5-58 DRAFT—DO NOT CITE OR QUOTE
-------
onto diesel particles, the gases and vapors can be transported and deposited deeper into the lungs,
and because they are more concentrated on the particle surface, the resultant cytotoxic effects or
3 physiological responses may be enhanced. Nitrogen dioxide adsorbed onto carbon particles
4 caused pulmonary parenchyma! lesions in mice, whereas NO2 alone produced edema and
5 inflammation but no lesions (Boren, 1964). Exposure to formaldehyde and acrolein adsorbed
6 onto carbon particles (1 to 4 jam) resulted in the recruitment of PMNs to trachea! and
7 intrapulmonary epithelial tissues but not when the aldehydes were tested alone (Kilburn and
8 McKenzie, 1978).
9 Madden et al. (2000) observed that O3 exposure increased the bioactivity of DPM. DPM,
10 pre-exposed to O3 for 48 h, was instilled into the lungs of laboratory rats. Lung inflammation
11 and injury were examined 24 h after instillation by lung lavage. DPM pre-exposed to 0.1 PPM
12 O3 was more potent in increasing neutrophilia, lavage total protein, and LDH compared to
13 unexposed DPM. Treatment of DPM with higher concentrations of O3 (1.0 PPM) decreased the
14 bioactivity of the particles.
15 There is no direct evidence that diesel exhaust, at concentrations found in the ambient
16 environment, interacts with other substances in the exposure environment or the physiological
17 status of the exposed subject other than unpaired resistance to respiratory tract infections.
Although there is experimental evidence that gases and vapors can be adsorbed onto
carbonaceous particles, enhancing the toxiciry of these particles when deposited in the lung, there
20 is no evidence for an increased health risk from such interactions with DPM under urban
21 atmospheric conditions. Likewise, there is no experimental evidence in laboratory animals that
22 the youth or preexisting emphysema of an exposed individual enhances the risk of exposure to
23 diesel exhaust.
24
25 5.4. COMPARATIVE RESPONSIVENESS AMONG SPECIES TO THE
26 HISTOPATHOLOGIC EFFECTS OF DIESEL EXHAUST
27 There is some evidence indicating that species may differ in pulmonary responses to
28 diesel exhaust. Mauderly (1994) compared the pulmonary histopathology of rats and mice after
29 18 mo of exposure to diesel exhaust. There was less aggregation of macrophages in rats. Diffuse
30 septal thickening was noted in the mice, but there were few inflammatory cells, no focal fibrosis,
31 little epithelial hyperplasia, and no epithelial metaplasia, as was observed in rats. Heinrich et al.
32 (1986a) reported that wet lung weight of hamsters increased only 1.8-fold following chronic
33 exposure to diesel exhaust, compared with an increase of 3.4-fold in rats. Smaller increases in
34 neutrophils, lactic acid dehydrogenase, collagen, and protein supported the conclusion of a lesser
|5 inflammatory response in Syrian hamsters. The histopathologic changes in the lungs of Chinese
36 hamsters after 6 mo exposure to diesel exhaust, on the other hand, was similar to that of rats
7/25/00 5-59 DRAFT—DO NOT CITE OR QUOTE
-------
1 (Pepelko and Peirano, 1983). Guinea pigs respond to chronic diesel exhaust exposure with a
2 well-defined epithelial proliferation, but it is based on an eosinophilic response in contrast to the
3 neutrophil-based responses in other species. Epithelial hyperplasia and metaplasia were quite
4 striking in the terminal and respiratory bronchioles of cats exposed for 27 mo to diesel exhaust
5 (Plopper et al., 1983). This study is of particular interest because the terminal airways of cats are
6 more similar to those of humans than rodent species are. It should be noted, however, that
7 exposure concentrations were very high (12 mg/m3) for most of the period. Lewis et al. (1989)
8 exposed rats and cynomolgus monkeys 8 h per day, 5 days per week for 2 years to diesel exhaust
9 at a particle concentration of 2 mg/m3. Unfortunately, this exposure rate was sufficiently low that
10 few effects were noted in either species other than focal accumulations of particles, primarily in
11 the alveolar macrophages, interstitium, and lymphoid tissue. It is apparent that species do vary in
12 their pulmonary responses to diesel exhaust exposure, despite the difficulty in making direct
13 comparisons because of differences in exposure regimes, lifespans, and pulmonary anatomy.
14 Most species do respond, however, suggesting that humans are likely to be susceptible to
15 induction of pulmonary pathology during chronic exposure to DE at some level.
16
17 5.5. DOSE-RATE AND PARTICULATE CAUSATIVE ISSUES
18 The purpose of animal toxicological experimentation is to elucidate mechanisms of action
19 and identify the hazards and dose-response effects posed by a chemical substance or complex
20 mixture and to extrapolate these effects to humans for subsequent health assessments. The
21 cardinal principle in such a process is that the intensity and character of the toxic action are a
22 function of the dose of the toxic agent(s) that reaches the critical site of action. The considerable
23 body of evidence reviewed clearly denotes that major noncancerous health hazards may be
24 presented to the lung following the inhalation of diesel exhaust. Based on pulmonary function
25 and histopathological and histochemical effects, a determination can be made concerning which
26 dose/exposure rates of diesel exhaust (expressed in terms of the DPM concentration) result in
27 injury to the lung and which appear to elicit no effect. The inhalation of poorly soluble particles,
28 such as those found in diesel exhaust, increases the pulmonary paniculate burden. When the
29 dosing rate exceeds the ability of the pulmonary defense mechanisms to achieve a steady-state
30 lung burden of particles, there is a slowing of clearance and the progressive retention of particles
31 in the lung that can ultimately approach a complete cessation of lung clearance (Morrow, 1988).
32 This phenomenon, which is reviewed in Section 3.4, has practical significance both for the
33 interpretation of experimental inhalation data and for the prevention of disease in humans
34 exposed to airborne particles.
35 The data for exposure intensities that cause adverse pulmonary effects demonstrate that
36 they are less than the exposure intensities reported to be necessary to induce lung tumors. Using
7/25/00 5-60 DRAFT—DO NOT CITE OR QUOTE
-------
the most widely studied laboratory animal species and the one reported to be the most sensitive
to tumor induction, the laboratory rat, the no-adverse-effect exposure intensity for adverse
3 pulmonary effects was 56 mg-h-m"3/week (Brightwell et al., 1986). The lowest-observed-effect
4 level for adverse pulmonary effects (noncancer) in rats was 70 mg-h-m"3/week (Lewis et al.,
5 1989), and for pulmonary tumors, 122.5 mg-h-m'Vweek (Mauderly et al., 1987a). The results
6 clearly show that noncancerous pulmonary effects are produced at lower exposure intensities
7 than are pulmonary tumors. Such data support the position that inflammatory and proliferative
8 changes in the lung may play a key role in the etiology of pulmonary tumors in exposed rats
9 (Mauderly et al., 1990b).
10 Adults who have a preexisting condition that may predispose their lungs to increased
11 particle retention (e.g., smoking or high paniculate burdens from nondiesel sources),
12 inflammation (e.g., repeated respiratory infections), epithelial proliferation (e.g., chronic
13 bronchitis), and fibrosis (e.g., silica exposure), as well as infants and children, because of their
14 developing pulmonary and immunologic systems, may have a greater susceptibility to the toxic
15 actions of diesel exhaust. It should be noted that both the developing lung and a model of a
16 preexisting disease state have been studied with regard to their effect on the lungs' response to
17 diesel exhaust (Mauderly et al., 1990a, 1987b). Mauderly et al. (1987b) showed that diesel did
not affect the developing lung more severely than the adult rat lung, and in fact, that clearance
T9 was faster in the younger lung. Mauderly et al. (1990a) compared the pulmonary response to
20 inhalation of diesel exhaust in rats with elastase-induced emphysema with normal rats. They
21 found that respiratory tract effects were not more severe in emphysematous rats and that the lung
22 burden of particles was less in the compromised rat. These studies provide limited evidence that
23 some factors that are often considered to result in a wider distribution of sensitivity among
24 members of the population may not have this effect with diesel exposure. However, these studies
25 have no counterpart in human studies and extrapolation to humans remains uncertain.
26 There is also the issue of whether the noncancerous health effects related to exposure to
27 diesel exhaust are caused by the carbonaceous core of the particle or substances adsorbed onto
28 the core, or both.
29 Current understanding, derived primarily from studies in rats, suggests that much of the
30 toxicity resulting from the inhalation of diesel exhaust relates to the carbonaceous core of the
31 particles. Several studies on inhaled aerosols demonstrate that lung reactions characterized by an
32 appearance of particle-laden AMs and their infiltration into the alveolar ducts, adjoining alveoli,
33 and tracheobronchial lymph nodes; hyperplasia of Type II cells; and the impairment of
34 pulmonary clearance mechanisms are not limited to exposure to diesel particles. Such responses
have also been observed in rats following the inhalation of coal dust (Lewis et al., 1989;
36 Karagianes et al., 1981), titanium dioxide (Heinrich et al., 1995; Lee et al., 1985), CB (Nikula et
7/25/00 5-61 DRAFT—DO NOT CITE OR QUOTE
-------
1 al., 1995; Heinrich et al., 1995), titanivim tetrachloride hydrolysis products (Lee et al., 1986),
2 quartz (Klosterkotter and Biinemann, 1961), volcanic ash (Wehner et al., 1986), amosite (Bolton
3 et al., 1983), and manmade mineral fibers (Lee et al., 1988) among others. In more recent
4 studies, animals have been exposed to CB that is similar to the carbon core of the diesel exhaust
5 particle. Nikula et al. (1995) exposed rats for 24 mo to CB or diesel exhaust at target exposure
Q concentrations of 2.5 and 6 mg/m3 (exposure rates of 200 or 520 mg-lrm'3/week). Both
7 concentrations induced AM accumulation, epithelial proliferation, inflammation, and fibrosis.
8 They observed essentially no difference in potency of nonneoplastic or in tumor responses based
9 on a regression analysis.
10 Dungworth et al. (1994) reported moderate to severe inflammation characterized by
11 multifocal bronchoalveolar hyperplasia, alveolar histiocytosis, and focal segmental fibrosis in
12 rats exposed to CB for up to 20 mo at exposure rates of 510 to 540 mg'rrm~3/week. The
13 observed lung pathology reflects notable dose-response relationships and usually evolves in a
14 similar manner. With increasing dose, there is an increased accumulation and aggregation of
15 particle-laden AMs, Type II cell hyperplasia, a foamy (degenerative) macrophage response,
16 alveolar proteinosis, alveolar bronchiolization, cholesterol granulomas, and often squamous cell
17 carcinomas and bronchioalveolar adenomas derived from metaplastic squamous cells in the areas
18 of alveolar bronchiolization.
19 Heinrich et al. (1995) compared effects of diesel exposure in rats and mice with exposure
20 to titanium dioxide or carbon black. Exposures to TiO2 and carbon black were adjusted during
21 the exposure to result in a similar lung burden for the three types of particles. At similar lung
22 burdens in the rat, DPM, TiO2, and CB had nearly identical effects on lung weights and on the
23 incidence of lesions, both noncancer and cancer. Also, a similar effect on clearance of a labeled
24 test aerosol was measured for the different particles. A comparison of the effect of DPM, TiO2,
25 and carbon black exposures in mice also showed a similar effect on lung weight, but noncancer
26 effects were not reported and no significant increase in tumors was observed.
27 Murphy et al. (1998) compared the lexicological effects of DPM with three other
28 particles chosen for their differing morphology and surface chemistry. One mg each of well-
29 characterized crystalline quartz, amorphous silica, CB, and DPM was administered to laboratory
30 , rats by a single intratracheai instiiiaiion. The laboratory rats were sacrificed at 48 h, and 1, 6, and
31 12 weeks after instillation. Crystalline quartz produced significant increases in lung permeability,
32 persistent surface inflammation, progressive increases in pulmonary surfactant and activities of
33 epithelial marker enzymes up to 12 wk after primary exposure. Amorphous silica did not cause
34 progressive effects but did produce initial epithelial damage with permeability changes that
35 regressed with time after exposure. By contrast, CB had little if any effect on lung permeability,
36 epithelial markers, or inflammation. Similarly, DPM produced only minimal changes, although
7/25/00 5^62 DRAFT—DO NOT CITE OR QUOTE
-------
the individual particles were smaller and differed in surface chemistry from CB. The authors
concluded that DPM is less damaging to the respiratory epithelium than is silicon dioxide, and
3 that the surface chemistry of the particle is more important than ultrafine size in explaining
4 biological activity.
5 These experiments provide strong support for the idea that diesel exhaust toxicity results
6 from a mechanism that is analogous to that of other relatively inert particles in the lung. This
7 qualitative similarity exists along with some apparent quantitative differences in the potency of
8 various particles for producing effects on the lung or on particle clearance.
9 The exact relationship between toxicity and particle size within the ultrafine particle
10 mode, including DPM (BeruBe et al., 1999), remains unresolved. Studies reviewed in the PM
11 CD (U.S. Environmental Protection Agency, 1996) suggest a greater inherent potential toxicity of
12 inhaled ultrafine particles. Exposure to ultrafine particles may increase the release of
13 proinflammatory mediators that could be involved in lung disease. For example, Driscoll and
14 Maurer (1991) compared the effects of fine (0.3 \im) and ultrafine (0.02 |im) TiO2 particles
15 instilled into the lungs of laboratory rats. Although both size modes caused an increase in the
16 numbers of AMs and PMNs in the lungs, and release of TNF and fibronectin by AMs, the
17 responses were greater and more persistent with the ultrafine particles. While fine particle .
exposure resulted in a minimally increased prominence of particle-laden macrophages associated
with alveolar ducts, ultrafine particle exposure produced a somewhat greater prominence of
20 macrophages, some necrosis of macrophages, and slight interstitial inflammation of the alveolar
21 duct region. Moreover, collagen increased only with exposure to ultrafine particles.
22 Oberdorster et al. (1992) compared the effects of fine (0.25 Jim) and ultrafine (0.02 \im)
23 TiO2 particles instilled into the lungs of laboratory rats on various indicators of inflammation.
24 Instillation of ultrafine particles increased the number of total cells recovered by lavage,
25 decreased the percentage of AMs, and increased the percentage of PMNs and protein. Instillation
26 with fine particles did not cause statistically significant effects. Thus, the ultrafine particles had
27 greater pulmonary inflammatory potency than did larger sizes of this material. The investigators
28 attributed the enhanced toxicity to greater interaction of the ultrafine particles with their large
29 surface area, with alveolar and interstitial macrophages, which resulted in enhanced release of
30 inflammatory mediators. They suggested that ultrafine particles of low in vitro solubility appear
31 to enter the interstitium more readily than do larger sizes of the same material, which accounted
32 for the increased contact with macrophages in this compartment of the lung. Driscoll and Maurer
33 (1991) noted that the pulmonary retention of ultrafine TiO2 particles instilled into rat lungs was
34 greater than for the same mass of fine-mode TiO2 particles. Thus, the available evidence tends to
suggest a potentially greater toxicity for inhaled ultrafine particles.
7/25/00 5-63 DRAFT—DO NOT CITE OR QUOTE
-------
1 Particle size, volume, surface area, and composition may be the critical elements in the
2 overload phenomenon following exposure to particles, which could explain those quantitative
3 differences. The overloaded AMs secrete a variety of cytokines, oxidants, and proteolytic
4 enzymes that are responsible for inducing particle aggregation and damaging adjacent epithelial
5 tissue (Oberdorster, 1994). For a more detailed discussion of mechanism, see Chapter 3.
6 The principal noncancerous health hazard to humans posed by exposure to diesel exhaust
7 is a structural or functional injury to the lung, on the basis of the laboratory animal data. Such
8 effects are demonstrable at dose rates or cumulative doses of DPM lower than those reported to
9 be necessary to induce lung tumors. An emerging human health issue concerning short-term
10 exposure to ambient DE/DPM is the potential for allergenic responses in several studies.
11 Heightened allergenic responses including increased cytokine production as well as increased
12 numbers of inflammatory cells have been detected in nasal lavage from humans exposed to
13 inhaled or instilled DE/DPM. In individuals already allergic to ragweed, exposure to DE/DPM
14 with the allergen was observed to result in an enhanced allergenic response, particularly IgE
15 production. Current knowledge indicates that the carbonaceous core of diesel particles is the
16 major causative factor in the injury to the lung and that other factors such as the cytotoxicity of
17 adsorbed substances on the particles also may play a role. The lung injury appears to be
18 mediated through effects on pulmonary AMs. Because noncancerous pulmonary effects occur at
19 lower doses than tumor induction does in the rat, and because these effects may be cofactors in
20 the etiology of diesel exhaust-induced tumors, noncancerous pulmonary effects must be
21 considered in the total evaluation of diesel exhaust, notably the particulate component.
22
23 5.6. SUMMARY AND DISCUSSION
24 5.6.1. Effects of Diesel Exhaust on Humans
25 The most readily identified acute noncancer health effect of diesel exhaust on humans is
26 its ability to elicit subjective complaints of eye, throat, and bronchial irritation and
27 neurophysiological symptoms such as headache, lightheadedness, nausea, vomiting, and
28 numbness and tingling of the extremities. Studies of the perception and offensiveness of the odor
29 of diesel exhaust and a human volunteer study in an exposure chamber have demonstrated that
30 the time of onset of the human subjective symptoms is inversely related to increasing
31 concentrations of diesel exhaust and the severity is directly related to increasing concentrations of
32 diesel exhaust. In one study in which a diesel engine was operated under varying load
33 conditions, a dilution factor of 140 to 475 was needed to reduce the exhaust level to an
34 odor-detection threshold level.
35 A public health issue is whether short-term exposure to diesel exhaust might result in an
36 acute decrement in ventilatory function and whether the frequent repetition of such acute
•?/*>
5/00 5-64 DRAFT—DO NOT CITE OR QUOTE
-------
respiratory effects could result in chronic lung function impairment. One convenient means of
studying acute decrements in ventilatory function is to monitor differences in pulmonary function
3 in occupationally exposed workers at the beginning and end of a workshift. In studies of
4 underground miners, bus garage workers, dockworkers, and locomotive repairmen, increases in
5 respiratory symptoms (cough, phlegm, and dyspnea) and decreases in lung function (FVC, FEV15
6 PEFR, and FEF25_75) over the course of a workshift were generally found to be minimal and not
7 statistically significant. In a study of acute respiratory responses in diesel bus garage workers,
8 there was an increased reporting of cough, labored breathing, chest tightness, and wheezing, but
9 no reductions in pulmonary function were associated with exposure to diesel exhaust.
10 Pulmonary function was affected in stevedores over a workshift exposure to diesel exhaust but
11 normalized after a few days without exposure to diesel exhaust fumes. In a third study, there was
12 a trend toward greater ventilatory function changes during a workshift among coal miners, but
13 the decrements were similar in miners exposed and not exposed to diesel exhaust.
14 Smokers appeared to demonstrate larger workshift respiratory function decrements and
15 increased incidents of respiratory symptoms. Acute sensory and respiratory symptoms were
16 earlier and more sensitive indicators of potential health risks from diesel exposure than were
17 decrements hi pulmonary function. Studies on the acute health effects of exposure to diesel
exhaust in humans, experimental and epidemiologic, have failed to demonstrate a consistent
pattern of adverse effects on respiratory morbidity; the majority of studies offer, at best,
20 equivocal evidence for an exposure-response relationship. The environmental contaminants have
21 frequently been below permissible workplace exposure limits; in those few cases where health
22 effects have been reported, the authors have failed to identify conclusively the individual or
23 collective causative agents in the diesel exhaust.
24 Chronic effects of diesel exhaust exposure have been evaluated in epidemiologic studies
25 of occupationally exposed workers (metal and nonmetal miners, railroad yard workers,
26 stevedores, and bus garage mechanics). Most of the epidemiologic data indicate an absence of an
27 excess risk of chronic respiratory disease associated with exposure to diesel exhaust. In a few
28 studies, a higher prevalence of respiratory symptoms, primarily cough, phlegm, or chronic
29 bronchitis, was observed among the exposed. These increased symptoms, however, were usually
30 not accompanied by significant changes in pulmonary function. Reductions in FEV, and FVC
31 and, to a lesser extent, FEF50 and FEF7S, also have been reported. Two studies detected
32 statistically significant decrements in baseline pulmonary function consistent with obstructive
33 airway disease. One study of stevedores had a limited sample size of 17 exposed and
34 11 controls. The second study in coal miners showed that both underground and surface workers
^fc at diesel-use mines had somewhat lower pulmonary performance than their matched controls.
36 The proportion of workers in or at diesel-use mines, however, showed equivalent evidence of
7/25/00 5-65 DRAFT—DO NOT CITE OR QUOTE
-------
1 obstructive airway disease, and for this reason the authors of the second paper felt that factors
2 other than diesel exposure might have been responsible. A doubling of minor restrictive airway
3 disease was also observed in workers in or at diesel-use mines. These two studies, coupled with
4 other reported nonsignificant trends in respiratory flow-volume measurements, suggest that
5 exposure to diesel exhaust may impair pulmonary function among occupational populations.
6 Epidemiologic studies of the effects of diesel exhaust on organ systems other than the pulmonary
7 system are scant. Whereas a preliminary study of the association of cardiovascular mortality and
8 exposure to diesel exhaust found a fourfold higher risk ratio, a more comprehensive
9 epidemiologic study by the same investigators found no significant difference between the
10 observed and expected number of deaths caused by cardiovascular disease.
11 Caution is warranted in the interpretation of results from the epidemiologic studies that
12 have addressed noncarcinogenic health effects from exposure to diesel exhaust. These
13 investigations suffer from myriad methodological problems, including (1) incomplete
14 information on the extent of exposure to diesel exhaust, necessitating in some studies estimations
15 of exposures from job titles and resultant misclassification; (2) the presence of confounding
16 variables such as smoking or occupational exposures to other toxic substances (e.g., mine dusts);
17 and (3) the short duration and low intensity of exposures. These limitations restrict drawing
18 definitive conclusions as to the cause of any noncarcinogenic diesel exhaust effect, observed or
19 reported.
20 It is also apparent that at some level of exposure DE as measured by DPM has the
21 potential to induce systemic and pulmonary inflammatory responses in healthy humans and in
22 stimulating allergen-induced allergic airway disease in sensitive humans.
23
24 5.6.2. Effects of Diesel Exhaust on Laboratory Animals
25 Laboratory animal studies of the toxic effects of diesel exhaust have involved acute,
26 subchronic, and chronic exposure regimens, hi acute exposure studies, toxic effects appear to
27 have been associated primarily with high concentrations of carbon monoxide, nitrogen dioxide,
28 and aliphatic aldehydes. In short- and long-term studies, toxic effects have been associated with
29 exposure to the complex exhaust mixture. Effects of diesel exhaust in various animal species are
30 summarized in Tables i-2 to 5-15. In short-term studies, health effects are not readily apparent,
31 and when found, are mild and result from concentrations of about 6 mg/m3 DPM and durations of
32 exposure approximating 20 h/day. There is ample evidence, however, that short-term exposures
33 at lower levels of diesel exhaust affect the lung, as indicated by an accumulation of DPM,
34 evidence of inflammatory response, AM aggregation and accumulation near the terminal
35 bronchioles, Type II cell proliferation, and the thickening of alveolar walls adjacent to AM
36 aggregation. Little evidence exists, however, from short-term studies that exposure to diesel
7/25/00 5-66 DRAFT—DO NOT CITE OR QUOTE
-------
exhaust impairs lung function. Chronic exposures cause lung pathology that results in altered
pulmonary function and increased DPM retention in the lung. Exposures to diesel exhaust have
3 also been associated with increased susceptibility to respiratory tract infection, neurological or
4 behavioral changes, an increase in banded neutrophils, and morphological alterations in the liver.
5
6 5.6.2.1. Effects on Survival and Growth
7 The data presented in Table 5-3 show limited effects on survival in mice and rats and
8 some evidence of reduced body weight in rats following chronic exposures to concentrations of
9 1.5 mg/m3 DPM or higher and exposure durations of 16 to 20 h/day, 5 days/week for 104 to
10 130 weeks. Increased lung weights and lung to body-weight ratios in rats, mice, and hamsters;
11 an increased heart to body weight ratio in rats; and decreased lung and kidney weights in cats
12 have been reported following chronic exposure to diesel exhaust. No evidence was found of an
13 effect of diesel exhaust on other body organs (Table 5-4). The lowest-observed-effect level in
14 rats approximated 1 to 2 mg/m3 DPM for 7 h/day, 5 days/week for 104 weeks.
15
16 5.6.2.2. Effects on Pulmonary Function
17 Pulmonary function impairment has been reported in rats, hamsters, cats, and monkeys
exposed to diesel exhaust and included lung mechanical properties (compliance and resistance),
diffusing capacity, lung volumes, and ventilatory performance (Table 5-5). The effects generally
20 appeared only after prolonged exposures. The lowest exposure levels (expressed in terms of
21 DPM concentrations) that resulted in impairment of pulmonary function occurred at 2 mg/m3 in
22 cynomolgus monkeys (the only level tested), 1.5 and 3.5 mg/m3 in rats, 4.24 and 6 mg/m3 in
23 hamsters, and 11.7 mg/m3 in cats. Exposures in monkeys, cats, and rats (3.5 mg/m3) were for
24 7 to 8 h/day, 5 days/week for 104 to 130 weeks. While this duration is considered to constitute a
25 lifetime study in rodents, it is a small part of the lifetime of a monkey or cat. Exposures in
26 hamsters and rats (1.5 mg/m3) varied in hours per day (8 to 20) and weeks of exposure (26 to
27 130). In all species but the monkey, the testing results were consistent with restrictive lung
28 disease; alteration in expiratory flow rates indicated that 1.5 mg/m3 DPM was a LOAEL for a
29 chronic exposure (Gross, 1981). Monkeys demonstrated evidence of obstructive airway disease.
30 The nature of the pulmonary impairment is dependent on the dose of toxicants delivered to and
31 retained in the lung, the site of deposition and effective clearance or repair, and the anatomy and
32 physiology of the affected species; these variables appear to be factors in the disparity of the
33 airway disease in monkey versus the other species tested.
34
7/25/00 5-67 DRAFT—DO NOT CITE OR QUOTE
-------
1 5.6.23. Histopathological and Histochemical Effects
2 Histological studies have demonstrated that chronic exposure to diesel exhaust can result
3 in effects on respiratory tract tissue (Table 5-6). Typical findings include alveolar histiocytosis,
4 AM aggregation, tissue inflammation, increase in PMNs, hyperplasia of bronchiolar and alveolar
5 Type II cells, thickened alveolar septa, edema, fibrosis, and emphysema. Lesions in the trachea
6 and bronchi were observed in some studies. Associated with these histopathological findings
7 were various histochemical changes in the lung, including increases in lung DNA, total protein,
8 alkaline and acid phosphatase, glucose-6-phosphate dehydrogenase; increased synthesis of
9 collagen; and release of inflammatory mediators such as leukotriene LTB and prostaglandin
10 PGF2et. Although the overall laboratory evidence is that prolonged exposure to DPM results in
11 histopathological and histochemical changes in the lungs of exposed animals, some studies have
12 also demonstrated that there may be a threshold of exposure to DPM below which pathologic
13 changes do not occur. These no-observed-adverse-effect levels for histopathological effects were
14 reported to be 2 mg/m3 for cynomolgus monkeys (the only concentration tested), 0.11 to
15 0.35 mg/m3 for rats, and 0.25 mg/m3 DPM for guinea pigs exposed for 7 to 20 h/day, 5 to
16 5.5 days/week for 104 to 130 weeks.
17
18 5.6.2.4. Effects on Airway Clearance
19 The pathological effects of DPM appear to be strongly dependent on the relative rates of
20 pulmonary deposition and clearance (Table 5-7). Clearance of particles from the alveolar region
21 of the lungs is a multiphasic process involving phagocytosis by AMs. Chronic exposure to DPM
22 concentrations of about 1 mg/m3 or above, under varying exposure durations, causes pulmonary
23 clearance to be reduced, with concomitant focal aggregations of particle-laden AMs, particularly
24 in the peribronchiolar and alveolar regions, as well as in the hilar and mediastinal lymph nodes.
25 The exposure concentration at which focal aggregates of particle-laden AMs occur may vary
26 from species to species, depending on rate of uptake and pulmonary deposition, pulmonary
27 clearance rates, the relative size of the AM population per unit of lung tissue, the rate of
28 recruitment of AMs and leukocytes, and the relative efficiencies for removal of particles by the
29 mucociliary and lymphatic transport system. The principal means by which PM clearance is
3G FcuUCcu IS uiTGug.li a ucCiT£35c ill LUC LuiiCtiOli Cl pUiiGOiISry Aivj.5. impairment OJL paltlCiC
31 clearance seems to be nonspecific and applies primarily to dusts that are persistently retained in
32 the lungs. Lung dust levels of approximately 0.1 to 1 mg/g lung tissue appear to produce this
33 effect in the Fischer 344 rat (Health Effects Institute, 1995). Morrow (1988) suggested that the
34 inability of particle-laden AMs to translocate to the mucociliary escalator is correlated to an
35 average composite particle volume per AM in the lung. When this particle volume exceeds
36 approximately 60 jam3 per AM in the Fischer 344 rat, impairment of clearance appears to be
7/25/00 5-68 DRAFT—DO NOT CITE OR QUOTE
-------
initiated. When the particulate volume exceeds approximately 600 um3 per cell, evidence
suggests that AM-mediated particulate clearance virtually ceases, agglomerated particle-laden
3 macrophages remain in the alveolar region, and increasingly nonphagocytized dust particles
4 translocate to the pulmonary interstitium. Data for other laboratory animal species and humans
5 are, unfortunately, limited.
6
7 5.6.2.5. Neurological and Behavioral Effects
8 Behavioral effects have been observed in rats exposed to diesel exhaust from birth to
9 28 days of age (Table 5-14). Exposure caused a decreased level of spontaneous locomotor
10 activity and a detrimental effect on learning in adulthood. In agreement with the behavioral
11 changes was physiological evidence for delayed neuronal maturation. Exposures were to
12 6 mg/m3 DPM for 8 h/day, 7 days/week from birth to about 7, 14,21, or 28 days of age.
13
14 5.6.2.6. Effects on Immunity and Allergenicity
15 Several laboratory animal studies have indicated that exposure to DPM can reduce an
16 animal's resistance to respiratory infection. This effect, which can occur even after only 2 or 6 h
17 of exposure to DE containing 5 to 8 mg/m3 DPM, does not appear to be caused by direct
impairment of the lymphoid or splenic immune systems; however, in one study of influenza virus
infection, interferon levels and hemaglutinin antibody levels were adversely affected in the
20 exposed mice.
21 As with humans, there are animal data suggesting that DPM is a possible factor in the
22 increasing incidence of allergic hypersensitivity. The effects have been demonstrated primarily
23 in acute human and laboratory animal studies and appear to be associated with both the
24 nonextractable carbon core and the organic fraction of DPM. It also appears that synergies with
25 DPM may increase the efficacy of known airborne allergens. Both animal and human cell culture
26 studies indicate that DPM also has the potential to act as an adjuvant.
27
28 5.6.2.7. Other Noncancer Effects
29 Essentially no effects (based on the weight of evidence of a number of studies) were
30 noted for reproductive and teratogenic effects in mice, rats, rabbits, and monkeys; clinical
31 chemistry and hematology in the rat, cat, hamster, and monkeys; and enzyme induction in the rat
32 and mouse (Tables 5-11 through 5-13 and 5-15).
33
34 5.6.3. Comparison of Filtered and Unfiltered Diesel Exhaust
The comparison of the toxic responses in laboratory animals exposed to whole diesel
36 exhaust or filtered exhaust containing no particles demonstrates across laboratories that diesel
7/25/00 5-69 DRAFT—DO NOT CITE OR QUOTE
-------
1 particles are the principal etiologic agent of noncancerous health effects in laboratory animals
2 exposed to diesel exhaust (Table 5-16). Whether the particles act additively or synergistically
3 with the gases cannot be determined from the designs of the studies. Under equivalent exposure
4 regimens, hamsters have lower levels of retained DPM in their lungs than rats and mice do and
5 consequently less pulmonary function impairment and pulmonary pathology. These differences
6 may result from a lower intake rate of DPM, lower deposition rate and/or more rapid clearance
7 rate, or lung tissue that is less susceptible to the cytotoxicity of DPM. Observations of a
8 decreased respiration in hamsters when exposed by inhalation favor lower intake and deposition
9 rates.
10
11 5.6.4. Interactive Effects of Diesel Exhaust
12 There is no direct evidence that diesel exhaust interacts with other substances in an
13 exposure environment, other than an impaired resistance to respiratory tract infections. Young
14 animals were not more susceptible. In several ways, animals with laboratory-induced
15 emphysema were more resistant. There is experimental evidence that both inorganic and organic
16 compounds can be adsorbed onto carbonaceous particles. When such substances become
17 affiliated with particles, these substances can be carried deeper into the lungs where they might
18 have a more direct and potent effect on epithelial cells or on AM ingesting the particles. Few
19 specific studies to test interactive effects of diesel exhaust with atmospheric contaminants, other
20 than coal dust, have been conducted. Coal dust and DPM had an additive effect only.
21
22 5.6.5. Conclusions
23 Conclusions concerning the principal human hazard from exposure to diesel exhaust are
24 as follows:
25 • Some occupational studies of acute exposure to diesel exhaust during work shifts
26 suggest that increased acute sensory and respiratory symptoms (cough, phlegm,
27 chest tightness, wheezing) are more sensitive indicators of possible health risks
28 from exposure to diesel exhaust than pulmonary function decrements (which were
29 consistently found not to be significantly associated with diesel exhaust exposure).
30 • Allergeiiic effects also have been demonstrated under short-term exposure
31 scenarios to either diesel exhaust or DPM. The evidence indicates that the
32 immunological changes appear to be due to the DPM component of diesel exhaust
33 and that the imrnimological changes are caused by both the non extractable carbon
34 core and the adsorbed organic fraction of the diesel particle. The toxicological
35 significance of these effects has yet to be resolved.
7/25/00 5-70 DRAFT—DO NOT CITE OR QUOTE
-------
• Noncancer effects in humans from long-term chronic exposure to DPM are not
evident. Noncancer effects from long-term exposure to DPM of several
3 laboratory animal species, conducted to assess the pathophysiologic effects of
4 DPM in humans showed pulmonary histopathology and chronic inflammation.
5
6 Although the mode of action of DE is not clearly evident for any of the effects documented
7 in this chapter, the respiratory tract effects observed under acute scenarios are suggestive of an
8 irritant mechanism, while lung effects observed in chronic scenarios indicate an underlying
9 inflammatory response. Current knowledge indicates that the carbonaceous core of the diesel
10 particle is the causative agent of the lung effects, with the extent of the injury being mediated at
11 least in part by a progressive impairment of AMs. It is noted that lung effects occur in response
12 to DE exposure hi several species and occur in rats at doses lower than those inducing particle
13 overload and a tumorigenic response (see above); it follows that lung effects such as
14 inflammation and fibrosis are relevant in the development of risk assessments for DE.
7/25/00 5-71 DRAFT—DO NOT CITE OR QUOTE
-------
Table 5-1. Human studies of exposure to diesel exhaust
Study
Description
Findings
Acute exposures
Kahn et al.
(1988)
El Batawi and
Noweir (1966)
Battigelli
(1965)
Katz et al.
(1960)
Hare and
Springer (1971)
Hare et al.
(1974)
Linnell and
Scott (1962)
Rudell et al.
(1990,1994)
Rudell et al.
(1996)
Battigelli
(1965)
Wade and
VJ,,,..—,,,,, nOQIN
l^VrVVAiAlMI )^LSS-*f
O*3iz-S!iiHr^1^7 **t
al. (1994)
13 cases of acute exposure, Utah and
Colorado coal miners.
161 workers, two diesel bus garages.
Six subjects, eye exposure chamber,
three dilutions.
14 persons monitoring diesel exhaust
in a train tunnel.
Volunteer panelists who evaluated
general public's response to odor of
diesel exhaust.
Odor panel under highly controlled
conditions determined odor threshold
for diesel exhaust.
Eight healthy nonsmoking subjects
exposed for 60 min in chamber to
diesel exhaust (3.7 ppm NO, 1.5 ppm
NO2, 27 ppm CO, 0.5 mg/m3
formaldehyde, particles (4.3
x 106/cm3). Exercise, 10 of each 20
min (75 W).
Volunteers exposed to diesel exhaust
for 1 h while doing light work
Exposure concentrations uncertain.
13 volunteers exposed to three
dilutions of diesel exhaust for 15 min
to 1 h.
Three railroad workers acutely
sx^ossd to dies?! pv.haust.
Vnlnnteers challenged bv a nasal
spray of 0.30 mg DPM.
Acute reversible sensory irritation, headache,
nervous system effects, bronchoconstriction were
reported at unknown exposures.
Eye irritation (42%), headache (37%), dizziness
(30%), throat irritation (19%), and cough and
phlegm (11%) were reported in this order of
incidence by workers exposed in the service and
repair of diesel-powered buses.
Time to onset was inversely related and severity of
eye irritation was associated with the level of
exposure to diesel exhaust
Three occasions of minor eye and throat irritation;
no correlation established with concentrations of
diesel exhaust components.
Slight odor intensity, 90% perceived, 60% objected;
slight to moderate odor intensity, 95% perceived,
75% objected; moderate odor intensity, 100%
perceived, almost 95% objected.
In six panelists, the volume of air required to dilute
raw diesel exhaust to an odor threshold ranged from
a factor of 140 to 475.
Odor, eye and nasal irritation in 5/8 subjects. BAL
findings: small decrease in mast cells, lymphocyte
subsets and macrophage phagocytosis; small
increase in PMNs.
Unpleasant smell along with irritation of eyes and
nose reported. Airway resistance increased.
Reduction of particle concentration by trapping did
rr* -a*^--*—.....I*.,.
iiGl aiicci i wouiio.
No significant effects on pulmonary resistance were
observed as measured by plethysmography.
The workers developed symptoms of asthma.
Enhancement of IgE production reported due to a
dramatic increase in IgE-secreting cells.
//25/00
5-72
DRAFT—DO NOT CITE OR QUOTE
-------
Table 5-1. Human studies of exposure to diesel exhaust (continued)
Study
Description
Findings
Takenaka et al.
(1995)
Diaz-Sanchez et
al. (1996)
Diaz-Sanchez et
al. (1997)
Salvi et al.
(1999)
Salvi et al.
(2000)
Reger(1979)
Ames et al.
(1982)
Volunteers challenged by a nasal
spray of 0.30 mg DPM.
Volunteers challenged by a nasal
spray of 0.30 mg DPM.
Ragweed-sensitive volunteers
challenged by a nasal spray of 0.30
mg DPM alone or in combination
with ragweed allergen.
Volunteers exposed to diluted diesel
exhaust (DPM 300 |Ag/m3) for 1 h
with intermittent exercise.
Volunteers exposed to diluted diesel
exhaust (DPM 300 Jig/m3) for 1 h.
DPM extracts enhanced interleukin-4 plus
monoclonal antibody-stimulated IgE production as
much as 360%, suggesting an enhancement of
ongoing IgE production rather than inducing
germline transcription or isotype switching.
A broad increase in cytokine expression predicted to
contribute to enhanced local IgE production.
Ragweed allergen plus DPM-stimulated ragweed-
specific IgE to a much greater degree than ragweed
alone, suggesting DPM may be a key feature in
stimulating allergen-induced respiratory allergic
disease.
No changes in pulmonary function, but
significant increases in neutrophils, B
lymphocytes, histamine, and fibronectin in
airway lavage fluid.
Bronchial biopsies 6 h after exposure showed
significant increase in neutrophils, mast cells,
CD4+ and CD8+ T lymphocytes; upregulation
of ICAM-1 and VCAM-1; increases in the
number of LFA-1+ in bronchial tissue.
• Significant increases in neutrophils and platelets
observed in peripheral blood.
• DPM enhanced gene transcription of IL-8 in
bronchial tissue and bronchial wash cells
• Increased expression of growth-regulated
oncogene-cc and IL-8 in bronchial epithelium;
trend towards increased IL-5 mRNA gene
transcripts.
Studies of cross-shift changes
Five or more VC maneuvers by each
of 60 coal miners exposed to diesel
exhaust at the beginning and end of a
workshift.
Pulmonary function of 60 diesel-
exposed compared with 90 non-
diesel-exposed coal miners over
workshift.
FEV,, FVC, and PEFR were similar between diesel
and non-diesel-exposed miners. Smokers had an
increased number of decrements over shift than
nonsmokers.
Significant workshift decrements occurred in miners
in both groups who smoked; no significant
differences in ventilatory function changes between
miners exposed to diesel exhaust and those not
exposed.
7/25/00
5-73
DRAFT—DO NOT CITE OR QUOTE
-------
Table 5-1. Human studies of exposure to diesel exhaust (continued)
Study
Description
Findings
Jorgensen and
Svensson
(1970)
Gamble et al.
(1979)
Gamble et al.
(1987a)
Ulfvarson et al.
(1987)
Battigelli et al.
(1964)
Gamble et al.
(1987b)
240 iron ore miners matched for
diesel exposure, smoking, and age
were given bronchitis questionnaires
and spirometry pre- and
postworkshift.
200 salt miners performed before- and
after-workshift spirometry. Personal
environmental NO2 and inhalable
particle samples were collected.
232 workers in 4 diesel bus garages
administered acute respiratory
questionnaire and before and after
workshift spirometry. Compared to
lead/acid battery workers previously
found to be unaffected by their
exposures.
Workshift changes in pulmonary
function were evaluated in crews of
roll-on/ roll-off ships and car ferries
and bus garage staff. Pulmonary
function was evaluated in six
volunteers exposed to diluted diesel
exhaust, 2.1 ppm NO2, and 0.6 mg/m3
paniculate matter.
Cross-sectional and longitudinal studies
Among underground (surrogate for diesel exposure)
miners, smokers, and older age groups, frequency of
bronchitis was higher. Pulmonary function was
similar between groups and subgroups except for
differences accountable to age.
Smokers had greater but not significant reductions in
spirometry than ex- or nonsmokers. NO2 but not
paniculate levels significantly decreased FEV1,
FEFjj, FEFjo, and FEF75 over the workshift.
Prevalence of burning eyes, headache, difficult or
labored breathing, nausea, and wheeze were higher
in diesel bus workers man in comparison population.
Pulmonary function was affected during a workshift
exposure to diesel exhaust, but it normalized after a
few days with no exposure. Decrements were
greater with increasing intervals between exposures.
No effect on pulmonary function was observed in the
experimental exposure study.
210 locomotive repairmen exposed to
diesel exhaust for an average of 9.6
years in railroad engine houses were
compared with 154 railroad yard
workers of comparable job status but
no exposure to diesel exhaust.
283 male diesel bus garage workers
from four garages in two cities were
examined for impaired pulmonary
function (FVC, FEV,, and flow rates).
Study population with a mean tenure
of 9 ± 10 years S.D. was compared to
a nonexposed blue-collar population.
No significant differences in VC, FEV,, peak flow,
nitrogen washout, or diffusion capacity or in the
prevalence of dyspnea, cough, or sputum were found
between the diesel exhaust-exposed and nonexposed
groups.
Analyses within the study population showed no
association of respiratory symptoms with tenure.
Reduced FEV, and FEFSO (but not FEF75) were
associated with increasing tenure. The study
population had a higher incidence of cough, phlegm,
and wheezing unrelated to tenure. Pulmonary
function was not affected in the total cohort of
diesel-exposed but was reduced with 10 or more
years of tenure.
5-74 DRAFT—DO NOT CITE OR QUOTE
-------
Table 5-1. Human studies of exposure to diesel exhaust (continued)
Study
Description
Findings
Purdham et al. Respiratory symptoms and pulmonary
(1987) function were evaluated in 17
stevedores exposed to both diesel and
gasoline exhausts in car ferry
operations; control group was 11 on-
site office workers.
Reger et al. Differences in respiratory symptoms
(1982) and pulmonary function were assessed
in 823 coal miners from 6 diesel-
equipped mines compared to 823
matched coal miners not exposed to
diesel exhaust.
Ames et al. Changes in respiratory symptoms and
(1984) function were measured during a 5-
year period in 280 diesel-exposed and
838 nonexposed U.S. underground
coal miners.
Attfield (1978) Respiratory symptoms and function
were assessed in 2,659 miners from
21 underground metal mines (1,709
miners) and nonmetal mines (950
miners). Years of diesel usage in the
mines were surrogate for exposure to
diesel exhaust.
Attfield et al. Respiratory symptoms and function
(1982) were assessed in 630 potash miners
from 6 potash mines through a
questionnaire, chest radiographs, and
spirometry. A thorough assessment of
the environment of each mine was
made concurrently.
No differences between the two groups for respira-
tory symptoms. Stevedores had lower baseline lung
function consistent with an obstructive ventilatory
defect compared with controls and those of Sydney,
Nova Scotia, residents. Caution in interpretation is
warranted because of small sample size. No
significant changes in lung function over workshift
or difference between two groups.
Underground miners in diesel-use mines reported
more symptoms of cough and phlegm and had lower
pulmonary function. Similar trends were noted for
surface workers at diesel-use mines. Pattern was
consistent with small airway disease but factors other
than exposure to diesel exhaust thought to be
responsible.
No decrements in pulmonary function or increased
prevalence of respiratory symptoms were found
attributable to diesel exhaust. In fact, 5-year
incidences of cough, phlegm, and dyspnea were
greater in miners without exposure to diesel exhaust
than in miners exposed to diesel exhaust.
Questionnaire found an association between an
increased prevalence of cough and aldehyde
exposure; this finding was not substantiated by
spirometry data. No adverse symptoms or
pulmonary function decrements were related to
exposure to NO2, CO, CO2, dust, or quartz.
No obvious association indicative of diesel exposure
was found between health indices, dust exposure,
and pollutants. Higher prevalences of cough and
phlegm but no differences in FVC and FEV, were
found in these diesel-exposed potash workers when
compared with predicted values from a logistic
model based on blue-collar staff working in
nondusty jobs.
7/25/00
5-75
DRAFT—DO NOT CITE OR QUOTE
-------
Table 5-1. Human studies of exposure to diesel exhaust (continued)
Study
Description
Findings
Gamble et al.
(1983)
Gamble and
Jones(1983)
Edling and
Axelson (1984)
Edling et al.
(1987)
Respiratory morbidity was assessed in
259 miners in 5 salt mines by
respiratory symptoms, radiographic
findings, and spirometry. Two mines
used diesels extensively, two had
limited use, and one used no diesels in
1956, 1957, 1963, or 1963 through
1967. Several working populations
were compared with the salt-mine
cohort
Same as above. Salt miners were
grouped into low-, intermediate-, and
high-exposure categories based on
tenure in jobs with diesel exposure.
Pilot study of 129 bus company
employees classified into 3 diesel-
exhaust exposure categories: clerks
(0), bus drivers (1), and bus garage
workers.
Cohort of 694 male bus garage
employees followed from 1951
through 1983 was evaluated for
mortality from cardiovascular disease.
Subcohorts categorized by levels of
exposure were clerks (0), bus drivers
(1), and bus garage employees (2).
After adjustment for age and smoking, salt miners
showed no symptoms or increased prevalence of
cough, phlegm, dyspnea, or air obstruction
(FEV,/FVC) compared with aboveground coal
miners, potash workers, or blue-collar workers.
FEV,, FVC, FEFjo, and FEF75 were uniformly lower
for salt miners in comparison with all the comparison
populations. No changes in pulmonary function
were associated with years of exposure or
cumulative exposure to inhalable particles or NO2.
A statistically significant dose-related association of
phlegm and diesel exposure was noted. Changes in
pulmonary function showed no association with
diesel tenure. Age- and smoking-adjusted rates of
cough, phlegm, and dyspnea were 145%, 169%, and
93% of an external comparison population.
Predicted pulmonary function indices showed small
but significant reductions; there was no dose-
response relationship.
The most heavily exposed group (bus garage
workers) had a fourfold increase in risk of dying
from cardiovascular disease, even after correction
for smoking and allowing for 10 years of exposure
and 14 years or more of induction latency time.
No increased mortality from cardiovascular disease
was found among the members of these five bus
companies when compared with the general
population or grouped as subcohorts with different
levels of exposure.
7/25/00
5-7f»
DRAFT—DO NOT CITE OR QUOTE
-------
Table 5-2. Short-term effects of diesel exhaust on laboratory animals
-/I
—-,
3
J\
Ij
•J
ri
v
>
T)
3
•5
mS
z
•)
i
i
2
—4
Species/sex
Rat, F344, M;
Mouse, A/J, M; Hamster,
Syrian, M
Rat, F344, M, F; Mouse,
CD-I,M,F
Cat, Inbred, M
Ral, Sprague-
Dawley, M
Guinea Pig,
Hartley, M, F
Rat, F344,
M
Guinea Pig, Hartley, M
Guinea Pig, Hartley, M
Exposure period
20 h/day
7 days/week
10- 13 weeks
7 h/day. -
5 days/week
19 weeks
20 h/day
7 days/week
4 weeks
20 h/day
7 days/week
4 weeks
20 h/day
7 days/week
4 weeks
20 h/day
5.5 days/week
4 weeks
30min
3h
Particles
(mg/m1)
1.5
O.I9umMMD
0.21
1.0
4.4
6.4
6.4
6.8'
6.8'
6.0
6.8umMMD
l-2mgDPM
Intranasally
1
3.2
CxT
(mg-h/m1)
2, 100 to 2,730
140
665
2,926
3,584
3,584
3,808
3,808
2,640
0.5
1.6
CO NO, SO,
(ppm) (ppm) (ppm) Effects
6.9 0.49 — Increase in lung wt; increase in
thickness of alveolar walls;
minimal species difference
— — — No effects on lung function in rats
— — — (not done in mice); increase in
— — — PMNs and proteases and AM
aggregation in both species
14.6 2.1 2.1 Few effects on lung function; focal
pneumonitis or alveolitis
16.9 2.49 2.10 Decreased body wt; arterial blood
16.1* 2.76' 1.86* pH reduced; vital capacity, total
, „ „ . „ lung capacities increased
(<0.01 ppm Oj)'
16.7 2.9 1.9 Exposure started when animals
were 4 days old; increase in
(<0.01 ppm O,)' pulmonary How; bardycardia
— — — Macrophage aggregation; increase
in PMNs; Type II cell
proliferation; thickened alveolar
walls
— — — Augmented increases in nasal
airway resistance and vascular
permeability induced by a
histamine aerosol
5.9 1 .4 0. 1 3 Similar results to those reported in
12.9 4.4 0.34 the previous study using intranasal
challenge
Study
Kaplan etal. (1982)
Mauderly et al. (|981)
Pepelkoetal. (1980a)
Pepelko (1982a)
Wiester etal. (1980)
White and Garg (1981)
.••••
Kobayashi and Ito( 1995)
Kobayashi etal. (1997)
w
-------
7/25/00
Table 5-2. Short-term effects o
Species/sex Exposure period
Guinea Pig, Hartley, M, F 20 h/day
7 days/week
8 weeks
f diesel exhaust on laboratory animals (continued)
Particles
(mg/m1)
6.3
CxT
(rag h/m3)
7,056
CO
(ppm)
17.4
NO,
(ppm)
2.3
SO,
(ppm)
2.1
Effects
Increase in relative lung wt. AM
aggregation; hypertrophy of goblet
cells; focal hyperplasia of alveolar
epithelium
Study
Wiesteretal. (1980)
Ul
ij
oo
Mouse ICR M
Rat, Sprague-Dawley,
M
6 weeks
24 h
100 ug DPM
intranasally
5-100 ug/IO'
AM/mLof
DPM
'Irradiated exhaust.
PMN = Polymorphonuclear leukocyte.
AM = Alveolar inacrophage.
DPM aggravated ovalbumin-
induced airway inflammation and
provided evidence that DPM can
enhance manifestations of allergic
asthma
Unchanged, but not organic-free
DPM enhanced production of
proinflammatory cytokines
Takanoctal. (1997)
Yang etal. (1997)
O
o
Z!
O
H
O
i— i
H
W
O
O
H
m
-------
NJ
Table 5-3. Effects of chronic exposures to diesel exhaust on survival and growth of laboratory animals
D
j)
^j
0
O
>
r)
1
3
D
z!
D
H
^
1
n
D
0
Species/sex
Rat, F344, M, F;
Monkey, Cynomolgus, M
Rat, F344, M;
Guinea Pig, Hartley, M
Hamster, Chinese, M
Rat, Wistar, M
Rat, F344, M, F;
Mouse, CD-I, M, F
Rat, Wistar, F;
Mouse, MMRJ, F
Rat, F344
M, F
Rat'
F344/Jcl.
Rat, Wistar, F;
Mouse, NMRI, F
(7 mg/m' only)
Exposure
period
7 h/day
5 days/week
1 04 weeks
20 h/day
5 days/week
106 weeks
8 h/day
7 days/week
26 weeks
6 h/day
5 days/week
87 weeks
7 h/day
5 days/week
130 weeks
19 h/day
5 days/week
104 weeks
16 h/day
5 days/week
104 weeks
16 h/day
6 days/week
130 weeks
18 h/day
5 days/week
24 mo
Particles
(mg/m')
2.0
0.23-0.36 um MMD
0.25
0.75
1.5
0.19 um MMD
6.0
12.0
8.3
0.71 umMMD
0.35
3.5
7.1
0.25 nm MMD
4.24
0.35 Mm MMD
0.7
2.2
6.6
0.11"
0.41"
1.08"
2.3111
3.72"
0.2-0.3 Mm MMD
0.84
2.5
6.98
CxT
(ragh/m5)
7,280
2,650
7,950
15,900
8,736
17,472
21,663
1,592
15,925
31,850
41,891
5,824
18,304
54,912
1,373
5,117
13,478
28,829
46,426
7,400
21,800
61,700
CO
(ppm)
11.5
2.71
4.4'
7.1'
—
—
50.0
2.9
16.5
29.7
12.5
—
—
32.0
1.23
2.12
3.96
7.10
12.9
2.6
8.3
21.2
NO,
(ppm)
1.5
O.lb
0.27"
0.5"
—
—
4.0-6.0
0.05
0.34
0.68
1.5
—
—
—
0.08
0.26
0.70
1.41
3.00
0.3
1.2
3.8
SO,
(ppm)
0.8
—
—
—
—
—
—
—
—
1.1
—
—
—
0.38
1.06
2.42
4.70
4.57
0.3
1.1
3.4
Effects
No effects on growth or survival
Reduced body weight in rats at
1.5 mg/m'
No effect on growth
No effect on growth or mortality
rates
No effect on growth or mortality
rates
Reduced body wts; increased
mortality in mice
Growth reduced at 2.2 and
6.6 mg/m'
Concentration-dependent
decrease in body weight; earlier
deaths in females exposed to 3.72
mg/m', stabilized by 15 mo
Reduced body weight in rats at
2.5 and 6.98 mg/m1 and no effect
in mice
Study
Lewis et al.
(1989)
Schreck et al.
(1981)
Vinegar et al.
(I981a,b)
Karagianes
etal. (1981)
Mauderly etal.
(1984, 1987a)
Heinrich et al.
(1986a)
Brightwell et al.
(1986)
Research
Committee for
HERP Studies
(1988)
Heinrich et al.
(1995)
/O
C
O
-------
7/25/00
Table 5-3. Effects of chronic exposures to diesel exhaust on survival and growth of laboratory animals (continued)
Species/sex
Mice.NMRI.F;
C57BL/6N, F
Rats, F344, M
Mouse, CD-I,
M,F
Exposure
period
ISh/day
5 days/week
13. 5 mo
(NMR1)
24 mo
(C57BL/N)
16h/day
5 days/week
23 mo
7 h/day
5 days/week
t04 weeks
Particles
(mg/m')
6.98
2.44
6.33
0.35
3.5
7.1
0.25 urn MOD
C*T
(rng-h/m1)
35,500 -NMR1
38,300 -C57
19,520
50,640
1,274
12,740
25,844
CO
(ppra)
14.2
—
3
17
30
NO,
(ppra)
2.3
—
0.1
0.3
0.7
SO,
(ppm) Effects
2.8 Reduced body weight in NMRJ
mice but not in C57BL/6N mice
— Reduced survival in 6.33 mg/mj
— after 300 days. Body weight
significantly lower at 6.33 mg/m'
— No effect on growth or mortality
— rates
Study
Heinrich et al.
(1995)
Nikulaetal.
(1995)
Mauderly et al.
(1996)
o
o
z;
o
H
O
*-H
H
m
o
5*3
'Estimated from graphically depicted mass concenlration data.
'Estimated from graphically presented mass concentration data for NO, (assuming 90% NO and 10% N02).
'Data for t( sis with light-duty engine; similar results with heavy-duty engine.
dLight-duty engine.
eHeavy-du y engine.
o
H
m
-------
Table 5-4. Effects of chronic exposures to diesel exhaust on organ weights and organ-to-body-weight ratios
1>^
o
L/1
oo
D
$>•
•D
H
j
1
O
0
trT,J
2!
O
H
n
H
tn
O
?o
O
c
o
H
m
Species/sex
Rat, F344, M;
Mouse, A/J, M;
Hamster, Syrian,
M
Rat, F344, M, F
Rat, F344, M
Rat, F344, F
Rat, F344; M
Guinea Pig,
Hartley, M
Hamster, Chinese,
M
Rat, Wistar, F;
Hamster, Syrian,
M, F
Mouse, NMR1, F
Rat, F344, M, F;
Hamster, Syrian,
M, F
Cat, inbred, M
Mouse, NMRI, F
(7 mg/m' only)
Exposure
period
20 h/day
7 days/week
12-13 weeks
7 h/day
5 days/week
52 weeks
20 h/day
5.5 days/ week
36 weeks
7 h/day
5 days/week
104 weeks
20 h/day
5.5 days/ week
78 weeks
8 h/day
7 days/week
26 weeks
19 h/day
5 days/week
120- 140 weeks
16 h/day
5 days/week
1 04 weeks
8 h/day
7 days/week
124 weeks
18 h/day
5 days/week
24 mo
Particles
(mg/m')
1.5
0.19 urn MMD
2.0
0.23-0.36 urn
MMD
0.25
1.5
019 urn MMD
2.0
0.23-0.36 urn
MMD
0.25
0.75
1.5
0.19 Mm MMD
6.0
12.0
4.24
0.35 urn MMD
0.7'
2.2b
6.6
6.01
12.0b
0.84
2.5
6.98
CxJ CO
(mg-h/rn') (ppm)
2,520-2,730 —
3,640 12.7
990 —
5,940 —
7,280 11.5
2,145 —
6,435 —
12,870 —
8,736 —
17,472 —
48,336-56,392 12.5
5,824 —
18,304 —
54,912 32.0
41,664 20.2
83,328 33.2
7,400 2.6
21,800 8.3
61,700 21.2
NO, SO,
(ppm) (ppm) Effects
— — No effect on liver, kidney, spleen, or
heart weights
1.6 0.83 No effects on weights of lungs, liver,
heart, spleen, kidneys, and testes
— — Increase in relative lung weight at
— — 1.5 mg/m' only initially seen at
12 weeks
1.5 0.81 No effects on heart weights
— — No effects on heart mass
— —
— —
— — Increase in lung weight and lung/body
— — weight ratio
1.5 1.1 Increase in rat, mouse, and hamster
lung weight and dry weights
— — Increase in lung weight concentration
— — related in rats; heart weight/body
— — weight ratio greater at 6.6 mg/m'
2.7 2.7 Decrease in lung and kidney weights
4.4 5.0
0.3 0.3 Increased rat and mouse lung weight at
1 .2 1.1 7 mg/m' from 6 mo and at 2.5 mg/m'
3.8 3.4 at 22 and 24 mo
Study
Kaplan etal. (1982)
Green et al. (1983)
Misiorowski et al.
(1980)
Vallyath an etal. (1986)
Penney etal. (1981)
Vinegar etal. (1981a,b)
Heinrich et al.
(I986a,b)
Stober(1986)
Brightwell et al. (1986)
Pepelkoetal.(1980b,
1981)
Moorman etal. (1985)
Heinrich etal. (1995)
-------
7/25/00
Table 5-1.
Species/se*
Mouse, NMRI.F;
C57BL/6N. F
Rats,F344,M
Rat
Mouse
Effects of chronic
Exposure
period
18h/day
5 days/week
l3.5mo(NMRJ)
24 mo
(C57BL/N)
16h/day
5 days/week
23 mo
exposures
Particles
(mg/mj)
6.98
2.44
6.33
0.8
2.5
6.98
6.98
4.5
to diesel exhaust on organ weights and
CxT CO NO, SO,
(mgh/mj) (ppm) (ppm) (ppra)
35,500 -NMRI 14.2 2.3 2.8
38,300 -C57
19,520 — — —
50,640 — — —
organ-to-body-weight ratios (continued)
Effects
Increased lung weight
Increase in lung weight was significant
at 2 and 6 mg/m'
Increased lung weight in rats and mice
at 3.5 and 7. 1 mg/mj
Study
Heinrichetal. (1995)
Nikulaetal.(1995)
Henderson el al. (1988)
K)
O
o
2
o
H
o
H
w
o
M to 61 weeks of exposure.
b62 to 124 weeks of exposure.
c
o
H
m
-------
TCble 5-5. Effects of diesel exhaust on pulmonary function of laboratory animals
••w
yi
s\
o
>J
3
*3
>
n
H
3
3
2
—4
H
~)
— 4
d
*J
D
O
D
— H
3
3
Species/sex
Rat, F344, M, F
Monkey,
Cynomolgus, M
Rat, F344, M
Rat, Wislar, F
Hamster, Chinese, M
Rat, F344,
M, F
Rat, F344, M, F;
Hamster, Syrian, M, F
Hamster, Syrian, M, F
Rat, Wislar, F
Cat, inbred, M
M to 61 weeks exposure.
Exposure
period
7 h/day
5 days/week
104 weeks
7 h/day
5 days/week
104 weeks
20 h/day
5.5 days/week
87 weeks
7-8 h/day
5 days/week
104 weeks
8 It/day
7 days/week
26 weeks
7 h/day
5 days/week
130 weeks
16 h/day
5 days/ week
104 weeks
19 h/day
5 days/week
120 weeks
19 h/day
5 days/week
140 weeks
8 h/day
7 days/week
124 weeks
Particles
(mg/m])
2.0
0.23-0.36 um
MMD
2.0
0.23-0.36 um
MMD
1.5
0.1 9 pm MMD
3.9
0.1 umMMD
6.0
12.0
0.35
3.5
7.1
0.23-0.26 um
MMD
0.7
2.2
6.6
4.24
0.35 urn MMD
4.24
0.35 um MMD
6.01
12.0"
C*T
(rng-h/m1)
7,280
7,280
14,355
14,196-16,224
8.736
17,472
1,593
15,925
31,850
5,824
18,304
54,912
48,336
56,392
41,664
83,328
CO
(ppm)
11.5
11.5
7.0
18.5
—
—
2.9
16.5
29.7
—
—
—
12.5
12.5
20.2
33.3
NO,
(ppm)
1.5
1.5
0.5
1.2
—
—
0.05
0.34
0.68
—
—
—
1.5
1.5
2.7
4.4
SO,
(ppm) Effects
0.8 No effect on pulmonary function
0.8 Decreased expiratory flow; no effect
on vital or diffusing capacities
— Increased functional residual capacity,
expiratory volume, and flow
3. 1 No effect on minute volume,
compliance, or resistance
— Decrease in vital capacity, residual
— volume, and diffusing capacity;
increase in static deflation lung
volume
— Diffusing capacity, lung compliance
— reduced at 3.5 and 7. 1 mg/mj
—
— Large number of pulmonary function
— changes consistent with obstructive
— and restrictive airway diseases at
6.6 mg/m1 (no specific data provided)
1 . 1 Significant increase in airway
resistance
1 . 1 Decrease in dynamic lung compliance;
increase in airway resistance
2. 1 Decrease in vital capacity, total lung
5.0 capacity, and diffusing capacity after
2 vears: no effect on expiratory flow
Study
Lewis etal. (1989)
Lewis etal. (1989)
Gross (1981)
Heinrich et al. (1982)
Vinegar etal. (1980,
1981a,b)
Mauderly etal. (1988)
McClellan etal. (1986)
Brightwell et al. (1986)
Heinrich etal. (1986a)
Heinrich etal. (1986a)
Pepelkoetal. (1980b,
1981)
Moorman etal. (1985)
-------
-J
o
M
^
0
J\
io
^
O
>
n
3
Z
3
-}
— i
-i
rl
5
:>
•4
3
Table 5-6. Histopathological effects of diesel exhaust in the lungs of laboratory animals
Species/sex
Rat, F344, M;
Mouse, A/.(, M;
Hamster, Syrian, M
Monkey, Cynomolgus,
M
Rat, F344, M, F
Rat, Sprajjue-Dawley,
M; Mouse, A/HEJ, M
Hamster, Chinese, M
Hamster, Syrian, M, F
Rai, Wistar, M
Rat, F344, F
Rat, F3-I4, M, F;
Mouse, CD-I,
M,F
Exposure
period
20 h/day
7 days/week
12-13 weeks
7 h/day
5 days/week
104 weeks
7 h/day
5 days/week
104 weeks
8 h/day
7 days/week
39 weeks
8 h/day
5 days/week
26 weeks
7-8 h/dny
5 days/week
120 weeks
6 h/day
5 days/week
87 weeks
8 h/day
7 days/week
104 weeks
7 h/day
5 days/week
130 weeks
Particles
(mg/m1)
1.5
0.19 urn MOD
2.0
0.23-0.36 urn
MOD
2.0
0.23-0.36 urn
MOD
6.0
6.0
12.0
3.9
0.1 urn MOD
8.3
0.71 Jim MOD
4.9
0.35
3.5
7.1
0.23 um MOD
CxT
(mg-h/m1)
2,520-2,730
7,280
3,640
13,104
6,240
12,480
16,380-18,720
21,663
28,538
1,592
15,925
31,850
CO NO, SO,
(ppm) (ppm) (ppm) Effects
— — — Inflammatory changes, increase in lung
weight, increase in thickness of alveolar
walls
11.5 1.5 0.8 AM aggregation; no fibrosis,
inflammation, or emphysema
11.5 1.5 0.8 Multifocal histiocytosis, inflammatory
changes, Type II cell proliferation,
fibrosis
— — — Increase in lung protein content and
collagen synthesis but a decrease in
overall lung protein synthesis in both
species; prolylhydroxylase activity
increased in rats in utero
— — — Inflammatory changes, AM accumu-
— — — lation, thickened alveolar lining, Type II
cell hyperplasia,edema, increase in
collagen
18.5 1.2 3.1 Inflammatory changes, 60%
adenomatous cell proliferation
50.0 4.0-6.0 — Inflammatory changes, AM aggregation,
alveolar cell hypertrophy, interstitial
fibrosis, emphysema (diagnostic method-
ology not described)
7.0 1.8 13.1 Type II cell proliferation, inflammatory
changes, bronchial hyperplasia, fibrosis
2.9 0.05 — Alveolar and bronchiolar epithelial
16.5 0.34 — metaplasia in rats at 3. 5 and 7.0 mg/m1,
29.7 0.68 — fibrosis at 7.0 mg/m' in rats and mice,
inflammatory changes
Study
Kaplan etal. (1982)
Lewis etal. (1989)
Bhatnagar et al,
(1980)
Pepelko(1982a)
Bhatnagar et al.
(1980)
Pepe!ko(l982a)
Pepelko(1982b)
Heinrich et al. (1982)
Karagianes et al.
(1981)
I wai etal. (1986)
Mauderly et al.
(1987a)
Henderson et al.
(1988)
-------
Table 5-6. Histopathological effects of diesel exhaust in the lungs of laboratory animals (continued)
O
O
oo
O
O
z
O
H
n
i— (
H
m
o
G
o
H
w
Species/sex
Rats, SPF 344
Exposure
period
7h/day
Sdays/week
104 weeks
Particles
(mg/m1)
2 mg/m1 coal
dust (CD)
2 mg/m] DPM
1 mg/m1
CD + 1 mg/m'
DPM
Cxj CO NO, SO,
(mg-h/m1) (ppm) (ppm) (ppm) Effects
— — — — • Assessed pharmacological responses
of rat airway smooth muscle in vitro
• Maximal contractile responses to
acetylcholine of tissues from CD-,
DPM-, and CD + DPM- exposed
animals significantly increased;
effects of CD and DPM were additive
Study
Fedenetal.(1985)
Maximal relaxation response to
isoproterenol increased significantly
by CD + DPM exposure, but not by
individual treatments
The results indicate that chronic
exposure to CD, DPM, and CD +•
DPM produce differential
modifications in the behavior of rat
airway smooth muscle
Rat, Wistar, F;
Mouse, NMRI, F
(7 mg/m! only)
Mouse, NMRI, F;
C57BL/6N, F
Mouse
Rat, M, F,
F344/Jcl.
18h/day
5 days/week
24 mo
I8h/day
5 days/week
13. 5 mo (NMRI)
24 mo
(C57BL/N)
16n/day
6 days/week
130 weeks
0.8
2.5
6.98
6.98
4.5
0.1 11
0.4 11
I.081
2.31'
3.72"
7,400
21,800
61,700
35,500 -NMRI
38,300 -C57
1,373
5,117
13,478
28,829
46,336
2.6
8.3
21.2
14.2
1.23
2.12
3.96
7.10
12.9
0.3
1.2
3.8
2.3
0.08
0.26
0.70
1.41
3.00
0.3
1.1
3.4
2.8
0.38
1.06
2.42
4.70
4.57
Bronchioalveolar hyperplasia, interstitial
fibrosis in all groups. Severity and
incidence increase with exposure
concentration
No increase in tumors. Noncancer
effects not discussed
No increase in tumors
Noncancer effects not discussed
Inflammatory changes Type II cell
hyperplasia and lung tumors seen at
>0.4 mg/m'; shortening and loss of cilia
in trachea and bronchi
Heinrichetal. (1995)
Research Committee
for HER? Studies
(1988)
-------
-J
o
s\
5
yi
Q
J\
3
0
>
t"4
n
3
•^
-y
1
j
Table 5-6. Histopathological effects of diesel exhaust in the lungs of laboratory animals (continued)
Species/sex
Moust:, NMRI, F
Rat, Wistar, F
Guinea Pig, Hartley, M
Cat, inbred, M
Rat, P344, M
Mouse, CD- 1.M.F
'Light-duty engine.
bHenvy-duty engine.
M to 61 weeks exposure.
Exposure
period
19 h/day
5 days/wesk
120 week;,
19 h/day
5 days/week
140 weeks
20 h/day
5.5 days/week
104 weeks
8 h/day
7 days/week
124 weeks
16 h/day
5 days/week
23 mo
7 h/day
5 days/week
104 weeks
Particles
(mg/m1)
4.24
4.24
0.25'
0.15
}.5
6.0
6.0'
12.0"
2.44
6.33
0.35
3.5
7.1
0.25 urn MDD
CXT
(mgh/m3)
48,336
56,392
2,860
8,580
17,160
68,640
41,664
83,328
19.520
50,640
1,274
12,740
25,844
CO NO, SO,
(ppm) (ppm) (ppm) Effects
12.5 1.5 1.1 Inflammatory changes, bronchiole-
alveolar hyperplasia, alveolar lipo-
proteinosis, fibrosis
12.5 1.5 1.1 Thickened alveolar septa; AM
aggregation; inflammatory changes;
hyperplasia; lung tumors
— — — Minimal response at 0.25 and
— — — ultrastructural changes at 0.75 mg/m';
— — — thickened alveolar membranes; cell
— — — proliferation; fibrosis at 6.0 mg/m1;
increase in PMN at 0.75 mg/m1 and
1.5 mg/m'
20.2 2.7 2.1 Inflammatory changes, AM aggregation,
33.2 4.4 5.0 bronchiolar epithelial metaplasia, Type II
cell hyperplasia, peribronchiolar fibrosis
— — — AM hyperplasia, epithelial hyperplasia,
— — — inflammation, septal fibrosis,
bronchoalveolar metaplasia
3 0.1 — Exposure-related increase in lung soot,
17 0.3 — pigment-laden macrophages, lung
30 0.7 — lesions.
Bronchiolization in alveolar ducts at
7. 1 mg/m'
Study
Heinrich et al.
(1986a)
Heinrich et al.
(1986a)
BamhartetaI.(198I,
1982)
Vostaletal.(1981)
Wallace et al. (1987)
Plopperetal. (1983)
Hyde et al. (1985)
Nikulaetal (1995)
Mauderly et al.
(1996)
d62 to 12'l weeks of exposure.
_< AM = Al veolar macrophage.
fl
jQ
D
-H
3
H
n
PMN = Folymorphonuclear leukocyte.
-------
^J
to
o
o
Table 5-7. Effects of exposure to diesel exhaust on the pulmonary defense mechanisms of laboratory animals
Species/sex
Exposure
period
Particles
(mg/m5)
CxT
(mg-h/ni1)
CO
(ppm)
NO,
(ppm)
SO,
(ppm)
Effects
Study
oo
-j
a
o
z
o
H
o
m
o
&
O
C
O
Guinea Pig,
Hartley
Rat, F344, M
Rat, F344, M
20 h/day
5.5
days/week
8 weeks
7 h/day
5 days/week
104 weeks
20 h/day
5.5
days/week
26,48, or
52 weeks
0.25
1.5
0.19 urn MOD
2.0
0.23-0.36 |im
MOD
0.25°
0.75"
1.5"
0.19 urn MOD
220
1,320
7,280
Alveolar macrophage status
2.9 —
7.5 —
No significant changes in absolute numbers
ofAMs
11.5
715-8,580
2.9
4.8
7.5
1.5 0.81 Little effect on viability, cell number,
oxygen consumption, membrane integrity,
lyzomal enzyme activity, or protein content
of AMs; decreased cell volume and ruffling
of cell membrane and depressed
luminescence of AM
— — AM cell counts proportional to
— — concentration of DPM at 0.75 and
— — 1.5 mg/m1; AM increased in lungs in
response to rate of DPM mass entering lung
rather than total DPM burden in lung;
increased PMNs were proportional to
inhaled concentrations and/or duration of
exposure; PMNs affiliated with clusters of
aggregated AM rather than DPM
Chen et. al. (1980)
Castranovaetal. (1985)
Strom (1984)
Vostaletal. (1982)
Rat F344/CH,
M, F
Mouse, CD, M, F
Rat, Wistar, F
Rat, F344/CM, M
7 h/day
5 days/week
104 weeks
(rat),
78 weeks
(mouse)
18 h/day
5 days/week
24 mo
7 h/day
5 days/week
24 mo
0.35
3.5
7.0
0.25 urn MOD
0.8
2.5
7.1
3.49
1,274C
I2,740C
25,480C
7,400
21,800
61,700
12,704
2.9
16.5
29.7
2.6
8.3
21.2
9.8
0.05
0.34
0.68
0.3
11
3.4
1.2
— Significant increases of AM in rats and
— mice exposed to 7.0 mg/m1 DPM for 24
— and 1 8 mo, respectively, but not at
concentrations of 3.5 or 0.35 mg/m1 DPM
for the same exposure durations; PMNs
increased in a dose-dependent fashion in
both rats and mice exposed to 3.5 or
7.0 mg/m1 DPM and were greater in mice
than in rats
— Changes in differential cell counts in lung
— lavage
—
— Significantly reduced AM in lavage at 24
mo
Henderson etal.( 1988)
Heinrichetal.(1995)
Mauderly et al. (1990a)
-------
^J
sj
-/»
-^
o
J\
1
»
X
3
tf
t>
n
3
z
3
~i
—^
n
D
JO
— i
D
T|
Table 5-7. Effects of exposure
(continued)
Exposure
Species/set period
Rat, M, F 7 h/day
5 days/week
1 2 weeks
Rat, Wistiir, F 18 h/day
5 days/week
24 mo
Rat, F344, M, 7 h/day
developing 0-6 5 days/week
mo 6 mo
adult 6-(2 mo
Rat F34«, M, F 7 h/day
5 days/week
18 weeks
Rat,F34't,M 7 h/day
5 days/week
26-104
weeks
Rat, Sprague- 4-6 h/day
Dawley, M 7 days/week
O.I 1014.3
weeks
Particles
(mg/m3)
0.2
1.0
4.5
0.25um MOD
0.8
2.5
7.1
3.55
0.15
0.94
4.1
<0.5 um MOD
2.0
0.23-0.36 nm
MOD
0.9
8.0
17.0
to diesel exhaust on
CxT CO
(mgb/m') (ppm)
84 —
420 —
1,890 —
7,400 2.6
21,800 8.3
61.700 21.2
3,321 7.9
94.5 —
592 —
2,583 —
1.820-7,280 11.5
2.5-10,210 —
the pulmonary defense mechanisms of laboratory animals
NO, SO,
(ppm) (ppm) Effects
Clearance
— — Evidence of apparent speeding of tracheal
— — clearance at the 4.5 mg/m' level after 1
— — week of TC macroaggregated-albumin
and reduced clearance of tracer aerosol in
each of the three exposure levels at 12
weeks; indication of a lower percentage of
ciliated cells at the 1 .0 and 4.5 mg/m'
levels
0.3 0.3 Significant increase in clearance half-time
1.2 1.1 of inhaled labeled aerosols in all groups at
3.8 3.4 3-18 mo
9.5 Clearance of 2 um, aluminosilicate
particles. Half-time significantly increased
in adult, not different in developing rats
— — Lung burdens of DPM were concentration-
— — related; clearance half-time of DPM almost
— — double in 4. 1 mg/m1 group compared to
0. 1 5 mg/m1 group
1.5 0.8 No difference in clearance of "Fe,O<
particles 1 day after tracer aerosol
administration; 120 days after exposure
tracer aerosol clearance was enhanced; lung
burden of DPM increased significantly
between 12 and 24 mo of exposure
5.0 0.2 Impairment of tracheal mucociliary
2.7 0.6 clearance in a concentration-response
8.0 1.0 manner
Study
Wolff and Gray (1980)
Heinrich et al. (1995)
Mauderly et al. (19876)
Griffisetal. (1983)
Lewis eta). (1989)
Battigelli et al. (1966)
-------
Ul
o
o
oo
o
z
o
H
O
O
c
o
H
m
Table 5-7. Effects of exposure to diesel exhaust on the pulmonary defense mechanisms of laboratory animals
(continued)
Species/sex
Rat, F344,
M, F
Rat, F344/CM, M
Exposure
period
7h/day
5 days/week
130 weeks
7h/day
5 days/week
24 mo
Particles
(mg/m1)
0.35
3.5
7.1
0.25 urn MOD
3.49
CxT CO
(mg-h/m1) (ppm)
1,593 2.9
15,925 16.5
31,850 29.7
12,704 9.8
NO, SO,
(ppm) (ppra)
0.1 —
0.3 —
0.7 —
1.2 —
Effects
No changes in trachea! mucociliary
clearance after 6, 12, 18, 24, or 30 mo of
exposure; increases in lung clearance half-
times as early as 6 mo at 7.0 mg/m1 level
and 18 mo at 3.5 mg/m3 level; no changes
seen at 0.35 mg/m3 level; after 24 mo of
diesel exposure, long-term clearance
half-times were increased in the 3.5 and
7.0 mg/mj groups
Doubling of long-term clearance half-time
for clearance of 1.0 um aluminosilicate
particles. Less effect on clearance in
animals with experimentally induced
emphysema
Study
Wolff et al. (1987)
MauderlyetaJ. (1990a)
Mlcrobial-induced mortality
Mice CD-I, F
7h/day
5 days/week
4, 12, or
26 weeks
2.0
0.23-0.36 nm
MOD
280-1,820 11.5
1.5 0.8
Mortality similar at each exposure duration
when challenged with Ao/PR/8/34
influenza virus; in mice exposed for 3 and
6 mo, but not 1 mo, there were increases in
Hahon etal. (1985)
the percentages of mice having lung
consolidation, higher virus growth,
depressed interferon levels, and a fourfold
reduction in hemagglutinin antibody levels
Mice,CR/CD-l,F 8 h/day 5.3 to 7.9
7 days/week
2 h up to
46 weeks
11-20,350 19
to
22
1.8
to
3.6
0.9
to
2.8
Enhanced susceptibility to lethal effects of
S. pyogenes infections at all exposure
durations (2 and 6 h; 8, 15, 16, 307, and
321 days); inconclusive results with
S. typhimurium because of high mortality
rates in controls; no enhanced mortality
when challenged with A/PR8-3 influenza
virus
Campbell etal. (1980,
1981)
'Chronic exposure lasted 52 weeks.
bChronic exposure lasted 48 weeks.
'Calculated for 104-week exposure.
DPM = Diesel participate matter.
AM = Alveolar macrophage.
PMN = Polymorphonuclear leukocyte.
-------
o
Ui
o
o
Table 5-8. Effects of inhalation of diesel exhaust on the immune system of laboratory animals
D
O
2
O
H
n
HH
a
o
c
c
H
Speci*s/se»
Guinea Pig,
Hartley, M
Rat, l;344, M
Rat, F344;
Mouse, CD-I .
Mouse,
BAI.B/C.M
Exposure
period
20 h/day
5.5 days/week
4 or 8 weeks
7 h/day
5 days/week
52 or 104 weeks
7 h/day
5 days/week
104 weeks
12 h/day,
7 days/week,
Particles
(mg/mj)
1.5
0.19 nm
MOD
2.0
0.23-0.36 nm
MOD
0.35
3.5
7.1
0.25 urn
MOD
3.0
6.0
CxT
(mg-h/m')
660 or 7,280
3, 640 or
7,280
1,274
12,740
25,480
756
1,512
CO
(ppm)
7.5
11.5
2.9
16.5
29.7
—
—
NO,
(ppm)
1.5
0.05
0.34
0.68
2.8
4.1
SO,
(ppm) Effects
— No alterations in numbers of B, T, and null
lymphocytes or cell viability among lymphocytes
isolated from tracheobronchial lymph nodes, spleen,
or blood
0.8 Neither humoral immunity (assessed by enumerating
antibody-producing cells) nor cellular immunity
(assessed by the lymphocyte blast transformation
assay) were markedly affected
— Total number of anti-sheep red blood cell IgM AFC
— in the lung-associated lymph nodes was elevated in
— rats exposed to 3.5 or 7.0 mg/m' DPM (no such
effects in mice); total number of AFC per 106
lymphoid cells in lung-associated lymph nodes and
level of specific IgM, IgO, or IgA in rat sera were not
altered
1.7 Spleen weights in mice exposed to diesel exhaust
2.7 (6 mg/m1) increased significantly. Serum anti-OA IgE
Study
Dziedzic
(1981)
Mentnech et
al. (1984)
Bice et al.
(1985)
Fujimaki et
al. (1997)
Mouse,
C3H/Heu, M
3 weeks
Mice administered OA
intrunasally before,
immediately after, and
3 weeks after exposure
0.4 urn
antibody tilers in mice exposed to 6 mg/m'
significantly higher than control. Antigen-stimulated
IL-4 and IL-10 production increased while IFN-g
production decreased significantly in spleen cells
from diesel exhaust-exposed (6 mg/m') mice
stimulated with OA in vitro. Diesel exhaust
inhalation may affect antigen-specific IgE antibody
production through alteration of the cytokine
network.
12 h/day,
for 12 weeks, flefore
exposure mice injected IP
with OA. After 3 weeks
and every 3 weeks
thereafter, mice
challenged with OA
aerosol.
1.0
3.0
1,008
3,024
1.42
4.02
0.87
1.83
Diesel exhaust + antigen challenge induced airway
hyperresponsiveness and inflammation with increased
eosinophils, mast cells, and goblet cells.
Diesel exhaust alone induced airway
hyperresponsiveness, but not eosinophilic infiltration
or increased goblet cells. Diesel exhaust inhalation
enhanced airway hyperresponsiveness and airway
inflammation caused by OA sensitization.
Miyabara et
al. (I998a)
-------
-o
U)
o
o
Table 5-8. Effects of inhalation of diesel exhaust on the immune system of laboratory animals (continued)
Species/sex
Mouse,
C3H/HeN,
M
Exposure Particles
period (mg/rn1)
12 h/day, 3.0
for 5 weeks. After 7 days
mice injected IP with OA.
At end of exposure mice
challenged with OA
aerosol for IS minutes.
C*T CO NO, SO,
(mg-h/ra1) (ppm) (ppm) (ppm) Effects
1,260 — 4.08 1.26 Diesel exhaust alone increased neutrophils and
macrophages in BAL fluid; after diesel exhaust + OA
challenge eosinophils increased.
OA alone increased eosinophils but the increase was
enhanced by diesel exhaust.
Diesel exhaust + OA, but not diesel exhaust alone,
increased goblet cells, respiratory resistance,
production of OA-specific IgE and Igl in the serum,
and overexpression of IL-5 in lung tissue.
Study
Miyabara et
al.(!998b)
Mouse,
ICR
(murine model
of allergic
asthma)
12 h/day, 7days/week,
40 weeks.
After 16 weeks sensitized
to OA and challenged
with OA aerosol for
6 min, at 3-week intervals
during the last 24 weeks
of exposure.
0.3
1.0
3.0
1,008
3,360
10,080
Diesel exhaust exposure enhanced allergen-related
recruitment to the submucosal layers of the airways
and the bronchoafveolar space, and increased GM-
CSF and IL-5 in the lung in a dose-dependent
manner. Increases in eosinophil recruitment and local
cytosine expression accompanied by goblet cell
proliferation in the bronchial epithelium and airway
hyperresponsiveness to inhaled acetylcholine. Mice
exposed to clean air or DE without allergen
provocation showed no eosinophil recruitment to the
submucosal layers of the airways nor to the
bronchoalveolar space, and few goblet cells in the
bronchial epithelium. Daily inhalation of DE may
enhance allergen-related respiratory diseases such as
allergic asthma, and effect may be mediated by the
enhanced local expression of IL-5 and GM-CSF.
Takano et al.
(1998a)
DPM = Diesel paniculate matter.
AFC = Antibody-forming cells.
-------
to
Ui
o
o
Table 5-9. Effects of diesel participate matter on the immune response of laboratory animals
Model
Treatment
Effects
Reference
WD
rO
o
t-H
H
w
O
C
Mous:,
BDFI F
Mous:,
ICR, w/w, VI
Mouse,
A/J.Iv
Mouse,
BDF,, M
Mouse,
BALB/C,
nu/nu, 7
Mouse,
BALB/:A, F
Mouse,
ICR.M
Intratracheal instill it on of DPM, once/week
for 16 weeks
Mice immunized intrinasally with Der f II +
pyrene. or Der f II H DPM 1 times at 2-week
intervals
Mice were administered 25 :ng of each of
5 Tine particles (Kaiito loam dust, fly ash, CB,
DPM, and aluminun hydroxide [alum])
inlranasally and exposed to aerosolized
Japanese cedar pollen allergens (JCPA) for
intervals up to 18 w't
Inoculated OA with DPM or CB into hind
footpad measured response rsing popliteal
lymph node assay
Intranasal administration of DPM. Mice
immunized with OA or OA combined with
DPM or CB
Intratracheal instillation of OA, DPM, or OVA
and DPM combined, once/week for 6 wk
OA-O"albumin.
DPM - Diesel paniculate matter.
CB-Carbonolack.
Suzuki etal. (1996)
Maejimaetal.
(1997)
Intranasally delivered doses of DPM as low as 1 mg exerted an adjuvant activity for IgE antibody Takafuji et al.
production. (1987)
Infiltration of inflammatory cells, proliferation of goblet cells, increased mucus secretion, respiratory Sagai et al. (1996)
resistance, and airway constriction. Increased eosinophils in the submucosa of the proximal bronchi
and medium bronchioles. Eosinophil infiltration suppressed by pretreatment with PEG-SOD. Bound
sialic acid, an index of mucus secretion, in bronchial alveolar lavage fluids increased, but was
suppressed by PEQ-SOD. Increased respiratory resistance suppressed by PEG-SOD. Oxygen radicals
produced by instilled DPM may cause features characteristic of bronchial asthma in mice.
IgE antibody responses to Der f II enhanced in mice immunized with Der f 11+ pyrene or Der f H +
DPM compared with Der f II alone. Response was dose related. DPM and pyrene contained in DPM
have adjuvant activity on IgE and IgGl antibody production in mice immunized with house dust mite
allergen.
Measurements were made of JCPA-specific IgE and IgG antibody tilers, the protein-adsorbing capacity
of each type of particle, and nasal rubbing movements (a parameter of allergic rhinitis in mice). The
increases in anti-JPCA IgE and IgG antibody liters were significantly greater in mice treated with
particles and aerosolized JCPA than in mice treated with aerosolized JCPA alone. In a subsequent
experiment, the mice received the particles as before, but about 160,000 grains of Japanese cedar pollen
(JCP) were dropped onto the tip of the nose of each mouse twice a week for 16 wk. After 18 wk there
were no significant differences in the anti-JCPA IgE and IgG production, nasal rubbing, or
histopathological changes. The workers concluded that the nature of the particle, the ability of the
particle to absorb antigens, and/or particle size is not related to the enhancement of IgE antibody
production or symptoms of allergic rhinitis. However, IgE antibody production did appear to occur
earlier in mice treated with particles than in mice immunized with allergens alone.
Increased response (increased weight, cell numbers, cell proliferation) and longer response observed Luvik et al. (1997)
with DPM and OA, compared to DPM or OA alone. Response was specific and not an unspecific
inflammatory response. CB was slightly less potent than DPM. Nonextractable carbon core
contributes, substantially to adjuvant activity of DPM.
Increased response to antigen in animals receiving DPM or CB. Increased number of responding Nilsen et al. (1997)
animals and increased serum ami OA IgE antibody. Both DPM and CB have adjuvant activity for IgE
production. DPM response more pronounced than CB, indicating both organic matter adsorbed to
DPM and the nonextractable carbon core responsible for adjuvant activity.
Respiratory resistance (Ris) measured 24 h after the final instillation. Rrs after acetylcholine challenge Takano et al.
was significantly greater in the mice treated with OVA and DPM than other treatments. DPM can (1998b)
enhance airway responsiveness associated with allergen exposure.
PEG-SOD - Polyethyleneglycol-conjugated superoxide dismutase.
IL-4 - Interleukin-4.
IL-S - Interleukin-5.
IL-10 - lnterleukin-10.
1FN - Interferon-g.
GM-CSF -Granulocyte-colony stimulating factor.
IP - Intraperitoneally
-------
^ Table 5-10. Effects of exposure to diesel exhaust
^ Exposure
O Species/sex period
Rat, F344, M, F 7 h/day
5 days/week
52 weeks
Hamster, Syrian 7-8 h/day
5 days/week
22 weeks
Cat, inbred, M 8 li/day
7 days/week
124 weeks
Particles
(mg/m1)
2.0
0.23-0.36 urn
MOD
4.0
8.0
11.0
6.0'
12.0b
CxT
(mg-h/rn')
3,640
3,080-9,680
41,664
83,328
CO
(ppm)
12.7
12.0
19.0
25.0
20.2
33.3
on the liver of laboratory animals
NO,
(ppm)
1.6
0.5
1.0
1.5
2.7
4.4
SO,
(ppm)
0.83
3.0
6.0
7.0
2.1
5.0
Effects
No changes in absolute liver weight or
liver/body weight ratio
Enlarged sinusoids, with activated Kupffer's
cells and slight changes of nuclei; fatty
deposits; mitochondria, loss of cristae and
pleomorphic character; gap junctions between
hepatocytes had wide range in structural
diversity
No change in the absolute liver weight
Study
Green etal. (1983)
Meiss etal. (1981)
Plopper etal. (1983)
OJ
O
O
Z
3
o
H
M
M to 61 weeks of exposure.
b62 to 124 weeks of exposure.
o
H
m
-------
Table 5-11. Effects of exposure to diesel exhaust on the hematological and cardiovascular systems of laboratory animals
N^
J\
^.
3
J\
b
t^
3
«
>
T)
H
-i
M/
D
z!
D
H
"1
-H
T)
^
?0
-H
3
Speci :s/se>
Monk:y,
Cynoinolgus, M
Rat, F 144, t/(, F
Guinea Pig,
Hartley, M, F
Hamster, Syian,
M, F
Rat, F344;
Guinea I'ig,
Hartley
Rat, Wistar, M
Rat, F3«.44/Jcl,
M, F
Rat, F344
Cat, lnbr:d, M
Exposure
period
7 h/day
5 days/week
104 weeks
7 h/ilay
5 days/week
104 weeks
20 h/day
7 days/week
8 weeks
7-8 h/day
5 days/week
75 weeks
20 h/day
5.5 days/week
78 weeks
6 h/day
5 days/week
78 weeks
1 eh/cay
6 days/week
130 weeks
16 h/day
5 days/week
104 weeks
8 h/day
7 days/week
124 weeks
'Nonirrat iated diesel exhaust.
Particles
(mg/m1)
2
C.23-0.36 Mm MOD
2
0.23-0.36 Mm MOD
6.3'
6.8"
3.9
0.1 Mm MOD
0.25
0.75
1.5
0.19 Mm MOD
8.3
0.71 Mm MOD
O.llc
0.41C
1.08C
2.31°
3.72"
0.; Mm MOD
0.7
2.2
6.6
6.0'
I2.0f
CxT
(mg-h/m1)
7,280
7,280
7,056
7,616
10,238-11,700
2,145
6,435
12,870
19,422
1,373
5,117
13,478
28,829
46,426
5,824
18,304
54,912
41.664
83,328
CO
(ppm)
11.5
11.5
17.4
16.7
18.5
3.0
4.8
6.9
50.0
1.23
2.12
3.96
7.10
12.9
—
—
32.0
20.2
33.3
NO,
(ppm)
1.5
1.5
2.3
2.9
1.2
0.11
0.27
0.49
4-6
0.08
0.26
0.70
1.41
3.00
—
—
—
2.7
4.4
SO,
(ppm)
0.8
0.8
2.1
1.9
3.1
—
—
—
—
0.38
1.06
2.42
4.70
4.57
—
—
—
2.1
5.0
Effects
Increased MCV
Increase in banded neutrophils; no effect on
heart or pulmonary arteries
No effect on heart mass or ECG; small
decrease in heart rate (IE only)
At 29 weeks, lower erythrocyte count;
increased MCV; reduced leukocyte count
No changes in heart mass or hematology at
any exhaust level or duration of exposure in
either species
3% increase in COHb
At higher concentrations, RBC, Hb, Hct
slightly elevated; MCV and mean
corpuscular hemoglobin and concentration
were lowered
Increases in RBC. Hb, Hct, and WBC,
primarily banded neutrophils; suggestion of
an increase in prothrombin time; increased
heart/body weight and right
ventricular/heart ratios and decreased left
ventricular contractility in 6.6 mg/m1 group
Increases in banded neutiophils; significant
at 12 mo, but not 24 mo
Study
Lewis eta). (1989)
Lewis eta). (1989)
Vallyathan et al.
(1986)
Wiester eta). (1980)
Heinrich et al.
(1982)
Penney etal. (1981)
Karagianes et al.
(1981)
Research Committee
for HERP Studies
(1988)
Brightwell et al.
(1986)
Pepelko and Peirano
(1983)
dHeavy-duty engine.
el tn £1 u/pplfc nf pvnrtcnrp
'Light-duty engine.
Key: MC'V = Mean corpuscular volume.
'62 to 124 weeks of exposure.
-------
-J
to
o
o
Table 5-12. Effects of chronic exposures to diesel exhaust on serum chemistry of laboratory animals
Ui
O
o
2
o
H
o
Species/sex
Rat, F344, M, F
Hamster, Syrian, M, F
Rat, F344/JcL, M, F
Rat, F344; Hamster,
Syrian
Cat inbred, M
Exposure
period
7 h/day
5 days/week
104 weeks
7-8 h/day
5 days/week
75 weeks
16 h/day
6 days/week
130 weeks
16 h/day
5 days/week
104 weeks
8 h/day
7 days/week
124 weeks
Particles
(mg/m5)
2.0
0.23
0.36 Mm MDD
3.9
0.1 urn MDD
O.ll1
0.41'
1.08'
2.311
3.72"
O.I 9-0.28 urn
MDD
0.7
2.2
6.6
6.0'
12.0"
CxT
(mg-b/m1)
7,280
10,238-11,700
1,373
5,117
13,478
28,829
46,426
5,824
18,304
54,912
41,664
83,328
CO
(ppm)
115
18.5
1.23
2.12
3.96
7.10
12.9
—
—
32.0
20.2
33.3
NO,
(ppm)
1.5
1.2
0.08
0.26
3.96
7.10
3.00
—
—
—
2.7
4.4
SO,
(ppm)
0.8
3.1
0.38
1.06
2.42
4.70
4.57
—
—
—
2.1
5.0
Effects
Decreased phosphate, LDH, SCOT, and SGPT;
increased sodium in females but not males
After 29 weeks, increases in SGOT, LDH, alkaline
phosphatase, gamma-glutamyl transferase, and BUN
Lower cholinesterase activity in males in both the
light-and heavy-duty series and elevated gamma
globulin and electrolyte levels in males and females
in both series
Rats, 6.6 mg/m1, reduction in blood glucose, blood
proteins, triglycerides, and cholesterol; increase in
BUN, alkaline phosphate alamine, and aspartate
aminotransferases (SGPT and SGOT); hamsters, 6.6
. mg/m', decrease in potassium, LDH, aspartate amino-
transferase; increase in albumin and gamma-glutamyl
Iransferase
BUN unaltered; SGOT and SGPT unaffected; LHD
increase after 1 year of exposure
Study
Lewis et al.
(1989)
Heinrich et al.
(1982)
Research
Committee for
HERP Studies
(1988)
Brightwell et al.
(1986)
Pepelko and
Peirano (1983)
"Light-duty engine.
"Heavy-duty engine.
cl to 61 weeks of exposure.
d62 to 124 weeks of exposure.
Key: LDH = Lactate dehydrogenase.
SGOT = Serum glutamic-oxaloacetic transaminase.
BUN = Blood urea nitrogen.
SGPT = Serum glutamic-pyruvic transaminase.
o
c
o
H
m
-------
\JD
O
o
>Z
O
H
O
^H
H
cn
o
c
o
H
W
Species/sex
Rat, F344, M
Mouse, CD-I, F
Rat, Spiague-
Dawley M
Rat, F344, M
Rat, F344, F
Rat, F344, M
Mouse, /v/J, IV
Exposure
period
7h/day
5 days/week
4 weeks
20 h/day
7 days/week
1-7 weeks
20 h/day
5.5 days/week
4, 13, 26, or
39 weeks
20 h/day
5.5 days/week
4, 13, 26, or
39 weeks
7 h/day
5 days/week
12, 26, or
104 weeks
20 h/day
5.5 days/week
8-53 weeks
8 days/week
7 days/week
26 or 35 weeks
Particles C x t CO
(mg/m1) (mg-h/nt1) (PPro)
2.0 280 11.5
0.2-0.36 nm mdd
6.3 882-6,174 17.4
0.75 330-6,435 4.8
1.5 7.5
O.I 9 urn mdd
0.75 330-6,435 4.8
1.5 7.5
O.I 9 urn mdd
2.0 840-7,280 11.5
0.23-0.36 urn mdd
0.25 220-8,745 2.9
1.5 7.5
0.1 9 urn mdd
6.0 17.4 17.4
NO, SO,
(ppm) (ppm) Effects
— — Intratracheal administration of DPM extract required
doses greater than 6 mg/m' before the lung AHH was
barely doubled; liver AHH activity was unchanged
1.5 0.8 Mice inoculated intranasally with influenza virus had
smaller increases in ethylmorphine demethylase
activity on days 2 to 4 postvirus infection and abolition
of day 4 postinfection increase in NADPH-dependent
cytochrome c reductase
2.3 2. 1 AHH induction occurred in lung, liver, and prostate
gland but not in testes; maximum significant activities
occurred at different times; liver has greatest overall
activity, percent increase highest in prostate; expoxide
hydrase activity was unaffected
— — Inhalation exposure had no significant effect on liver
— — AHH activity; lung AHH activity was slightly reduced
after 6-mo exposure to 1.5 mg/m1 DPM; an ip dose of
dp extract, estimated to be equivalent to inhalation
— — exposure, had no effect on AHH activity in liver and
— — lungs; cyt. P-50 was unchanged in lungs and liver
following inhalation or ip administration
1.5 0.8 No effect on B[a]p hydrolase or 7-exthoxycoumarin
deethylase activities in the liver
— — After 8 weeks, no induction of cyt. P-450, cyt. P-448,
— — or NADPH-dependent cyt. c reductase; after 1 year of
exposure, liver microsomal oxidation of B[o]p was not
increased; 1 year of exposure to either 0.25 or
1.5 mg/m1 DPM impaired lung microsomal metabolism
ofB[a]p
2.3 2. 1 No differences in lung and liver AHH activities and
liver P-448, P-450 levels
Study
Chen (1986)
Rabovskyetal. (1986)
Lee etal. (1980)
Chen and Vostal (1981)
Rabovsky et al. (1984)
Navarro etal. (1981)
Pepelko and Peirano
(1983)
AHH =: aryl hydrocarbon hydroxylase.
B[a]p =: bemo[a]pyrene.
-------
7/25/00
Table
Species/sex
Rat, Sprague-
Dawley, M
Rat, Spraguc
Dawley, F
Rat, Spragtie-
Dawley, F
5-14. Effects
Exposure
period
8h/day
7 days/week
1 -4 weeks
20 h/day
7 days week
6 weeks
8 or 20 h/day
7 days/week
3, 4, 6, or
16 weeks
of chronic
Particles
(mg/m3)
6
6
6
exposures to diesel
CxT
(tngh/rn5)
336-1,344
5,040
1,008-13,440
exhaust
CO
(ppm)
19
19
19
on behavior
NO,
(ppm)
2.5
2.5
2.5
and
so,
(ppm)
1.8
1.8
1.8
neurophysiology
Effects
Somatosensory and visual evoked
potentials revealed longer pulse
latencies in pups exposed neonatally
Reduction in adult SLA and in
neonatal pivoting
Reduction in SLA in adults; neonatal
• exposures for 20 or 8 h/day caused
reductions in SLA. Neonatal
exposures for 20 h/day for 17 days
resulted in a slower rate of a
bar-pressing task to obtain food
Study
Laurie and Boyes
(1980, 1981)
Laurie etal. (1978)
Laurie et al. (1980)
I
O
z:
o
H
O
i— t
H
trt
SLA = Spontaneous locomotor activity.
O
H
IT)
-------
7/25/00
Ul
1
*o
oo
j>
I*
O
0
2
O
H
n
>— t
a
o
*
XD
C
H
Table 5-1S. Effects of chronic exposures to diesel exhaust on reproduction and development in laboratory animals
Spec;es/se:i
Mouse,
[C5VBL]'
6XC3HJF,, M
Rat, Sprague-
Dawley, ?
Rabt it, New
Zealiind Albino,
F
Monkey,
Cyncmoljsus, M
Mouse,
A/Stnng, M
MOUSJ, CD-I,
M, F
Exposure Particles C * T CO
period (nig/m]) (rng-h/m9) (ppm)
5 days 50, 100, or — —
200 mg/kg
in com oil;
i.p. injection
8h/day 6 571 20
7 days/wee!<
1 .7 weeks
8 a/day 6 638 20
7 days/weak
1.9 weeks
7h/day 2 7,280 11.5
5 days/week
104 weeks
8h/day 6 10,416- 20
7 days/week 12,768
31 or
38 weeks
8h/day 12 4,032-18,816 33
7 days/week
6 to 28
weeks
NO, SO,
(ppm) (ppm) Effects
— — Dose-related increase in
sperm abnormalities;
decrease in sperm number at
highest dose; testicular
weights unaffected
2.7 2.1 No signs of maternal toxicity
or decreased fertility; no
skeletal or visceral
teratogenic effects in 20-day-
old fetuses
2.7 2. 1 No adverse effects on
maternal weight gain or
fertility; no skeletal or
visceral teratogenic effects in
the fetuses
1 .5 0.8 No effects on sperm motility,
velocity, density,
morphology, or incidence of
abnormalities
2.7 2. 1 No effect on sperm
morphology; high rate of
spontaneous sperm
abnormalities may have
masked small effects
4.4 5.0 Overall fertility and survival
rates were unaffected in the
three-generation reproductive
study; only consistent change
noted, an increase in lung
weights, was diagnosed as
anthracosis
Study
Quinto and De
Marinis(1984)
Werchowski et al.
(1980a)
Pepelko and Peirano
(1983)
Werchowski et al.
(1980a)
Pepelko and Peirano
(1983)
Lewis etal. (1989)
Pereiraetal. (1981)
Pepelko and Peirano
(1983)
-------
K)
O
O
Table 5-16. Composition of exposure atmospheres in studies comparing unfiltered and filtered diesel exhaust8
Ul
H
a
O
2
O
H
O
H-H
H
tfl
O
Species/sex
Rat, Wislar, F;
Hamster, Syrian
Rat, F344, F
Rat, F344, M, F;
Hamster, Syrian,
M, F
Rat, Wistar, F;
Hamster, Syrian, F;
Mouse NMR1, F
Mouse, NMRI.F,
C57BL/6N, F
Exposure'
period
7h/day
5 days/week
104 weeks
8h/day
7 days/week
104 weeks
16h/day
5 days/week
1 04 weeks
19 li/day
5 days/week
120to
140 weeks
18h/day
5 days/week
23 mo
(NMRI)
24 mo
(C57BL/6N)
Uf
F
C
Uf
P
C
Uf
Uf
Uf
Fd
C
Uf
Fd
C
Uf
F
C
Particles
(mg/m3)
3.9
—
—
4.9
—
—
0.7
2.2
6.6
—
—
4.24
—
—
4.5
0.01
0.01
C* t
(nig-h/m1)
14,196
28,538
5,824
18,304
54,912
48,336
56,392
40,365
CO NO,
(ppm) (ppm)
18.5 1.2
18.0 1.0
— —
7.0 1.8
— —
— —
— —
— —
32.0 —
32.0 —
1.0 —
12.5 1.5
11.1 1.2
0.16 —
14.2 2.3
14.2 2.9
0.2 0.01
SO,
(ppm) Effects
3.1 No effect on pulmonary function or heart rate in rats; increases in
2.8 pulmonary adenomatous proliferations in hamsters, UF
— significantly higher than F or C
13.1 Body weight decrease after 6 mo in UF, 1 8 mo in f; lung/body
— rate weight rate higher in both groups at 24 mo; at 2 years,
— fibrosis and epithelial hyperplasia in lungs of uf; nominal lung
and spleen histologic changes
— Uf: elevated red and white cell counts, hematocrit and hemo-
— globin; increased heart/body weight and right ventricular/heart
— weight ratios; lower left ventricular contractility; changes in blood
— chemistry; obstructive and restrictive lung disease; F: no effects
—
3. 1 Uf: decreased body wt in rats and mice but not hamsters; increas-
1 .02 ed mortality, mice only; decreased lung compliance and increased
— airway resistance, rats and hamsters; species differences in lung
lavage enzymes and cell counts and lung histopathology and
collagen content, most pronounced in rats; F: no effect on
glucose-6-phosphate dehydrogenase, total protein, and lung
collagen
2.8 Uf: increased lung wet weight starting at 3 mo
2.4
0.1 F: no noncancer effects reported
Study
Heinrich et at.
(1982)
Iwaietal. (1986)
Brightwell et al.
(1986)
Heinrich et al.
(1986a)
Heinrich et al.
(1995)
'Man values.
bUF= unfiltered whole exhaust, F = filtered exhaust, C = control.
'Reported to have the same component concentrations as the unfiltered, except particles were present in undetectable amounts.
'Concentrations reported for high concentration level only.
-------
1 5.7. REFERENCES
2
3 Ames, RG; Attfield, MD; Hankinson, JL; et al. (1982) Acute respiratory effects of exposure to diesel emissions in
4 coal miners. Am Rev Respir Dis 125:39-42.
5
6 Ames, RG; Reger, RB; Hall, DS. (1984) Chronic respiratory effects of exposure to diesel emissions in coal mines.
7 Arch Environ Health 39:389-394.
8
9 Attfield, MD. (1978) The effect of exposure to silica and diesel exhaust in underground metal and nonmetal miners.
10 In: Industrial hygiene for mining and tunneling: proceedings of a topical symposium; November; Denver, CO.
11 Kelley, WD, ed. Cincinnati, OH: The American Conference of Governmental Industrial Hygienists, Inc.; pp.
12 129-135.
13
14 Attfield, MD; Trabant, GD; Wheeler, RW. (1982) Exposure to diesel fumes and dust at six potash mines. Ann Occup
15 Hyg 26:817-831.
16
17 Barnhart, MI; Chen, S-T; Salley, SO; et al. (1981) Ultrastructure and morphometry of the alveolar lung of guinea
18 pigs chronically exposed to diesel engine exhaust: six month's experience. J Appl Toxicol 1:88-103.
19
20 Barnhart, MI; Salley, SO; Chen, S-T; et al. (1982) Morphometric ultrastructural analysis of alveolar lungs of guinea
21 pigs chronically exposed by inhalation to diesel exhaust (DE). In: Toxicological effects of emissions from diesel
22 engines: proceedings of the Environmental Protection Agency diesel emissions symposium; October, 1981; Raleigh,
23 NC. Lewtas, J., ed. New York: Elsevier Biomedical; pp. 183-200. (Developments in toxicology and environmental
24 science: v. 10).
25
26 Battigelli, MC. (1965) Effects of diesel exhaust. Arch Environ Health 10:165-167.
27
28 Battigelli, MC; Mannella, RJ; Hatch, TF. (1964) Environmental and clinical investigation of workmen exposed to
29 diesel exhaust in railroad engine houses. Ind Med Surg 33:121-124.
30
31 Battigelli, MC; Hengstenberg, F; Mannela, RJ; et al. (1966) Mucociliary activity. Arch Environ Health 12:460-466.
32
33 Bayram, H; Devalia, JL; Sapsford, RJ; et al. (1998a) The effect of diesel exhaust particles on cell function and
34 release of inflammatory mediators from human bronchial epithelial cells in vitro. Am J Respir Cell Mol Biol
35 18:441-448.
36
37 Bayram, H; Devalia, JL; Khair, O; et al. (1998b) Comparison of ciliary activity and inflammatory mediator release
38 from bronchial epithelial cells of nonatopic nonasthmatic subjects and atopic asthmatic patients and the effect of
39 diesel exhaust particles in vitro. J Allergy Clin Immunol 102(5): 771-782.
40
41 Be"ruBe, KA; Jones, TP; Williamson, BJ; et al. (1999) Physicochemical characterisation of diesel exhaust particles:
42 factors for assessing biological activity. Atmos Environ 33:1599-1614.
43
44 Bhatnagar, RS; Hussain, MZ; Sorensen, KR; et ai. (1980) Biochemical alterations in lung connective tissue in rats
45 and mice exposed to diesel emissions. In: Health effects of diesel engine emissions: proceedings of an international
46 symposium, v. i; i>remher 1979; Cincinnati, OH. Pepslkc, V/E; Darker, RM; Clarke, NA, cds. Cincinnati,
47 OH: U.S. Environmental Protection Agency, Health Effects Research Laboratory; pp. 557-570; EPA report no.
48 EPA/600/9-80/057a. Available from: NTIS, Springfield, VA; PBS 1-173809.
49
50 Bice, DE; Mauderly, JL; Jones, RK; et al. (1985) Effects of inhaled diesel exhaust on immune responses after lung
51 immunization. Fundam Appl Toxicol 5:1075-1086.
52
53 Blomberg, A; SaLnsbury. C: Rudell. B; et al. (1998) Nacal cavity lining fluid ascorbic acid concentration increases in
54 healthy human volunteers following short term exposure to diesel exhaust. Free Radical Res 28:59-67.
7/25/00 5-100 DRAFT—DO NOT CITE OR QUOTE
-------
1 Boland, S; Baeza-Squiban, A; Fournier, T; et al. (1999) Diesel exhaust panicles are taken up by human airway
2 epithelial cells in vitro and alter cytokine production. Am J Physiol 276:L604-L613.
4 Bolton, RE; Vincent, JH; Jones, AD; et al. (1983) An overload hypothesis for pulmonary clearance of UICC amosite
5 fibres inhaled by rats. Br J Ind Med 40:264-272. :,
6
7 Boren, HG. (1964) Carbon as a carrier mechanism for irritant gases. Arch Environ Health 8:119-124.
8
9 Brightwell, J; Fouillet, X; Cassano-Zoppi, AL; et al. (1986) Neoplastic and functional changes in rodents after
10 chronic inhalation of engine exhaust emissions. In: Carcinogenic and mutagenic effects of diesel engine exhaust:
11 proceedings of the international satellite symposium on toxicological effects of emissions from diesel engines; July;
12 Tsukuba Science City, Japan. Ishinishi, N; Koizumi, A; McClellan, RO; et al., eds. Amsterdam, Holland: Elsevier
13 Science Publishers B. V.; pp. 471-485. (Developments in toxicology and environmental science: v. 13).
14
15 Brunekreef, B; Janssen, NA; de Hartog, J.; et al. (1997) Air pollution from truck traffic and lung function in children
16 living near motorways. Epidemiology 8(3):298-303.
17
18 Brunekreef, B; Janssen, NA; van Vliet, P; et al. (2000) Traffic related air pollution and its effect on respiratory
19 health of children living near motorways. Presented at: PM 2000: Paniculate matter and health-the scientific basis
20 for regulatory decision-making. Sponsored by the Air and Waste Management Association , January 24-28, 2000.
21
22 Campbell, KI; George, EL; Washington, IS, Jr. (1980) Enhanced susceptibility to infection in mice after exposure to
23 dilute exhaust from light duty diesel engines. In: Health effects of diesel engine emissions: proceedings of an
24 international symposium, v. 2; December 1979; Cincinnati, OH. Pepelko, WE; Danner, RM; Clarke, NA, eds.
25 Cincinnati, OH: U.S. Environmental Protection Agency, Health Effects Research Laboratory; pp. 772-785; EPA
26 report no. EPA/600/9-80/057b. Available from: NTIS, Springfield, VA; PB81-173817.
27
k8 Campbell, KI; George, EL; Washington, IS, Jr. (1981) Enhanced susceptibility to infection in mice after exposure to
dilute exhaust from light duty diesel engines. Environ Int 5:377-382.
30
31 Castranova, V; Bowman, L; Reasor, MJ; et al. (1985) The response of rat alveolar macrophages to chronic inhalation
32 of coal dust and/or diesel exhaust. Environ Res 36:405-419.
33
34 Chan, TL; Lee, PS; Hering, WE. (1981) Deposition and clearance of inhaled diesel exhaust panicles in the
35 respiratory tract of Fischer rats. J Appl Toxicol 1:77-82.
36
37 Chaudhari, A; Dutta, S. (1982) Alterations in tissue glutathione and angiotensin converting enzyme due to inhalation
38 of diesel engine exhaust. J Toxicol Environ Health 9:327-337.
39
40 Chaudhari, A; Fairer, RG; Dutta, S. (1980) Effect of exposure to diesel exhaust on pulmonary prostaglandin
41 dehydrogenase (PGDH) activity. In: Health effects of diesel engine emissions: proceedings of an international
42 symposium, v. 1; December 1979; Cincinnati, OH. Pepelko, WE; Danner, RM; Clarke, NA, eds. Cincinnati, OH:
43 U.S. Environmental Protection Agency, Health Effects Research Laboratory; pp. 532-537; EPA report no.
44 EPA/600/9-80-057a. Available from: NTIS, Springfield, VA; PB81-173809.
45
46 Chaudhari, A; Fairer, RG; Dutta, S. (1981) Effect of exposure to diesel exhaust on pulmonary prostaglandin
47 dehydrogenase (PGDH) activity. J Appl Toxicol 1:132-134.
48
49 Chen, KC. (1986) Induction of aryl hydrocarbon hydroxylase in rat tissue following intratracheal instillation of diesel
50 paniculate extract and benzo[a]pyrene. J Appl Toxicol 6:259-262.
51
52 Chen, KC; Vostal, JJ. (1981) Aryl hydrocarbon hydroxylase activity induced by injected diesel paniculate extract vs
5 3 inhalation of diluted diesel exhaust. J Appl Toxicol 1:127-131.
Chen, S; Weller, MA; Bamhart, MI. (1980) Effects of diesel engine exhaust on pulmonary alveolar macrophages.
56 Scanning Electron Microsc 3:327-338.
7/25/00 5-101 DRAFT—DO NOT CITE OR QUOTE
-------
1 Devalia, JL; Bayram, H; Abdelaziz, MM; et al. (1999) Differences between cytokine release from bronchial
2 epithelial cells of asthmatic patients and non-asthmatic subjects: effect of exposure to diesel exhaust particles. Int
3 Arch Allergy Immunol 118:437-439.
4
5 Diaz-Sanchez, D. (1997) The role of diesel exhaust particles and their associated polyaromatic hydrocarbons in the
6 induction of allergic airway disease. Allergy (Copenhagen) 52(suppl. 38):52-56.
7
8 Diaz-Sanchez, D; Dotson, AR; Takenaka, H; et al. (1994) Diesel exhaust particles induce local IgE production in
9 vivo and alter the pattern of IgE messenger RNA isofonns. J Clin Invest 94(4): 1417-1425.
10
11 Diaz-Sanchez, D; Tsien, A; Casillas, A; et al. (1996) Enhanced nasal cytokine production in human beings after in
12 vivo challenge with diesel exhaust particles. J Allergy Clin Immunol 98:114-123.
13
14 Diaz-Sanchez, D; Tsien, A; Fleming, J; et al. (1997) Combined diesel exhaust particulate and ragweed allergen
15 challenge markedly enhances human in vivo nasal ragweed-specific IgE and skews cytokine production to a T helper
16 cell 2-type pattern. J Immunol 158:2406-2413.
17
18 Diaz-Sanchez, D; Garcia, MP; Wang, M; et al. (1999) Nasal challenge with diesel exhaust particles can induce
19 sensitization to a neoallergen in the human mucosa. J Allergy Clin Immunol 104(6): 1183-1188.
20
21 Driscoll, KE; Maurer, JK. (1991) Cytokine and growth factor release by alveolar macrophages: potential biomarkers
22 of pulmonary toxicity. Toxicol Pathol 19:398-405.
23
24 Dungworth, DL; Mohr, U; Heinrich, U; et al. (1994) Pathologic effects of inhaled particles in rat lungs: associations
25 between inflammatory and neoplastic processes. In: Toxic and carcinogenic effects of solid particles in the
26 respiratory tract: [proceedings of the 4th international inhalation symposium]; March 1993; Hannover, Germany.
27 Mohr, U; Dungworth, DL; Mauderly, JL, et al., eds. Washington, DC: International Life Sciences Institute Press; pp.
28 75-98.
29
30 Dziedzic, D. (1981) Differential counts of B and T lymphocytes in the lymph nodes, circulating blood and spleen
31 after inhalation of high concentrations of diesel exhaust. J Appl Toxicol 1:111-115.
32
33 Edling, C; Axelson, O. (1984) Risk factors of coronary heart disease among personnel in a bus company. Int Arch
34 Occup Environ Health 54:181 -183.
35
36 Edling, C; Anjou, CG; Axelson, O; et al. (1987) Mortality among personnel exposed to diesel exhaust. Int Arch
37 Occup Environ Health 59:559-565.
38
39 El Batawi, MA; Noweir, MH. (1966) Health problems resulting from prolonged exposure to air pollution in diesel
40 bus garages. Ind Health 4:1-10.
41
42 Fedan, JS; Frazier, DG; Moorman, WJ; et al. (1985) Effects of a two-year inhalation exposure of rats to coal dust
43 and/or diesel exhaust on tension responses of isolated airway smooth muscle. Am Rev Respir Dis 131:651-655.
44
45 Fujimaki, H; Nohara, O; Ichinose, T; et al. (1994) IL-4 production in mediastinal lymph node ceils in mice
46 intratracheally instilled with diesel exhaust particulates and antigen. Toxicolony 92(l-3):261-268.
47
48 Fujimaki, H; Saneyoshi, K; Shiraishi, F; et al. (1997) Inhalation of diesel exhaust enhances antigen-specific IgE
49 antibody production in mice. Toxicology 116:227-233.
50
51 Gamble, JF; Jones, WG. (1983) Respiratory effects of diesel exhaust in salt miners. Am Rev Respir Dis
52 128:389-394.
53
54 Gamble, J; Jones, W; Hudak, J; et al. (1979) Acute changes in pulmonary function in salt miners. Presented at:
55 Industrial hygiene for mining and tunneling: proceedings of a topical symposium; November 1978, Denver, CO.
56 Cincinnati, OH: American Conference of Governmental Industrial Hygienists, Inc.; pp. 119-128.
7/25/00 5-102 DRAFT—DO NOT CITE OR. QUOTE
-------
Gamble, J; Jones, W; Hudak, J. (1983) An epidemiological study of salt miners in diesel and nondiesel mines. Am J
Ind Med 4:435-458.
4 Gamble, J; Jones, W; Minshall, S. (1987a) Epidemiological-environmental study of diesel bus garage workers: acute
5 effects of NO2 and respirable paniculate on the respiratory system. Environ Res 42:201-214.
6
7 Gamble, J; Jones, W; Minshall, S. (1987b) Epidemiological-environmental study of diesel bus garage workers:
8 chronic effects of diesel exhaust on the respiratory system. Environ Res 44:6-17.
9
10 Green, FHY; Boyd, RL; Danner-Rabovsky, J; et al. (1983) Inhalation studies of diesel exhaust and coal dust in rats.
11 Scand J Work Environ Health 9:181-188.
12
13 Griffis, LC; Wolff, RK; Henderson, RF; et al. (1983) Clearance of diesel soot particles from rat lung after a
14 subchronic diesel exhaust exposure. Fundam Appl Toxicol 3:99-103.
15
16 Gross, KB. (1981) Pulmonary function testing of animals chronically exposed to diluted diesel exhaust. J Appl
17 Toxicol 1:116-123.
18
19 Hahon, N; Booth, JA; Green, F; et al. (1985) Influenza virus infection in mice after exposure to coal dust and diesel
20 engine emissions. Environ Res 37:44-60.
21
22 Hare, CT; Springer, KJ. (1971) Public response to diesel engine exhaust odors [final report]. San Antonio, TX:
23 Southwest Research Institute; report no. AR-804. Available from: NTIS, Springfield, VA; PB-204012.
24
25 Hare, CT; Springer, KJ; Somers, JH; et al. (1974) Public opinion of diesei odor. In: Automotive engineering
26 congress; February-March; Detroit, MI. New York: Society of Automotive Engineers; SAE technical paper no.
27 740214.
Hashimoto, S; Gon, Y; Takeshita, I; et al. (2000) Diesel exhaust particles activate p38 MAP kinase to produce
30 interleukin 8 and RANTES by human bronchial epithelial cells and N-acetylcysteine attenuates p38 MAP kinase
31 activation [In Process Citation]. Am J Respir Crit Care Med 161(l):280-285.
32
33 Hatch, GE; Boykin, E; Graham, JA; et al. (1985) Inhalable particles and pulmonary host defense: in vivo and in vitro
34 effects of ambient air and combustion particles. Environ Res 36:67-80.
35
36 Health Effects Institute. (1995) Diesel exhaust: a critical analysis of emissions, exposure, and health effects: a special
37 report of the Institute's Diesel Working Group. Cambridge, MA: Health Effects Institute.
38
39 Heinrich, U; Peters, L; Funcke, W; et al. (1982) Investigation of toxic and carcinogenic effects of diesel exhaust in
40 long-term inhalation exposure of rodents. In:Toxicological effects of emissions from diesel engines:
41 proceedings of the Environmental Protection Agency diesel emissions symposium; October 1981; Raleigh, NC.
42 Lewtas, J., ed. New York: Elsevier Biomedical; pp. 225-242. (Developments in toxicology and environmental
43 science: v. 10).
44
45 Heinrich, U; Muhle, H; Takenaka, S; et al. (1986a) Chronic effects on the respiratory tract of hamsters, mice, and
46 rats after long-term inhalation of high concentrations of filtered and unfiltered diesel engine emissions. J Appl
47 Toxicol 6:383-395.
48
49 Heinrich, U; Pott, F; Rittinghausen, S. (1986b) Comparison of chronic inhalation effects in rodents after long-term
50 exposure to either coal oven flue gas mixed with pyrolized pitch or diesel engine exhaust. In: Carcinogenic and
51 mutagenic effects of diesel engine exhaust: proceedings of the international satellite syposium on toxicological
52 effects of emissions from diesel engines; July; Tsukuba Science City, Japan. Ishinishi, N; Koizumi, A; McClellan,
53 RO; et al., eds. Amsterdam, Holland: Elsevier Science Publishers BV; pp. 441-457. (Developments in toxicology
and environmental science: v. 13).
7/25/00 5-103 DRAFT—DO NOT CITE OR QUOTE
-------
1 Heinrich, U; Fuhst, R; Rittinghausen, S; et al. (1995) Chronic inhalation exposure of Wistar rats and two different
2 strains of mice to diesel engine exhaust, carbon black, and titanium dioxide. Inhal Toxicol 7:533-556.
3
4 Henderson, RF; Pickrell, JA; Jones, RK; et al. (1988) Response of rodents to inhaled diluted diesel exhaust:
5 biochemical and cytological changes in bronchoalveolar lavage fluid and in lung tissue. Fundam Appl Toxicol
6 11:546-567.
7
8 Henderson, RF; Leung, HW; Harmsen, AG; et al. (1988) Species differences in release of arachidonate metabolites
9 in response to inhaled diluted diesel exhaust. Toxicol Lett 42(3):325-32.
10
11 Hiruma, K; Terada, N; Hanazawa, T; et al. (1999) Effect of diesel exhaust on guinea pig nasal mucosa. Ann Otol
12 Rhinol Laryngol 108(6):582-588.
13
14 Hyde, DM; Plopper, CG; Weir, AJ; et al. (1985) Peribronchiolar fibrosis in lungs of cats chronically exposed to
15 diesel exhaust. Lab Invest 52:195-206.
16
17 Ikeda, M; Suzuki, M; Watarai, K; et al. (1995) Impairment of endothelium-dependent relaxation by diesel exhaust
18 particles in rat thoracic aorta. Jpn J Pharmacol 68:183-189.
19
20 Ikeda, M; Watarai, K; Suzuki, M; et al. (1998) Mechanism of pathophysiological effects of diesel exhaust particles
21 on endothelial cells. Environ Toxicol Pharmacol 6:117-123.
22
23 International Life Sciences Institute (ILSI). (2000) ILSI Risk Science Institute workshop: The relevance of the rat
24 lung response to particle overload for human risk assessment. Gardner, DE, ed. Inhal Toxicol: 12(1-2).
25
26 Iwai, K; Udagawa, T; Yamagishi, M; et al. (1986) Long-term inhalation studies of diesel exhaust on F344 SPF rats.
27 Incidence of lung cancer and lymphoma. In: Carcinogenic and mutagenic effects of diesel engine exhaust:
28 proceedings of the international satellite symposium on toxicological effects of emissions from diesel engines; July;
29 Tsukuba Science City, Japan. Ishinishi, N; Koizumi, A; McClellan, RO; et al., eds. Amsterdam, Holland: Elsevier
30 Science Publishers B. V.; pp. 349-360. (Developments in toxicology and environmental science: v. 13).
31
32 Jorgensen, H; Svensson, A. (1970) Studies on pulmonary function and respiratory tract symptoms of workers in an
33 iron ore mine where diesel trucks are used underground. J Occup Med 12:348-354.
34
35 Kahn, G; Orris, P; Weeks, J. (1988) Acute overexposure to diesel exhaust: report of 13 cases. Am J Ind Med
36 13:405-406.
37
38 Kaplan, HL; MacKenzie, WF; Springer, KJ; et al. (1982) A subchronic study of the effects of exposure of three
39 species of rodents to diesel exhaust. In: Toxicological effects of emissions from diesel engines: proceedings of the
40 Environmental Protection Agency diesel emission symposium; October, 1981; Raleigh, NC. Lewtas, J., ed. New
41 York: Elsevier Biomedical; pp. 161-182. (Developments in toxicology and environmental science: v. 10).
42
43 Karagianes, MT; Palmer, RF; Busch, RH. (1981) Effects of inhaled diesel emissions and coal dust in rats. Am Ind
44 HygAssocJ 42:382-391.
45
46 Katz, M; Rennie, RP; Jegier, Z. (1960) Air pollution and associated health hazards from diese! locomotive traffic in
47 a railroad tunnel. Occup Health Rev 11:2-15.
48
49 Kilbum, KH; McKenzie, WN. (1978) Leukocyte recruitment to airways by aldehyde-carbon combinations that
50 mimic cigarette smoke. Lab Invest 38:134-142.
51
52 Klosterkotter, W; Bunemann, G. (1961) Animal experiments on the elimination of inhaled dust. In: Inhaled particles
53 and vapours: proceedings of an international symposium; March-April 1960; Oxford, United Kingdom. Davies, C.
54 N., ed. New York: Pergamon Press; pp. 327-341.
7/25/00 5-104 DRAFT—DO NOT CITE OR QUOTE
-------
Knox, KB; Suphioglu, C; Taylor, P; et al. (1997) Major grass pollen allergen Lol p 1 binds to diesel exhaust
particles: implications for asthma and air pollution. Clin Exp Allergy 27:246-251.
4 Kobayashi, T; Ito, T. (1995) Diesel exhaust particulates induce nasal mucosal hyperresponsiveness to inhaled
5 histamine aerosol. Fundam Appl Toxicol 27:195-202.
6
7 Kobayashi, T; Dceue, T; Ito, T; et al. (1997) Short-term exposure to diesel exhaust induces nasal mucosal
8 hyperresponsiveness to histamine in guinea pigs. Fundam Appl Toxicol 3 8:166-172.
9
10 Laurie, RD; Boyes, WK. (1980) Neurophysiological alterations due to diesel exhaust exposure during the neonatal
11 life of the rat. In: Health effects of diesel engine emissions: proceedings of an international symposium, v. 2;
12 December 1979; Cincinnati, OH. Pepelko, WE; Banner, KM; Clarke, NA, eds. Cincinnati, OH: U.S. Environmental
13 Protection Agency, Health Effects Research Laboratory; pp. 713-727; EPA report no. EPA-600/9-80/057b.
14 Available from: NTIS, Springfield, VA; PBS 1-173817.
15
16 Laurie, RD; Boyes, WK. (1981) Neurophysiological alterations due to diesel exhaust exposure during the neonatal
17 life of the rat. Environ Int 5:363-368.
18
19 Laurie, RD; Lewkowski, JP; Cooper, GP; et al. (1978) Effects of diesel exhaust on behavior of the rat. Presented at:
20 71st annual meeting of the Air Pollution Control Asscociation; June; Houston, TX. Pittsburgh, PA: Air Pollution
21 Control Association.
22
23 Laurie, RD; Boyes, WK; Wessendarp, T. (1980) Behavioral alterations due to diesel exhaust exposure. In: Health
24 effects of diesel engine emissions: proceedings of an international symposium, v. 2; December 1979; Cincinnati, OH.
25 Pepelko, WE; Danner, KM; Clarke, NA, eds. Cincinnati, OH: U.S. Environmental Protection Agency, Health Effects
26 Research Laboratory; pp. 698-712; EPA report no. EPA-600/9-80-057b. Available from: NTIS, Springfield, VA;
27 PB81-173817.
Lee, IP; Suzuki, K; Lee, SD; et al. (1980) Aryl hydrocarbon hydroxylase induction in rat lung, liver, and male
30 reproductive organs following inhalation exposure to diesel emission. Toxicol Appl Pharmacol 52:181 -184.
31
32 Lee, KP; Trochimowicz, HJ; Reinhardt, CF. (1985) Pulmonary response of rats exposed to titanium dioxide (TiOJ
• 33 by inhalation for two years. Toxicol Appl Pharmacol 79:179-192.
34
35 Lee, KP; Kelly, DP; Schneider, PW; et al. (1986) Inhalation toxicity study on rats exposed to titanium tetrachloride
36 atmospheric hydrolysis products for two years. Toxicol Appl Pharmacol 83:30-45.
37
38 Lee, KP; Kelly, DP; O'Neal, FO; et al. (1988) Lung response to ultrafine Kevlar aramid synthetic fibrils following
39 2-year inhalation exposure in rats. Fundam Appl Toxicol 11:1-20.
40
41 Levins, PL. (1981) Review of diesel odor and toxic vapor emissions. Washington, DC: U.S. Department of
42 Transportation, National Highway Traffic Safety Administration; report no. DOT-TSC-NHTSA-81-9.
43
44 Lewis, TR; Green, FHY; Moorman, WJ; et al. (1989) A chronic inhalation toxicity study of diesel engine emissions
45 and coal dust, alone and combined. J Am Coll Toxicol 8:345-375.
46
47 Linnell, RH; Scott, WE. (1962) Diesel exhaust composition and odor studies. J Air Pollut Control Assoc
48 12:510-515,545.
49
50 L0vik, M; H0gseth, AK; Gaarder, PI; et al. (1997) Diesel exhaust particles and carbon black have adjuvant activity
51 on the local lymph node response and systemic IgE production to ovalbumin. Toxicology 121:165-178.
52
53 Madden, MC; Richards, JE; Dailey, LA; et al. (2000) Effect of ozone on diesel exhaust particle toxicity. Toxicol
~ Appl Pharmacol (submitted)
7/25/00 5-105 DRAFT—DO NOT CITE OR QUOTE
-------
1 Maejima, K; Tamura, K; Taniguchi, Y; et al. (1997) Comparison of the effects of various fine particles on IgE
2 antibody production in mice inhaling Japanese cedar pollen allergens. J Toxicol Environ Health 52:231-248.
3
4 Mauderly, JL. (1994) Noncancer pulmonary effects of chronic inhalation exposure of animals to solid particles. In:
5 Toxic and carcinogenic effects of solid particles in the respiratory tractrjproceedings of the 4th international
6 inhalation symposium]; March 1993; Hannover, Germany. Mohr, U; Dungworth, DL; Mauderly, JL; et al., eds.
7 Washington, DC: International Life Sciences Institute Press; pp. 43-55.
8
9 Mauderly, JL; Benson, JM; Bice, DE; et al. (1981) Observations on rodents exposed for 19 weeks to diluted diesel
10 exhaust In: Inhalation Toxicology Research Institute annual report 1980-1981. Albuquerque, NM: Lovelace
11 Biomedical and Environmental Research Institute; pp. 305-311; report no. LMF-91.
12
13 Mauderly, JL; Benson, JM; Rice, DE; et al. (1984) Life span study of rodents inhaling diesel exhaust: effects on
14 body weight and survival. In: Inhalation Toxicology Research Institute annual report. Guilmette, RA; Medinsky,
15 MA, eds. Albuquerque, NM: Inhalation Toxicology Research Institute; pp. 287-291; report no. LMF-113. Available
16 from: NTIS, Springfield, VA; DE85-009283.
17
18 Mauderly, JL; Bice, DE; Carpenter, RL; et al. (1987a) Effects of inhaled nitrogen dioxide and diesel exhaust on
19 developing lung. Cambridge, MA: Health Effects Institute; research report no. 8.
20
21 Mauderly, JL; Jones, RK; Griffith, WC; et al. (1987b) Diesel exhaust is a pulmonary carcinogen in rats exposed
22 chronically by inhalation. Fundam Appl Toxicol 9:208-221.
23
24 Mauderly, JL; Gillett, NA; Henderson, RF; et al. (1988) Relationships of lung structural and functional changes to
25 accumulation of diesel exhaust particles. In: Inhaled particles VI: proceedings of an international symposium and
26 workshop on lung dosimetry; September 1985; Cambridge, United Kingdom. Dodgson, J; McCallum, RI; Bailey,
27 MR; et al., eds. Ann Occup Hyg 32(suppl. l):659-669.
28
29 Mauderly, JL; Bice, DE; Cheng, YS; et al. (1990a) Influence of preexisting pulmonary emphysema on susceptibility
30 of rats to diesel exhaust. Am Rev Respir Dis 141:1333-1341.
31
32 Mauderly, JL; Cheng, YS; Snipes, MB. (1990b) Particle overload in toxicology studies: friend or foe? In:
33 Proceedings of the particle lung interaction symposium; May; Rochester, NY. J Aerosol Med 3(suppl. 1):
34 S-169-S-187.
35
36 Mauderly, JL; Banas, DA; Griffith, WC; et al. (1996) Diesel exhaust is not a pulmonary carcinogen in CD-I mice
37 exposed under conditions carcinogenic to F344 rats. Fundam Appl Toxicol 30:233-242.
38
39 McClellan, RO; Bice, DE; Cuddihy, RG; et al. (1986) Health effects of diesel exhaust. In: Aerosols: research, risk
40 assessment and control strategies: proceedings of the second U.S.-Dutch international symposium; May 1985;
41 Williamsburg, VA. Lee, SD; Schneider, T; Grant, LD; et al., eds. Chelsea, MI: Lewis Publishers, Inc.; pp. 597-615.
42
43 Meiss, R; Robenek, H; Schubert, M; et al. (1981) Ultrastructural alterations in the livers of golden hamsters
44 following experimental chronic inhalation of diluted diesel exhaust emission. Int Arch Occup Environ Health
45 48:147-157.
46
47 Mentnech, MS; Lewis, DM; Olenchock, SA; et al. (1984) Effects of coal dust and diesel exhaust on immune
48 competence in rats. J Toxicol Environ Health 13:31-41.
49
50 Misiorowski, RL; Strom, KA; Vosta!, JJ; et al. (1980) Lung biochemistry of rats chronically exposed to diesel
51 particulates. In: Health effects of diesel engine emissions: proceedings of an international symposium, v. 1;
52 December 1979. Cincinnati, OH. Pepelko, WE; Danner, RM; Clarke, NA, eds. Cincinnati, OH: U.S. Environmental
53 Protection Agency, Health Effects Research Laboratory; pp. 465-480; EPA report no. EPA/600/9-80/057a.
54 Available from: NTIS, Springfield, VA; PB81-173809.
7/25/00 5-106 DRAFT—DO NOT CITE OR QUOTE
-------
Miyabara, Y; Ichinose, T; Takano, H; et al. (1998a) Diesel exhaust inhalation enhances airway hyperresponsiveness
in mice. Int Arch Allergy Immunol 116:124-131.
Miyabara, Y; Takano, H; Ichinose, T; et al. (1998b) Diesel exhaust enhances allergic airway inflammation and
5 hyperresponsiveness in mice. Am J Resp Crit Care Med 157:1138-1144.
6
7 Moorman, WJ; Clark, JC; Pepelko, WE; et al. (1985) Pulmonary fuction responses in cats following long-term
8 exposure to diesel exhaust J Appl Toxicol 5:301-305.
9
10 Morrow, PE. (1988) Possible mechanisms to explain dust overloading of the lungs. Fundam Appl Toxicol
11 10:369-384.
12
13 Muranaka, M; Suzuki, S; Koizumi, K; et al. (1986) Adjuvant activity of diesel-exhaust particulates for the
14 production of IgE antibody in mice. J Allergy Clin Immunol 77(4):616-23.
15
16 Murphy, SA; BeiuB6, KA; Pooley, FD; et al. (1998) The response of lung epithelium to well characterised fine
17 particles. Life Sci 62:1789-1799.
18
19 Murphy, SA; et al (1999) Bioreactivity of carbon black and diesel exhaust particles to primary Clara and type II
20 epithelial cell cultures. Occup Environ Med 56:813-819.
21
22 Navarre, C; Charboneau, J; McCauley, R. (1981) The effect of in vivo exposure to diesel exhaust on rat hepatic and
23 pulmonary microsomal activities. J Appl Toxicol 1:124-126.
24
25 Nikula, KJ; Snipes, MB; Barr, EB; et al. (1995) Comparative pulmonary toxicities and carcinogenicities of
26 chronically inhaled diesel exhaust and carbon black in F344 rats. Fundam Appl Toxicol 25:80-94.
27
f Nikula, KJ; Avila, KJ; Griffith, WC; et al. (1997) Sites of particle retention and lung tissue responses to chronically
inhaled diesel exhaust and coal dust in rats and cynomolgus monkeys. In: Proceedings of the sixth international
meeting on the toxicology of natural and man-made fibrous and non-fibrous particles; September 1996; Lake Placid,
31 NY. Driscoll, KE; Oberdorster, G, eds. Environ Health Perspect Suppl 105(5): 1231 -1234.
32
33 Nilsen, A; Hagemann, R; Eide, I. (1997) The adjunct activity of diesel exhaust particles and carbon black on
34 systemic IgE production to ovalbumin in mice after intranasal instillation. Toxicology 124:225-232.
35
36 Oberdorster, G. (1994) Extrapolation of results from animal inhalation studies with particles to humans? In: Toxic
37 and carcinogenic effects of solid particles in the respiratory tract: [proceedings of the 4th international inhalation
38 symposium]; March 1993; Hannover, Germany. Mohr, U; Dungworth, DL; Mauderly, JL; et al., eds. Washington,
39 DC: International Life Sciences Institute Press; pp. 335-353.
40
41 Oberdorster, G; Ferin, J; Gelein, R; et al. (1992) Role of the alveolar macrophage in lung injury: studies with
42 ultrafine particles. Environ Health Perspect 97:193-199.
43
44 Odor Panel of the CRC-APRAC Program Group.on Composition of Diesel Exhaust. (1979) Development and
45 evaluation of a method for measurement of diesel exhaust odor using a diesel odor analysis system (DOAS). New
46 York: Coordinating Research Council, Inc., Air Pollution Research Advisory Committee; CRC-APRAC project no.
47 CAPI-1-64.
48
49 Ohtoshi, T; Takizawa, H; Okazaki, H; et al. (1998) Diesel exhaust particles stimulate human airway epithelial cells
50 to produce cytokines relevant to airway inflammation in vitro. J Allergy Clin Immunol 101:778-785.
51
52 Ormstad, H; Johansen, BV; Gaarder, PI. (1998) Airborne house dust particles and diesel exhaust particles as allergen
53 carriers. Clin Exp Allergy 28:702-708.
Pattle, RE; Stretch, H; Burgess, F; et al. (1957) The toxicity of fumes from a diesel engine under four different
56 running conditions. Br J Ind Med 14:47-55.
7/25/00 5-107 DRAFT—DO NOT CITE OR QUOTE
-------
1 Penney, DG; Baylerian, MS; Fanning, KE; et al. (1981) A study of heart and blood of rodents inhaling diesel engine
2 exhaust particulates. Environ Res 26:453-462.
3
4 Pepelko, WE. (1982a) Effects of 28 days exposure to diesel engine emissions in rats. Environ Res 27:16-23.
5
6 Pepelko, WE. (1982b) EPA studies on the lexicological effects of inhaled diesel engine emissions. In: lexicological
7 effects of emissions from diesel engines: proceedings of the Environmental Protection Agency diesel emissions
8 symposium; October 1981; Raleigh, NC. Lewtas, J, ed. New York: Elsevier Biomedical; pp. 121-142.
9 (Developments in toxicology and environmental science: v. 10).
10
1 1 Pepelko, WE; Peirano, WB. (1983) Health effects of exposure to diesel engine emissions: a summary of animal
1 2 studies conducted by the U.S. Environmental Protection Agency's Health Effects Research Laboratories at
1 3 Cincinnati, Ohio. J Am Coll Toxicol 2:253-306.
14
1 5 Pepelko, WE; Mattox, JK; Yang, YY; et al. (1980a) Pulmonary function and pathology in cats exposed 28 days to
1 6 diesel exhaust J Environ Pathol Toxicol 4:449-458.
17
1 8 Pepelko, WE; Mattox, J; Moorman, WJ; et al. (1980b) Pulmonary function evaluation of cats after one year of
1 9 exposure to diesel exhaust. In: Health effects of diesel engine emissions: proceedings of an international symposium,
20 v. 2; December 1979; Cincinnati, OH. Pepelko, WE; Danner, RM; Clarke, NA, eds. Cincinnati, OH: U.S.
21 Environmental Protection Agency, Health Effects Research Laboratory; pp. 757-765; EPA report no.
22 EPA/600/9-80-057b. Available from: NTIS, Springfield, VA; PB81-173817.
23
24 Pepelko, WE; Mattox, J; Moorman, WJ; et al. (1981) Pulmonary function evaluation of cats after one year of
25 exposure to diesel exhaust. Environ Int 5:373-376.
26
27 Pereira, MA; Sabharwal, PS; Gordon, L; et al. (1981) The effect of diesel exhaust on sperm-shape abnormalities in
28 mice. Environ Int 5:459-460.
29
30 Plopper, CG; Hyde, DM; Weir, AJ. (1983) Centriacinar alterations in lungs of cats chronically exposed to diesel
3 1 exhaust. Lab Invest 49:39 1 -399.
32
33 Purdham, JT; Holness, DL; Pilger, CW. (1987) Environmental and medical assessment of stevedores employed in
34 ferry operations. Appl Ind Hyg 2: 133-139.
35
36 Quinto, I; De Marinis, E. (1984) Sperm abnormalities in mice exposed to diesel paniculate. Mutat Res 130:242.
37
38 Rabovsky, J; Petersen, MR; Lewis, TR; et al. (1984) Chronic inhalation of diesel exhaust and coal dust: effect of age
39 and exposure on selected enzyme activities associated with microsomal cytochrome P-450 in rat lung and liver. J
40 Toxicol Environ Health 14:655-666.
41
42 Rabovsky, J; Judy, DJ; Kodak, DJ; et al. (1986) Influenza virus-induced alterations of cytochrome P-450 enzyme
43 activities following exposure of mice to coal and diesel particulates. Environ Res 40: 136-144.
44
45 Reger, R. (1979) Ventilatcry function changes over a work shift for coal miners exposed to diesel emissions. In:
46 Tcxicc logical and carcinogenic health hazards in the workplace: proceedings of the first annual NIOSK scientific
47 symposium; 1978; Cincinnati, OH. Bridbord, K; French, J, eds. Park Forest South, IL: Pathotox Publishers, Inc.; pp.
48 346-347.
49
50 Reger, R; Hancock, J; Hankinson, J; et al. (1982) Coal miners exposed to diesel exhaust emissions. Ann Occup Hyg
51 26:799-815.
52
53 Research Committee for HERP Studies. (1988) Diesel exhaust and health risks: results of the HERP studies.
54 Tsukuba, Ibaraki, Japan: Japan Automobile Research Institute, Inc.
7/25/00 5-108 DRAFT— DO NOT CITE OR QUOTE
-------
Rudell, B; Sandstrdm, T; Stjernberg, N; et al. (1990) Controlled diesel exhaust exposure in an exposure chamber
pulmonary effects investigated with bronchoalveolar lavage. J Aerosol Sci 21(suppl. I):s411-S414.
4 Rudell, B; SandstrOm, T; Hammarstrom, U; et al. (1994) Evaluation of an exposure setup for studying effects of
5 diesel exhaust in humans. Int Arch Occup Environ Health 66:77-83.
6
7 Rudell, B; Ledin, MC; Hammarstrom, U; et al. (1996) Effects on symptoms and lung function in humans
8 experimentally exposed to diesel exhaust. Occup Environ Med 53:658-662.
9
10 Sagai, M; Furuyama, A; Ichinose, T. (1996) Biological effects of diesel exhaust particles (DEP). III. Pathogenesis of
11 asthma like symptoms in mice. Free Radic Biol Med 21:199-209.
12
13 Salvi, S; Blomberg, A; Rudell, B; et al. (1999) Acute inflammatory responses in the airways and peripheral blood
14 after short-term exposure to diesel exhaust in healthy human volunteers. Am J Respir Crit Care Med 159:702-709.
15
16 Salvi, SS; Nordenhall, C; Blomberg, A; et al. (2000) Acute exposure to diesel exhaust increases IL-8 and GRO-alpha
17 production in healthy human airways. Am J Respir Crit Care Med 161(2):550-557.
18
19 Saneyoshi, K; Nohara, O; Imai, T; et al. (1997) IL-4 and IL-6 production of bone marrow-derived mast cells is
20 enhanced by treatment with environmental pollutants. Int Arch Allergy Immunol 114(3):237-245.
21
22 Schneider, DR; Felt, BT. (1981) Effect of diesel paniculate exposure on adenylate and guanylate cyclase of rat and
23 guinea pig liver and lung. J Appl Toxicol 1:135-139.
24
25 Schreck, RM; Soderholm, SC; Chan, TL; et al. (1981) Experimental conditions in GMR chronic inhalation studies of
26 diesel exhaust. J Appl Toxicol 1:67-76.
27
Steerenberg, PA; Zonnenberg, JAJ; Dormans, JAMA; et al. (1998) Diesel exhaust particles induced release of
interleukin 6 and 8 by (primed) human bronchial epithelial cells (BEAS 2B) in vitro. Exp Lung Res 24:85-100.
JO
31 Stober, W. (1986) Experimental induction of tumors in hamsters, mice and rats after long-term inhalation of filtered
32 and unfiltered diesel engine exhaust. In: Carcinogenic and mutagenic effects of diesel engine exhaust: proceedings of
33 the international satellite symposium on lexicological effects of emissions from diesel engines; July; Tsukuba
34 Science City, Japan. Amsterdam,: Ishinishi, N; Koizumi, A; McClellan, RO; et al. Elsevier Science Publishers B.
35 V.; pp. 421-439. (Developments in toxicology and environmental science: v. 13).
36
37 Strom, KA. (1984) Response of pulmonary cellular defenses to the inhalation of high concentrations of diesel
38 exhaust. J Toxicol Environ Health 13:919-944.
39
40 Suzuki, T; Kanoh, T; Kanbayashi, M; et al. (1993) The adjuvant activity of pyrene in diesel exhaust on IgE antibody
41 production in mice. Arerugi 42(8):963-8.
42
43 Suphioglu, C; Singh, MB; Taylor, P; et al. (1992) Mechanism of grass-pollen-induced asthma. Lancet
44 339(8793):569-572.
45
46 Suzuki, T; Kanoh, T; Ishimori, M; et al. (1996) Adjuvant activity of diesel exhaust particulates (DEP) in production
47 of anti-IgE and anti-IgGl antibodies to mite allergen in mice. J Clin Lab Immunol 48:187-199.
48
49 Takafuji, S; Suzuki, S; Koizumi, K; et al. (1987) Diesel-exhaust particulates inoculated by the intranasal route have
50 an adjuvant activity for IgE production in mice. J Allergy Clin Immunol 79:639-645.
51
52 Takano, H; Yoshikawa, T; Ichinose, T; et al. (1997) Diesel exhaust particles enhance antigen-induced airway
53 inflammation and local cytokine expression in mice. Am J Respir Crit Care Med 156:36-42.
Takano, H; Ichinose, T; Miyabara, Y; et al. (1998a) Inhalation of diesel exhaust enhances allergen-related eosinophil
56 recruitment and airway hyperresponsiveness in mice. Toxicol Appl Pharmacol 150:328-337.
7/25/00 5-109 DRAFT—DO NOT CITE OR QUOTE
-------
1 Takano, H; Ichinose, T; Miyabara, Y; et al. (1998b) Diesel exhaust particles enhance airway responsiveness
2 following allergen exposure in mice. Immunopharmacol Immunotoxicol 20:329-336.
3
4 Takenaka, H; Zhang, K; Diaz-Sanchez, D; et al. (1995) Enhanced human IgE production results from exposure to the
5 aromatic hydrocarbons from diesel exhaust: direct effects on B-cell IgE production. J Allergy Clin Immunol
6 95:103-115.
7
8 Terada, N; Maesako, K; Hiruma, K; et al. (1997). Diesel exhaust particulates enhance eosinophil adhesion to nasal
9 epithelial cells and cause degranulation. Int Arch Allergy Immunol 114(2): 167-174.
10
11 Terada, N; Hamano, N; Maesako, K; et al. (1999) Diesel exhaust particulates upregulate histamine receptor mRNA
12 and increase histamine-induced IL-8 and GM-CSF production in nasal epithelial cells and endothelial cells. Clin Exp
13 Allergy 29(l):52-59.
14
15 Turk, A. (1967) Selection and training of judges for sensory evaluation of the intensity and character of diesel
16 exhaust odors. Cincinnati, OH: U.S. Department of Health, Education, and Welfare, National Center for Air
17 Pollution Control; Public Health Service publication no. 999-AP-32. Available from: NTIS, Springfield, VA;
18 PB-174707.
19
20 U.S. Environmental Protection Agency (U.S. EPA). (1996) Air quality criteria for paniculate matter. Research
21 Triangle Park, NC: National Center for Environmental Assessment-RTF Office; report nos.
22 EPA/600/P-95/001aF-cF. 3v. Available from: NTIS, Springfield, VA; PB96-168224.
23
24 Ulfvarson, U; Alexandersson, R; Aringer, L; et al. (1987) Effects of exposure to vehicle exhaust on health. Scand J
2 5 Work Environ Health 13:505-512.
26
27 Vallyathan, V; Virmani, R; Rochlani, S; et al. (1986) Effect of diesel emissions and coal dust inhalation on heart and
28 pulmonary arteries of rats. Br J Ind Med 14:47-55.
29
30 van Vliet, P; Knape, M; de Hartog, J; et al. (1997) Motor vehicle exhaust and chronic respiratory symptoms in
31 children living near freeways. Environ Res 74(2):122-132.
32
. 33 Vinegar, A; Carson, Al; Pepelko, WE. (1980) Pulmonary function changes in Chinese hamsters exposed six months
34 to diesel exhaust. In: Health effects of diesel engine emissions: proceedings of an international symposium, v. 2;
35 December, 1979; Cincinnati, OH. Pepelko, W E; Danner, R M; Clarke, N A, eds. Cincinnati, OH: U.S.
36 Environmental Protection Agency, Health Effects Research Laboratory; pp. 749-756; EPA report no.
37 EPA-600/9-80-057b. Available from: NTIS, Springfield, VA; PB81-173817.
38
39 Vinegar, A; Carson, A; Pepelko, WE; et al. (198 la) Effect of six months of exposure to two levels of diesel exhaust
40 on pulmonary function of Chinese hamsters. Fed Proc 40:593.
41
42 Vinegar, A; Carson, A; Pepelko, WE. (1981b) Pulmonary function changes in Chinese hamsters exposed six months
43 to diesel exhaust. Environ Int 5:369-3 71.
44
45 Vostal, JJ; Chan, TL; Garg, BD; et al. (1981) Lymphatic transport of inhaled diesel particles in the lungs of rats and
46 guinea pigs exposed to diluted diesel exhaust. F.nviron Int 5:339-347.
47
48 Vostal, JJ; White, HJ; Strom, KA; et al. (1982) Response of the pulmonary defense system to diesel paniculate
49 exposure. In: Toxicological effects of emissions from diesel engines: proceedings of the Environmental Protection
50 Agency diesel emissions symposium; October 1981; Raleigh, NC. Lewtas, J., ed. New York: Elsevier Biomedicai;
51 pp. 201-221. (Developments in toxicology and environmental science: v. 10).
52
53 WP ILSI. (2000) The relevance of the rat lung response to particle overload for human risk assessment: a workshop
54 consensus report. L-J-.a! Toxicol 12(1/2): 1-17.
55
7/25/00 5-110 DRAFT—DO NOT CITE OR QUOTE
-------
1 Wade, JF, III; Newman, LS. (1993) Diesel asthma: reactive airways disease following overexposure to locomotive
2 exhaust J Occup Med 35:149-154.
Wallace, MA; Salley, SO; Bamhart, MI. (1987) Analysis of the effects of inhaled diesel exhaust on the alveolar
5 intravascular and interstitial cellular components of rodent lungs. Scanning Microsc 1:13 87-1395.
6
7 Watanabe, N; Oonuki, Y. (1999) Inhalation of diesel engine engine exhaust affects spermatogenesis in growing male
8 rats. Environ Health Perspect 107:539-544.
9
10 Wehner, AP; Dagle, GE; Clark, ML; et al. (1986) Lung changes in rats following inhalation exposure to volcanic ash
11 for two years. Environ Res 40:499-517.
12
13 Werchowski, KM; Chaffee, VW; Briggs, GB. (1980a) Teratologic effects of long-term exposure to diesel exhaust
14 emissions (rats). Cincinnati, OH: U.S. Environmental Protection Agency, Health Effects Research Laboratory; EPA
15 report no. EPA-600/1-80-010. Available from: NTIS, Springfield, VA; PB80-159965.
16
17 Werchowski, KM; Henne, SP; Briggs, GB. (1980b) Teratologic effects of long-term exposure to diesel exhaust
18 emissions (rabbits). Cincinnati, OH: U.S. Environmental Protection Agency, Health Effects Research Laboratory;
19 EPA report no. EPA-600/1 -80/011. Available from: NTIS, Springfield, VA; PB80-168529.
20
21 White, HJ; Garg, BD. (1981) Early pulmonary response of the rat lung to inhalation of high concentration of diesel
22 particles. J Appl Toxicol 1:104-110.
23
24 Wiester, MJ; Iltis, R; Moore, W. (1980) Altered function and histology in guinea pigs after inhalation of diesel
25 exhaust. Environ Res 22:285-297.
26
27 Wolff, RK; Gray, RL. (1980) Tracheal clearance of particles, hi: Inhalation Toxicology Research Institute annual
28 report: 1979-1980. Diel, JH; Bice, DE; Martinez, BS, eds. Albuquerque, NM: Lovelace Biomedical and
Environmental Research Institute; p. 252; report no. LMF-84.
31 Wolff, RK; Henderson, RF; Snipes, MB; et al. (1987) Alterations in particle accumulation and clearance in lungs of
32 rats chronically exposed to diesel exhaust. Fundam Appl Toxicol 9:154-166.
33
34 Yang, HM; Ma, JY; Castranova, V; et al. (1997) Effects of diesel exhaust particles on the release of interleukin-1
35 and tumor necrosis factor-alpha from rat alveolar macrophages. Exp Lung Res 23:269-284.
7/25/00 5-111 DRAFT—DO NOT CITE OR QUOTE
-------
6. QUANTITATIVE APPROACHES TO ESTIMATING HUMAN
NONCANCER HEALTH RISKS OF DIESEL EXHAUST
1 6.1. INTRODUCTION
2 As discussed earlier in this document (Chapter 2, Section 2.2.2), diesel exhaust (DE)
3 consists of a complex mixture of gaseous pollutants and particles. In attempting to estimate
4 potential health risks associated with human exposure to DE, researchers have focused attention
5 mostly on the paniculate matter (PM) components, based, in part, on comparisons of relative
6 toxicity of unfiltered versus filtered DE (with gaseous components removed), as discussed in
7 Chapter 5.
8 Diesel particulate matter (DPM) consists mainly of: (a) elemental carbon (EC) particles;
9 (b) soluble organic carbon, including 5-ring or higher polycyclic aromatic hydrocarbons (PAHs)
10 such as benzo-a-pyrene (BaP), and other 3- or 4-ring compounds distributed between gas and
11 particle phases; and (c) metallic compounds. DPM also typically contains small amounts of
12 sulfate/sulfuric acid and nitrates, trace elements, and water, plus some unidentified components.
13 DPM is almost entirely fine particles <1.0 urn, with many very small ultrafine particles (i.e.,
14 O.lOum).
15 Health concerns have long focused on diesel particles, which have very large surface
T0 areas that allow for adsorption of organics from the diesel combustion process and adsorb
17 additional compounds during transport in ambient air. The small particles and large surface area
18 likely provide an enhanced potential for subcellular interactions with important cellular
19 components of respiratory tissues once the particles are inhaled by humans or other species. Also
20 of growing health concern in recent years is the potential for enhanced toxic effects of ultrafine
21 particles (compared with particles of the same chemical composition but of larger size).
22 Although many questions remain regarding specific aspects of DPM "aging," these fine and
23 ultrafine particles are viewed as likely important toxic components of the overall mix of
24 combustion-related fine particles typically found in most urban airsheds.
25 One approach for deriving quantitative estimates of potential human health risks
26 associated with ambient (nonoccupational) DE exposures is to treat the DE constituent DPM as a
27 lexicologically important component of ambient fine particle mixes and to assume that
28 quantitative estimates of risk for ambient fine particle exposure effects in general also apply to
29 DPM specifically. Risk estimates or exposure guidance derived for ambient fine particles in
30 general would presumably then represent a plausible upper limit for levels of risk potentially
31 associated with DE measured as DPM (given that the latter is one of numerous constituents of
^fc typical ambient fine particle mixes). The bases for deriving risk estimates for fine particles
33 recently used by EPA in setting new ambient air fine particle (PM2 5) standards are concisely
7/25/00 6-1 DRAFT—DO NOT CITE OR QUOTE
-------
1 summarized and their relationships to ambient air DPM discussed in the next two sections
2 (Sections 6.2 and 6.3).
3 Another approach to evaluating noncancer risks of ambient DE exposures is to combine
4 key elements from evaluation of specific DPM noncancer effects in animals and humans
5 (described in Chapter 5) with use of quantitative dosimetry models (described in Chapter 3), for
6 estimating DPM concentrations to which humans may be exposed throughout their lives (i.e.,
7 chronically) without experiencing any untoward or adverse noncancer effects. This can be
8 accomplished via analysis of dose-response relationships where the adverse response is
9 considered as a function of a corresponding measure of dose. Chapter 5 is replete with dose-
10 response information on adverse (but nonlethal) noncancer health effects observed in long-term
11 (chronic/lifetime) exposure studies to DE in general and to DPM in particular, albeit in animals.
12 Chapter 3 presents methods that convert external exposure concentrations of DPM in animal
13 studies to estimates of a human equivalent concentration (HEC). Later sections (6.4 to 6.6) of
14 this chapter assess and integrate this information to derive a chronic reference concentration
15 (RfC), using a well-established Agency method for developing dose-response assessments of
16 noncancer effects for toxic air pollutants other than those identified below in Section 6.2 as being
17 regulated by National Ambient Air Quality Standards (NAAQS).
18 Estimates of DE levels associated with effects occurring under less than lifetime exposure
19 scenarios (such as acute) are not addressed in this chapter. Acute studies of DE exposure,
20 discussed in Chapter 5, are accompanied by scant dose-response information, with single-
21 exposure studies for various specialized endpoints (e.g., allergenicity/adjuvancy) and other
22 multiple-exposure-level studies reporting only data on mortality. Based on currently available
23 methodologies, these studies do not yet appear to provide a sufficient basis from which to derive
24 a dose-response assessment for an acute DE exposure scenario.
25
26 6.2. DEVELOPMENT OF THE PM2 5 NAAQS
27 Historically, EPA has developed NAAQS to protect sensitive human population groups
28 against adverse health effects associated with ambient exposures to certain widespread air
29 pollutants, including PM, ozone (O3), carbon monoxide (CO), sulfur dioxide (SO2), nitrogen
30 dioxide (NG2), and lead (Pb). The U.S. Clean Air Act, as amended in 1977 and 1990, requires
31 that EPA periodically review and revise as appropriate the criteria (scientific bases) and
32 standards for a given pollutant or class of pollutants (e.g., PM) regulated by NAAQS.
33 The original total suspended particulate (TSP) NAAQS set in 1971 included both
34 inhalable and noninhalable particles, ranging in size up to 25-50 jam. A later periodic review of
35 the PM criteria and NAAQS led to the setting in 1987 of "PM10" NAAQS (150 ng/m3, 24-h;
36 50 |ag/m3. annual average) aimed at protecting against health effects of inhalable particles
7/25/00 6-2 DRAFT—DO NOT CITE OR QUOTE
-------
1 (^ 10.0 um) capable of penetrating to lower (thoracic) regions of the human respiratory tract and
f depositing in tracheobronchial and alveolar tissue of the lung (Federal Register, 1987). As for
the most recently completed PM NAAQS review, assessment of the latest available scientific
4 information characterized in the EPA document Air Quality Criteria for Paniculate Matter or
5 "PM CD" (U.S. EPA, 1996a) and additional exposure and risk analyses in an associated EPA
6 PM Staff Paper (U.S. EPA, 1996b) led EPA to promulgate decisions to retain, in modified form,
7 the 1987 PM10 NAAQS and to add new PM2 5 NAAQS (65 ug/m3, 24 h; 15 ug/m3, annual
8 average) to protect against adverse health effects associated with exposures to fine particles
9 (Federal Register, 1997).
10 The 1997 PM NAAQS decisions were based, in part, on important distinctions
11 highlighted in the PM CD between fine and coarse ambient air particles with regard to size,
12 chemical composition, sources, and transport. Also of key importance was the assessment and
13 interpretation of new epidemiologic findings on airborne particle health effects. The PM CD
14 (U.S. EPA, 1996a) and Staff Paper (U.S. EPA, 1996b) highlighted more than 80 newly published
15 community epidemiology studies, of which more than 60 found significant associations between
16 increased mortality and/or morbidity risks and various ambient PM indicators. The main
17 findings of concern were community epidemiology results showing ambient PM exposures to be
18 statistically associated with increased mortality (especially among people over 65 years of age
^P and those with preexisting cardiopulmonary conditions) and morbidity (indexed by increased
20 hospital admissions, respiratory symptom rates, and decrements in lung function). As noted in
21 the PM CD, several viewpoints emerged on how best to interpret the epidemiology findings:
22 (a) reported PM-related effects are attributable to PM components (per se) of the air pollution
23 mixture and reflect independent PM effects; (b) PM exposure indicators serve as surrogate
24 measures of complex ambient air pollution mixtures, with reported PM-related effects
25 representing those of the overall mixture; or (c) PM can be viewed both as a surrogate indicator
26 and as a specific cause of the observed health effects. See Appendix C for a summary overview
27 of key epidemiologic findings supporting the 1997 NAAQS decisions.
28 As indicated in Appendix C, time-series mortality studies reviewed in the 1996 PM CD
29 (U.S. EPA, 1996a) provide strong evidence that ambient PM air pollution is associated with
30 increases in daily human mortality. These studies provided evidence that such effects occur at
31 routine ambient PM levels, extending to 24-h concentrations below the 150 ug/m3 level of the
32 PMIO NAAQS set in 1987. Overall, as noted in Table C-l of Appendix C, the PM10 relative risk
33 estimates derived from the recent PM10 total mortality studies suggest that a 24-h average 50
34 (ig/m3 PM,0 increase in acute exposure has an effect on the order of RR = 1.025 to 1.05 in the
^fc general population. Higher relative risks are indicated for the elderly and for those with
36 preexisting respiratory conditions, both of which represent subpopulations at special risk for
7/25/00 6-3 DRAFT—DO NOT CITE OR QUOTE
-------
1 mortality implications of acute exposures to air pollution, including PM. Results are very similar
2 over a range of statistical models used in the analyses, and are not artifacts of the methods by
3 which the data were analyzed. Figure C-l in Appendix C illustrates the coherence and
4 consistency of the PMIO epidemiology findings.
5 The PM CD (U.S. EPA, 1996a) also highlighted that a growing body of evidence
6 suggests that fine particles are more strongly related than inhalable coarse particles to excess
7 morality in both acute and chronic exposure studies. Such evidence notably includes the results
8 of analyses of the type illustrated in Figure C-2 of Appendix C, where a stronger linear
9 relationship is seen between acute (24-h) exposure estimates for fine particles (<2.5 um) and
10 increased mortality risks than for acute exposure estimates for thoracic coarse particles (PMI5.2 5).
11 Table C-2 of Appendix C summarized results from a wide array of U.S. and Canadian studies
12 that showed increased risks of mortality and morbidity to be related to short-term (24-h) ambient
13 fine particle exposures. On the basis of such studies, EPA proposed (Federal Register, 1996) and
14 then later promulgated (Federal Register, 1997) the new 24-h PM2 5 NAAQS of 65 ug/m3.
15 More importantly for present purposes, EPA also promulgated a long-term PM2 5 NAAQS
16 of 15 ug/m3 (annual average) to protect against effects of chronic exposures to ambient fine
17 particles (which include DPM as a notable constituent for which extensive toxicologic evidence
18 highlights the importance of chronic exposure effects). Appendix C discusses two key
19 prospective studies of long-term PM exposure effects that were of particular importance: the
20 Harvard Six Cities Study (Dockery et al., 1993) and the American Cancer Society (ACS) Study
21 (Pope et al., 1995). These two studies agree in their findings of strong associations between fine
22 particles and excess mortality. The RR estimates for total mortality are large and highly
23 significant in the Six Cities study. With their 95% confidence intervals, the RR for 50 ug/m3
24 PMI5 is 1.42 (1.16, 2.01), the RR for 25 ug/m3 PM25 is 1.31 (1.11,1.68), and the RR for 15
25 ug/m3 SO4 is 1.46 (1.16, 2.16). The ACS study estimates for total mortality are smaller, but also
26 more precise: RR = 1.17 for 25 ug/m3 PM25 (1.09, 1.26), and RR = 1.10 for 15 ug/m3 SO4 (1.06,
27 1.16). Both studies used Cox regression models and were adjusted for similar sets of individual
28 covariates. In each case, however, caution must be applied in use of the stated quantitative risk
29 estimates, given that the lifelong cumulative exposures of the study cohorts (especially in the
30 dirtiest cities) included distinctly higher past PM exposures than those indexed by the more
31 current PM measurements used to estimate chronic PM exposures in the study. Thus, somewhat
32 lower risk estimates than the published ones may well apply.
33 An additional line of evidence concerning long-term effects may be seen in comparing
34 some specific causes of death in the prospective cohort studies. Appendix C tabulates relative
35 risk estimates for total mortality, lung cancer deaths, cardiopulmonary deaths, and other deaths in
36 the Six Cities study and the ACS study. The relative risks for the most versus least polluted
6-4 DRAFT—DO NOT CITE OR QUOTE
-------
1 cities in the two studies are very similar for total, cardiopulmonary, and other causes of mortality.
f However, for lung cancer, statistically significant increased relative risk was only found for
sulfates in the ACS study, and not for PM2 5 in either study. The credibility of the air pollution-
4 related findings of the two studies is enhanced by both generating very similar elevated risk
5 estimates for smokers versus nonsmokers for cardiopulmonary and cancer mortality.
6 The PM CD (U.S. EPA, 1996a) also discussed early results reported for another
7 prospective cohort study of long-term PM exposure effects, i.e., the Adventist Health Study of
8 Smog (AHSMOG). As noted in the PM CD, Abbey et al. (1991) reported no significant
9 associations between any mortality or morbidity endpoints and TSP levels, except for respiratory
10 cancers and female cancers (any site). Follow-up analyses reported by Abbey et al. (1995)
11 considered exposures to PM10 (estimated from site-specific regressions on TSP), PM2 5 (estimated
12 from visibility), sulfates (SO4), and visibility per se (extinction coefficient). No significant
13 associations with nonexternal mortality were reported, and only high levels of TSP or PM]0 were
14 associated with airways obstructive disease or bronchitis symptoms. Further follow-up analyses
15 of the same California AHSMOG database have been reported recently by Abbey et al. (1999).
16 These latter analyses (not considered in the 1996 PM CD or 1997 PM NAAQS decisions) do
17 provide evidence indicative of increased risk of mortality from contributing nonmalignant
18 respiratory causes being associated with long-term PM exposures. Other AHSMOG analyses
^P reported by Abbey et al. (1999) and Beeson et al. (1998) also provide suggestive indications of
20 increased risk of lung cancer mortality being associated with long-term PM10 exposures.
21 The chronic exposure studies, taken together, suggest that there may be increases in
22 mortality in disease categories that are consistent with long-term exposure to airborne fine
23 particles. At least some fraction of these deaths are likely a consequence of cumulative
24 long-term exposure effects beyond the additive impacts of acute exposure episodes, in terms of
25 immediate harvesting of seriously health-compromised individuals in danger of near-future
26 death. If this is correct, then at least some individuals may experience some reduction of life as a
27 consequence of PM exposure. This issue, of better quantifying the life-shortening consequences
28 of ambient PM exposure, is being addressed more explicitly by research studies underway since
29 completion of the 1996 PM CD (U.S. EPA, 1996b).
30 The PM Staff Paper (U.S. EPA, 1996b) drew upon the quantitative epidemiology
31 information concisely summarized above to derive a rationale for selection of an annual-average
32 PM25 standard of 15 ug/m3. First, major reliance was placed on several acute (24-h) exposure
33 studies showing significantly increased risks of daily mortality (Schwartz et al., 1996) and
34 morbidity indexed by hospital admissions (Thurston et al., 1994) and respiratory symptoms/lung
^^ function decrements in children (Neas et al., 1995) in relationship to fine particle indicators
36 (PM;, 5, PM2,). It was judged that such effects of short-term exposures to fine PM were most
7/25/00 6-5 DRAFT—DO NOT CITE OR QUOTE
-------
1 strongly related to fine particle levels above the annual-average concentrations for the cities
2 evaluated in each of these studies. More specifically, statistically significant increases in relative
3 risks for daily mortality or morbidity were most clearly observed in these studies to be associated
4 with 24-h fine particle concentrations in cities with annual mean fine particle concentrations that
5 exceeded 15 ug/m3, as described in Federal Register (1996).
6 Selection of 15 ug/m3 as an acceptable level for an annual-average fine particle (PM2 5)
7 NAAQS was further supported by the findings of the long-term fine PM exposure studies, e.g.,
8 the Harvard Six Cities and ACS studies. The first noticeable increment in mortality risk
9 demonstrated by the Harvard Six Cities study occurred for Watertown (Boston), with mean
10 annual-average PM2 5 around 15 ug/m3, and more clearly increased risks were evident for the
11 other three cities, with PM2 5 annual-average values around 20 ug/m3 or higher. This comported
12 well with evidence of increased risks of mortality in the ACS study, which were also most clearly
13 attributable to PM exposures in excess of PM2 5 annual median values of 18 ug/m3 or more, and
14 with the findings of fine particle-related respiratory symptom and lung function decrement risks
15 observed by Razienne et al. (1996).
16
17 6.3. DEMAND THE PM2 s NAAQS
18 Chapter 2 of this document, as well as the PM CD (U.S. EPA, 1996a), documents the
19 extent to which DPM may be contributory to ambient PM2 5 concentrations. In some urban
20 situations, the annual average fraction of PM2 5 attributable to DPM (according to mass
21 concentrations) is about 35% on the high end, although the proportion appears to be more
22 typically in the range of about 10% (see Table 2-23 and Section 2.4.2.1). The actual contribution
23 of DPM to toxicologic effects of ambient PM2 s, however, may be disproportionately large
24 (compared with DPM's mass contribution), because of the large numbers of ultrafine particles in
25 DPM emissions and consequent increased surface area for possible interactions with other
26 ambient air toxicants and pathophysiological impacts on subcellular components of lung tissues.
27 One approach to dealing with DPM would be (a) to view DPM as an exceptionally toxic
28 component of ambient fine particle mixes in general; (b) to treat any increased mortality and/or
29 morbidity risks attributable to ambient fine particle exposures (as assessed above) as if they were
30 wholly due tc DPM; and (c) to assume, therefore, that any characterization of health risk
31 attributable to ambient fine particles would represent an upper-limit estimate of risk possibly
32 assignable to DPM. The findings of high particle counts for ultrafine particles in DPM and
33 possible consequent disproportionally enhanced impacts on ambient PM2 5 particle numbers (and
34 any associated enhanced toxicity due to this) may support taking such an approach.
35 Another alternative would be to assume that DPM's potency is essentially equal to other
36 fine particle constituents typically comprising ambient PM, 5. Only very limited specific
7/25/00 6-6 DRAFT—DO NOT CITE OR QUOTE
-------
information can be cited as empirically supporting such an assertion! Some laboratory animal
studies indicate, for example, that DPM is no more potent at eliciting pulmonary pathology than
3 are other poorly soluble particles such as talc, titanium dioxide, or carbon black in rats or talc or
4 titanium dioxide in mice. This information provides some, but by no means definitive, support
5 for the notion that there is no clear basis to substantiate that DPM is more potent in eliciting
6 pulmonary pathology than any other poorly soluble particle that may be present in ambient PM2 5.
7 In that case, a reasonably logical approach would be to attribute risks associated with ambient
8 fine particles in a roughly proportional way to constituent particles of different chemical
9 composition that typically make up ambient fine particle mixes. Then, one could estimate that
10 keeping ambient DPM exposures below an approximate range of 1.5 ng/m3 (10% * 15 ug/m3) to
11 5 [ig/m3 (35% x 15 ng/m3) would provide roughly equivalent protection against DE effects as
12 does the 15 ng/m3 PM25 annual average NAAQS for fine particle effects in general.
13 Deriving a guidance value for DPM by apportioning the PM2 5 standard as above
14 represents a very generalized, nonspecific approach to estimating a safe air level for DE as
15 indexed by DPM. That approach relies principally on the accuracy of the apportionment of DPM
16 from PM2 5, is limited by assumptions such as that of equal particle potency, and is not based on
17 more detailed consideration of specific aspects of the DPM toxicity data. Given the uncertainties
inherent in most dose-response assessment processes, it may be informative to evaluate yet
another approach to quantifying potential risk associated with ambient DPM exposure on the
20 basis of the robust and specific database documented and evaluated in Chapter 5. A data-specific
21 approach for DPM could then complement the above apportionment estimates derived from more
22 general, ambient fine particle data; apportionment-derived values from PM, 5 (as noted above)
23 could serve as a rough benchmark by which to judge the credibility of estimates derived from the
24 DPM data-specific approach. That is, other procedures performed independently of the PM25
25 database should yield values in the range of general, nonspecific estimates derived from
26 evaluation of PM2 5 risks. This would presumably be the case for RfC values derived in
27 Section 6.5 below, based on application of the RfC methodology summarized in the next section.
28
29 6.4. THE INHALATION REFERENCE CONCENTRATION APPROACH
30 Historically, approaches such as the Acceptable Daily Intake (ADI) were developed
31 whereby effect levels, such as no-observed-adverse-effect-levels (NOAELs) or lowest-observed-
32 adverse-effect-levels (LOAELs) from human or animal data, were combined with certain "safety
33 factors" to accommodate areas of uncertainty in order to make quantitative estimates of a safe-
34 dose, i.e., that at which no adverse effect would likely occur. In response to the National
^P Academy of Sciences (N AS) report entitled "Risk Assessment in the Federal Government:
36 Managing the Process" (National Research Council, 1983), EPA developed two approaches
7/25/00 6-7 DRAFT—DO NOT CITE OR QUOTE
-------
1 similar to the ADI, i.e., the oral reference dose (RfD) (Barnes and Dourson, 1988) and the
2 parallel inhalation reference concentration (RfC) with its formal methodology (U.S. EPA, 1994).
3 Similar to ADIs in intent, the RfD/C approach is used for dose-response assessment for
4 noncancer effects based upon explicitly delineated rigorous methodology adhering to the
5 principles set forth in the 1983 NRC report. The RfC methodology includes comprehensive
6 guidance on a number of complex issues, including consistent application to effect levels of
7 "uncertainty factors" (UFs) rather than the ADI "safety factors" for consideration of uncertainty.
8 Basically, these approaches attempt to estimate a likely subthreshold concentration in the human
9 population. Use of the RfD/C approach is one of the principal current agency methods for
10 deriving dose-response assessments.
11 A chronic RfC is defined as:
12 An estimate (with uncertainty spanning perhaps an order of magnitude) of a continuous
13 inhalation exposure to the human population (including sensitive subgroups) that is likely
14 to be without an appreciable risk of deleterious noncancer effects during a lifetime.
15 The RfC approach involves the following general steps:
16
17 • identification of a critical effect relevant to humans, i.e., the effect that occurs at the
18 lowest exposure/dose in human or animal studies;
19 • selection of appropriate dose-response data to derive a point of departure for
20 extrapolation of a key study (or studies) that provides a NOAEL, LOAEL, or
21 benchmark concentration (BMCLJ1;
22 • Obtain HECs when animal exposure-response data are used (via use of
23 PBPK/dosimetry models);
24 • application of UFs to the point of departure (e.g., NOAEL, LOAEL, BMCLX) to
25 address extrapolation uncertainties (e.g., interindividual variation, interspecies
26 differences, adequacy of database); and
27 • characterization of the confidence of the dose-response assessment and resultant
28 RfC.
29
'BMCLX is denned as the iuwcr 95% confidence limit of the dose that will result in a level ot "x" response
(e.g., BMCLIO is the lower 95% confidence limit of a dose for a 10% increase in a particular response).
7/25/00 6-8 DRAFT—DO NOT CITE OR QUOTE
-------
1 The basic quantitative formula for derivation of an RfC, given in Equation 6-1, has as its
basic components an effect level, here a NOAEL, expressed in an HEC and UFs. The units of an
3 RfCaremg/m3.
4 Alternatively, the numerator in Equation 6-1 may be a LOAEL or BMCLX. The
5 benchmark concentration (BMC) approach and its application in this assessment are documented
6 in Appendix B. Also, a modifying factor (MF) may be used in the denominator of this equation
7 to account for scientific uncertainties in the study chosen as the basis for the RfC. Further
8 specifics of RfC derivation procedures are discussed as they are used in the following sections.
9 All such procedures are described in detail in the RfC Methodology (U.S. EPA, 1994).
10
11 6.5. CHRONIC REFERENCE CONCENTRATION FOR DIESEL EXHAUST
12 As concluded in Chapter 5, chronic respiratory effects are the principal noncancer human
13 hazard from long-term environmental exposure to DE. Other effects (e.g., neurological, liver)
14 are observed in animal studies at higher exposures than the respiratory effects. Thus, the
respiratory effects are considered the "critical effect" for the derivation of a chronic RfC for DE.
The human and animal data for the immunological effects of DE are considered inadequate for
17 dose-response evaluation.
18 The evidence for chronic respiratory effects is based mainly on animal studies showing
19 consistent findings of inflammatory, histopathological (including fibrosis), and functional
20 changes in the pulmonary and tracheobronchial regions of laboratory animals, including the rat,
21 mouse, hamster, guinea pig, and monkey. Occupational studies of DE provide some
22 corroborative evidence of possible respiratory effects (e.g., respiratory symptoms and possible
23 lung function changes), although those studies are generally deficient in exposure-response
24 information.
25 Mode-of-action information about respiratory effects from DE exposure indicates that, at
26 least in rats, the pathogenic sequence following the inhalation of DPM begins with the
27 phagocytosis of diesel particles by alveolar macrophages (AMs). These activated AMs release
28 chemotactic factors that attract neutrophils and additional AMs. As the lung burden of DPM
29 increases, there are aggregations of particle-laden AMs in alveoli adjacent to terminal
30 bronchioles, increases in the number of Type II cells lining particle-laden alveoli, and the
presence of particles within alveolar and peribronchial interstitial tissues and associated lymph
nodes. The neutrophils and AMs release mediators of inflammation and oxygen radicals, and
7/25/00 6-9 DRAFT—DO NOT CITE OR QUOTE
-------
1 particle-laden macrophages are functionally altered, resulting in decreased viability and impaired
2 phagocytosis and clearance of particles. The latter series of events may result in pulmonary
3 inflammation, fibrosis, and eventually lesions like those described in the studies reviewed in
4 Chapter 5. Although information describing the possible pathogenesis of respiratory effects in
5 humans is not available, the effects reported in studies of humans exposed to DE are not
6 inconsistent with the findings in controlled animal studies.
7 There are several reasons the dose-response data in rats are considered appropriate for use
8 in characterizing noncancer health effects in humans and deriving a chronic RfC for DE. First,
9 similar noncancer respiratory effects are seen in other species (mouse, hamster, guinea pig, and
1 0 monkey). Second, rats and humans exhibit similar noncancer responses (macrophage response
1 1 and interstitial fibrosis) to other particles such as coal mine dust, silica, and beryllium (Haschek
1 2 and Witschi, 1991; Oberdorster, 1994). Third, an expert panel convened by ILSI recommends
1 3 that response data on persistent inflammatory processes may be used to assess non-neoplastic
1 4 responses of poorly soluble particles such as DPM (ILSI, 2000).
15
16 6.5.1. Principal Studies for Dose-Response Analysis: Chronic, Multiple-Dose Level Rat
1 7 Studies
1 8 The experimental protocols and results from the long-term repeated-exposure chronic
1 9 studies demonstrating and characterizing the critical effect of pulmonary fibrotic changes and
20 inflammation are discussed in Chapter 5. Salient points of these studies, including species/sex
21 of the test species, the exposure regime and concentrations reported in mg DPM/m3, and effect
22 levels, are abstracted in Table 6-1 for further consideration. The effect levels are designated as
23 N for no-observed-adverse-effect-level, A for adverse-effect-level, and BMCL10 for the
24 benchmark concentration at 10% incidence (see Appendix B).
25 The purpose of many of the chronic studies listed in this table was not the elucidation of
26 the concentration-response character of DPM. The studies of Heinrich et al. (1982, 1986) in
27 hamsters, mice, and rats; of Iwai et al. (1986) in rats; of Heinrich et al. (1995) in mice; of Lewis
28 et al. (1989) in monkeys; and of Pepelko (1982a) in rats are all single dose-level studies that have
29 as their genesis mechanistic or species-comparative purposes. As discussed in Chapter 5, many
30 of these studies provide valuable supporting information for designation of the critical effect of
31 pulmonary histopathology. The lack of any clear dose-response information, however, precludes
32 consideration of these studies as the basis for RfC derivation.
33 Likewise, multiple-level exposure chronic studies involving species other than rats, i.e.,
34 hamsters (Pepelko, 1 982b), cats (Plopper et al., 1 983), and guinea pigs (Barnhart et al., 1 98 1 ,
35 1982), provide cross-species corroboration of the critical effects of pulmonary histopathology
36 and inflammatory alteration.
-I /") c /r\r\ £. ] C\ r~»r> A t?T
7/25/l/U O-1'J L/KAr i —
-------
The remaining studies showing exposure-response relationships in rats for the critical
effects include those of Ishinishi et al. (1986, 1988), Mauderly et al. (1987a), Heinrich et al.
3 (1995), and Nikula et al. (1995). As described in Chapter 5, all of these studies were conducted
4 and reported in a thorough, exhaustive manner on the critical effects and little, if any, basis exists
5 for choosing one over another for purposes of RfC derivation. One way of taking advantage of
6 this high degree of methodological and scientific merit would be to array data from all these
7 studies and their effect levels (NOAEL, LOAEL, BMCLJ subsequent to normalization of the
8 exposure conditions, i.e., conversion of the exposure regimes via the model of Yu et al. (1991) to
9 yield an HEC. This exercise would result in an interstudy concentration-response continuum that
10 would further facilitate the choice of a concentration to fulfill the purposes of an RfC.
11
12 6.5.2. HEC Derivation
13 Pharmacokinetic, or PK, models can be used to project across species the concentrations
14 of a toxicant that would result in equivalent internal doses. When used for these purposes, PK
15 models may be termed dosimetric models. Chapter 3 reviewed and evaluated a number of
16 dosimetric models applicable to DPM. The model developed by Yu et al. (1991) accounts for
17 species differences in deposition efficiency, normal and particle overload lung clearance rates,
respiratory exchange rates, and particle transport to lung-associated lymph nodes. Of the models
considered in Chapter 3 and currently available, that of Yu et al. (1991) is the only recent model
20 parameterized for both animals and humans that is capable of performing animal-to-human
21 extrapolation; a major assumption of this model is that the phenomenon of particle overload
22 would occur in humans at the same lung burdens (expressed as mass per unit surface area) as in
23 rats. This assumption allows for the development of a diesel-particle-specific human retention
24 model, thereby allowing for extrapolation from exposures in rat studies to exposures in humans.
25 Chapter 3 and Appendix A further discuss the model and its limitations, and document its use in
26 this assessment. Note that this procedure would address species differences in dose (i.e.,
27 toxicokinetics), although not necessarily comparative response, or toxicodynamics, the second
28 aspect of uncertainty in interspecies extrapolation.
29 A principal and critical decision in utilizing any dosimetric model is the measure of dose.
30 DPM is composed of an insoluble carbonaceous core with a surface coating of relatively soluble
31 organic constituents. Because macrophage accumulation, pulmonary histopathology, and
32 reduced clearance have been observed in rodents exposed to high concentrations of chemically
33 inert particles (Morrow, 1992), the toxicity of DPM may be considered to result from the
34 carbonaceous core rather than the associated organics. However, the organic component of
^B diesel particles, consisting of a large number of PAHs and heterocyclic compounds and their
36 derivatives (Chapter 2), has been implicated in toxicity. The model of Yu et al. (1991) considers
7/25/00 6-11 DRAFT—DO NOT CITE OR QUOTE
-------
1 the interspecies kinetics of organics (as slowly and fast-cleared) desorbed from the carbonaceous
2 core. Other guidance on choice of dosimetrics for poorly soluble particles such as DPM states
3 that some estimate of lung burden is necessary, that the aerosol exposure parameters such as
4 MMAD, og, particle surface area, and density are characterized such that different dose metrics
5 may be considered as new mechanistic insights are developed (ILSI, 2000). The whole particle,
6 as characterized in this assessment and utilized in the model of Yu et al. (1991), meets this
7 recommended guidance, and therefore ug/m3 of DPM is used as the measure of dose.
8 The input data required to run the dosimetric model of Yu et al. (1991) include the
9 particle size characterization, expressed as mass median aerodynamic diameter (MMAD), and
10 the geometric standard deviation (og). Simulation data presented by Yu and Xu (1986) show that
11 across a range of MMAD and og inclusive of the values reported in these studies, the pulmonary
12 deposition fraction differs by no more than 20%. The minimal effect of even a large distribution
13 of particle size on deposition probably results because the particles are still mostly in the
14 submicron range, where deposition is influenced primarily by diffusion. It has also been shown,
15 however, that the particle characteristics in a DE exposure study depend very much on the
16 procedures used to generate the chamber atmosphere. Because of the rapid coagulation of
17 particles, the volume and temperature of the dilution gas are especially important. The
18 differences reported in particle sizes and distributions in various studies are relatively small and
19 likely reflect different analytical methods as well as real differences in the exposure chambers.
20 Because the particle diameter and size distribution were not reported in the two lowest exposure
21 concentrations in the Ishinishi studies, it was decided to use a representative DPM particle size of
22 MMAD = 0.2 urn and og = 2.3 (values typically reported for DPM) for modeling of lung burden.
23 For consistency, the lung burdens for the other studies were also calculated using this
24 assumption. The difference in the HEC using the default particle size compared with the actual
25 reported particle size is no more than 4% in the Ishinishi studies (Ishinishi et al., 1986, 1988) and
26 19% in the Mauderly et al. (1987b) study.
27 The foregoing discussion addresses, in part, the variability in outcomes that may be
28 predicted from the Yu et al. (1991) model from deposition of DPM. Variability in output of the
29 model (lung soot burden) was also examined by Yu and Yoon (1990). who studied dependency
30 on tidal volume, respiration rate and clearance (in terms of the overall particle transport rate,
31 ^A(l))- Analysis of the output dependency indicated that the model output is sensitive but not
32 overly so for these determinative parameters. A ± 20% change in values for 1A(1), for example,
33 were estimated to result in a 16%-26% change in soot burden at a 0.1 mg/m3 continuous diesel
34 exposure for 10 years. For a ± 10% change in tidal volume, the model projected changes in soot
35 burden ranging from 14% to 22% for this samp pvnncnre scenario. That the changes in the
36 model outcome were comparable to changes in the input parameters such as tidal volume is an
7/25/00 6- i 2 DRAFT—DO NOT CITE OR QUOTE
-------
indication that the variability of the model applied to the human population would be the
variability of these physiologic parameters in the human population. Variability within the
3 human population is often addressed by applying safety or uncertainty factors, usually in the
4 range of 10 (Renwick and Lazarus, 1998; U.S. EPA, 1994). This matter will be discussed further
5 below.
6 As discussed in Chapter 3, evidence from Kuempel et al. (2000) suggests that the Yu
7 model may underpredict the lung dust burdens in humans, as judged from occupational data
8 obtained from coal miners (Freedman and Robinson 1988), ostensibly because of the lack of an
9 interstitial compartment in the Yu model. However, further investigation is needed to ascertain
10 whether transfer of particles to the interstitium would also describe the clearance and retention
11 processes in the lungs of humans with exposures to particles at lower environmental
12 concentrations, or to submicron particles such as DPM.
13 HECs were obtained for the dose levels and exposure scenarios presented in the studies of
14 Mauderly et al. (1987b), of Ishinishi et al. (1986,1988), of Nikula et al. (1995), and of Heinrich
15 et al. (1995), the specifics of which are shown in Appendix A. The HECs are arrayed ordinally
16 according to their effect level (NOAEL, LOAEL, BMCL10) in Table 6-2.
17 Further inspection of Table 6-2 shows that calculating and ordering the HECs created a
concentration-response continuum based on an internal dose that blends from HECs with no
observed adverse effects at concentrations as low as 0.032 mg/m3 to HECs that are associated
20 with an adverse effect level that first appears definitively in the continuum probably at
21 0.33 mg/m3.
22 Inspection of the interstudy dose-response continuum in Table 6-2 to elucidate a point of
23 departure for an RfC entails some interpretation. Exposures at the lower end of this table show
24 that elevated chronic exposures to DPM consistently result in AELs. Conversely, entries in the
25 upper portion of this table show that low-level chronic exposures to DPM have minimal, if any,
26 effects within the capability of these studies to detect them. Intermediate chronic exposures from
27 0.128 mg/m3 to 0.9 mg/m3, however, are less clear, and effect levels and exposures either have no
28 or few observable effects, or effects that are minimally adverse. In choosing from among levels
29 (e.g., NO AELs, LO AELs, BMCLxs) as a point of departure for derivation of an RfC, the
30 methodology (U.S. EPA, 1994) provides guidance for choice of a highest no-effect level below
31 an effect level; the interim guidance for the BMC suggests that for use as a point of departure, a
32 benchmark (e.g., BMCL,0) should be within the range of the observable response data so as to
33 avoid excessive extrapolation, and take the shape of the dose-response curve into consideration
34 (Barnes et al., 1995; U.S. EPA, 1995b). The highest no-effect HECs (NOAELHEC) in this table
^P are 0.128 mg/m3 and 0.144 mg/m3 from the Ishinishi et al. (1988) study, nearly fivefold above
36 other no-effect levels of 0.032 and 0.038 mg/m3. The lower BMCL!0 (0.37 mg/m3) is at nearly
7/25/00 6-13 DRAFT—DO NOT CITE OR QUOTE
-------
1 the same concentration as the lowest LOAEL of 0.33 mg/m3 and thus may be too high an
2 estimate for use as a point of departure, possibly because of excessive extrapolation
3 (Appendix B). This BMCLIO, generated directly from a modeled dose-response curve for
4 chronic inflammation, lends credence to these NOAELs as being associated with the dose-
5 response curve at incidences of considerably less than 10%. The value of 0.144 mg/m3 is chosen
6 as the point of departure for development of the RfC because it is the highest NOAHI^c among
7 the available NOAELs.
8
1 6.5.3. Consideration of Uncertainty Factors for the RfC
2 Uncertainty for the DE assessment exists in the following areas: inter-individual variability
3 and animal-to-human extrapolation. Each shall be addressed in this section.
4 Considerable qualitative but little, if any, quantitative information exists regarding
5 subgroups that could be sensitive to any respiratory tract effects of DPM. The population
6 assumed in this assessment consists of individuals of average health in their adult years. It is
7 acknowledged that exposure to DPM could be additive to many other daily or lifetime exposures
8 to airborne organic compounds and nondiesel ambient PM. It is also likely that individuals who
9 predispose their lungs to increased particle retention through smoking or other high paniculate
10 burdens, who have existing respiratory tract inflammation or infections, or who have chronic
11 bronchitis, asthma, or fibrosis could be more susceptible to adverse impacts from DPM exposure
12 (Chapter 5). Also, infants and children could have a greater susceptibility to the acute/chronic
13 toxicity of DPM because of their greater breathing frequency and consequent potential for greater
14 particle deposition in the respiratory tract. Increased respiratory symptoms and decreased lung
15 function in children versus ambient PM levels, of which DPM is a part, have been observed (U.S.
16 EPA, 1996a). Likewise, a number .of factors may modify normal lung clearance, including,
17 aging, gender, and disease. Although the exact role of these factors is not resolved, all would
18 influence the particle dose to the lung tissue from inhalation exposure. Activity patterns related
19 to occupation and habitation in the proximity of major roadways are certain to be contributory for
20 some subgroups in receiving higher DPM exposures (Chapter 2). In the absence of DE-specific
21 data, this assessment utilises a default UF value of 10 to account tor possible inter-individual
22 human variability (U.S. EPA, 1994; Renwick and Lazarus, 1998).
23 In terms of animal-to-human extrapolation, this dose-response assessment utilizes data
24 from the rat to predict human response. To account for interspecies differences in toxico-
25 dynamics and kinetics, a default UF of 10 is typically used when there is no information about
26 which test animal species best represents humans. For DE, available data indicate that the rat
27 appears more sensitive to the inflammatory effects than humans. Furthermore, the toxicokinetic
7/25/00 6-14 DRAFT—DO NOT CITE OR QUOTE
-------
differences were accounted for by the use of a dosimetry model (Yu et al., 1991), hence, a UF is
not needed.
3 In summary, the application of a factor of 10 to the HEC of 0.144 mg/m3 would be
4 prudent to address the issue of human variability in response to effects from exposure to DPM.
5 Use of other UFs is not considered netessary. It should be noted that, given the emerging
6 research on DE-induced immunological effects, it may be necessary at a later date to reconsider
7 the basis for selection of the critical effect and UFs for derivation of the DE RfC.
8
9 6.5.4. Derivation of the RfC for DE
10 On the basis of the above analysis, the value of 0.144 mg/m3 DPM was selected as the
11 basis of the RfC evaluation. This value was derived from concentrations in rat chronic studies
12 that were modeled to obtain HECs. The pulmonary effects, histopathology and inflammation,
13 were determined to be the critical noncancer effects. Response data on inflammation was also
14 suggested by a specific scientific working group as a satisfactory surrogate for fibrogenic
15 responses in assessing the pulmonary responses of poorly soluble particles such as DPM (ILSI,
16 2000). Sufficient documentation from other studies showed that there is no effect in the
17 extrathoracic (nasopharyngeal) region of the respiratory system or in other organs at the lowest
« levels that produce pulmonary effects in chronic exposures. Application of the dosimetric model
of Yu et al. (1991) to the exposure value from Ishinishi et al. (1988) of 0.46 mg/m3 yielded an
20 HEC of 0.144 mg/m3. Application of the UF for intraspecies variability would yield the
21 following RfC:
22
NOAELHEC-UF = RfC
23
0.144 mg/m3- 10 = 0.0144 mg/m3 = 14//g/m3.
24
25 6.6. CHARACTERIZATION OF THE NONCANCER ASSESSMENT FOR DE
26 Adverse health effects from short-term acute (high-level) exposures to DE such as
27 occupational reports of decreases in lung function, wheezing, chest tightness, increases in airway
28 resistance, and reports in laboratory animals of inflammatory airway changes and lung function
29 changes are acknowledged but are not quantitatively assessed. Thus, the focus of this dose-
30 response assessment of is on the adverse noncancer health consequences of a lifetime low-level
31 continuous air exposure of humans to DE.
32 This assessment uses the whole particle, termed DPM, as the key index or measure of DE
^B dose. DPM includes any and all adsorbed organics, among which are a large number of PAHs,
34 heterocyclic compounds, and their derivatives (Chapter 2), as well as the carbon core. It is not
7/25/00 6-15 DRAFT—DO NOT CITE OR QUOTE
-------
1 possible to separate the carbon core from the adsorbed organics to compare the toxicity. The
2 dosimetric model used in the derivation of the RFC (Yu et al., 1991) is consistent with this
3 designation as it considers DPM as well as the adsorbed organics as two types, slowly cleared
4 and fast-cleared. Some studies with diesel do occasionally report levels of accompanying
5 gaseous components of DE (NOX, CO, etc.), but nearly all report particle concentration and
6 characteristics.
7 Adverse responses occurring in the rat lung have been used in this assessment as the basis
8 for characterizing non-neoplastic human lung responses. The basis for this presumption includes
B the fact that humans and rats exhibit similar responses to poorly soluble particles such as DPM
10 (ILSI, 2000). Also, similar noncancer effects are seen in other species. Thus, when viewed
11 across species (including humans), the non-neoplastic pulmonary effects of inflammation and
12 fibrosis used in this assessment are dissociable from the cancer response and are of likely
13 relevance to humans.
14 As a part of the RfC methodology (U.S. EPA, 1994), dose-response assessments are
1 5 assigned levels of confidence that are intended to reflect the strengths and limitations of the
16 assessment as well as to indicate the likelihood of the assessment changing with any additional
17 information. Confidence levels of either low, medium, or high are assigned both to the study (or
18 studies) used in the assessment to characterize the critical effects and to the overall toxicological
13 database of the substance. An overall confidence level is also assigned to the entire assessment
20 and is usually limited to and the same as the confidence in the database. An assessment with a
21 substance having a database as extensive as DE would normally be characterized as having high
22 confidence. The critical effects are characterized using not one but multiple long-term chronic
23 studies conducted independently of one another (Table 6-2). The exhaustive manner in which
24 these studies were conducted and reported imparts a high degree of confidence.
25 The toxicological database for DE is relatively complete. Both developmental and
26 reproductive areas are addressed. Ancillary studies that address mechanistic aspects of DE
27 toxicity, either as the whole particle with adsorbed organics, or segregated as a poorly soluble
28 particle and extracted organics. are available and used in this assessment. Although only limited
29 human data are available, extensive consideration has been given to the relevancy of the animal
30 studies to the human condition. A major point to consider in assigning confidence in this
31 assessment, and a reason that it may change in the future, is the emerging issue of allergenicity
32 caused or exacerbated by DE. Although information to evaluate allergenicity in parallel to the
33 present effects (pulmonary inflammation and histopathology) is currently lacking, future efforts
34. to elucidate and characterize this effect may well be a driver to make a reevaluation of DE
3R annronriatp Out of consideration of the relevance of (and information lacking on) allergenicity
36 effects associated with DE, and the possibility that the current RfC could change as a
7/25/00 6-16 DRAFT—DO NOT CITE OR QUOTE
-------
consequence of this information becoming available from the scientific community, the database
and overall confidence in the current RfC for DE is regarded as medium.
3 In the introductory portion of this chapter, DPM is acknowledged as a subtraction of
4 PM2 s. It was proposed that apportionment of DPM contributions in relationship to the PM2 5
5 standard and the NAAQS of 15 ng/m3 could itself conceivably serve both as a general
6 nonspecific estimate of a reasonably defensible guideline for DE measured as DPM and as a
7 reasonable bounding estimate for any value(s) derived from any approach taken in formulating a
8 dose-response assessment specific for DE. In evaluating the entirety of the disparate DE
9 database, including many chronic studies from several different species, a myriad of possible DE-
10 specific toxicological endpoints, and dose extrapolation models, application of the RfC method
11 produced a value of 14 ug/m3. As the accuracy of the RfC is part of its definition ("...within an
12 order of magnitude... "), this dose-response estimate could be considered to be no different from
13 the apportionment estimate of 1.5-5 ug/m3 or from the NAAQS of 15 ug/m3. This congruence of
14 estimates attests to the reasonableness of the data used and the judgments made in the RfC
15 process, as well as enhancing the overall confidence in these estimates regarding the toxicity of
16 DE and its potential health risk for the human population.
17
6.7. SUMMARY
Table 6-3 summarizes the key data and factors used in the dose-response analysis leading
20 to the derivation of the RfC for DE. The DE RfC of 14 ug/m3 of DPM is a chronic exposure
21 likely to be without an appreciable risk of adverse human health effects.
22 Given the perspective of RfC values being by definition "within an order of magnitude"
23 of actual values likely to be associated with low probability of adverse health effects occurring
24 with lifetime chronic exposures of sensitive human populations, the DE RfC value of 14 ug/m3
25 appears to be reasonably concordant with (a) the annual PM2S NAAQS of 15 ug/m3 serving as an
26 upper bound for possible allowable DE health risks, and/or (b) the 1.5-5 ug/m3 apportionment for
27 DE contributions implied to be inherent within the PM2 5 NAAQS established to protect against
28 adverse effects of ambient air fine-particle mixes typical of the current U.S. environment.
29 The estimated air concentration of 14 ug/m3 (the RfC, a lifetime exposure to DE
30 measured as DPM) is well above the ambient air levels that are reported in most rural areas but
31 could be below that reported under short-term conditions in some urban scenarios such as busy
32 intersections or bus stops (see Table 2-23, Chapter 2). Aspects of time-averaging concentrations
33 are also not part of this assessment, although readers and users may wish to consider this in
34 relation to the 14 ug/m3 air level.
7/25/00 6-17 DRAFT—DO NOT CITE OR QUOTE
-------
Table 6-1. Histopathological effects of diesel exhaust in the lungs of laboratory
animals
Study
Species/sex
Exposure
period
Particles
(mg/mj)
NOAEL, AEL,
or BMCL,,
(mg/mj)
Effects
Lewis etal. (1989)
Bhatnagar et al. (1980)
Pepelko(1982a)
Pepelko(1982b)
Heinrich etal. (1982)
Iwai etal. (1986)
Mauderly et al. (1987a)
Henderson et al. (1988)
Heinrich et al. (1995)
Ishnishi etal., (1986,
1988)
Heinrich etal.,(1986)
Monkey, 7 h/day
Cynomolgus, M 5 days/wk
104 wks
Rat, F344, M, F
Hamster,
Chinese, M
Hamster, Syrian,
M, F
Rat, F344, F
Rat, F344, M, F;
Mouse, CD-I,
M, F
Rat, Wistar, F;
Mouse, NMRI, F
(7 mg/m3 only)
Hamster, Syrian,
M. F; Mouse,
NMRI. F; Rat,
Wistar, F
7 h/day
5 days/wk
104 wks
8 h/day
5 days/wk
26 wks
7-8 h/day
5 days/wk
120 wks
8 h/day
7 days/wk
104 wks
7 h/day
5 days/wk
130 wks
18 h/day
5 days/wk
24 mo
Mouse, NMRI,
F;
C57BL/6N, F
Rat, M, F,
F344, /Jcl.
18 h/day
5 days/wk
13.5 mo
(NMRI)
24 mo
(C57BL/N)
16 h/day
6 days/wk
130 wks
19 h/day
5 days/wk
120 wks
2.0
2.0
6.0
12.0
3.9
4.9
0.35
3.5
7.1
0.8
2.5
7.0
7.0
0.11*
0.41'
1.08'
2.32'
0.46"
0.96"
!.84b
4.24
N
A
A
A
A
A
N
N
A
A
N
A
A
A
A
AM aggregation; no fibrosis,
inflammation, or emphysema
Multifocal histiocytosis;
inflammatory changes; Type II cell
proliferation; fibrosis
Inflammatory changes; AM
accumulation; thickened alveolar
lining; Type II cell hyperplasia;
edema; increase in collagen
Inflammatory changes, 60%
adenomatous cell proliferation
Type II cell proliferation;
inflammatory changes; bronchial
hyperplasia; fibrosis
Alveolar and bronchiolar epithelial
metaplasia in rats at 3.5 and
7.0 mg/m3; fibrosis at 7.0 mg/m3 in
rats and mice; inflammatory changes.
Little quantitative data given
Bronchioalveolar hyperplasia,
interstitial fibrosis in all groups.
Severity and incidence increase with
exposure concentration. Text only
given
No increase in tumors. Noncancer
effects not discussed
Inflammatory changes; Type II cell
hyperplasia and lung tumors seen at
>0.4 mg/m3; shortening and loss of
cilia in trachea and bronchi. Data
given in text only
Inflammatory changes; thickened
alveolar septa; bronchioloalveolai
hyperplasia; alveolar lipoproteinosis;
emphysema (diagnostic methodology
not described); hyperplasia; lung
tumors
7/25/00
6-18
DRAFT—DO NOT CITE OR QUOTE
-------
Table 6-1. Histopathological effects of diesel exhaust in the lungs of laboratory
animals (continued)
Study
Barnhart et al. (1981,
1982);Vostaletal.
(1981)
Plopperetal. (1983)
Hyde etal. (1985)
Nikulaetal. (1995)
Exposure
Species/sex period
Guinea pig, 20 h/day
Hartley, M 5.5 days/wk
104 wks
Cat, inbred, M 8 h/day
7 days/wk
124 wks
Rat,F344, M 16 h/day
5 days/wk
23 mo
Particles
(mg/m3)
0.25
0.75
1.5
6.0
6.0=
12.0"
2.44
6.33
NOAEL, AEL,
or BMCL,,
(mg/m1)
N
A
A
A
A
A
A
A
BMCL,0<
Effects
Minimal response at 0.25 and
ultrastructural changes at 0.75 mg/m3;
thickened alveolar membranes; cell
proliferation; fibrosis at 6.0 mg/m3;
increase in PMN at 0.75 mg/m3 and
1.5 mg/m3
Inflammatory changes; AM
aggregation; bronchiolar epithelial
metaplasia; Type II cell hyperplasia;
peribronchiolar fibrosis
AM hyperplasia, epithelial
hyperplasia, inflammation, septal
fibrosis, bronchoalveolar metaplasia
'Light-duty engine.
'Heavy-duty engine.
"1 to 61 weeks exposure.
d62 to 124 weeks of exposure.
'See Appendix C.
AM = Alveolar macrophage.
PMN = Polymorphonuclear leukocyte.
7/25/00
6-19
DRAFT—DO NOT CITE OR QUOTE
-------
Table 6-2. Human equivalent continuous concentrations (HECs) calculated with
the model of Yu et al. (1991) from long-term repeated exposure rat studies of DPM
exposure. Effect levels are based on the critical effects of pulmonary
histopathology and inflammation as reported in the individual studies
Study
Ishinishietal.(1988)(LDc)
Mauderly et al. (1987a)
Ishinishietal.(1988)(LD)
Ishinishi et al. (1988) (HD)
Heinrichetal.(1995)
Nikulaetal. (1995)
Ishinishi etal. (1988) (HD)
Ishinishi etal. (1988) (LD)
Nikulaetal. (1995)
Mauderly etal. (1987a)
Nikulaetal. (1995)
Ishinishi et al. (1988) (HD)
Heinrich etal. (1995)
Ishinishi etal. (1988) (LD)
Mauderly et al. (1987a)
Ishinishi etal. (1988) (HD)
Exposure concentration
(mg/m3)
0.11
0.35
0.41
0.46
0.84
2.44 & 6.3
0.96
1.18
2.44 & 6.3
3.47
2.44
1.84
2.5
2.32
7.08
3.72
Effect level*
NOAEL
NOAEL
NOAEL
NOAEL
LOAEL
BMCL,0 -inflam
LOAEL
LOAEL
BMCL10 - fibrosis
LOAEL
LOAEL
AEL
AEL
AEL
AEL
AEL
HEC
(mg/m3)
0.032
0.038
0.128
0.144
0.33
0.37
0.883
1.25
1.3
1.375
1.95
2.15
2.35
2.75
3.05
4.4
"NOAEL: no-observed-adverse-effect level; LOAEL: lowest-observed-adverse-effect level; AEL: adverse-effect
level; BMCL]0; lower 95% confidence estimate of the concentration of DPM associated with a 10% incidence of
chronic pulmonary inflammation (inflam) or fibrosis (see Appendices A and C for more specifics).
bThe duration-adjusted value from the laboratory animal exposure concentrations from hours/day, days/week to a
continuous 24 hr/day, 7 day/week exposure concentration.
CL/HD = light/heavy duty diesel engine.
-7 /"> c ir
11 ^.J/
6-20
-------
Table 6-3. Decision summary for the quantitative noncancer RfC assessment for
continuous exposure to diesei particulate matter (PPM)
Quantitative assessment for noncancer effects from
lifetime exposure to PPM 14 jig/m3
Critical effect Pulmonary inflammation and
histopathology in rats
Principal study Array of 4 chronic rat studies
Designated basis for quantitation (in laboratory animals) 0.46 mg/m3, a NOAEL
NOAELHEC (Human Equivalent Concentration) 0.144 mg/m3
Adjustments for human-to-sensitive-human (Uncertainty 10
Factor, UF)
NOAEW/UF 0.144 mg/m3/10= 14 ug/m3
7/25/00 6-21 DRAFT—DO NOT CITE OR QUOTE
-------
1 6.8. REFERENCES
2
3 Abbey, DE; Mills, PK; Petersen, FF; et al. (1991) Long-term ambient concentrations of total suspended participates
4 and oxidants as related to incidence of chronic disease in California Seventh-Day Adventists. Environ Health
5 Perspect 94:43-50.
6
7 Abbey, DE; Hwang, BL; Burchette, RJ; et al. (1995) Estimated long-term ambient concentrations of PM10 and
8 development of respiratory symptoms in a nonsmoking population. Arch Environ Health 50:139-152.
9
10 Abbey, DE; Nishino, N; McDonnell, WF; et al. (1999) Long-term inhalable particles and other air pollutants related
11 to mortality in nonsmokers. Am J Respir Crit Care Med 159:373-382.
12
13 Barnes, DG; Dourson, M. (1988) Reference dose (RfD): description and use in health risk assessments. Regul
14 Toxicol Pharmacol 8:471 -486.
15
16 Barnes, DG; Daston, GP; Evans, JS; et al. (1995) Benchmark dose workshop: criteria for use of a benchmark dose to
1 7 estimate a reference dose. Regul Toxicol Pharmacol 21:296-306.
18
19 Barnhart, MI; Chen, S-T; Salley, SO; et al. (1981) Ultrastructure and morphometry of the alveolar lung of guinea
20 pigs chronically exposed to diesel engine exhaust: six months' experience. J Appl Toxicol 1:88-103.
21
22 Barnhart, MI; Salley, SO; Chen, S-T; et al. (1982) Morphometric ultrastructural analysis of alveolar lungs of guinea
23 pigs chronically exposed by inhalation to DE (DE). In: Toxicological effects of emissions from diesel engines:
24 proceedings of the Environmental Protection Agency diesel emissions symposium; Lewtas, J., ed. October, 1981;
25 Raleigh, NC. New York: Elsevier Biomedical; pp. 183-200. (Developments in toxicology and environmental
26 science: v. 10).
27
28 Beeson, WL; Abbey, DE; Knutsen, SF. (1998) Long-term concentrations of ambient air pollutants and incident lung
29 cancer in California adults: results from the AHSMOG study. Environ Health Perspect 106:813-823.
30
31 Bhatnagar, RS; Hussain, MZ; Sorensen, KR; et al. (1980) Biochemical alterations in lung connective tissue in rats
32 and mice exposed to diesel emissions. In: Health effects of diesel engine emissions: proceedings of an international
33 symposium Pepelko, WE; Danner, RM; Clarke, NA., eds., v. 1; December 1979; Cincinnati, OH. Cincinnati, OH:
34 U.S. Environmental Protection Agency, Health Effects Research Laboratory; pp. 557-570; EPA report no.
35 EPA-600/9-80-057a. Available from: NTIS, Springfield, VA; PB81-173809.
36
37 Dockery, DW; Pope, CA, III; Xu, X; et al. (1993) An association between air pollution and mortality in six U.S.
38 cities. N Engl J Med 329:1753-1759.
39
40 Federal Register. (1987) Revisions to the national ambient air quality standards for particulate matter. F. R. (July 1)
41 52:24634-24669.
42
43 Federal Register. (1996) National ambient air quality standards for particulate matter; proposed rule. F. R.
44 (December 15) 61:65637-65713.
45
46 Federal Register. (1997) National ambient air quality standards fcr particuiate matter; fir.a! rale. F. R. (July 18)
47 62:38652-38752.
48
49 Freedman, AP; Robinson, SE. (1988) Noninvasive magnetopneumographic studies of lung dust retention and
50 clearance in coal miners. In: Respirable dust in the mineral industries: health effects, characterization, and control:
51 proceedings of the international symposium on respirable dust in the mineral industries; October 1986; University
52 Park, PA. Frantz, RL; Ramani, RV., eds. University Park, PA: Pennsylvania State University Press; pp. 181-186.
53
54 Haschek, WM; Witschi, HP. (1991) Respiratory system. In: Haschek, WM; Rousseaux, CG., eds. Handbook of
bb toxicologic pathology. New York: Academic Press, pp. 761-827.
7/25/00 6-22 DRAFT—DO NOT CITE OR QUOTE
-------
1 Heinrich, U; Peters, L; Funcke, W; et al. (1982) Investigation of toxic and carcinogenic effects of DE in long-term
f inhalation exposure of rodents. In: Toxicological effects of emissions from diesel engines: proceedings of the
Environmental Protection Agency diesel emissions symposium; October 1981; Raleigh, NC. Lewtas, J., ed. New
York: Elsevier Biomedical; pp. 225-242. (Developments in toxicology and environmental science: v. 10).
5
6 Heinrich, U; Muhle, H; Takenaka, S; et al. (1986) Chronic effects on the respiratory tract of hamsters, mice, and rats
7 after long-term inhalation of high concentrations of filtered and unfiltered diesel engine emissions. J Appl Toxicol
8 6:383-395.
9
10 Heinrich, U; Fuhst, R; Rittinghausen, S; et al. (1995) Chronic inhalation exposure of Wistar rats and two different
11 strains of mice to diesel engine exhaust, carbon black, and titantium dioxide. Inhal Toxicol 7:533-556.
12
13 Henderson, RF; Pickrell, JA; Jones, RK; et al. (1988) Response of rodents to inhaled diluted DE: biochemical and
14 cytological changes in bronchoalveolar lavage fluid and in lung tissue. Fundam Appl Toxicol 11:546-567.
15
16 Hyde, DM; Plopper, CG; Weir, AJ; et al. (1985) Peribronchiolar fibrosis in lungs of cats chronically exposed to DE.
17 Lab Invest 52: 195-206.
18
19 International Life Sciences Institute Risk Science Institute Workshop Participants (ILSI). (2000) The relevance of
20 the rat lung response to particle overload for human risk assessment: a workshop consensus report. In: ILSI Risk
21 Science Institute Workshop: the relevance of the rat lung response to particle overload for human risk assessment;
22 March, 1998. Gardner, DE., ed. Inhal Toxicol 12:1-17.
23
24 Ishinishi, N; Kuwabara, N; Nagase, S; et al. (1986) Long-term inhalation studies on effects of exhaust from heavy
25 and light duty diesel engines on F344 rats. In: Carcinogenic and mutagenic effects of diesel engine exhaust:
26 proceedings of the international satellite symposium on toxicological effects of emissions from diesel engines; July;
27 Tsukuba Science City, Japan. Ishinishi, N; Koizumi, A; McClellan, RO; et al., eds. Amsterdam, Holland: Elsevier
Science Publishers B. V; pp. 329-348. (Developments in toxicology and environmental science: v. 13).
30 Ishinishi, N; Kuwabara, N; Takaki, Y; et al. (1988) Long-term inhalation experiments on DE. In: DE and health
31 risks: results of the HERP studies. Tsukuba, Ibaraki, Japan: Japan Automobile Research Institute, Inc., Research
32 Committee for HERP Studies; pp. 11-84.
33
34 Iwai, K; Udagawa, T; Yamagishi, M; et al. (1986) Long-term inhalation studies of DE on F344 SPF rats. Incidence
35 of lung cancer and lymphoma. In: Ishinishi, N; Koizumi, A; McClellan, RO; et al., eds. Carcinogenic and mutagenic
36 effects of diesel engine exhaust: proceedings of the international satellite symposium on toxicological effects of
37 emissions from diesel engines; July; Tsukuba Science City, Japan. Amsterdam, Holland: Elsevier Science Publishers
38 B. V; pp. 349-360. (Developments in toxicology and environmental science: v. 13).
39
40 Kuempel, ED. (2000) Comparison of human and rodent lung dosimetry models for particle clearance and retention.
41 Drug Chem Toxicol 23: 203-222.
42
43 Lewis, TR; Green, FHY; Moorman, WJ; et al. (1989) A chronic inhalation toxicity study of diesel engine emissions
44 and coal dust, alone and combined. J. Am. Coll. Toxicol. 8:345-375.
45
46 Mauderly, JL; Jones, RK; Griffith, WC; et al. (1987a) DE is a pulmonary carcinogen in rats exposed chronically by
47 inhalation. Fundam. Appl. Toxicol. 9:208-221.
48
49 Mauderly, JL; Bice, DE; Carpenter, RL; et al. (1987b) Effects of inhaled nitrogen dioxide and DE on developing
50 lung. Cambridge, MA: Health Effects Institute; research report no. 8.
51
52 Morrow, PE. (1992) Dust overloading of the lungs: update and appraisal. Toxicol Appl Pharmacol 113: 1-12.
5L
National Research Council. (1983) Risk assessment in the federal government: managing the process. Washington,
DC: National Academy Press.
7/25/00 6-23 DRAFT—DO NOT CITE OR QUOTE
-------
1 Neas, LM; Dockery, DW; Koutrakis, P; et al. (1995) The association of ambient air pollution with twice daily peak
2 expiratory flow rate measurements in children. Am. J. Epidemiol. 141:111-122.
3
4 Nikula, KJ; Snipes, MB; Barr, EB; et al. (1995) Comparative pulmonary toxicities and carcinogenicities of
5 chronically inhaled DE and carbon black in F344 rats. Fundam Appl Toxicol 25:80-94.
6
7 OberdQrster, G. (1994) Extrapolation of results from animal inhalation studies with particles to humans? In: Toxic
8 and carcinogenic effects of solid particles in the respiratory tract: [proceedings of die 4th international inhalation
9 symposium]; March 1993; Hannover, Germany. Mohr, U; Dungworth, DL; Mauderly, JL; et al, eds. Washington,
10 DC: International Life Sciences Institute Press; pp. 335-353.
11
12 Pepelko, WE. (1982a) Effects of 28 days exposure to diesel engine emissions in rats. Environ Res 27:16-23.
13
14 Pepelko, WE. (1982b) EPA studies on the toxicological effects of inhaled diesel engine emissions. In: Toxicological
15 effects of emissions from diesel engines: proceedings of the Environmental Protection Agency diesel emissions
16 symposium; October 1981; Raleigh, NC. Lewtas, J., ed. New York: Elsevier Biomedical; pp. 121-142.
1 7 (Developments in toxicology and environmental science: v. 10).
18
1 9 Plopper, CG; Hyde, DM; Weir, AJ. (1983) Centriacinar alterations in lungs of cats chronically exposed to DE Lab
20 Invest 49:391-399.
21
22 Pope, CA, III; Thun, MJ; Namboodiri, MM; et al. (1995) Paniculate air pollution as a predictor of mortality in a
23 prospective study of U.S. adults. Am J Respir Crit Care Med 151:669-674.
24
25 Raizenne, M; Neas, LM; Damokosh, Al; et al. (1996) Health effects of acid aerosols on North American children:
26 pulmonary function. Environ Health Perspect 104:506-514.
27
28 Renwick, AG; Lazarus, NR. (1998) Human variability and noncancer risk assessment—an analysis of the default
29 uncertainty factor. Regul Toxicol Pharmacol 27:3-20.
30
31 Schwartz, J; Dockery, DW; Neas, LM. (1996) Is daily mortality associated specifically with fine particles? J Air
32 Waste Manage Assoc 46:927-939.
33
34 Thurston, GD; Ito, K; Hayes, CG; et al. (1994) Respiratory hospital admissions and summertime haze air pollution in
35 Toronto, Ontario: consideration of the role of acid aerosols. Environ Res 65:271-290.
36
37 U.S. Environmental Protection Agency (U.S. EPA). (1994) Methods for derivation of inhalation reference
38 concentrations and application of inhalation dosimetry [draft final]. Research Triangle Park, NC: Office of Health
39 and Environmental Assessment, Environmental Criteria and Assessment Office; report no. EPA/600/8-88/066F.
40
41 U.S. EPA. (1995a) Integrated Risk Information System (IRIS). Online. Office of Health and Environmental
42 Assessment, Environmental Criteria and Assessment Office, Cincinnati, OH.
43
44 U.S. EPA. (1995b) The use of the benchmark dcse approach in health risk assessment. Washington, DC: U.S. E?A,
45 Risk Assessment Forum; report no. EPA/630/R-94/007. Available from: NTIS, Springfield, VA;
46 PB95-2I3765/XAB.
47
48 U.S. EPA. (1996a) Air quality criteria for particulate matter. Research Triangle Park, NC: National Center for
49 Environmental Assessment-RTP Office; report nos. EPA/600/P-95/OOlaF-cF. 3v.
50
51 U.S. EPA. (1996b) Review of the national ambient air quality standards for particulate matter: policy assessment of
52 scientific and technical information. OAQPS staff paper. Research Triangle Park, NC: Office of Air Quality
53 Planning and Standards; report no. EPA/452/R-96-013. Available from: NTIS, Springfield. VA: PB97-115406REB.
54
7/25/00 6-24 DRAFT—DO NOT CITE OR QUOTE
-------
Vostal, JJ; Chan, TL; Garg, BD; et al. (1981) Lymphatic transport of inhaled diesel particles in the lungs of rats and
guinea pigs exposed to diluted DE. Environ Int 5:339-347.
Yu, CP; Xu, GB. (1986) Predictive models for deposition of DE particulates in human and rat lungs. Aerosol Sci
5 Technol 5:337-347.
6
7 Yu, CP; Yoon, KJ. (1990) Retention modeling of DE particles in rats and humans. Amherst, NY: State University of
8 New York at Buffalo (Health Effects Institute research report no. 40).
9
10 Yu, CP; Yoon, KJ; Chen, YK. (1991) Retention modeling of DE particles in rats and humans. J Aerosol Med
11 4:79-115.
7/25/00 6-25 DRAFT—DO NOT CITE OR QUOTE
-------
7. CARCINOGENICITY OF DIESEL EXHAUST
"l 7.1. INTRODUCTION
2 Initial health hazard concerns regarding the potential carcinogenicity of diesel exhaust
3 were based on the reported induction of skin papillomas by diesel particle extracts (Kotin et al.,
4 1955), evidence for mutagenicity of extracts (Huisingh et al., 1978), evidence that components of
5 diesel extract act as weak tumor promoters (Zamora et al., 1983), and the knowledge that diesel
6 particles and their associated organics are respirable. During the 1980s, both human
7 epidemiologic studies and long-term animal cancer bioassays were initiated. In 1981, Waller
8 published the first epidemiologic investigation, a retrospective mortality study of London
9 transport workers. Since then a large number of cohort and case-control studies have been
10 carried out with railroad workers, dockworkers, truck drivers, construction workers, miners, and
11 bus garage employees. During 1986 and 1987, several chronic animal cancer bioassays were
12 published. These studies and numerous laboratory investigations carried out since then have
13 been directed toward assessing the carcinogenic potential of whole exhaust, evaluating the
14 importance of various exhaust components in the induction of cancer, and understanding the
15 mode of action and implications of deposition, retention, and clearance of diesel exhaust
particles.
18 7.1.1. Overview
.19 This chapter evaluates the carcinogenic potential of diesel exhaust in both humans
20 (Section 7.2) and animals (Section 7.3), determines likely mode/s of action (Section 7.4), and
21 provides an overall weight of evidence (Section 7.5) for carcinogenicity in humans. This
22 assessment focuses on diesel exhaust, although diesel particles comprise a portion of ambient
23 paniculate matter (PM). In 1998, diesel emissions constituted 72% (521,000 tons) of mobile
24 sources PM)0 and 18% of total PM10 in ambient air (excluding natural and miscellaneous
25 emissions). Diesel emissions made up 77% (473,000 tons) of mobile source PM2.5 emissions,
26 and 23% of total PM25 in ambient air (excluding natural and miscellaneous emissions) in 1998.
27 Ambient PM, notably PM10, has been known for many years to potentially affect human health;
28 these effects have been evaluated in a separate document (U.S. EPA, 1996a). This document is
29 also undergoing revision.
30
31
7/25/00 7-1 DRAFT—DO NOT CITE OR QUOTE
-------
1 7.1.2. Ambient PM-Lung Cancer Relationships
2 A quick overview of the data regarding exposure to ambient PM and lung cancer is
3 provided as background information. With DE being part of ambient PM, the question of what is
4 seen in the ambient PM data is of interest, since insight about the ambient PM exposure lung
5 cancer relationships may contribute to evaluation of DE-specific epidemiologic data.
6 Chapter 5 noted that (a) DPM, consisting mostly of fine particles (<1.0 urn diameter),
7 represents a lexicologically important component of typical ambient fine particle mixes, and (b)
8 health risk estimates for ambient fine particles would, logically, likely represent an upper limit
9 for estimates of the health risks associated with exposures to DPM as a subset of ambient fine
10 PM. Chapter 5 (and Appendix C) went on to summarize key epidemiologic findings from
11 studies of ambient PM noncancer effects, which provided important inputs to the setting, in
12 1997, of new ambient fine particle standards (PM2.5 NAAQS) to protect against mortality and
13 morbidity effects of airborne fine particles. Several large-scale prospective studies (Harvard Six
14 City Study; American Cancer Society or ACS Study; Adventist Health Study of Smog or
15 AHSMOG) were highlighted in Chapter 5 and Appendix C as providing important evidence
16 regarding associations between chronic exposures to ambient fine particles and increased risks of
17 noncancer mortality/morbidity effects (e.g., cardiorespiratory-related deaths or hospital
18 admissions). The same studies also evaluated relationships between chronic PM exposures and
19 lung cancer mortality and/or incidence, evaluations of much pertinence here to consideration of
20 ambient PM cancer risks as possibly representing upper limits for DPM-related cancer risks.
21 The Harvard Six City Study (Dockery et al., 1993), of approximately 8,000 adults in six
22 cities comprising a transect across the northcentral and northeastern United States, found
23 markedly increased relative risks (RR) of lung cancer mortality for current (RR = 8.00, 95% CL
24 2.97-21.6) and former (RR = 2.54, CL 0.90-7.18) smokers. Also, elevated but statistically
25 nonsignificant associations of lung cancer mortality risks (RR = 1.37, CL 0.81 -2.31) were found
26 by the Six City Study analyses (which included data for both males and females) to be related to
27 ambient fine particles indexed by a range of annual mean PM2 5 concentrations from the least to
28 the most polluted of the six cities.
29 The ACS Study (Pope et al., 1995), of 550,000 adults in 151 cities across all U.S.
30 geographic regions, also found markedly elevated lung cancer risks for current smokers (RR =
31 9.73, CL 5.96-15.9) and somewhat elevated and statistically significant lung cancer risk (RR =
32 1.36, CL 1.11-1.66) associations with a range (19.9 mg/m3) of annual average sulfate (SO4)
33 concentrations as one index of chronic exposures to ambient fine particles, in combined analyses
34 of data for both maies and females. However, in further analyses of subgroups broken out by sex
3b and smoking status ^and thus having smaller sample sizes in each than for the above overall
5/00 7-2 DRAFT—DO NOT CITE OR QUOTE
-------
combined analyses), only the lung cancer mortality risks for male "ever-smokers" (RR = 1.44,
CL 1.14-1.83) were statistically significant in relation to sulfate concentrations as the fine
3 particle indicator in the 151 cities. Note that the analogous adjusted risk ratios (and 95% CL) for
4 the most polluted versus least polluted cities in terms of sulfate levels were statistically
5 nonsignificant for male "never-smokers" (RR = 1.36, CL 0.40 - 4.66), for female "ever-smokers"
6 (RR =1.10, CL 0.72-1.68) and female "never-smokers" (RR = 1.61; CL 0.66 -3.92). Also, lung
7 cancer mortality risks (RR = 1.03; CL 0.80-1.33) were not statistically significantly associated
8 with ambient PM25 concentrations (across a range of 24.5 ng/m3 from least to most polluted of a
9 subset of 50 of the 151 cities) in overall combined analyses of data for both males and females.
10 Nor were the relative risk ratios statistically significant for smaller sample size subgroups broken
11 out by sex and smoking status in relation to PM2.5 concentrations, as a second index of ambient
12 airborne fine PM. Hence, the ACS Study provides only very limited evidence hinting at a
13 possible lung cancer mortality association with one indicator (sulfates) of ambient fine particles,
14 but not with another such index (PM2 s).
15 In the first of an ongoing series of reports on AHSMOG data analyses, Abbey et al.
16 (1991) described the results of initial analyses related to the AHSMOG evaluation of air
17 pollution effects on the health of 6,338 nonsmoking, long-term California adult residents. Of a
variety of health endpoints evaluated, only respiratory symptoms and female cancers (any site)
but not respiratory cancer for either sex, were reported by Abbey et al. (1991) to be associated
20 with concentrations of total suspended particulate (TSP) matter (which includes not only fine
21 particles indexed by PM25 but also larger coarse mode particles ranging up to 25-50 mm). Later
22 follow-up analyses (Abbey et al., 1995) considered chronic exposures to PM,0 (estimated from
23 TSP data), PM2.5 (estimated from visibility data), and SO4, but found no statistically significant
24 associations with nonexternal mortality. Subsequent AHSMOG analyses reported out by Abbey
25 et al. (1999) and Beesan et al. (1998) do, however, hint at possible associations of increased risk
26 of lung cancer mortality and/or incidence in males with ambient PM exposures. More
27 specifically, chronic exposures to ambient PM,0 (which includes both fine particle <2.5 um and
28 coarse particles 2.5 to 10 um in size) were reported to be significantly associated with markedly
29 increased lung cancer mortality risks in the nonsmoking AHSMOG males (RR = 23.39, CL 2.55-
30 60.10), but not for the females (RR = 9.8; CL 0.34-9.52). Male lung cancer mortality was also
31 reported to be significantly associated with numbers of days per year that PM10 exceeded 100
32 mg/m3 (RR = 1.055, CL 0.66-1.69). Other analyses of AHSMOG data were reported by Beeson
33 et al. (1998) also showing statistically significant associations of increased lung cancer incidence
34 (especially PM,0 > 100 ng/m3) for males, but not for females.
7/25/00 7-3 DRAFT—DO NOT CITE OR QUOTE
-------
1 In summary, the three key prospective cohort studies (discussed hi more detail hi
2 Appendix C) provide an equivocal array of results with regard to possible associations between
3 chronic exposures to ambient PM and lung cancer mortality and/or incidence. Only the ACS
4 Study found a statistically significant association of increased risk of lung cancer with one
5 indicator of ambient fine particles (sulfates), but not with another such indicator (PM2.5)-the
6 latter being consistent with Harvard Six City Study results for PM2.5. Also, the AHSMOG results
7 hint at possible increased lung cancer risks hi males, but not females, hi relation to PM10 levels.
8 Overall, then, these studies are not conclusive and appear, at best, only to provide some
9 indication of possible associations between increased lung cancer risk and chronic ambient fine
10 PM exposures.
11
12 7.2. EPIDEMIOLOGIC STUDIES OF THE CARCINOGENICITY OF EXPOSURE TO
13 DIESEL EXHAUST
14 An increased risk from malignancies of the lung, bladder, and lymphatic tissue has been
15 reported in populations potentially exposed to diesel emissions. Isolated authors have reported
16 other malignancies, including testicular cancer (Garland et al., 1988), gastrointestinal cancer
17 (Balarajan and McDowall, 1988; Guberan et al., 1992), and prostate cancer (Aronsen et al.,
18 1996). A detailed review of lung cancer studies is presented in this section. A detailed review of
19 other health effect studies is not presented because findings are equivocal.
20 Excess risk of bladder cancer has been reported in several studies (Howe et al., 1980;
21 Wynder et al., 1985; Hoar and Hoover et al., 1985; Silverman et al., 1983; Vineis and Magnani
22 1985; Silverman et al., 1986; Jensen et al., 1987; Steenland et al., 1987; Isocovich et al., 1987;
23 Risen et al., 1988; Iyer et al., 1990; Steineck et al., 1990; Cordier et al., 1993; Notani et al.,
24 1993). Very few studies found significant excesses after adjustment for cigarette smoking. Most
25 studies failed to show any association between exposure to diesel exhaust and occurrence of
26 bladder cancer. Some authors have reported excess mortality from lymphohematopoietic system
27 cancers in people potentially exposed to diesel fumes. Rushton and Alderson (1983) and Howe
28 and Lindsay (1983) found increased mortality from lymphatic neoplasms. Balarajan and
29 McDowall (1983) found raised mortality for malignant lyrnphornas. Flcdin et al. (1987)
30 observed increased risk for multiple myeloma, and Bender et al. (1989) reported excess mortality
31 from leukemia. Because evidence for bladder cancer and lymphohematopoietic cancer was
32 found to be equivocal, detailed reviews of these studies are not presented here.
33 In this section, various mortality and morbidity studies of lung cancer from potential
34 exposure to diesel engine emissions are reviewed. Although an attempt was made to cover all
35 the relevant aiuuics, a. number of studies are not included lor several reasons. The change from
7/25/00 7-4 DRAFT—DO NOT CITE OR QUOTE
-------
steam to diesel engines in locomotives began after World War II. By 1946 about 10% of the
locomotives in service were diesel, by 1952 55% were diesel, and dieselization was about 95%
"3 complete by 1959 (Garshick et al., 1988). Therefore, exposure to diesel exhaust was less
4 common, and the follow-up period for studies conducted prior to 1960 (Raffle, 1957; Commins
5 et al., 1957; Kaplan, 1959) was not long enough to cover the long latency period of lung cancer.
6 The usefulness of these studies in evaluating the carcinogenicity of diesel exhaust is greatly
7 reduced; thus, they are not considered here.
8 On the other hand, the trucking industry changed to diesel trucks by the 1960s. In the
9 1960s sales of diesel-powered Class 8 trucks (long-haul trucks) were 48% of the market, and by
10 the 1970s sales had risen to 85%. Thus, studies conducted among truck drivers prior to the
11 1970s may reflect exposures to gasoline exhaust as well as diesel exhaust. Hence, studies with
12 ambiguous exposures or studies that examined several occupational risk factors were excluded
13 because they would have contributed little to the evaluation of the carcinogenicity of diesel
14 exhaust (Waxweiler et al., 1973; Williams et al., 1977; Ahlberg et al., 1981; Stem et al., 1981;
15 Buiatti et al., 1985; Gustafsson et al., 1986; Siemiatycki et al., 1988). A study by Coggon et al.
16 (1984) was excluded because occupational information abstracted from death certificates had not
17 been validated; this would have resulted in limited information.
Several types of studies of the health effects of exposure to diesel engine emissions are
reviewed in this chapter, such as cohort studies, case-control studies, and studies that conducted
20 meta-analysis. In the cohort studies, cohorts of heavy construction equipment operators, railroad
21 and locomotive workers, bus garage employees, and miners were studied retrospectively to
22 determine increased mortality and morbidity resulting from exposures to varying levels of diesel
23 emissions in the workplace. The evaluation of each study presents the study population,
24 methodology used for the study, i.e., data collection and verification, analysis, results, and a
25 critique of the study. There are some methodologic limitations that are common to studies with
26 similar design. The total evidence, including limitations, is discussed at the end of the chapter in
27 the summary and discussion section.
28
29 7.2.1. Cohort Studies
30 7.2.1.1. Waller (1981): Trends in Lung Cancer in London in Relation to Exposure to Diesel
31 Fumes
32 A retrospective mortality study of a cohort of London transport workers was conducted to
33 determine if there was an excess of deaths from lung cancer that could be attributed to diesel
34 exhaust exposure. From nearly 20,000 male employees in the early years, those aged 45 to 64
^P were followed for the 25-year period between 1950 and 1974 (the actual number of employees is
7/25/00 7-5 DRAFT—DO NOT CITE OR QUOTE
-------
1 not given in the paper), constituting a total of 420,700 man-years at risk. These workers were
2 distributed among five job categories: drivers, garage engineers, conductors, motormen or
3 guards, and engineers (works). Lung cancer were ascertained from death certificates of
4 individuals who died while still employed, or if retired, following diagnosis. Expected death
5 rates were calculated by applying greater London death rates to the population at risk within each
6 job category. Data were calculated in 5-year periods and 5-year age ranges, and the results were
7 combined to obtain the total expected deaths in the required age range for the calendar period. A
8 total of 667 cases of lung cancer was reported, compared with 849 expected, to give a cancer
9 mortality ratio of 79%. In each of the five job categories, the observed numbers were below
10 those expected. Engineers in garages had the highest mortality ratio, 90%, motormen and guards
11 had a mortality ratio of 87%, and both the bus drivers and conductors had mortality ratios of
12 75%. The engineers in the central works had a mortality ratio of 66%. These mortality ratios did
13 not differ significantly from each other. Environmental sampling was done at one garage, on one
14 day in 1979, for benzo[a]pyrene concentrations and was compared with corresponding values
15 recorded in 1957. Concentrations of benzo[a]pyrene recorded in 1957 were at least 10 times
16 greater than those measured in 1979.
17 This study failed to find any association between diesel exhaust and occurrence of lung
18 cancer, which may be due to several methodologic limitations. The lung cancer deaths were
19 ascertained while the workers were employed (the worker either died of lung cancer or retired
20 after lung cancer was diagnosed). Although man-years at risk were based on the entire cohort,
21 no attempt was made to trace or evaluate the individuals who had resigned from the London
22 transport company for any other reason. Hence, information on resignees who may have had
23 significant exposure to diesel exhaust, and on lung cancer deaths among them, was not available
24 for analysis. This may have led to a dilution effect, resulting in underascertainment of observed
25 lung cancer deaths and underestimation of mortality ratios. Eligibility criteria for inclusion in the
26 cohort, such as starting date and length of service with the company, were not specified.
27 Therefore, there may not have been sufficient latency for the development of lung cancer. Use of
28 greater London population death rates to obtain expected number of deaths may have resulted in
29 a deficit in mortality ratios reflecting the "healthy worker effect." Investigators did not
30 categorize the five job categories either by qualitative or quantitative levels of diesel exhaust
31 exposure; neither did they use an internal comparison group to derive risk estimates.
32 The age range considered for this study was limited (45 to 64 years of age) for the period
33 between 1950 and 1974. It is not clear whether this age range was applied to calendar year 1950
34 or 1974, or ai trie iriiupoint of Lhc 25-year follow-up period. No analyses were presented either
35 by latency or by duration of employment (surrogate for exposure). Tne environmental survey
7/oc/nn 7-£ r>PAPT T-\/~\ XT/-VT
-------
based on benzo[a]pyrene concentrations suggests that the cohort in its earlier years was exposed
to much higher concentrations of environmental contaminants than currently exist. It is not clear
3 when the reduction in benzo[a]pyrene concentration occurred, because there are no
4 environmental readings available between 1957 and 1979. It is also important to note that the
5 concentrations of benzo[a]pyrene inside the garage in 1957 were not very different from those
6 outside the garage, thus indicating that exposure for garage workers was not much different from
7 that of the general population. Thus, this study fails to provide any negative association between
8 the diesel exhaust exposure and the occurrence of lung cancer.
9
10 7.2.1.2. Howe et al (1983): Cancer Mortality (1965 to 1977) in Relation to Diesel Fumes and
11 Coal Exposure in a Cohort of Retired Railroad Workers
12 This is a retrospective cohort study of the mortality experience of 43,826 male pensioners
13 of the Canadian National Railroad (CNR) between 1965 and 1977. Members of this cohort
14 consisted of male CNR pensioners who had retired before 1965 and who were known to be alive
15 at the start of that year, as well as those who retired between 1965 and 1977. The records were
16 obtained from a computer file that is regularly updated and used by the company for payment of
17 pensions. To receive a pension, each pensioner must provide, on a yearly basis, evidence that he
is alive. Specific cause of death among members of this cohort was ascertained by linking these
records to the Canadian Mortality Data Base, which contains records of all deaths registered in
20 Canada since 1950. Of the 17,838 deaths among members of the cohort between 1965 and 1977,
21 16,812 (94.4%) were successfully linked to a record in the mortality file. A random sample
22 manual check on unlinked data revealed that failure to link was due mainly to some missing
23 information on the death records.
24 Occupation at time of retirement was used by the Department of Industrial Relations to
25 classify workers into three diesel fume and coal dust exposure categories: (1) nonexposed, (2)
26 possibly exposed, and (3) probably exposed. Person-years of observation were calculated and
27 classified by age at observation in 5-year age groups (35 to 39,40 to 44,..., 80 to 84, and ^85
28 years). The observed deaths were classified by age at death for different cancers, for all cancers
29 combined, and for all causes of death combined. Standard mortality ratios (SMRs) were then
30 calculated using rates of the Canadian population for the period between 1965 and 1977. The
31 relative risks were calculated using the three exposure categories: nonexposed, possibly exposed,
32 and probably exposed.
33 Both total mortality (SMR = 95, pO.OOl) and all cancer deaths (SMR = 99, /?>0.05)
34 were close to that expected for the entire cohort. Analysis by exposure to diesel fume levels in
^f the three categories (nonexposed, possibly exposed, and probably exposed) revealed an increased
7/25/00 7-7 DRAFT—DO NOT CITE OR QUOTE
-------
1 relative risk for lung cancer among workers with increasing exposure to diesel fumes. The
2 relative risk for nonexposed workers was presumed to be 1.0; for those possibly exposed, the
3 relative risk was significantly elevated to 1.2 (p=0.013); and for those probably exposed, it was
4 significantly elevated to 1.35 (p=0.001). The corresponding rates for exposure to varying levels
5 of coal dust were very similar at 1.00, 1.21 (p=0.012), and 1.35 (p=0.001), respectively. The
6 trend tests were highly significant for both exposures (p<0.001). Analysis performed after the
7 exclusion of individuals who worked in the maintenance of steam engines, and hence were
8 exposed to high levels of asbestos, yielded a risk of lung cancer of 1.00, 1.21, and 1.33 for those
9 nonexposed, possibly exposed, and probably exposed to diesel exhaust, respectively, with a
10 highly significant trend (p<0.001).
11 An analysis done on individuals who retired prior to 1950 showed the relative risk of lung
12 cancer among nonexposed, possibly exposed, and probably exposed to be 1.00,0.70, and 0.44,
13 respectively, based on fewer than 15 deaths in each category. A similar analysis of individuals
14 who retired after 1950 found the results in the same categories to be 1.00,1.23, and 1.40,
15 respectively. Although retirement prior to 1950 indicated exposure to coal combustion fumes
16 alone, retirement after 1950 shows the results of mixed exposure to coal combustion fumes and
17 diesel fumes. As there was considerable overlap between occupations involving probable
18 exposure to diesel fumes and probable exposure to coal, and as most members of the cohort were
19 employed during the years in which the transition from coal to diesel occurred, it was difficult to
20 distinguish whether lung cancer was associated with exposure to coal combustion fumes or diesel
21 fumes or a mixture of both.
22 Although this study showed a highly significant dose-response relationship between
23 diesel fumes and lung cancer, it has some methodological limitations. There were concurrent
24 exposures to both diesel fumes and coal combustion fumes during the transition period;
25 therefore, misclassification of exposure may have occurred, because only occupation at
26 retirement was available for analysis. It is possible that the elevated response observed for lung
27 cancer was due to the combined effects of exposure to both coal dust/coal combustion products
23 and diesel fumes and riot just one or the other. However, deaths due to lung cancer were not
29 elevated among workers who retired prior to the 1950s and thus would have been primarily
30 exposed to coal dust/coal combustion products. Furthermore, it should be noted that so far coal
31 dust has not been demonstrated to be a pulmonary carcinogen in studies of coal miners. This
32 study was restricted to deaths among retired workers; therefore, it is unclear if a worker who
33 developed lung cancer when actively employed and filed for a disability claim instead of
34 retirement claim would be included in the study or not. Thus, it is possible that workers with
3t> heavy exposure might have been excluded from the study. Neither information on duration of
7/25/00 7-8 DRAFT—DO NOT CITE OR QUOTE
-------
employment in diesel work, nor coal dust-related jobs other than those held at retirement, nor
details of how the exposure categories were created was provided. Therefore, it was not possible
to evaluate whether this omission would have led to an under- or overestimate of the true relative
4 risk. Although information on potential confounders such as smoking is lacking, the use of an
5 internal comparison group to compute the relative risks minimizes the potential for confounding
6 by smoking, as there is no reason to assume different smoking patterns among individuals
7 exposed to diesel exhaust versus those not exposed. Despite these limitations, this study
8 provides suggestive evidence toward a causal association between exposure to diesel exhaust and
9 excess lung cancer.
10
11 7.2.1.3. Rushton et aL (1983): Epidem iological Survey of Maintenance Workers in the
12 London Transport Executive Bus Garages and Chiswick Works
13 This is a retrospective mortality cohort study of male maintenance workers employed for
14 at least 1 continuous year between January 1, 1967, and December 31, 1975, at 71 London
15 transport bus (also known as rolling stock) garages and at Chiswick Works. The following
16 information was obtained from computer listings: surname with initials, date of birth, date of
17 joining company, last or present job, and location of work. For those individuals who left their
[8 job, date of and reason for leaving were also obtained. For those who died in service or after
retirement, and for men who had resigned, full name and last known address were obtained from
20 an alphabetical card index in the personnel department. Additional tracing of individuals who
21 had left was carried out through social security records. The area of residence was assumed to be
22 close to their work; therefore place of work was coded as residence. One hundred different job
23 titles were coded into 20 broader groups. These 20 groups were not ranked for diesel exhaust
24 exposure, however. The reason for leaving was coded as died in service, retired, or other. The
25 underlying cause of death was coded using the eighth revision of the International Classification
26 of Diseases (ICD). Person-years were calculated from date of birth and dates of entry to and exit
27 from the study using the man-years computer language program. The workers were then
28 subdivided into 5-year age and calendar period groups. The expected number of deaths was
29 calculated by applying the 5-year age and calendar period death rates of the comparison
30 population with the person-years of corresponding groups. The mortality experience of the male
31 population in England and Wales was used as the comparison population. Significance values
32 were calculated for the difference between the observed and expected deaths, assuming a Poisson
33 distribution.
34 The person-years of observation totaled 50,008 and were contributed by 8,490 individuals
in the study, with a mean follow-up of 5.9 years. Only 2.2% (194) of the men were not traced.
7/25/00 7-9 DRAFT—DO NOT CITE OR QUOTE
-------
1 Observed deaths from all causes were significantly lower than expected (O = 495, /K0.001).
2 Observed deaths from all neoplasms and cancer of the lung were approximately the same as
3 those expected. The only significant excess observed, for cancer of the liver and gall bladder at
4 Chiswick Works, was based on four deaths (p<0.05). A few job groups showed a significant
5 excess of risks for various cancers. All the excess deaths observed for the various job groups,
6 except for the general hand category, were based on very small numbers (usually fewer than five)
7 and merited cautious interpretation. Only a notable excess in the general hand category for lung
8 cancer was based on as many as 48 cases (SMR = 133,/K0.03).
9 This mortality study did not demonstrate any cancer excess. Details of work history were
10 not obtained to permit any analysis by diesel exhaust exposure. The study's limitations,
11 including small sample size, short duration of follow-up (average of only 6 years), and lack of
12 sufficient latency period, make it inadequate to draw any conclusions.
13
14 7.2.1.4. Wong et aL (1985): Mortality Among Members of a Heavy Construction Equipment
15 Operators Union With Potential Exposure to Diesel Exhaust Emissions
16 This retrospective mortality study was conducted on a cohort of 34,156 male members of
17 a heavy construction equipment operators union with potential exposure to diesel exhaust
18 emissions. Study cohort members were identified from records maintained at Operating
19 Engineers' Local Union No. 3-3A in San Francisco, CA. This union has maintained both work
20 and death records on all its members since 1964. Individuals with at least 1 year of membership
21 in this union between January 1, 1964, and December 31, 1978, were included in the study.
22 Work histories of the cohort were obtained from job dispatch computer tapes. The study foliow-
23 up period was January 1964 to December 1978. Death information was obtained from a trust
24 fund, which provided information on retirement dates, vital status, and date of death for those
25 who were entitled to retirement and death benefits. Approximately 50% of the cohort had been
26 union members for less than 15 years, whereas the other 50% had been union members for 15
27 years or more. The average duration of membership was 15 years. As of December 31, 1978,
23 29,046 (85%) cohort members were alive, 3,345 (9.8%) were dead, and 1,765 (5.2%) remained
29 untraced. Vital status of 10,505 members who had left the union as of December 31, 1978, was
30 ascertained from the Social Security Administration. Death certificates were obtained from
31 appropriate State health departments. Altogether, 3,243 deaths (for whom death certificates were
32 available) in the cohort were coded using the seventh revision of the ICD. For 102 individuals,
33 death certificates could not be obtained, only the date of death; these individuals were included in
34 the calculation of the SMR for ail causes of death but were deieted from the cause-specific SMR
35 analyses, hxpected deaths and SMKs were calculated using the U.S. national age-sex-race
7/25/00 7-10 DRAFT—DO NOT CITE OR QUOTE
-------
cause-specific mortality rates for 5-year time periods between 1964 and 1978. The entire cohort
population contributed to 372,525.6 person-years in this 5-year study period.
A total of 3,345 deaths was observed, compared with 4,109 expected. The corresponding
4 SMR for all causes was 81 (p=0.01), which is consistent with the "healthy worker effect." A
5 total of 817 deaths was attributed to malignant neoplasms, slightly fewer than the 878 expected
6 based on U.S. white male cancer mortality rates (SMR = 93, /7=f0.05). Mostly there were SMR
7 deficits for cause-specific cancers, including lung cancer for the entire cohort (SMR = 99, O =
8 309). The only significant excess SMR was observed for cancer of the liver (SMR = 167, O =
9 23,/?<0.05).
10 Analysis by length of union membership as a surrogate of duration for potential exposure
11 showed statistically significant increases in SMRs of cancer of the liver (SMR = 424,/?<0.01) in
12 the 10- to 14-year membership group and of the stomach (SMR = 248,/?<0.05) in the 5- to 9-
13 year membership group. No cancer excesses were observed in the 15- to 19-year and 20+-year
14 membership groups. Although the SMR for cancer of the lung had a statistically significant
15 deficit in the less-than-5-year duration group, it showed a positive trend with increasing length of
16 membership, which leveled off after 10 to 14 years.
17 Cause-specific mortality analysis by latency period showed a positive trend for SMRs of
18 all causes of death, although all of them were statistically significant deficits, reflecting the
^B diminishing "healthy worker effect." This analysis also demonstrated a statistically significant
20 SMR excess for cancer of the liver (10- to 19-year group, SMR = 258). The SMR for cancer of
21 the lung showed a statistically significant deficit for a <10-year latency but showed a definite
22 positive trend with increasing latency.
23 In addition to these analyses of the entire cohort, similar analyses were carried out in
24 various subcohorts. Analyses of retirees, 6,678 individuals contributing to 32,670 person-years,
25 showed statistically significant increases (pO.Ol) in SMRs for all cancers; all causes of death;
26 cancers of the digestive system, large intestine, respiratory system, and lung; emphysema; and
27 cirrhosis of the liver. The other two significant excesses (p<0.01) were for lymphosarcoma and
28 reticulosarcoma and nonmalignant respiratory diseases. Further analysis of the 4,075 retirees
29 (18,678 person-years) who retired at age 65 or who retired earlier but had reached the age of 65
30 revealed statistically significant SMR increases (p<0.05) for all cancers, cancer of the lung, and
31 lymphosarcoma and reticulosarcoma.
32 To analyze cause-specific mortality by job held (potential exposure to diesel exhaust
33 emissions), 20 functional job titles were used, which were further grouped into three potential
34 categories: high exposure, low exposure, and unknown exposure. A person was classified in a
job title if he ever worked on that job. Based on this classification system, if a person had ever
7/25/00 7-11 DRAFT—DO NOT CITE OR QUOTE
-------
1 worked in a high-exposure job title he was included in that group, even though he may have
2 worked for a longer time in a low-exposure group or in an unknown exposure group.
3 Information on length of work in any particular job, hence indirect information on potential
4 length of exposure, was not available either.
5 For the high-exposure group a statistically significant excess was observed for cancer of
6 the lung among bulldozer operators who had 15 to 19 years of membership and 20+ years of
7 follow-up (SMR = 343, /K0.05). This excess was based on 5 out of 495 deaths observed in this
8 group of 6,712 individuals, who contributed 80,328 person-years of observation.
9 The cause-specific mortality analysis in the low-exposure group revealed statistically
10 significant SMR excesses in individuals who had ever worked as engineers. These excesses were
11 for cancer of the large intestine (SMR = 807, O = 3,/><0.05) among those with 15 to 19 years of
12 membership and length of follow-up of at least 20 years, and cancer of the liver (SMR = 872, O
13 = 3,/?<0.05) among those with 10 to 14 years of membership and length of follow-up of 10 to 19
14 years. There were 7,032 individuals who contributed to 78,403 person-years of observation in
15 the low-exposure group.
16 For the unknown exposure group, a statistically significant SMR was observed for motor
17 vehicle accidents only (SMR = 174,0 = 21, p<0.05). There were 3,656 individuals who
18 contributed to 33,388 person-years of observation in this category.
19 No work histories were available for those who started their jobs before 1967 and for
20 those who held the same job prior to and after 1967. This group comprised 9,707 individuals
21 (28% of the cohort) contributing to 104,448 person-years. Statistically significant SMR excesses
22 were observed for all cancers (SMR = 112, O = 339, p<0.05) and cancer of the lung (SMR =119,
23 O = 141, /?<0.01). A significant SMR elevation was also observed for cancer of the stomach
24 (SMR =199, O = 30,p<0.01).
25 This study demonstrates a statistically significant excess for cancer of the liver but also
26 shows statistically significant deficits in cancers of the large intestine and rectum. It may be, as
27 the authors suggested, that the liver cancer cases actually resulted from metastases from the large
28 intestine and/or rectum, as tumors of these sites will frequently metastasize to the liver. The
29 excess in liver cancer mortality and the deficits in mortality that are due to cancer of the large
30 intestine and rectum could also, as the authors indicate, be due to misclassification. Both
31 possibilities have been considered by the investigators in their discussion.
3 2 Cancer of the lung showed a positive trend with length of membership as well as with
33 latency, although none of the SMRs were statistically significant except for workers without any
34 WGik hibturicb. The individuals wilhoui any work histories may have been the ones who were in
35 Lhcir jobs fur the longest period of time, because workers whhout job histories included those
7/25/00 7-12 DRAFT—DO NOT CITE OR QUOTE
-------
who had the same job before and after 1967 and thus may have worked 12 to 14 years or longer.
If they had belonged to the category in which heavy exposure to diesel exhaust emissions was
3 very common for this prolonged time, then the increase in lung cancer, as well as stomach
4 cancer, might be linked to diesel exhaust. Further information on those without work histories
5 should be obtained if possible, because such information may be quite informative with regard to
6 the evaluation of the carcinogenicity of diesel exhaust.
7 The study design is adequate, covers about a 15-year observation period, has a large
8 enough population, and is appropriately analyzed; however, it has too many limitations to permit
9 any conclusions. First, no exposure histories are available; one has to make do with job histories,
10 which provide limited information on exposure level. Any person who ever worked at the job, or
11 any person working at the same job over any period of time, is included in the same category;
12 this would have a dilution effect, because extremely variable exposures were considered in the
13 study. Second, the length of time worked in any particular job is not available. Third, work
14 histories were not available for 9,707 individuals, who contributed 104,448 person-years, a large
15 proportion of the study cohort (28%). These individuals happen to show the most evidence of a
16 carcinogenic effect. Confounding by alcohol consumption for cancer of the liver and smoking
17 for emphysema and cancer of the lung was not ruled out. Fourth, 15 years' follow-up may not
provide sufficient latency to observe excess lung cancer. Last, although 34,156 members were
eligible for the study, the vital status of 1,765 individuals was unknown. Nevertheless, they were
20 still considered in the denominator of all the analyses. The investigators fail to mention how the
21 person-year calculation for these individuals was handled. Also, some of the person-years might
22 have been overestimated, as people may have paid the dues for a particular year and then left
23 work. These two causes of overestimation of the denominator may have resulted in some or all
24 the SMRs being underestimated.
25
26 7.2.1.5. Edling et al. (1987): Mortality Among Personnel Exposed to Diesel Exhaust
27 This retrospective cohort mortality study of bus company employees investigated a
28 possible increased mortality of cardiovascular diseases and cancers from diesel exhaust exposure.
29 The cohort comprised all males employed at five different bus companies in southeastern
30 Sweden between 1950 and 1959. Based on information from personnel registers, individuals
31 were classified into one or more categories and could have contributed person-years at risk in
32 more than one exposure category. The study period was from 1951 to 1983; information was
33 collected from the National Death Registry, and copies of death certificates were obtained from
34 the National Bureau of Statistics. Workers who died after age 79 were excluded from the study
because diagnostic procedures were likely to be more uncertain at higher ages (according to
7/25/00 7-13 DRAFT—DO NOT CITE OR QUOTE
-------
1 investigators). The cause-, sex-, and age-specific national death rates in Sweden were applied to
2 the 5 -year age categories of person-years of observation to determine expected deaths for all
3 causes, malignant diseases, and cardiovascular diseases. A Poisson distribution was used to
4 calculate p- values and confidence limits for the ratio of observed to expected deaths. The total
5 cohort of 694 men (after loss of 5 men to follow-up) was divided into three exposure categories:
6 (1) clerks with lowest exposure, (2) bus drivers with moderate exposure, and (3) bus garage
7 workers with highest exposure.
8 The 694 men provided 20,304 person-years of observation, with 195 deaths compared
9 with 237 expected. A deficit in cancer deaths largely accounted for this lower-than-expected
1 0 mortality in the total cohort. Among subcohorts, no difference between observed and expected
1 1 deaths for total mortality, total cancers, or cardiovascular causes was observed for clerks (lowest
1 2 diesel exposure), bus drivers (moderate diesel exposure), and garage workers (high diesel
1 3 exposure). The risk ratios for all three categories were less than 1 except for cardiovascular
1 4 diseases among bus drivers, which was 1.1.
1 5 When the analysis was restricted to members who had at least a 1 0-year latency period
1 6 and either any exposure or an exposure exceeding 10 years, similar results were obtained, with
1 7 fewer neoplasms than expected, whereas cardiovascular diseases showed risk around or slightly
1 8 above unity.
1 9 Five lung cancer deaths were observed among bus drivers who had moderate diesel
20 exhaust exposure, whereas seven were expected. The only other lung cancer death was observed
2 1 among bus garage workers who had the highest diesel exhaust exposure. This study's major
22 limitations, including small size and poor data on diesel exhaust exposure, make it inadequate to
23 draw any conclusions.
24
2 5 7.2.1.6. Boffetta and Stellman (1988): Diesel Exhaust Exposure and Mortality Among Males
2 6 in the American Cancer Society Prospective Study
27 Boffetta and Stellman conducted a mortality analysis of 461,981 males with known
28 smoking history and vital status at the end of the first 2 years of follow-up. The analysis was
29 restricted to males aged 40 to 79 years in 1982 who enrolled in the American Cancer Society's
30 prospective mortality study of cancer. Mortality was analyzed in relation to exposure to diesel
3 1 exhaust and to employment in selected occupations related to diesel exhaust exposure. In 1982,
32 more than 77,000 American Cancer Society volunteers enrolled more than 1.2 million men and
33 women from all 50 States, the District of Columbia, and Puerto Rico in a long-term cohort study,
ancer Prevention Study II (CPS-H). Enrcllees \vere usually friends, neighbors, or relatives
the C
_r»l -- --- 1 --- 4. ----- ---- 11 - „_<. ,,,„„ U,, -rn«*;Ur ~,^,,,~r- ,,,;*U „+ 1ao r-t ^n« **» «~w ~~v **u V.»»V*H .~
7/25/00 7-14 DRAFT— DO NOT CITE OR QUOTE
-------
1 years of age or older. Subjects were asked to fill out a four-page confidential questionnaire and
f return it in a sealed envelope. The questionnaire included history of cancer and other diseases;
use of medications and vitamins; menstrual and reproductive history; occupational history; and
4 information on diet, drinking, smoking, and other habits. The questionnaire also included three
5 questions on occupation: (1) current occupation, (2) last occupation, if retired, and (3) job held
6 for the longest period of time, if different from the other two. Occupations were coded to an ad
7 hoc two-digit classification in 70 categories. Exposures at work or in daily life to any of the 12
8 groups of substances were also ascertained. These included diesel engine exhausts, asbestos,
9 chemicals/acids/solvents, dyes, formaldehyde, coal or stone dusts, and gasoline exhausts.
10 Volunteers checked whether their enrollees were alive or dead and recorded the date and place of
11 all deaths every other year during the study. Death certificates were then obtained from State
12 health departments and coded by a trained nosologist according to a system based on the ninth
13 revision of the ICD.
14 The data were analyzed to determine the mortality for all causes and lung cancer in
15 relation to diesel exhaust exposure, mortality for all causes and lung cancer in relation to
16 employment in selected occupations with high diesel exhaust exposure, and mortality from other
17 causes in relation to diesel exhaust exposure. The incidence-density ratio was used as a measure
of association, and test-based confidence limits were calculated by the Miettinen method. For
stratified analysis, the Mantel-Haenszel method was used for testing linear trends. Although data
20 on 476,648 subjects comprising 939,817 person-years of risk were available for analysis, 3% of
21 the subjects (14,667) had not given any smoking history, and 20% (98,026) did not give
22 information on diesel exhaust exposure and were therefore excluded from the main diesel
23 exhaust analysis. Among individuals who had provided diesel exhaust exposure history, 62,800
24 were exposed and 307,143 were not exposed. Comparison of the population with known
25 information on diesel exhaust exposure with the excluded population with no information on
26 diesel exhaust exposure showed that the mean ages were 54.7 and 57.7 years, the nonsmokers
27 were 72.4% and 73.2%, and the total mortality rates per 1,000 per year were 23.0% and 28.8%,
28 respectively.
29 All-cause mortality was elevated among railroad workers (relative risk [RR] = 1.43, 95%
30 confidence interval [CI] = 1.2, 1.72), heavy equipment operators (RR = 1.7, 95% CI = 1.19,
31 2.44), miners (RR = 1.34, 95% CI = 1.06, 1.68), and truck drivers (RR = 1.19, 95% CI = 1.07,
32 1.31). The age-adjusted lung cancer relative risk was elevated significantly (RR = 1.41, 95% CI
33 = 1.19,1.66), which was slightly decreased to 1.31 (95% CI = 1.10, 1.54). For lung cancer
34 mortality the age- and smoking-adjusted risks were significantly elevated for miners (RR = 2.67,
95% CI = 1.63, 4.37) and heavy equipment operators (RR = 2.60, 95% CI =1.12, 6.06). Risks
7/25/00 7-15 DRAFT—DO NOT CITE OR QUOTE
-------
1 were also elevated, but not significantly, for railroad workers (RR = 1.59, 95% CI = 0.94, 2.69)
2 and truck drivers (RR = 1.24, 95% CI = 0.93, 1.66). These risks were calculated with the
3 Mantel-Haenszel method, controlling for age and smoking. Although the relative risk was
4 nonsignificant for truck drivers, a small dose-response effect was observed when duration of
5 diesel exhaust exposure was examined. For drivers who worked for 1 to 15 years, the relative
6 risk was 0.87, whereas for drivers who worked for more than 16 years, the relative risk was 1.33
7 (95% CI = 0.64, 2.75). Relative risks for lung cancer were not presented for other occupations.
8 Mortality analysis for other causes and diesel exhaust exposure showed a significant excess of
9 deaths (/?<0.05) in the following categories: cerebrovascular disease, arteriosclerosis,
10 pneumonia, influenza, cirrhosis of the liver, and accidents.
11 The main strength of this study is detailed information on smoking. The two main
12 methodologic concerns are the representativeness of the study population and the quality of
13 information on exposure. The sample, though very large, was composed of volunteers. Thus,
14 the cohort was healthier and less frequently exposed to important risk factors such as smoking
15 and alcohol. Self-administered questionnaires were used to obtain data on occupation and diesel
16 exhaust exposure. None of this information was validated. Nearly 20% of the individuals had an
17 unknown exposure status to diesel exhaust, and they experienced a higher mortality for all causes
18 and lung cancer than both the diesel exhaust exposed and unexposed groups. This could have
19 introduced a substantial bias in the estimate of the association. Given that all diesel exhaust
20 exposure occupations, such as heavy equipment operators, truck drivers, and railroad workers,
.21 showed elevated lung cancer risk, this study is suggestive of a causal association. It should be
22 noted that after adjusting for smoking, the RR reduced slightly from 1.41 to 1.31 and remained
23 significant, indicating that observed excess of lung cancer was associated mainly with diesel
24 exhaust exposure. This study did not find any association between exposure to diesel exhaust
25 and bladder cancer.
26
27 7.2.1.7. Garshick et aL (1988): A Retrospective Cohort Study of Lung Cancer and Diesel
28 Exhaust Exposure in Railroad Workers
29 An earlier case-control study of lung cancer and diesel exhaust exposure in US. railroad
30 workers by these investigators had demonstrated a relative odds of 1.41 (95% CI = 1.06, 1.88)
31 for lung cancer with 20 years of work in jobs with diesel exhaust exposure. To confirm these
32 results, a large retrospective cohort mortality study was conducted by the same investigators.
33 Data sources for the study were the work records of the U.S. Railroad Retirement Board (RRB).
34 The cohort was selected based en job titles in 1959, which was the year by which 95% of the
35 icccrnctives in tu£ wniteu states were uiesci povvercu. i_>icsci cxiiaust exposure was
7/25/00 7-16 DRAFT—DO NOT CITE OR QUOTE
-------
1 to be a dichotomous variable depending on yearly job codes between 1959 and death or
^fc retirement through 1980. Industrial hygiene evaluations and descriptions of job activities were
3 used to classify jobs as exposed or unexposed to diesel emissions. A questionnaire survey of 534
4 workers at one of the railroads where workers were asked to indicate the amount of time spent in
5 railroad locations, either near or away from sources of diesel exhaust, was used to validate this
6 classification. Workers selected for this survey were actively employed at the time of the survey,
7 40 to 64 years of age, started work between 1939 and 1949 in the job codes sampled in 1959, and
8 eligible for railroad benefits. To qualify for benefits, a worker must have had 10 years or more
9 of service with the railroad and should not have worked for more than 2 years in a nonrailroad
10 job after leaving railroad work. Workers with recognized asbestos exposure, such as repair of
11 asbestos-insulated steam locomotive boilers, passenger cars, and steam pipes, or railroad building
12 construction and repairs, were excluded from the job categories selected for study. However, a
13 few jobs with some potential for asbestos exposure were included in the cohort, and the analysis
14 was done both ways, with and without them.
15 The death certificates for all subjects identified in 1959 and reported by the RRB to have
16 died through 1980 were searched. Twenty-five percent of them were obtained from the RRB and
17 the remainder from the appropriate State departments of health. Coding of cause of death was
done without knowledge of exposure history, according to the eighth revision of the ICD. If the
underlying cause of death was not lung cancer, but was mentioned on the death certificate, it was
20 assigned as a secondary cause of death, so that the ascertainment of all cases was complete.
21 Workers not reported by the RRB to have died by December 31,1980, were considered to be
22 alive. Deceased workers for whom death certificates had not been obtained or, if obtained, did
23 not indicate cause of death, were assumed to have died of unknown causes.
24 Proportional hazard models were fitted that provided estimates of relative risk for death
25 caused by lung cancer using the partial likelihood method described by Cox, using the time
26 dimension being the time since first entry into the cohort. The model also controlled for the birth
27 year and the calender time. The 95% confidence intervals were constructed using the asymptotic
28 normality of the estimated regression coefficients of the proportional hazards model. Exposure
29 was analyzed by diesel exhaust-exposed jobs in 1959 and by cumulative number of years of
30 diesel exhaust exposure through 1980. Directly standardized rate ratios for deaths from lung
31 cancer were calculated for diesel exhaust exposed compared with unexposed for each 5-year age
32 group in 1959. The standardized rates were based on the overall 5-year person-year time
33 distribution of individuals in each age group starting in 1959. The only exception to this was
34 between 1979 and 1980, when a 2-year person-year distribution was used. The Mantel-Haenszel
7/25/00 7-17 DRAFT—DO NOT CITE OR QUOTE
-------
1 analogue for person-year data was used to calculate 95% confidence intervals for the
2 standardized rate ratios.
3 The cohort consisted of 55,407 workers, 1 9,396 of whom had died by the end of 1 980.
4 Death certificates were not available for 1 1 .7% of all deaths. Of the 1 7, 1 20 deaths for whom
5 death certificates were obtained, 48.4% were attributable to diseases of the circulatory system,
6 whereas 21% were attributable to all neoplasms. Of all neoplasms, 8.7% (1,694 deaths) were due
7 to lung cancer. A higher proportion of workers in the younger age groups, mainly brakemen and
8 conductors, were exposed to diesel exhaust, while a higher proportion of workers in the older age
9 groups were potentially exposed to asbestos. In a proportional hazards model, analyses by age in
10 1 959 found a relative risk of 1 .45 (95% CI = 1 . 1 1 , 1 .89) among the age group 40 to 44 years and
1 1 a relative risk of 1 .33 (95% CI = 1 .03, 1 .73) for the age group 45 to 49 years. Risk estimates in
1 2 the older age groups 50 to 54, 55 to 59, and 60 to 64 years were 1.2, 1.18, and 0.99, respectively,
1 3 and were not statistically significant. The two youngest age groups in 1959 had workers with the
1 4 highest prevalence and longest duration of diesel exhaust exposure and lowest exposure to
1 5 asbestos. When potential asbestos exposure was considered as a confounding variable in a
1 6 proportional hazards model, the estimates of relative risk for asbestos exposure were all near null
1 7 value and not significant. Analysis of workers exposed to diesel exhaust in 1959 (n = 42,535),
1 8 excluding workers with potential past exposure to asbestos, yielded relative risks of 1 .57 (95%
19 CI = 1 . 1 9, 2.06) and 1 .34 (95% CI = 1 .02, 1 .76) in the 1 959 age groups 40 to 44 years and 45 to
20 49 years. Directly standardized rate ratios were also calculated for each 1959 age group based on
2 1 diesel exhaust exposure in 1 959. The results confirmed those obtained by using the proportional
22 hazards model.
23 Relative risk estimates were then obtained using duration of diesel exhaust exposure as a
24 surrogate for dose. In a model that used years of exposure up to and including exposure in the
25 year of death, no exposure duration-response relationship was obtained. When analysis was done
26 by disregarding exposure in the year of death and 4 years prior to death, the risk of dying from
27 lung cancer increased with the number of years worked in a diesel-exhaust-exposed job. In this
28 analysis, exposure to diese! exhaust was analyzed by exposure duration groups and in a model
29 entering ^ge in 1. 959 as a continuous variable. The workers with greater than 15 years of
30 exposure had a relative risk of lung cancer of 1 .72 (95% CI = 1 .27, 2.33). The risk for 1 to 4
3 1 years of cumulative exposure was 1 .20 (95% CI = 1 .0 1 , 1 .44); for 5 to 9 years of cumulative
32 exposure, it was 1.24 (95% CI = 1.06, 1.44); and for 10 to 14 years of cumulative exposure, it
33 was 1. 32 (95% CI = 1.1 3, 1.56).
^ ^ T"1. _ — . ___ •.'••- — — •£" -«-l- ' — — J ___ 1- -
OH- 1 11C ICiUlli
-------
conducted by the same investigators in railroad workers dying of lung cancer from March 1981
through February 1982. This cohort study has addressed many of the weaknesses of the other
"3 epidemiologic studies. The large sample size (60,000) allowed sufficient power to detect small
4 risks and also permitted the exclusion of workers with potential past exposure to asbestos. The
5 stability of job career paths in the cohort ensured that of the workers 40 to 44 years of age in
Q 1959 classified as diesel exhaust-exposed, 94% of the cases were still in diesel exhaust-exposed
7 jobs 20 years later.
8 The main limitation of the study is the lack of quantitative data on exposure to diesel
9 exhaust. This'is one of the few studies in which industrial hygiene measurements of diesel
10 exhaust were done. These measurements were correlated with job titles to divide the cohort in
11 dichotomous exposure groups of exposed and nonexposed. This may have led to an
12 underestimation of the risk of lung cancer because exposed groups included individuals with low
13 to high exposure. The number of years exposed to diesel exhaust was used as a surrogate for
14 dose. The dose, based on duration of employment, was inaccurate because individuals were
15 working on steam and diesel locomotives during the transition period. It should be noted that the
16 investigators only included exposures after 1959; the duration of exposure prior to 1959 was not
17 known. If the categories of exposure to diesel exhaust had been set up as no, low, moderate, and
high exposure, the results would have been more meaningful, as would the dose-response
relationship. Another limitation of this study was its inability to examine the effect of years of
20 exposure prior to 1959 and latency. No adjustment for smoking was made in this study.
21 However, an earlier case-control study done in the same cohort (Garshick et al., 1987) showed no
22 significant difference in the risk estimate after adjusting for smoking. Despite these limitations,
23 the results of this study indicate that occupational exposure to diesel exhaust is associated with a
24 modest risk (1.5) of lung cancer.
25 The data of this study were reanalyzed by Crump et al. (1991), who found that the
26 relative risk can be positively or negatively related to the duration of exposure depending on how
27 age was controlled in a model. Garshick conducted some additional analyses (letter from E.
28 Garshick, Harvard Medical School, to Dr. Chao Chen, U.S. EPA, dated August 15, 1991) and
29 reported that the relationship between years of exposure, when adjusted for attained age, and
30 calendar year is flat to negative depending upon which model was used. They also found that in
31 the years 1977-1980 the death ascertainment was incomplete; approximately 20% to 70% of
32 deaths were missing depending upon the calendar year. Their analysis, based on job titles in
33 1959 and limited to deaths occurring through 1976, showed that the youngest workers still had
34 the highest risk of dying of lung cancer. Crump (1999) reported that the negative dose-response
continued to be upheld in his latest analysis. California EPA (CalEPA, 1998) found a positive
7/25/00 7-19 DRAFT—DO NOT CITE OR QUOTE
-------
1 dose-response by using age at 1959 but allowing for an interaction term of age and calendar year
2 in the model. A detailed discussion of divergent results observed by Crump and CalEPA can be
3 found in Chapter 8.
4
5 7.2.1.8. Gustavsson et aL (1990): Lung Cancer and Exposure to Diesel Exhaust Among Bus
6 Garage Workers
7 A retrospective mortality study (from 1952 to 1986), cancer incidence study (from 1958
8 to 1984), and nested case-control study were conducted among a cohort of 708 male workers
9 from five bus garages in Stockholm, Sweden, who had worked for at least 6 months between
10 1945 and 1970. Thirteen individuals were lost to follow-up, reducing the cohort to 695.
11 Information was available on location of workplace, job type, and beginning and ending
12 of work periods. Workers were traced through a computerized register of the living population,
13 death and burial books, and data from the Stockholm city archives.
14 For the cohort mortality analyses, death rates of the general population of greater
15 Stockholm were used. Death rates of occupationally active individuals, a subset of the general
16 population of greater Stockholm, were used as a second comparison group to reduce the bias
17 from "healthy worker effect." Mortality analysis was conducted using the "occupational
18 mortality analysis program" (OCMAP-PC). For cancer incidence analysis, the "epidemiology in
19 Linkoping" (EPILIN) program was used, with the incidence rates obtained from the cancer
20 registry.
21 For the nested case-control study, both dead and incident primary lung cancers identified
22 in the register of cause of deaths and the cancer register were selected. Six controls matched on
23 age ± 2 years, selected from the noncases at the time of the diagnosis of cases, were drawn at
24 random without replacements. Matched analyses were done to calculate odds ratios using
25 conditional logistic regression. The EGRET and Epilog programs were used for these analyses.
26 Diesel exhaust and asbestos exposure assessments were performed by industrial
27 hygienists based on the intensity of exposure to diesel exhaust and asbestos, specific for
28 workplace, work task, and calendar time period. A diesel exhaust exposure assessment was
29 based on (1) amount of emission (number of buses, engine siVe, mnning time, and type of fuel),
30 (2) ventilatory equipment and air volume of the garages, and (3) job types and work practices.
31 Based on detailed historical data and very few actual measurements, relative exposures were
32 estimated (these were not absolute exposure levels). The scale was set to 0 for unexposed and 1
33 for lowest exposure, with each additional unit increase corresponding to a 50% increase in
34 successive intensity (i.e., 1.5, 2.25, 3.38, and 5.06).
7-20 DRAFT—DO NOT CITE OR QUOTE
-------
Based on personal sampling of asbestos during 1987, exposures were estimated and time-
weighted annual mean exposures were classified on a scale of three degrees (0, 1, and 2).
3 Cumulative exposures for both diesel exhaust and asbestos were calculated by multiplying the
4 level of exposure by the duration of every work period. An exposure index was calculated by
5 adding for every individual contribution from all work periods for both diesel exhaust and
6 asbestos. Four diesel exhaust index classes were created: 0 to 10, 10 to 20,20 to 30, and >30.
7 The four asbestos index classes were 0 to 20, 20 to 40, 40 to 60, and >60. The cumulative
8 exposure indices were used for the nested case-control study.
9 Excesses were observed for all cancers and some other site-specific cancers using both
10 comparison populations for the cohort mortality study, but none of them was statistically
11 significant. Based on 17 cases, SMRs for lung cancer were 122 and 115 using Stockholm
12 occupationally active and general population, respectively. No dose-response was observed with
13 increasing cumulative exposure in the mortality study. The cancer incidence study reportedly
14 confirmed the mortality results (results not given).
15 The nested case-control study, on the other hand, showed increasing risk of lung cancer
16 with increasing exposure. Using 0 to 10 diesel exhaust exposure index as the comparison group
17 yielded RRs of 1.34 (95% CI = 1.09 to 1.64), 1.81 (95% CI = 1.20 to 2.71), and 2.43 (95% CI =
fl .32 to 4.47) for the diesel exhaust indices 10 to 20, 20 to 30, and >30, respectively. The study
was based on 17 cases and 6 controls for each case matched on age ± 2 years. Adjustment for
20 asbestos exposure did not change the lung cancer risk for diesel exhaust.
•21 The main strength of this study is the detailed exposure matrices constructed for both
22 diesel exhaust and asbestos exposure, although they were based primarily on job tasks and very
23 few actual measurements. There are a few methodological limitations to this study. The cohort
24 is small and there were only 17 lung cancer deaths; thus the power is low. Exposure or outcome
25 may be misclassified, although any resulting bias in the relative risk estimates is likely to be
26 toward unity, because exposure classification was done independently of the outcome. Although
27 the analysis by dose indices was done, no latency analysis was performed. Although data on
28 smoking were missing, it is unlikely to confound the results because this is a nested case-control
29 study; therefore, smoking is not likely to be different among the individuals irrespective of their
30 exposure status to diesel exhaust. Overall, this study provides some support to the excess lung
31 cancer results found earlier among populations exposed to diesel exhaust.
32
33 7.2.1.9. Hansen (1993): A Followup Study on the Mortality of Truck Drivers
This is a retrospective cohort mortality study of unskilled male laborers, ages 15 to 74
years, in Denmark, identified from a nationwide census file of November 9, 1970. The exposed
7/25/00 7-21 DRAFT—DO NOT CITE OR QUOTE
-------
1 group included all truck drivers employed in the road delivery or long-haul business (14,225).
2 The unexposed group included all laborers in certain selected occupational groups considered to
3 be unexposed to fossil fuel combustion products and to resemble truck drivers in terms of work-
4 related physical demands and various personal background characteristics (43,024).
5 Through automatic record linkage between the 1970 census register (the Central
6 Population Register 1 970 to 1 980) and the Death Certificate Register ( 1 970 to 1 980), the
7 population was followed for cause-specific mortality or emigration up to November 9, 1980.
8 Expected number of deaths among truck drivers was calculated by using the 5 -year age group
9 and 5-year time period death rates of the unexposed group and applying them to the person-years
1 0 accumulated by truck drivers. ICD Revision 8 was used to code the underlying cause of death.
1 1 Test-based CIs were calculated using Miettinen's method. A Poisson distribution was assumed
1 2 for the smaller numbers, and CI was calculated based on exact Poisson distribution (Ciba-Geigy).
1 3 Total person-years accrued by truck drivers were 138,302, whereas for the unexposed population,
1 4 they were 407,780. There were 627 deaths among truck drivers and 3,81 1 deaths in the
1 5 unexposed group. Statistically significant excesses were observed for all cancer mortality (SMR
16 = 121 , 95% CI = 104 to 140); cancer of respiratory organs (SMR = 160, 95% CI = 128 to 198),
1 7 which was due mainly to cancer of bronchus and lung (SMR = 160, 95% CI = 126 to 200); and
1 8 multiple myeloma (SMR = 439, 95% CI = 142 to 1,024). When lung cancer mortality was
1 9 further explored by age groups, excesses were observed in most age groups (30 to 39, 45 to 49,
20 50 to 54, 55 to 59, 60 to 64, and 65 to 74), but there were small numbers of deaths in each group
2 1 when stratified by age, and the excesses were statistically significant for the 55 to 59 (SMR =
22 229, O = 19, 95% CI = 138 to 358) and 60 to 64 (SMR = 227, O = 22, 95% CI = 142 to 344) age
23 groups only. No excess was observed for bladder cancer.
24 As acknowledged by the author, the study has quite a few methodologic limitations. The
25 exposure to diesel exhaust is assumed in truck drivers based on use of diesel-powered trucks, but
26 no validation of qualitative or quantitative exposure is attempted. It is also not known whether
27 any of these truck drivers or any other laborers had changed jobs after the census of November 9,
28 1970, thus creating potential misclassificatien bias in exposure to diesel exhaust. The truck
?9 drivers and the uoexposed laborers were from the same socioeconomic class and may have the
30 same smoking habits. Still, the lack of information on smoking data and a 36% rural population
3 1 (usually consuming less tobacco) in the unexposed group may potentially confound the lung
32 cancer results. However, a population survey carried out in 1988 showed very little difference in
33 smoking habits of residents of rural areas and the total Danish male population. The investigator
•-<« ~_ --- -v-^l -*- A' ___ ' * ___ '-- _____ ^» - — * --- ' ----- J ;— T~» ----- '- -•& -- T" -- 1 J 117__ TT __ J _: __
OM- icpuita uia.1 Ciicsci uu^is.5 wci€ uuiuuuccu 111 i^ciuucun anci vnjnu. »» 3i 11, cuiu
oo iy+\js Lac inajuriiy ui ine uaiiiMi nccl nai uccn cuiupubcu 01 uicbci tiuCKS. ^GuScCjuciitiy, CVCIi
7/25/00 7-22 DRAFT — DO NOT CITE OR QUOTE
-------
though the follow-up period is relatively short, the truck drivers may have had exposure to diesel
exhaust for 20 to 30 years. Therefore, the finding of excess lung cancer in this study is
consistent with the findings of other truck driver studies.
4
5 7.2.1.10. Saverin et al (1999): Diesel Exhaust and Lung Cancer Mortality in Potash Mining
6 This is a cohort mortality study conducted in male potash miners in Germany. The mines
7 began using mobile diesel-powered vehicles in 1969 and 1970. Miners who had worked
8 underground for at least 1 year after 1969 to 1991, when the mines were closed, were followed
9 from 1970 to 1994. A total of 5,981 individuals were identified from the medical records by a
10 team of medical personnel familiar with the mining technology. A total of 5,536 were eligible
11 for follow-up after 5.5% were excluded due to implausible or incomplete work history and 1.9%
12 were lost to follow-up. A subcohort of 3,258 miners who had worked for at least 10 years
13 underground (80% had held a single job) was also identified. The miners' biannual medical
14 examination records were used to extract the information about personal data, smoking data, and
15 pre-mining occupation, and to reconstruct a chronology of workplaces occupied by the worker
16 since hire for each person.
17 Exposure categories were defined as production, maintenance, and workshop, roughly
§ corresponding to high, medium, and low. Concentrations of total carbon, including elemental
and organics, were measured in the airborne fine dust in 1992. A total of 255 samples covering
20 all workplaces was obtained. Most were personal dust samples; some were area dust samples.
21 Cumulative exposure was calculated for each miner, for each year of observation, using the work
22 chronology and the work category. For the workshop category years of employment were
23 considered as exposure time; for production and maintenance years of employment was weighted
24 by a factor of 5/8, since these workers for an 8-hour shift worked for only 5 hours underground.
25 As neither the mining technology nor the type of machinery used had changed substantially from
26 1970 to 1992, the exposure measurements were considered to represent the exposures throughout
27 the study period. Accrued person-years were classified into cumulative exposures and were
28 expressed in intervals of 0.5 ymg/m3. Both the exposure data and the smoking data obtained
29 from the medical files were validated by personal interviews with 1,702 cohort members. Death
30 certificates were obtained from local health centers for 94.4% of deceased members. Autopsy
31 data were available for 13% of the deceased. Internal comparison was done between production
32 and workshop categories. Using East German general male population rates, SMRs were
33 computed for the total cohort as well as the subcohort. Analyses were done using Poisson and
34 Cox regression models.
7/25/00 7-23 DRAFT—DO NOT CITE OR QUOTE
-------
1 The concentrations of total carbon for production, maintenance, and workshop categories
2 were 0.39 mg/m3, 0.23 mg/m3, and 0.12 mg/m3, respectively. The cumulative exposure ranged
3 from 0.25 ymg/m3 to 6.25 ymg/m3. The regression analysis showed that the cohort's smoking
4 habits were homogenous and that smoking had an even distribution over cumulative exposure.
5 A total of 424 deaths were observed for the entire cohort (SMR = 54). The all-cancer
6 deaths were 133, of which 38 were from lung cancer (SMR = 78). Analysis for the subcohort
7 using the internal comparison group of low exposure (workshop category, mean cumulative
8 exposure = 2.12 ymg/m3) RR of 2.17 (95% CI = 0.79, 5.99) was found for the production
9 category (mean cumulative exposure = 4.38 ymg/m3). The relative risks for lung cancer for 20
10 years of exposure in the production category (highest exposure = cumulative exposure of 4.9
11 ymg/m3) were calculated using Poisson and Cox regression methods. RRs of 1.16 and 1.68 were
12 observed for the total cohort, while RRs of 1.89 and 2.7 were observed for the subcohort by
13 Poisson and Cox regression methods respectively.
14 The main strengths of the study are the information available on diesel exhaust exposure
1 5 and smoking. Although these potash miners were exposed to salt dust and nitric gases,
16 exposures to other confounders such as heavy metals and radon were absent. Smoking does not
17 seem to be a confounder in this study but cannot be completely ruled out. Unfortunately, the age
18 distribution of the cohort is not available. Since there were only 424 deaths in 25 years of
19 follow-up in this cohort of 5,536, it appears that the cohort is young. Although lung cancer risk
20 was elevated by twofold in the production category of the subcohort of miners who had worked
21 for at least 10 years underground at the same job for 80% of their time and did not have more
22 than 3 jobs, it was not statistically significant. The follow-up period for this study was 25 years,
23 but the cohort members could have entered the cohort any time between 1970 and 1990, as long
24 as they worked underground for a year, i.e., they could have worked in the mines for 1 year to 21
25 years. Thus, the authors may not have had enough follow-up or latency to observe the lung
26 cancer excess. Despite these limitations, the results of this study provide suggestive evidence for
27 the causal association between diesel exhaust and excess lung cancer.
28 Table 7-1 summarizes the above cohort studies.
29
30 7.2.2. Case-Control Studies of Lung Cancer
31 7.2.2.1. Hall and Wynder (1984): A Case-Control Study of Diesel Exhaust Exposure and
32 Lung Cancer
33 Hall and Wynder (1984) conducted a case-control study of 502 male lung cancer cases
34 and 502 controls without tobacco-related diseases thai examined mi association between
35 occupational diesei exhaust exposure and iung cancer. Histoiogicaiiy confirmed primary lung,
725/00 7-24 DRAFT—DO NOT CITE OR QUOTE
^ /o z ir\f\
-------
cancer patients who were 20 to 80 years old were ascertained from 18 participating hospitals in 6
U.S. cities 12 months prior to the interview. Eligible controls, patients at the same hospitals
3 without tobacco-related diseases, were matched to cases by age (± 5 years), race, hospital, and
4 hospital room status. The number of male lung cancer cases interviewed totaled 502, which was
5 64% of those who met the study criteria for eligibility. Of the remaining 36%, 8% refused, 21%
6 were too ill or had died, and 7% were unreliable. Seventy-five percent of eligible controls
7 completed interviews. Of these interviewed controls, 49.9% were from the all-cancers category,
8 whereas 50.1% were from the all-noncancers category. All interviews were obtained in hospitals
9 to gather detailed information on smoking history, coffee consumption, artificial sweetener use,
10 residential history, and abbreviated medical history as well as standard demographic variables.
11 Occupational information was elicited by a question on the usual lifetime occupation and was
12 coded by the abbreviated list of the U.S. Bureau of Census Codes. The odds ratios were
13 calculated to evaluate the association between diesel exhaust exposure and risk of lung cancer
14 incidence. Summary odds ratios were computed by the Mantel-Haenszel method after adjusting
15 for potential confounding by age, smoking, and socioeconomic class. Two-sided, 95%
16 confidence intervals were computed by Woolf s method. Occupational exposure to diesel
17 exhaust was defined by two criteria. First, occupational titles were coded "probably high
« exposure" as defined by the industrial hygiene standards established for the various jobs. The
job titles included under this category were warehousemen, bus and truck drivers, railroad
20 workers, and heavy equipment operators and repairmen. The second method used the National
21 Institute for Occupational Safety and Health (NIOSH) criteria to analyze occupations by diesel
22 exposure. In this method, the estimated proportion of exposed workers was computed for each
23 occupational category by using the NIOSH estimates of the exposed population as the numerator
24 and the estimates of individuals employed in each occupational category from the 1970 census as
25 the denominator. Occupations estimated to have at least 20% of their employees exposed to
26 diesel exhaust were defined as "high exposure," those with 10% to 19% of their employees
27 exposed were defined as "moderate exposure," and those with less than 10% of their employees
28 exposed were defined as "low exposure."
29 Cases and controls were compared with respect to exposure. The relative risk was 2.0
30 (95% CI = 1.2, 3.2) for those workers who were exposed to diesel exhaust versus those who were
31 not. The risk, however, decreased to a nonsignificant 1.4 when the data were adjusted for
32 smoking. Analysis by NIOSH criteria found a nonsignificant relative risk of 1.7 in the high-
33 exposure group. There were no significantly increased cancer risks by occupation either by the
first method or by the NIOSH method. To assess any possible synergism between diesel exhaust
exposure and smoking, the lung cancer risks were calculated for different smoking categories.
7/25/00 7-25 DRAFT—DO NOT CITE OR QUOTE
-------
1 The relative risks were 1.46 among nonsmokers and ex-smokers, 0.82 among current smokers of
2 <20 cigarettes/day, and 1.3 among current smokers of 20+ cigarettes/day, indicating a lack of
3 synergistic effects.
4 The major strength of this study is the availability of a detailed smoking history for all the
5 study subjects. However, this is offset by lack of diesel exhaust exposure measurements, use of a
6 poor surrogate for exposure, and lack of consideration of latency period. Information was
7 collected on only one major lifetime occupation, and it is likely that those workers who had more
8 than one major job may not have reported the occupation with the heaviest diesel exhaust
9 exposures. Furthermore, the exposure categories based on job titles were broad, and thus would
10 have made a true effect of diesel exhaust difficult to detect.
11
12 7.2.2.2. Dumber and Larsson (1987): Occupation and Male Lung Cancer, a Case-Control
13 Study in Northern Sweden
14 A case-control study of lung cancer was conducted in northern Sweden to determine the
15 occupational risk factors that could explain the large geographic variations of lung cancer
16 incidence in that country. The study region comprised the three northernmost counties of
17 Sweden, with a total male population of about 390,000. The rural municipalities, with 15% to
18 20% of the total population, have forestry and agriculture as dominating industries, and the urban
19 areas have a variety of industrial activities (mines, smelters, steel factories, paper mills, and
20 mechanical workshops). All male cases of lung cancer reported to the Swedish Cancer Registry
21 during the 6-year period between 1972 and 1977 who had died before the start of the study were
22 selected. Of 604 eligible cases, 5 did not have microscopic confirmation, and in another 5 the
23 diagnosis was doubtful, but these cases were included nevertheless. Cases were classified as
24 small-cell carcinomas, squamous cell carcinomas, adenocarcinomas, and other types. For each
25 case a dead control was drawn from the National Death Registry matched by sex, year of death,
26 age, and municipality. Deaths in controls classified as lung cancer and suicides were excluded.
27 A living control matched to the case by sex, year of birth, and municipality was also drawn from
28 the National Population Registry. Postal questionnaires were sent to close relatives of cases and
29 dead controls, and to living controls themselves to collect data on occupation, employment, and
30 smoking habits. Replies were received from 589 cases (98%), 582 surrogates of dead controls
31 (96%), and 453 living controls (97%).
32 Occupational data were collected on occupations or employment held for at least 1 year
33 and included type of industry, company name, task, and duration of employment.
34 Supplementary telephone interviews were performed if occupational data were lacking for any
35 period between age 2U and time of diagnosis. Data analysis involved calculation of the odds
-------
ratios by the exact method based on the hypergeometric distribution and the use of a linear
logistic regression model to adjust for the potential confounding effects of smoking. Separate
3 analyses were performed with dead and living controls, and on the whole there was good
4 agreement between the two control groups. A person who had been active for at least 1 year in a
5 specific occupation was in the analysis assigned to that occupation.
6 Using dead controls, the odds ratios adjusted for smoking were 1.0 (95% CI = 0.7, 1.5)
7 and 2.7 (95% CI = 1.0, 8.1) for professional drivers (;»1 year of employment) and underground
8 miners (* 1 year of employment), respectively. For 20 or more years of employment in those
9 occupations, the odds ratios adjusted for smoking were 1.2 (95% CI = 0.9, 2.6) and 9.8 (95% CI
10 = 1.5, 414). These were the only two occupations listed with potential diesel exhaust exposure.
11 An excess significant risk was detected for copper smelter workers, plumbers, electricians, and
12 asbestos workers, as well as concrete and asphalt workers. All the odds ratios were calculated by
13 adjusting for age, smoking, and municipality. A comparison with the live controls resulted in the
14 odds ratios being lower than those observed with dead controls, and none were statistically
15 significant in this comparison.
16 This study did not detect any excess risk of lung cancer for professional drivers, who,
17 among all the occupations listed, had the most potential for exposure to motor vehicle exhaust.
f However, it is not known whether these drivers were exposed exclusively to gasoline exhaust.
diesel exhaust, or varying degrees of both. An excess risk was detected for underground miners,
20 but it is not known if this was due to diesel emissions from engines or from radon daughters in
21 poorly ventilated mines. Although a high response rate (98%) was obtained by the postal
22 questionnaires, the use of surrogate respondents is known to lead to misclassification errors that
23 can bias the results in either direction.
24
25 7.2.2.3. Lerchen etal. (1987): Lung Cancer and Occupation in New Mexico
26 This is a population-based case-control study conducted in New Mexico that examined
27 the association between occupation and occurrence of lung cancer in Hispanic and non-Hispanic
28 whites. Cases involved residents of New Mexico, 25 through 84 years of age, and diagnosed
29 between January 1, 1980, and December 31, 1982, with primary lung cancer, excluding
30 bronchioalveolar carcinoma. Cases were ascertained through the New Mexico Tumor Registry,
31 which is a member of the Surveillance Epidemiology and End Results (SEER) Program of the
32 National Cancer Institute. Controls were chosen by randomly selecting residential telephone
33 numbers and, for those over 65 years of age, from the Health Care Financing Administration's
€ roster of Medicare participants. They were frequency-matched to cases for sex, ethnicity, and
10-year age category with a ratio of 1.5 controls per case. The 506 cases (333 males and 173
7/25/00 7-27 DRAFT—DO NOT CITE OR QUOTE
-------
1 females) and 771 controls (499 males and 272 females) were interviewed, with a nonresponse
2 rate of 11% for cases. Next of kin provided interviews for 50% and 43% of male and female
3 cases, respectively. Among controls, only 2% of the interviews were provided by next of kin for
4 each sex. Data were collected by personal interviews conducted by bilingual interviewers in the
5 participants' homes. A lifetime occupational history and a self-reported history of exposure to
6 specific agents were obtained for each job held for at least 6 months since age 12. Questions
7 were asked about the title of the position, duties performed, location and nature of industry, and
8 time at each job title. A detailed smoking history was also obtained. The variables on
9 occupational exposures were coded according to the Standard Industrial Classification scheme by
10 a single person and reviewed by another. To test the hypothesis about high-risk jobs for lung
11 cancer, the principal investigator created an a priori listing of suspected occupations and
12 industries by a two-step process involving a literature review for implicated industries and
13 occupations. The principal investigator also determined the appropriate Standard Industrial
14 Classification and Standard Occupational Codes associated with job titles. For four
15 agents—asbestos, wood dust, diesel exhaust, and formaldehyde—the industries and occupations
16 determined to have exposure were identified, and linking of specific industries and occupations
17 was based on literature review and consultation with local industrial hygienists.
18 The relative odds were calculated for suspect occupations and industries, classifying
19 individuals as ever employed for at least 1 year in an industry or occupation and defining the
20 reference group as those subjects never employed in that particular industry or occupation.
21 Multiple logistic regression models were used to control simultaneously for age, ethnicity, and
22 smoking status. For occupations with potential diesel exhaust exposure, the analysis showed no
23 excess risks for diesel engine mechanics and auto mechanics. Similarly, when analyzed by
24 exposure to specific agents, the odds ratio (OR) adjusted for age, smoking, and ethnicity was not
25 elevated for diesel exhaust fumes (OR = 0.6, 95% CI = 0.2, 1.6). Significantly elevated ORs
26 were found for uranium miners (OR = 2.8), underground miners (OR = 2.4), construction
27 workers, and welders (OR = 4.3). No excess risks were detected for the following industries:
23 shipbuilding, petroleum refining, printing, blast furnace, and steel mills. No excess risks were
29 detected for the following occupations: construction workers, painters, plumbers, paving
30 equipment operators, roofers, engineers and firemen, woodworkers, and shipyard workers.
31 Females were excluded from detailed analysis because none of the Hispanic female controls had
32 been employed in high-risk jobs; among the non-Hispanic white controls, employment in a high-
33 risk job was recorded for at least five controls for only two industries, construction and painting,
34 for which the OR were not significantly elevated. Therefore, the analyses were presented for
3t> maies oniy.
7/25/00 7-28 DRAFT—DO NOT CITE OR QUOTE
-------
Among the many strengths of this study are its population-based design, high
participation rate, detailed smoking history, and the separate analysis done for two ethnic groups,
3 southwestern Hispanic and non-Hispanic white males. The major limitations pertain to the
4 occupational exposure data. Job titles obtained from occupational histories were used as proxy
5 for exposure status, but these were not validated. Further, for nearly half the cases, next of kin
6 provided occupational histories. The authors acknowledge the above sources of bias but state
7 without substantiation that these biases would not strongly affect their results. They also did not
8 use a job exposure matrix to link occupations to exposures and did not provide details on the
9 method they used to classify individuals as diesel exhaust exposed based on reported
10 occupations. The observed absence of an association for exposure to asbestos, a well-established
11 lung carcinogen, may be explained by the misclassification errors in exposure status or by
12 sample size constraints (not enough power). Likewise, the association for diesel exhaust
13 reported by only 7 cases and 17 controls also may have gone undetected because of low power.
14 In conclusion, there is insufficient evidence from this study to confirm or refute an association
15 between lung cancer and diesel exhaust exposure.
16
17 7.2.2.4. Garshick et al. (1987): A Case-Control Study of Lung Cancer and Diesel Exhaust
Exposure in Railroad Workers
An earlier pilot study of the mortality of railroad workers by the same investigators
20 (Schenker et al., 1984) found a moderately high risk of lung cancer among workers exposed to
21 diesel exhaust compared with those who were not. On the basis of these findings the
22 investigators conducted a case-control study of lung cancer in the same population. The
23 population base for this case-control study was approximately 650,000 active and retired male
24 U.S. railroad workers with 10 years or more of railroad service who were born in 1900 or later.
25 The U.S. Railroad Retirement Board (RRB), which operates the retirement system, is separate
26 from the Social Security System, and to qualify for the retirement or survivor benefits the
27 workers had to acquire 10 years or more of service. Information on deaths that occurred between
28 March 1, 1981, and February 28, 1982, was obtained from the RRB. For 75% of the deceased
29 population, death certificates were obtained from the RRB, and, for the remaining 25%, they
30 were obtained from the appropriate State departments of health. Cause of death was coded
31 according to the eighth revision of the ICD. The cases were selected from deaths with primary
32 lung cancer, which was the underlying cause of death in most cases. Each case was matched to
33 two deceased controls whose dates of birth were within 2.5 years of the date of birth of the case
4 and whose dates of death were within 31 days of the date of death noted in the case. Controls
i
7/25/00 7-29 DRAFT—DO NOT CITE OR QUOTE
-------
1 were selected randomly from workers who did not have cancer noted anywhere on their death
2 certificates and who did not die of suicide or of accidental or unknown causes.
3 Each subject's work history was determined from a yearly job report filed by his
4 employer with the RRB from 1959 until death or retirement. The year 1959 was chosen as the
5 effective start of diesel exhaust exposure for this study since by this time 95% of the locomotives
6 in the United States were diesel powered. Investigators acknowledge that because the transition
7 to diesel-powered engines took place in the early 1950s, some workers had additional exposure
8 prior to 1959; however, if a worker had died or retired prior to 1959, he was considered
9 unexposed. Exposure to diesel exhaust was considered to be dichotomous for this study, which
10 was assigned based on an industrial hygiene evaluation of jobs and work areas. Selected jobs
11 with and without regular diesel exhaust exposure were identified by a review of job title and
12 duties. Personal exposure was assessed in 39 job categories representative of workers with and
13 without diesel exhaust exposure. Those jobs for which no personal sampling was done were
14 considered exposed or unexposed on the basis of similarities in job activities and work locations
15 and by degree of contact with diesel equipment. Asbestos exposure was categorized on the basis
16 of jobs held in 1959, or on the last job held if the subject retired before 1959. Asbestos exposure
17 in railroads occurred primarily during the steam engine era and was related mostly to the repair
18 of locomotive steam boilers that were insulated with asbestos. Smoking history information was
19 obtained from the next of kin.
20 Death certificates were obtained for approximately 87% of the 15,059 deaths reported by
21 the RRB, from which 1,374 cases of lung cancer were identified. Fifty-five cases of lung cancer
22 were excluded from the study for either incomplete data (20) or refusal by two States to use
23 information on death certificates to contact the next of kin. Successful matching to at least one
24 control with work histories was achieved for 335 (96%) cases s64 years of age at death and 921
25 (95%) cases s65 years of age at death. In both age groups, 90% of the cases were matched with
26 two controls. There were 2,385 controls in the study; 98% were matched within ± 31 days of the
27 date of death, whereas the remaining 2% were matched within 100 days. Deaths from diseases
28 of the circulatory system predominated among controls. Among the younger workers,
29 approximately 60% had exposure to diesel exhaust, whereas among elder workers, only 47%
30 were exposed to diesel exhaust.
31 Analysis by a regression model, in which years of diesel exhaust exposure were the sum
32 total of the number of years in diesel-exposed jobs, used as a continuous exposure variable,
33 yielded an odds ratio of lung cancer of 1.39 (95% CI = 1.05, 1.83) for >20 years of diesel exhaust
34 exposure in the ^64 years of age group. After adjustment for asbestos exposure and lifetime
35 smoking (pack-years), me odds ratio was I.4i (95% Cl = 1.06. 1.88). Both crude odds ratio and
7/25/00 7-30 DRAFT—DO NOT CITE OR QUOTE
-------
asbestos exposure as well as lifetime smoking-adjusted odds ratio for the ;>65 years of age group
were not significant. Increasing years of diesel exhaust exposure, categorized as ^20 diesel years
3 and 5 to 19 diesel years, with 0 to 4 years as the referent group, showed significantly increased
4 risk in the s64 years of age group after adjusting for asbestos exposure and pack-year category of
5 smoking. For individuals who had *20 years of diesel exhaust exposure, the odds ratio was 1.64
6 (95% CI = 1.18, 2.29), whereas among individuals who had 5 to 19 years of diesel exhaust
7 exposure, the odds ratio was 1.02 (95% CI = 0.72,1.45). In the *65 years of age group, only 3%
8 of the workers were exposed to diesel exhaust for more than 20 years. Relative odds for 5 to 19
9 years and *20 years of diesel exposure were less than 1 (p>0.01) after adjusting for smoking and
10 asbestos exposure.
11 Alternative models to explain past asbestos exposure were tested. These were variables
12 for regular and intermittent exposure groups and an estimate of years of exposure based on
13 estimated years worked prior to 1959. No differences in results were seen. The interactions
14 between diesel exhaust exposure and the three pack-year categories (<50, >50, and missing pack-
15 years) were explored. The cross-product terms were not significant. A model was also tested
16 that excluded recent diesel exhaust exposure occurring within the 5 years before death and gave
17 an odds ratio of 1.43 (95% CI = 1.06, 1.94), adjusted for cigarette smoking and asbestos
exposure, for workers with 15 years of cumulative exposure. For workers with 5 to 14 years of
cumulative exposure, the OR were not significant.
20 The many strengths of the study are consideration of confounding factors such as
. 21 asbestos exposure and smoking; classification of diesel exhaust exposures by job titles and
22 industrial hygiene sampling; exploration of interactions between smoking, asbestos exposure,
23 and diesel exhaust exposure; and good ascertainment (87%) of death certificates from the 15,059
24 deaths reported by the RRB.
25 The investigators also recognized and reported the following limitations: overestimation
26 of cigarette consumption by surrogate respondents, which may have exaggerated the contribution
27 of smoking to lung cancer risk, and use of the Interstate Commerce Commission (ICC) job
28 classification as a surrogate for exposure, which may have led to misclassification of diesel
29 exhaust exposure jobs with low intensity and intermittent exposure, such as railroad police and
30 bus drivers, as unexposed. These two limitations would result in underestimation of the lung
31 cancer risk. This source of error could have been avoided if diesel exhaust exposures were
32 categorized by a specific dose range associated with a job title that could have been classified as
33 heavy, medium, low, and zero exposure instead of a dichotomous variable. The use of death
34 certificates to identify cases and controls may have resulted in misclassification. Controls may
VI have had undiagnosed primary lung cancer, and lung cancer cases might have been secondary
7/25/00 7-31 DRAFT—DO NOT CITE OR QUOTE
-------
1 lesions misdiagnosed as primary lung cancer. However, the investigators quote a third National
2 Cancer Survey report in which the death certificates for lung cancer were coded appropriately in
3 95% of the cases. Last, as in all previous studies, there is a lack of data on the contribution of
4 unknown occupational or environmental exposures and passive smoking. Furthermore, the lung
5 cancer cases were selected between 1981 and 1982, a total of 22 years latency, which is probably
6 short. In conclusion, this study provides strong evidence that occupational diesel exhaust
7 emission exposure increases the risk of lung cancer.
8
9 7.2.2.5. Benhamou et al. (1988): Occupational Risk Factors of Lung Cancer in a French
10 Case-Control Study
11 This is a case-control study of 1,625 histologically confirmed cases of lung cancer and
12 3,091 matched controls, conducted in France between 1976 and 1980. This study was part of an
13 international study to investigate the role of smoking and lung cancer. Each case was matched
14 with one or two controls, whose diseases were not related, to tobacco use, sex, age at diagnosis
15 (±5 years), hospital of admission, and interviewer. Information was obtained from both cases
16 and controls on place of residence since birth, educational level, smoking, and drinking habits. A
17 complete lifetime occupational history was obtained by asking participants to give their
18 occupations from the most recent to the first. Women were excluded because most of them had
19 listed no occupation. Men who smoked cigars and pipes were excluded because there were very
20 few in this category. Thus, the study was restricted to nonsmokers and cigarette smokers.
21 Cigarette smoking exposure was defined by age at the first cigarette (nonsmokers, s20 years, or
22 >20 years), daily consumption of cigarettes (nonsmokers, <20 cigarettes a day, and ;>20 cigarettes
23 a day), and duration of cigarette smoking (nonsmokers, <35 years, and ^35 years). The data on
24 occupations were coded by a panel of experts according to their own chemical or physical
25 exposure determinations. Occupations were recorded blindly using the International Standard
26 Classification of Occupations. Data on 1,260 cases and 2,084 controls were available for
27 analysis. The remaining 365 cases and 1,007 controls were excluded because they did not satisfy
28 the required smoking status criteria.
29 A matched logistic regression analysis was performed *"o estimate the effect of each
30 occupational exposure after adjusting for cigarette status. Matched relative risk ratios were
31 calculated for each occupation with the baseline category, which consisted of patients who had
32 never been engaged in that particular occupation. The matched RR ratios, adjusted for cigarette
33 smoking for the major groups of occupations, showed that the risks were significantly higher for
34 production and related workers, transport equipment operators, and laborers (RR = 1.24, 95% CI
3b = 1 .U4, 1.47). On further analysis of this group, for occupations with potential diesel emission
7/25/00 7-32 DRAFT—DO NOT CITE OR QUOTE
-------
exposure, significant excess risks were found for motor vehicle drivers (RR = 1.42, 95% CI =
1.07, 1.89) and transport equipment operators (RR = 1.35, 95% CI = 1.05, 1.75). No interaction
3 with smoking status was found in any of the occupations. The only other significant excess was
4 observed for miners and quarrymen (RR = 2.14, 95% CI = 1.07,4.31). None of the significant
5 associations showed a dose-response relationship with duration of exposure.
6 This study was designed primarily to investigate the relationship between smoking (not
7 occupations or environmental exposures) and lung cancer. Although an attempt was made to
8 obtain complete occupational histories, the authors did not clarify whether, in the logistic
9 regression analysis, they used the subjects' first occupation, predominant occupation, last
10 occupation, or ever worked in that occupation as the risk factor of interest. The most important
11 limitation of this study is that the occupations were not coded into exposures for different
12 chemical and physical agents, thus precluding the calculation of relative risks for diesel
13 exposure. Using occupations as surrogate measures of diesel exposure, an excess significant risk
14 was obtained for motor vehicle drivers and transport equipment operators, but not for motor
15 mechanics. However, it is not known if subjects in these occupations worked with diesel engines
16 or nondiesel engines.
17
f 7.2.2.6. Hayes et aL (1989): Lung Cancer in Motor Exhaust-Related Occupations
This study reports the findings from an analysis of pooled data from three lung cancer
20 case-control studies that examine in detail the association between employment in motor
21 exhaust-related (MER) occupations and lung cancer risk adjusted for confounding by smoking
22 and other risk factors. The three studies were carried out by the National Cancer Institute in
23 Florida (1976 to 1979), New Jersey (1980 to 1981), and Louisiana (1979 to 1983). These three
24 studies were selected because the combined group would provide a sufficient sample to detect a
25 risk of lung cancer in excess of 50% among workers in MER occupations. The analyses were
26 restricted to males who had given occupational history. The Florida study was hospital based,
27 with cases ascertained through death certificates. Controls were randomly selected from hospital
28 records and death certificates, excluding psychiatric diseases, matched by age and county. The
29 New Jersey study was population based, with cases ascertained through hospital records, cancer
30 registry, and death certificates. Controls were selected from among the pool of New Jersey
31 licensed drivers and death certificates. The Louisiana study was hospital based (it is not
32 specified how the cases were ascertained), and controls were randomly selected from hospital
33 patients, excluding those with lung diseases and tobacco-related cancers.
A total of 2,291 cases of male lung cancers and 2,570 controls were eligible, and the data
on occupations were collected by next-of-kin interviews for all jobs held for 6 months or more,
7/25/00 7-33 DRAFT—DO NOT CITE OR QUOTE
-------
1 including the industry, occupation, and number of years employed. The proportion of next-of-
2 kin interviews varied by site from 50% in Louisiana to 85% in Florida. The coding schemes
3 were reviewed to identify MER occupations, which included truck drivers and heavy equipment
4 operators (cranes, bulldozers, and graders); bus drivers, taxi drivers, chauffeurs, and other motor
5 vehicle drivers; and automobile and truck mechanics. Truck drivers were classified as routemen
6 and delivery men and other truck drivers. All jobs were also classified with respect to potential
7 exposure to known and suspected lung carcinogens. OR were calculated by the maximum
8 likelihood method, adjusting for age by birth year, usual amount smoked, and study area.
9 Logistic regression models were used to examine the interrelationship of multiple variables.
10 A statistically significant excess risk was detected for employment of 10 years or more
11 for all MER occupations (except truck drivers) adjusted for birth cohort, usual daily cigarette use,
12 and study area. The odds ratio for lung cancer using data gathered by direct interviews was 1.4
13 (95% CI = 1.1, 2.0), allowing for multiple MER employment, and 2.0 (95% CI = 1.3, 3.0),
14 excluding individuals with multiple MER employment. OR for all MER employment, except
15 truck drivers who were employed for less than 10 years, were 1.3 (95% CI = 1.0, 1.7) and 1.3
16 (95% CI = 0.9, 1.8) including and excluding multiple MER employment, respectively. OR were
17 then derived for specific MER occupations and, to avoid the confounding effects of multiple
18 MER job classifications, analyses were also done excluding subjects with multiple MER job
19 exposures. Truck drivers employed for more than 10 years had an odds ratio of 1.5 (95% CI =
20 1.1,1.9). A similar figure was obtained excluding subjects with multiple MER employment. An
21 excess risk was not detected for truck drivers employed less than 10 years. The only other job
22 category that showed a statistically significant excess for lung cancer included taxi drivers and
23 chauffeurs who worked multiple MER jobs for less than 10 years (OR = 2.5, 95% CI = 1.4, 4.8).
24 For the same category, the risk for individuals working in that job for more than 10 years was 1.2
25 (95% CI = 0.5, 2.6). A statistically significant positive trend (p<0.05) with increasing
26 employment of <2 years, 2 to 9 years, 10 to 19 years, and 20+ years was observed for truck
27 drivers but not for other MER occupations. A statistically nonsignificant excess risk was also
28 observed for heavy equipment operators, bus drivers, taxi drivers and chauffeurs, and mechanics
29 employed for 10 years or more. All of the above-mentioned OR were derived, adjusted for birth
30 cohort, usual daily cigarette use, and State of residence. Exposure to other occupational suspect
31 lung carcinogens did not account for the excess risks detected.
32 Results of this large study provide evidence that workers in MER jobs are at an excess
33 risk of lung cancer that is not explained by their smoking habits or exposures to other lung
34 carcinogens. Because no information on type of engine had been collected, it was not possible to
35 determine If llic excess risk was due to exposure to diesel exhaust or gasoline exhaust or a
7/25/00 7-34 DRAFT—DO NOT CITE OR QUOTE
-------
mixture of the two. Among the study's other limitations are a possible bias due to
misclassification of jobs reported by the large proportion of next-of-kin interviews. Such a bias
3 would make the effect of diesel exhaust harder to detect due to broad categorization of jobs and
4 the problems in classifying individuals into uniform occupational groups based on the pooled
5 data in the three studies that used different occupational classification schemes.
6
7 7.2.2.7. Steenland et al (1990): A Case-Control Study of Lung Cancer and Truck Driving in
8 the Teamsters Union
9 Steenland et al. conducted a case-control study of lung cancer deaths in the Teamsters
10 Union to determine the risk of lung cancer among different occupations. Death certificates were
11 obtained from the Teamsters Union files in the central States for 10,485 (98%) male decedents
12 who had filed claims for pension benefits and who had died in 1982 and 1983. Individuals were
13 required to have 20 years' tenure in the union to be eligible to claim benefits. Cases comprised
14 all deaths (n = 1,288) from lung cancer, coded as ICD 162 or 163 for underlying or contributory
15 cause on the death certificate. The 1,452 controls comprised every sixth death from the entire
16 file, excluding deaths from lung cancer, bladder cancer, and motor vehicle accidents. Detailed
17 information on work history and potential confounders such as smoking, diet, and asbestos
exposure was obtained by questionnaire.' Seventy-six percent of the interviews were provided by
spouses and the remainder by some other next of kin. The response rate was 82% for cases and
20 80% for controls. Using these interview data and the 1980 census occupation and industry
21 codes, subjects were classified either as nonexposed or as having held other jobs with potential
22 diesel exhaust exposure. Data on job categories were missing for 12% of the study subjects. A
23 second work history file was also created based on the Teamsters Union pension application that
24 lists occupation, employer, and dates of employment. A three-digit U.S. census code for
25 occupation and industry was assigned to each job for each individual. This Teamsters Union
26 work history file did not have information on whether men drove diesel or gasoline trucks, and
27 the four principal occupations were long-haul drivers, short-haul or city drivers, truck mechanics,
28 and dockworkers. Subjects were assigned the job category in which they had worked the
29 longest.
30 The case-control analysis was done using unconditional logistic regression. Separate
31 analyses were conducted for work histories from the Teamsters Union pension file and from
32 next-of-kin interviews. Covariate data were obtained from next-of-kin interviews. Analyses
33 were also performed for two time periods: employment after 1959 and employment after 1964.
34 These two cut-off years reflect years of presumed dieselization: 1960 for most trucking
companies and 1965 for independent driver and nontrucking firms. Data for analysis could be
7/25/00 7-35 DRAFT—DO NOT CITE OR QUOTE
-------
1 obtained for 994 cases and 1,085 controls using Teamsters Union work history and for 872 cases
2 and 957 controls using next-of-kin work history. When exposure was considered as a
3 dichotomous variable, for both Teamsters Union and next-of-kin work history, no single job
4 category had an elevated risk. From the next-of-kin data, diesel truck drivers had an odds ratio of
5 1.42 (95% CI = 0.74, 2.47) and diesel truck mechanics had an odds ratio of 1.35 (95% CI = 0.74,
6 2.47). OR by duration of employment as a categorical variable were then estimated. For the
7 Teamsters Union work history data, when only employment after 1959 was considered, both
8 long-haul (p<0.04) and short-haul drivers (not significant) showed an increase in risk with
9 increased years of exposure. The length-of-employment categories for which the trends were
10 analyzed were 1 to 11 years, 12 to 17 years, and 18 years or more. Using 1964 as the cutoff date,
11 long-haul drivers continued to show a significant positive trend (p=0.04), with an odds ratio of
12 1.64 (95% CI = 1.05, 2.57) for those who worked for 13+ years, the highest category. Short-haul
13 drivers, however, did not show a positive trend when 1964 was used as the cutoff date. Similar
14 trend analysis was done for most next-of-kin data. A marginal increase in risk with increasing
15 duration of employment as a truck driver (p=0.12) was observed. For truck drivers who
16 primarily drove diesel trucks for 35 years or longer, the odds ratio for lung cancer was 1.89 (95%
17 CI = 1.04, 3.42). Similarly, the corresponding odds ratio was 1.34 (95% CI = 0.81, 2.22) for
18 both gasoline truck drivers and drivers who drove both types of trucks, and 1.09 (95% CI = 0.44,
19 2.66) for truck mechanics.
20 No significant interactions between age and diesel exhaust exposure or smoking and
•21 diesel exhaust exposure were observed. All the OR were adjusted for age, smoking, and asbestos
22 in addition to various exposure categories.
23 This is a well-designed and analyzed study. The main strengths of the study are the
24 availability of detailed records from the Teamsters Union, a relatively large sample size,
25 availability of smoking data, and measurements of exposures. The authors acknowledge some
26 limitations of this study, which include possible misclassifications of exposure and smoking
27 habits, as information was provided by next of kin; lack of sufficient latency to observe lung
28 cancer excess; and a small nonexposed group (n = 120). Also, they could not evaluate the
29 concordance between Teamsters Union and next-of-kin job categories easily because job
30 categories were defined differently in each data set. No data were available on levels of diesel
31 exposure for the different job categories. Despite these limitations, the positive findings of this
32 study, which are probably underestimated, provide a positive evidence toward causal association
33 between diesel exhaust exposure and excess lung cancer.
34
7/25/00 7-36 DRAFT—DO NOT CITE OR QUOTE
-------
1 7.2.2.8. Steenland et al (1998): Diesel Exhaust and Lung Cancer in the Trucking Industry:
£| Exposure-Response Analyses and Risk Assessment
3 Steenland et al. (1998) conducted an exposure-response analysis by supplementing the
4 data from their earlier case-control study of lung cancer and truck drivers in the Teamsters Union
5 (Steenland et al., 1990) with exposure estimates based on a 1990 industrial hygiene survey of
6 elemental carbon exposure, a surrogate for diesel exhaust in the trucking industry.
7 Study subjects were long-term Teamsters enrolled in the pension system who died during
8 the period 1982-1983. Using death certificate information, the researchers identified 994 cases
9 of lung cancer for the study period, and 1,085 non-lung-cancer deaths served as controls.
10 Subjects were divided into job categories based on the job each held the longest. Most had held
11 only one type of job. The job categories were short-haul driver, long-haul driver, mechanic,
12 dockworker, other jobs with potential diesel exposure, and jobs outside the trucking industry
13 without occupational diesel exposure. Smoking histories were obtained from next of kin. OR
14 were calculated for work in an exposed job category at any time and after 1959 (an estimated
15 date when the majority of heavy-duty trucks had converted to diesel) compared with work in
16 nonexposedjobs. OR were adjusted for age, smoking, and potential asbestos exposure. Trends
17 in effect estimates for duration of work in an exposed job were also calculated.
An industrial hygiene survey by Zaebst et al. (1991) of elemental carbon exposures in the
trucking industry provided exposure estimates for each job category in 1990. The elemental
20 carbon measurements were generally consistent with the epidemiologic results, in that mechanics
21 were found to have the highest exposures and relative risk, followed by long-haul and then
22 short-haul drivers, although dockworkers had the highest exposures and the lowest relative risks.
23 Past exposures were estimated assuming that they were a function of (1) the number of
24 heavy-duty trucks on the road, (2) the particulate emissions (grams/mile) of diesel engines over
25 time, and (3) leaks from truck exhaust systems for long-haul drivers. Estimates of past exposure
26 to elemental carbon, as a marker for diesel exhaust exposure, for subjects in the case-control
27 study were made by assuming that average 1990 levels for ajob category could be assigned to all
28 subjects in that category, and that levels prior to 1990 were directly proportional to vehicle miles
29 traveled by heavy-duty trucks and the estimated emission levels of diesel engines. A 1975
30 exposure level of elemental carbon in terms of micrograms per cubic meter was estimated by the
31 following equation: 1975 level = 1990 level*(vehicle miles 1975/vehicle miles 1990) (emissions
32 1975/emissions 1990). Once estimates of exposure for each year of work history were derived
33 for each subject, analyses were conducted by cumulative level of estimated carbon exposure.
34 Estimates were made for long-haul drivers (n = 1,237), short-haul drivers (n = 297),
dockworkers (n = 164), mechanics (n = 88), and those outside the trucking industry (n = 150).
7/25/00 7-37 DRAFT—DO NOT CITE OR QUOTE
-------
1 Logistic regression was used to estimate OR adjusted for five categories of age, race, smoking
2 (never, former-quitting before 1963, former-quitting in 1963 or later, current-with <1 pack per
3 day, and current-with 1 or more packs per day), diet, and reported asbestos exposure. A variety
4 of models for cumulative exposure were considered, including a log-linear model with
5 cumulative exposure, a model adding a quadratic term for cumulative exposure, a log transform
6 of cumulative exposure, dummy variables for quartile of cumulative exposure, and smoothing
7 splines of cumulative exposure. The estimates of rate ratios from logistic regression for specific
8 levels of exposure to elemental carbon were then used to derive excess risk estimates for lung
9 cancer after lifetime exposure to elemental carbon.
10 The survey found that mechanics had the highest current levels of diesel exhaust
11 exposures and dockworkers who mainly used propane-powered forklifts had the lowest exposure.
12 ORs of 1.69 and 0.93 were observed for the mechanics and dockworkers, respectively. The
13 finding of the highest lung cancer risk for mechanics and lowest for dockworkers is indicative of
14 causal association between the diesel exhaust exposure and development of lung cancer. The log
15 of cumulative exposure was found to be the best-fitting model and was a significant predictor (p
16 = 0.01). However, the risk among mechanics did not increase with increasing duration of
17 employment.
18 OR for quartile of cumulative exposure show a pattern of significantly increasing trends
19 in risk with increasing exposure, ranging between 1.08 and 1.72, depending on the exposure level
20 and lag structure used. The lifetime excess risk of lung cancer death (through age 75) for a male
21 truck driver was estimated to be in the range of 1.4%-2.3% (95% confidence limits ranged from
22 0.3% to 4.6%) above the background risk, depending on the emissions scenarios assumed. The
23 authors found that current exposures indicated that truck drivers are exposed to diesel exhaust at
24 levels about the same as ambient levels on the highways., which are about double the background
25 levels in urban air. They conclude that the data suggest a positive and significant increase in
26 lung cancer risk with increasing estimated cumulative exposure to diesel exhaust among workers
27 in the trucking industry. They assert that these estimates suggest that the lifetime excess risk for
28 lung cancer is 10 times higher than the OSHA standards, but caution that the results should be
29 viewed as exploratory.
30 The authors acknowledge that the increasing trend in risk with increasing estimates of
31 cumulative exposure is partly due to the fact that a component of cumulative dose is simple
32 duration of exposure, and that analyses by simple duration also exhibit a positive trend with
33 duration. This analysis essentially weights the duration by contrived estimates of exposure
34 intensity, and the authors acknowledge that this weighting depends on very broad assumptions.
7/25/00 7-38 DRAFT—DO NOT CITE OR QUOTE
-------
This is not an analysis of new data that provides independent estimates of relative risk for
diesel exhaust and lung cancer incidence. Instead, it is an attempt to convert the data from
3 Steenland's earlier study of lung cancer for the purpose of estimating a different risk metric,
4 "lifetime excess risk of lung cancer," by augmenting these data with limited industrial hygiene
5 data and rationalizations about plausible models for cumulative exposure.
6 The Health Effects Institute (HEI, 1999) and others have raised some questions about the
7 exposure estimations and control for confounding variables. EPA and NIOSH will address these
8 concerns in the year 2000. It should be noted that these concerns are about the use of these data
9 for quantitative risk assessment. As far as qualitative risk assessment is concerned, this study is
10 still considered to be positive and strong.
11
12 7.2.2.9. Boffetta et aL (1990): Case-Control Study on Occupational Exposure to Diesel
13 Exhaust and Lung Cancer Risk
14 This is an ongoing (since 1969) case-control study of tobacco-related diseases in 18
15 hospitals (six U.S. cities). Cases comprise 2,584 males with histologically confirmed primary
16 lung cancers. Sixty-nine cases were matched to 1 control, whereas 2,515 were matched to 2
17 controls. Controls were individuals who were diagnosed with non-tobacco-related diseases. The
« matching was done for sex, age (±2 years), hospital, and year of interview. The interviews were
conducted at the hospitals at the time of diagnosis. In 1985, the occupational section of the
20 questionnaire was modified to include the usual occupation and up to five other jobs as well as
21 duration (in years) worked in those jobs. After 1985, information was also obtained on exposure
22 to 45 groups of chemicals, including diesel exhaust at the workplace or during hobby activities.
23 A priori aggregation of occupations was categorized into low probability of diesel exhaust
24 exposure (reference group), possible exposure (19 occupations), and probable exposure (13
25 occupations). Analysis was conducted based on "usual occupation" on all study subjects, and
26 any occupation with sufficient cases was eligible for further analysis. In addition, cases enrolled
27 after 1985 for which there were self-reported diesel exhaust exposure and detailed work histories
28 were also analyzed separately.
29 Both matched and unmatched analyses were done by calculating the adjusted (for
30 smoking and education) relative odds using the Mantel-Haenzael method and calculating the test-
31 based 95% confidence interval using the Miettinen method. Unconditional logistic regression
32 was used to adjust for potential confounders (the PROC LOGIST of SAS). Linear trends for risk
33 were also tested according to Mantel.
Adjusted relative odds for possible and probable exposure groups as well as the truck
drivers were slightly below unity, none being statistically significant for the entire study
7/25/00 7-39 DRAFT—DO NOT CITE OR QUOTE
-------
1 population. Although slight excesses were observed for the self-reported diesel exhaust exposure
2 group and the subset of post-1985 enrollees for highest duration of exposure (for self-reported
3 exposure, occupations with probable exposure, and truck drivers), none was statistically
4 significant. Trend tests for the risk of lung cancer among self-reported diesel exhaust exposure,
5 probable exposure, and truck drivers with increasing exposure (duration of exposure used as
6 surrogate for increasing dose) were nonsignificant too. Statistically significant lung cancer
7 excesses were observed for cigarette smoking only.
8 The major strength of this study is availability of detailed smoking history. Even though
9 detailed information was obtained for the usual and five other occupations (1985), because it was
10 difficult to estimate or verify the actual exposure to diesel exhaust, duration of employment was
11 used as a surrogate for dose instead. The numbers of cases and controls were large; however, the
12 number of individuals exposed to diesel exhaust was relatively few, thus reducing the power of
13 the study. This study did not attempt latency analysis either. Due to these limitations, the
14 findings of this study are unable to provide either positive or negative evidence for a causal
15 association between diesel exhaust and occurrence of lung cancer.
16
17 7.2.2.10. Emmelin et al. (1993): Diesel Exhaust Exposure and Smoking: A Case-Referent
18 Study of Lung Cancer Among Swedish Dock Workers
19 This case-control study of lung cancer was drawn from a cohort defined as all male
20 workers who had been employed as dockworkers for at least 6 months between 1950 and 1974.
21 In the population of 6,573 from 20 ports, there were 90 lung cancer deaths (cases), identified
22 through Swedish death and cancer registers, during the period 1960 to 1982. Of these 90 deaths,
23 the 54 who were workers at the 15 ports for which exposure surrogate information was available
24 were chosen for the case-control study. Four controls, matched on port and age, were chosen for
25 each case from the remaining cohort who had survived to the time of diagnosis of the case. Both
26 live and deceased controls were included. The final analyses were done on 50 cases and 154
27 controls who had complete information on employment dates and smoking data. The smoking
28 strata were created by classifying ex-smokers as nonsmokers if they had not smoked for at least 5
29 years prior to the date of diagnosis of the case; otherwise they were classified as smokers.
30 Relative odds and regression coefficients were calculated using conditional logistic
31 regression models. Comparisons were made both with and without smoking included as a
32 variable, and the possible interaction between smoking and diesel exhaust was tested. Both the
33 weighted linear regressions of the adjusted relative odds and the regression coefficients were
34 used to test mortality irends with aii three exposure variables.
7/25/00 7-40 DRAFT—DO NOT CITE OR QUOTE
-------
Exposure to diesel exhaust was assessed indirectly by initially measuring (1) exposure
intensity based on exhaust emission, (2) characteristics of the environment in terms of
"3 ventilation, and (3) measures of proportion of time in higher exposed jobs. For exhaust
4 emissions, annual diesel fuel consumption at a port was used as the surrogate. For ventilation,
5 the annual proportion of ships with closed or semiclosed holds was used as the surrogate. The
6 proportion of time spent below decks was used as the surrogate for more exposed jobs. Although
7 data were collected for all three measures, only the annual fuel consumption was used for
8 analysis. Because every man was likely to rotate through the various jobs, the authors thought
9 using annual consumption of diesel fuel was the appropriate measure of exposure.
10 Consequently, in a second analysis, the annual fuel consumption was divided by the number of
11 employees in the same port that year to come up with the fuel-per-person measure, which was
12 further used to create a second measure, "exposed time." The "annual fuel" and exposed-time
13 data were entered in a calendar time-exposure matrix for each port, from which individual
14 exposure measures were created. A third measure, "machine time" (years of employment from
15 first exposure), was also used to compare the results with other studies. All exposure measures
16 were accumulated from the first year of employment or first year of diesel machine use,
17 whichever came later. The last year of exposure was fixed at 1979. All exposures up to 2 years
before the date of lung cancer diagnosis were omitted from both cases and matched controls. A
priori classification into three categories of low, medium, and high exposure was done for all
20 three exposure variables: machine time, fuel, and exposed time.
.21 Conditional logistic regression models, adjusting for smoking status and using low
22 exposures and/or nonsmokers as a comparison group, yielded positive trends for all exposure
23 measures, but no trend test results were reported, and only the relative odds for the exposed-time
24 exposure measure in the high-exposure group (OR = 6.8, 90% CI = 1.3 to 34.9) was reported as
25 statistically significant. For smokers, adjusting for diesel exhaust exposure level, the relative
26 odds were statistically significant and about equal for all three exposure variables: machine time,
27 OR = 5.7 (90% CI = 2.4 to 13.3); fuel, OR = 5.5 (90% CI = 2.4 to 12.7); and exposed time, OR =
28 6.2 (90% CI = 2.6 to 14.6). Interaction between diesel exhaust and smoking was tested by
29 conditional logistic regression in the exposed-time variable. Although there were positive trends
30 for both smokers and nonsmokers, the trend for smokers was much steeper: low, OR = 3.7 (90%
31 CI = 0.9 to 14.6); medium, OR = 10.7 (90% CI = 1.5 to 78.4); and high, OR = 28.9 (90% CI =
32 3.5 to 240), indicating more than additive interaction between these two variables.
33 In the weighted linear regression model with the exposed-time variable, the results were
34 similar to those using the logistic regression model. The authors also explored the smoking
variable further in various analyses, some of which suggested a strong interaction between diesel
7/25/00 7-41 DRAFT—DO NOT CITE OR QUOTE
-------
1 exhaust and smoking. However, with just six nonsmokers and' no further categorization of
2 smoking amount or duration, these results are of limited value.
3 The diesel exhaust exposure matrices created using three different variables are intricate.
4 Analyses by any of these variables yield essentially the same positive results and positive trends,
5 providing consistent support for a real effect of diesel exhaust exposure, at least in smokers.
6 However, methodological limitations to this study prevent a more definitive conclusion. The
7 numbers of cases and controls are small. There are very few nonsmokers; thus, testing the
8 effects of diesel exhaust exposure in them is futile. Lack of information on asbestos exposure, to
9 which dockworkers are usually exposed, may also confound the results. Also, no latency
10 analyses are presented. Overall, despite these limitations, this study supports the earlier findings
11 of excess lung cancer mortality among individuals exposed to diesel exhaust.
12
13 7.2.2.11. Swanson etaL (1993): Diversity in the Association Between Occupation and Lung
14 Cancer Among Black and White Men
15 This population-based case-control study of lung cancer was conducted in metropolitan
16 Detroit. The cases and controls for this study were identified from the Occupational Cancer
17 Incidence Surveillance Study (OCISS). A total of 3,792 incident lung cancer cases and 1,966
18 colon and rectal cancer cases used as controls, diagnosed between 1984 and 1987 among white
19 and black males aged 40 to 84 years, were selected for the study. Information was obtained by
20 telephone interview either with the individual or a surrogate about lifetime work history and
21 smoking history, as well as medical, demographic, and residential history. Occupation and
22 industry data were coded using the 1980 U.S. Census Bureau classification codes. The
23 investigators selected certain occupations and industries as having little or no exposure to
24 carcinogens and defined them as an unexposed group. Analysis was done using logistic
25 regression method and adjusting for age at diagnosis, pack-years of cigarette smoking, and race.
26 The results were presented by various occupations and industries; those with potential
27 exposures to diesel exhaust were drivers of heavy trucks and light trucks, farmers, and railroad
28 workers, respectively. Among white males, increasing iiing cancer risks were observed with
29 increasing duration of employment for drivers of heavy trucks, drivers of light trucks, and
30 farmers. Although none of the individual ORs were statistically significant, trend tests were
31 significant for all three occupations (psO.5). On the other hand, among black males increasing
32 lung cancer risks with increasing duration of employment were observed for farmers only, with
33 an OR of 10.4 (95% CI = 1.4, 77.1) reaching significance for employment of 20+ years. As for
34 the railroad industry, increasing iung cancer risks with increasing duration of employment were
15/00 7-42 DRAFT—DO NOT CITE OR QUOTE
-------
observed for both white and black males. The trend test was significant for white males only,
with an OR of 2.4 (95% CI = 1.1, 5.1) reaching significance for employment of 10+ years.
3 The main strengths of the study are large sample size, availability of lifetime work history
4 and smoking history, and the population-based study format, precluding selection bias. The
5 major limitation, as in other studies, is lack of direct information on specific exposures. The
6 interesting result of this study is lung cancer excesses observed in farmers, mainly among crop
7 farmers, who have potential exposure to diesel exhaust from their tractors in addition to
8 pesticides, herbicides, and other PM10. The authors point out that this is the first study to find
9 excess lung cancer in this occupation.
10
11 7.2.2.12. Hansen et aL (1998): Increased Risk of Lung Cancer Among Different Types of
12 Professional Drivers in Denmark
13 This is a population-based case-control study of lung cancer, conducted in professional
14 drivers in Denmark. The cases first diagnosed as primary lung cancer between 1970 and 1989
15 among males born between 1897 and 1966 were identified from the Danish Cancer Registry.
16 The registry provided the information on diagnosis from ICD-7, name, sex, and unique personal
17 identification number (PIDN). Information about past employment was obtained by linkage with
the nationwide pension fund. The fund keeps the records by name and PIDN about the date of
start and end of each job and unique company number of the employer. The records are kept
20 even after the employee has retired or died. Information about current employment was obtained
21 from the Danish Central Population Registry (CPR) by linkage with the PIDN.
22 Of 37,597 cases identified from the Registry, 8,853 did not have any employment
23 records. Controls (1:1) for 28,744 lung cancer cases with employment histories were selected
24 randomly from CPR, matched with the case by year of birth and sex. Furthermore, these controls
25 had to be alive, cancer free, and employed prior to the diagnosis of lung cancer in the
26 corresponding case. Employment histories were obtained for the controls in the same fashion as
27 cases from the pension fund. The employment record search resulted in a total of 1,640 lorry/bus
28 drivers and 426 taxi drivers. They were further divided into subgroups by their duration of
29 employment. Information about smoking in drivers was acquired from two national surveys
30 conducted in 1970-72 and 1983. No direct information on smoking was available in either cases
31 or controls. A separate case-control study of mesothelioma indirectly looked at asbestos
32 exposure among professional drivers. OR, adjusting for socioeconomic status and 95% CI, were
33 computed using conditional logistic regression (PECAN procedure in the statistical package
34 EPICURE).
7/25/00 7-43 DRAFT—DO NOT CITE OR QUOTE
-------
1 Significant ORs for lung cancer were found for lorry/bus drivers (OR =1.31, 95% CI =
2 1.17, 1.46), taxi drivers (OR = 1.64,95% CI = 1.22, 2.19), and unspecified drivers (OR = 1.39,
3 95% CI = 1.30, 1.51). Significant ORs were found for both lorry/bus drivers and taxi drivers by
4 duration of employment in 1-5 years and >5 years categories, with no lag time and with a 10-
5 year lag time. The OR remained the same for lorry/bus drivers in these employment categories
6 for no lag time and 10-year lag time. Among taxi drivers, on the other hand, the OR of 2.2 in >5
7 year employment in no-lag-time analysis increased to 3.0 in the 10-year lag time analysis. The
8 authors asserted that the higher risk seen in the taxi drivers may be due to higher exposure
9 attributable due to longer time spent in traffic congestion. The trend tests for increasing risk with
10 increasing duration of employment (surrogate for exposure) were statistically significant
11 (pO.OOl) for both lorry/bus drivers and taxi drivers in no-lag-time and 10-year lag time
12 analysis. All the ORs were adjusted for socioeconomic status.
13 The main strengths of the study are the large sample size, availability of information on
14 socioeconomic status, and detailed employment records. The main limitation, however, is lack
15 of information on what type of fuel these vehicles used. It is probably safe to assume that the
16 lorry/buses were diesel powered, whereas the taxis could be either diesel or gasoline powered. A
17 personal communication with Dr. Johnni Hansen confirmed that lorries, buses, and taxis have
18 been using diesel fuel since the beginning of the 1960s. Although direct adjustments were not
19 done for smoking and exposure to asbestos, indirect information on both these confounders
20 indicates that they are unlikely to explain the observed excesses and the increasing risk with
21 increasing duration of employment. Thus, the results of this study are strongly supportive of
22 diesel exhaust being associated with increased lung cancer.
23
24 7.2.2.13. Briiske-Hohlfeld et al. (1999): Lung Cancer Risk in Male Workers Occupationally
2 5 Exposed to Diesel Motor Emissions in Germany
26 This paper presents a pooled analysis of two case-control studies of lung cancer. The first
27 study, by Jockel et al. (1995,1998), was conducted between 1988 and 1993 and had 1,004 cases
28 and 1,004 controls matched for sex, age, and region of residence, selected randomly from the
29 compulsory municipal registries. The inclusion criteria for cases were: they should have been
30 born in or after 1913, should have been of German nationality, and should have been diagnosed
31 with lung cancer within 3 months prior to the interview. The second study, by Wichmann et al.
32 (1998), was ongoing when it was included in this study. The study span covered the years 1990
33 to 1996. By 1994 a total of 3,180 cases and 3,249 controls, randomly selected from the
34 compulsory population registries, were frequency matched on sex, age, arid region. Trie cases
35 were iess man 76 years old, were residents of me region and riving in Germany for more man 25
•7 /i c /nn
/ / ^—'i uu
7-44 DRAFT—DO NOT CITE OR QUOTE
-------
years, and had a diagnosis not more than 3 months old. Of 4,184 pooled cases and 4,253 pooled
controls, the analysis was conducted on 3,498 male cases and 3,541 male controls. A personal
3 interview was conducted with each study participant. Data were collected on basic demographic
4 information, detailed smoking history, and lifelong occupational history about jobs held and
5 industries worked in. The job titles and industries were classified into 33 and 21 categories,
6 respectively, using the German Statistical Office codes.
7 Based on job codes with potential exposure to diesel motor emission (DME), four
8 exposure groups were constituted. Group A comprised professional drivers of trucks, buses,
9 taxis, etc. Group B comprised other traffic-related jobs such as switchmen, diesel locomotive
10 drivers, and diesel forklift truck drivers. Group C comprised bulldozer operators, graders, and
11 excavators. Group D comprised full-time farm tractor drivers. Validation of the jobs was done
12 by written evaluation of the job task descriptions, which also avoided misclassification. The
13 following information was acquired for the construction of job task descriptions: (1) What were
14 your usual tasks at work and how often (in % of daily working hours) were they performed? (2)
15 What did you produce, manufacture, or transport? (3) Which material was used? (4) What kind
16 of machine did you operate? Some individuals had more than one job task with DME exposure.
17 The exposure assessment was done without knowing the status of the case/control.
For each individual, cumulative exposure was calculated for the complete work history
by categorizing the duration of exposure as >0-3, >3-10, >10-20, >20-30, >30 years, and
20 beginning and end of exposure. The first year of exposure was defined as <1945, 1946-1955,
21 and >1956 while the last year of exposure was defined as <1965,1966-1975, and >1976. For
22 professional drivers, hours driven per day were accumulated and were classified as "driving
23 hours."
24 A smoker was defined as any individual who had smoked regularly for at least 6 months.
25 Smoking information was acquired in series with the starting time, type of tobacco, amount
26 smoked, duration in years, and calender year of quitting. Asbestos exposure was estimated by
27 certain job-specific supplementary questions.
28 The cases and controls were post-hoc stratified into 6 age and 17 region categories. OR
29 adjusted for smoking and asbestos exposure were calculated by conditional logistic regression,
30 using "never exposed" workers as the reference group. The adjustment for cigarette smoking
31 was done by using pack-years as a continuous variable; adjustment for other tobacco products
32 was done by considering them as a binary variable. A total of 716 cases and 430 controls were
33 found to be ever exposed to DME. The smoking- and asbestos-adjusted OR of 1.43 (95% CI =
1.23. 1.67) for all DME exposed was reduced from the crude OR of 1.91. For the entire group
the various analyses yielded statistically significant ORs ranging from 1.25 to 2.31, adjusted for
36 smoking and asbestos exposure (West Germany, >10-20 years and >20-30 years of exposure,
7/25/00 7-45 DRAFT—DO NOT CITE OR QUOTE
-------
1 first year of exposure in 1946-1955 and 1956+, end of exposure in 1966-1975 and 1976+, and for
2 the job categories of Group A, B, and C). The risk increased with increasing years of exposure,
3 and for both the first year of exposure (<1945,1946-1955, and >1956) and end year of exposure
4 (<1965, 1966-1975, and >1976).
5 Separate analyses by four job categories (all the ORs were adjusted for smoking and
6 asbestos exposure) showed that for professional drivers (Group A) the overall OR was 1.25 (95%
7 CI = 1.05, 1.47). Significant ORs were found for various factors in West Germany only. The
8 factors were: >0-3 years and >10-20 years of exposure (OR = 1.69, 95% CI = 1.13, 2.53, and
9 OR = 2.02, 95% CI = 1.32, 3.08, respectively), beginning of exposure in 1956+ and end of
10 exposure in 1976+ (OR = 1.56, 95% CI = 1.21, 2.03, and OR= 1.5, 95% CI = 1.14, 1.98,
11 respectively), and 1,000-49,999 driving hours (OR = 1.54, 95% CI = 1.15, 2.07). None of the
12 ORs were significant in East Germany in this group.
13 For other traffic-related jobs (Group B) the overall OR was 1.53 (95% CI = 1.04, 2.24).
14 The ORs for beginning of exposure in 1956+ and end of exposure in 1976+ were OR = 1.71,
15 95% CI = 1.05, 2.78, and OR = 2.68, 95% CI = 1.47, 4.90, respectively. The risk increased with
16 increasing duration of exposure and was statistically significant for > 10-20 years (OR = 2.49)
17 and more than 20 years (OR = 2.88). No separate analyses for West Germany and East Germany
18 were presented in this category.
19 For heavy equipment operators (Group C) the overall OR of 2.31 (95% CI = 1.44, 3.7)
20 was highest among all the job categories. Significant ORs were observed for beginning exposure
.21 in 1946-1955 (OR = 2.83, 95% CI = 1.10, 7.23) and end exposure in 1966-1975 (OR = 3.74,
22 95% CI = 1.20, 11.64). The risk increased with increasing duration of exposure and was
23 statistically significant for more than 20 years of exposure (OR = 4.3). Although no separate
24 analyses for West Germany and East Germany were presented, investigators mentioned that for
25 this job group hardly any difference was seen between West Germany and East Germany.
26 For drivers of the farming tractors (Group D) the overall OR of 1.29 was not significant.
27 Risk increased with increasing duration of exposure and was significant for exposure of more
28 than 30 years (OR = 6.81, 95% CI = 1.17, 39.51). No separate analyses for West Germany and
29 East Germany were presented in this category
30 The professional drivers and the other traffic-related job categories probably have mixed
31 exposures to gasoline exhaust in general traffic. On the other hand, it should be noted that
32 exposure to DME among heavy equipment and farm tractor drivers is much higher and not as
33 mixed as in professional drivers. The heavy equipment drivers usually drive repeatedly through
34 their own equipment's exhaust. Therefore, the observed highest risk for lung cancer in this job
35 category establishes a direci link with the DME. The only other study thai found significantly
36 higher risk for heavy equipment operators (RR = 2.6) was conducted by Boffeta et al. (1988).
7/25/00 7-46 DRAFT—DO NOT CITE OR QUOTE
-------
Although the only significant excess was observed for farming tractor operators among
individuals with more than 30 years of exposure, a steady increase in risk was observed for this
3 job category with increasing exposure. The investigators stated that the working conditions and
4 the DME of tractors remained fairly constant over the years. This increase may be due mainly to
5 exposure to DME and, in addition, PMIO
6 This is a well-designed, well-conducted, and well-analyzed study. Its main strengths are
7 large sample size, resulting in good statistical power; inclusion of incident cases that were
8 diagnosed not more than 3 months prior to the interview; use of only personal interviews,
9 reducing recall bias; diagnosis ascertained by cytology or histology; and availability of lifelong
10 detailed occupational and smoking history. Exposure estimation for each individual was based
11 on job codes and industry codes, which were validated by written job descriptions to avoid
12 misclassification. The main limitation of the study is lack of data on actual exposure to DME.
13 The cumulative quantitative exposures were calculated based on time spent in each job with
14 potential exposure to DME and the type of equipment used. Thus, this study provides strong
15 evidence for a causal association between exposure to diesel exhaust and occurrence of lung
16 cancer.
17 Table 7-2 summarizes the above lung cancer case-control studies.
7.2.3. Summaries of Studies and Meta-Analyses of Lung Cancer
20 7.2.3.1. Cohen and Higgins (1995): Health Effects of Diesel Exhaust: Epidemiology
21 The Health Effects Institute (HEI) reviewed all published epidemiologic studies on the
22 health effects of exposure to diesel exhaust available through June 1993, identified by a
23 MEDLINE search and by reviewing the reference sections of published research and earlier
24 reviews. HEI identified 35 reports of epidemiologic studies (16 cohort and 19 case-control) of
25 the relation of occupational exposure to diesel emissions and lung cancer published between
26 1957 and 1993.
27 HEI reviewed the 35 reports for epidemiologic evidence of health effects of exposure to
28 diesel exhaust for lung cancer, other cancers, and nonmalignant respiratory disease. They found
29 that the data were strongest for lung cancer. The evidence suggested that occupational exposure
30 to diesel exhaust from diverse sources increases the rate of lung cancer by 20% to 40% in
31 exposed workers generally, and to a greater extent among workers with prolonged exposure.
32 They also found that the results are not explicable by confounding caused by cigarette smoking
33 or other known sources of bias.
Control for smoking was identified in 15 studies. Six studies (17%) reported relative risk
estimates less than 1; 29 studies (83%) reported at least relative risk indicating positive
7/25/00 7-47 DRAFT—DO NOT CITE OR QUOTE
-------
1 association. Twelve studies indicating a relative risk greater than 1 had 95% confidence
2 intervals, which excluded unity.
3 The authors conclude that epidemiologic data consistently show weak associations
4 between exposure to diesel exhaust and lung cancer. They find that the evidence suggests that
5 long-term exposure to diesel exhaust in a variety of occupational circumstances is associated
6 with a 1.2- to 1.5-fold increase in the relative risk of lung cancer compared with workers
7 classified as unexposed. Most of the studies that controlled for smoking found that the
8 association between increased risk of lung cancer and exposure to diesel exhaust persisted after
9 such controls were applied, although in some cases the excess risk was lower. None of the
10 studies measured exposure to diesel emissions or characterized the actual emissions from the
11 source of exposure for the time period most relevant to the development of lung cancer. Most
12 investigators classified exposure on the basis of work histories reported by subjects or their next
13 of kin, or by retirement records. Although these data provide relative rankings of exposure, the
14 absence of concurrent exposure information is the key factor that limits interpretation of the
15 epidemiologic findings and subsequently their utility in making quantitative estimates of cancer
16 risks.
17 This is a comprehensive and thorough narrative review of studies of the health effects of
18 diesel exhaust. It does not undertake formal estimation of summary measures of effect or
19 evaluation of heterogeneity in the results. The conclusion drawn about the consistency of the
20 results is based on the author's assessment of the failure of potential biases and alternative
21 explanations for the increase in risk to account for the observed consistency. In many if not most
22 studies, the quality of the data used to control confounding was relatively crude. Although the
23 studies do include qualitative assessment of whether control for smoking is taken into account,
24 careful scrutiny of the quality of the control or adjustment for smoking among the studies is
25 absent. This leaves open the possibility that prevalent residual confounding by inadequate
26 control for smoking in many or most studies may account for the consistent associations seen.
27
28 7.2.3.2. Bhatia et aL (1998): Diesel Exhaust Exposure and Lung Cancer
29 Bhatia et al. (1998) report a meta-analysis of 29 published1 cohort and case-control
30 studies of the relation between occupational exposure to diesel exhaust and lung cancer. A
31 search of the epidemiologic literature was conducted for all studies concerning lung cancer and
32 diesel exhaust exposure. Occupational studies involving mining were excluded because of
33 concern about the possible influence of radon and silica exposures. Studies in which the
'Ul 35 studies laentined in the literature search, 6 pairs 01 studies represented analyses of the same study
population, reducing the number of studies to 29.
-7/->c/nn 7.48 r\v> A CT r\r\ XTrvr r^TTC no
I I 4.JI \J\J ' « ^ Av'ANU-l.A A 1-XV^ i •* ^ A >-*A A Aw X-'AV
-------
minimum interval from time of first exposure to end of follow-up was less than 10 years, and
studies in which work with diesel equipment or engines could not be confirmed or reliably
"3 inferred, were excluded. When studies presented risk estimates for more than one specific
4 occupational category of diesel exhaust-exposed workers, the subgroup risk estimates were used
5 in the meta-analysis. Smoking-adjusted effect measures were used when present.
6 Of 29 studies 23 met the criteria for inclusion in the meta-analysis. The observed relative
7 risk estimates were greater than 1 in 21 of these studies; this result is unlikely to be due to
8 chance. The pooled relative risk weighted by study precision was 1.33 (95% CI = 1.24, 1.44),
9 indicating increased relative risk for lung cancer from occupational exposure to diesel exhaust.
10 Subanalyses by study design (case-control and cohort studies) and by control for smoking
11 produced results that did not differ from those of the overall pooled analysis. Cohort studies
12 using internal comparisons showed higher relative risks than those using external comparisons.
13 (See Figure 7-1.)
14 Bhatia and colleagues conclude that the analysis shows a small but consistent increase in
15 the risk for lung cancer among workers with exposure to diesel exhaust. The authors evaluate the
16 dependence of the relative risk estimate on the presence of control for smoking among studies,
17 and provide a table that allows assessment of whether the quality of the data contributing to
control for smoking is related to the relative risk estimates (albeit in a limited number of studies).
Bhatia et al. assert that residual confounding is not affecting the summary estimates or
20 conclusions for the folio whig reasons: (1) the pooled relative risks for studies adjusted for
21 smoking were the same as those for studies not adjusting for smoking; (2) in those studies giving
22 risk estimates adjusted for smoking and risk estimates not adjusted for smoking, there was only a
23 small reduction in the pooled relative risk from diesel exhaust exposure; and (3) in studies with
24 internal comparison populations, in which confounding is less likely, the pooled relative risk
25 estimate was 1.43.
26 The validity of this assessment depends on the adequacy of control for smoking in the
27 individual studies. If inadequate adjustment for smoking is employed and residual confounding
28 by cigarette smoking pertains in the result of the individual studies, then the comparisons and
29 contrasts of the pooled estimates the authors cite as reasons for dismissing the effect of residual
30 confounding by smoking will remain contaminated by residual confounding in the individual
31 studies. In fact, Bhatia et al. erroneously identify the treatment of the smoking data in the main
32 analysis for the 1987 report by Garshick et al. as a continuous variable representing pack-years of
33 smoking, whereas the analysis actually dichotomized the pack-years data into two crude dose
34 categories (above and below the 50 pack-years level). This clearly reduced the quality of the
adjustment for smoking, which already suffered from the fact that information on cumulative
7/25/00 7-49 DRAFT—DO NOT CITE OR QUOTE
-------
1 cigarette consumption was missing for more than 20% of the lung cancer cases. In this instance,
2 the consistency between the adjusted and unadjusted estimates of the relative risk for diesel
3 exhaust exposure may be attributable to failure of adjustment rather than lack of confounding by
4 cigarette smoking, and pooled estimates of association of diesel exhaust with lung cancer derived
5 in the meta-analysis would remain confounded. A similar problem exists for the Bhatia et al.
6 representation of the control for confounding in the study by Boffetta and Stellman (1988). Such
7 mischaracterizations may indicate an overstatement by Bhatia et al. that the association of DE
8 and lung cancer is insensitive to adjustment.
9 An evaluation of the potential for publication bias is presented that provides reassurance
10 that the magnitude of published effects is not a function of the precision or study power;
11 however, this assessment cannot rule out the possibility of publication bias.
12
13 7.2.3.3. Lipsett and Campleman (1999): Occupational Exposure to Diesel Exhaust and Lung
14 Cancer: A Meta-Analysis
15 Lipsett and Campleman (1999) conducted electronic searches to identify epidemiologic
16 studies published between 1975 and 1995 of the relationship of occupational exposure to diesel
17 exhaust and lung cancer. Studies were selected based on the following criteria: (1) Estimates of
18 relative risks and their standard errors must be reported or derivable from the information
19 presented. (2) Studies must have allowed for a latency period of 10 or more years for
20 development of lung cancer after onset of exposure. (3) No obvious bias resulted from
21 incomplete case ascertainment in follow-up studies. (4) Studies must be independent: that is, a
22 single representative study selected from any set of multiple analyses of data from the same
23 population. Studies focusing on occupations involving mining were excluded because of
24 potential confounding by radon, arsenic, and silica, as well as possible interactions between
25 cigarette smoking and exposure to these substances in lung cancer induction.
26 Thirty of the 47 studies initially identified as relevant met the specified inclusion criteria.
27 Several risk estimates were extracted from six studies reporting results from multiple mutually
28 exclusive diesel-related occupational subgroups. If a study reported effects associated with
29 several levels or durations of exposure, the effect reported for the highest level, or longest
30 duration of exposure was used. If estimates for several occupational subsets were reported, the
31 most diesel-specific occupation or exposure was selected. Adjusted risk estimates were used
32 when available.
33 Thirty-nine independent estimates of relative risk and standard errors were extracted.
34 Pooled estimates of relative risk were calculated using a laiidom-effects model. Among study
7/25/00 7-50 DRAFT—DO NOT CITE OR QUOTE
-------
populations most likely to have had substantial exposure to diesel exhaust, the pooled smoking-
adjusted relative risk was 1.47 (95% CI = 1.29, 1.67). (See Figure 7-2.)
The between-study variance of the relative risks indicated the presence of significant
4 heterogeneity in the individual estimates. The authors evaluated the potential sources of
5 heterogeneity by subset analysis and linear meta-regressions. Major sources of heterogeneity
6 included control for confounding by smoking, selection bias (a healthy worker effect), and
7 exposure patterns characteristic of different occupational categories. A modestly higher, pooled
8 relative risk was derived for the subset of case-control studies, which, unlike the cohort studies,
9 showed little evidence of heterogeneity.
10 An evaluation of the potential for publication bias is presented that provides reassurance
11 that the magnitude of published effects is not a function of the precision or study power;
12 however, this assessment cannot rule out the possibility of publication bias.
13 Although a relatively technical approach was used in deriving summary estimates of
14 relative risk and the evaluation of possible sources of variation in the relative risks in this meta-
15 analysis, this approach should not be confused with rigorous evaluation of the potential
16 weaknesses among the studies included in the analysis. The heterogeneity attributable to
17 statistical adjustment for smoking was evaluated on the basis of a dichotomous assessment of
« whether control for smoking could be identified in the studies considered. This does not reflect
the adequacy of the adjustment for smoking employed in the individual studies considered. The
20 potential for residual confounding by inadequate adjustment for the influence of smoking
21 remains in the summary estimate of the relative risk.
22
23 7.2.4. Summary and Discussion
24 Certain extracts of diesel exhaust have been demonstrated as both mutagenic and
25 carcinogenic in animals and in humans. Animal data suggest that diesel exhaust is a pulmonary
26 carcinogen among rodents exposed by inhalation to high doses over long periods of time. While
27 rat lung cancer response to diesel exhaust is not suitable for dose-response extrapolation to
28 humans, the positive lung cancer response doses imply a hazard for humans. Because large
29 working populations are currently exposed to diesel exhaust and because nonoccupational
30 ambient exposures currently are of concern as well, the possibility that exposure to this complex
31 mixture may be carcinogenic to humans has become an important public health issue.
32 Because diesel emissions become diluted in the ambient air, it is difficult to study the
33 health effects in the general population. Nonoccupational exposure to diesel exhaust is
34 worldwide in urban areas. Thus, "unexposed" reference populations used in occupational cohort
studies are likely to contain a substantial number of individuals who are nonoccupationally
7/25/00 7-51 DRAFT—DO NOT CITE OR QUOTE
-------
1 exposed to diesel exhaust. Furthermore, the "exposed" group in these studies is based on job
2 titles, which in most instances are not verified or correlated with environmental hygiene
3 measurement. The issue of health effect measurement is further complicated by the fact that
4 occupational cohorts tend to be healthy and have below-average mortality, usually referred to as
5 the "healthy worker effect." Hence, the usual standard mortality ratios observed in cohort
6 mortality studies are likely to be underestimations of true risk.
7 A major difficulty with the occupational studies considered here was measurement of
8 actual diesel exhaust exposure. Because all the cohort mortality studies were retrospective,
9 assessment of health effects from exposure to diesel exhaust was naturally indirect. In these
1 0 occupational settings, no systematic quantitative records of ambient air were available. Most
1 1 studies compared men in job categories with presumably some exposure to diesel exhaust with
1 2 either standard populations (presumably no exposure to diesel exhaust) or men in other job
1 3 categories from industries with little or no potential for diesel exhaust exposure. A few studies
1 4 have included measurements of diesel fumes, but there is no standard method for the
1 5 measurement. No attempt is made to correlate these exposures with the cancers observed in any
1 6 of these studies, nor is it clear exactly which extract should have been measured to assess the
1 7 occupational exposure to diesel exhaust. All studies have relied on the job categories or self-
1 8 report of exposure to diesel exhaust. Gustavsson et al. (1990), Emmelin et al. (1993), and
1 9 Briiske-Hohlfeld et al. (1 999) estimated exposure levels by getting detailed histories of job
20 tskas/categories and computing cumulative exposures, which unfortunately were not verifiable
2 1 due to of the lack of industrial hygiene data. In the studies by Garshick et al. (1 987, 1988), the
22 diesel-exhaust-exposed job categories were verified on the basis of an industrial hygiene survey
23 done by Woskie et al. (1988a,b). The investigators found that in most cases the job titles were
24 good surrogates for diesel exhaust exposure. Also, in the railroad industry, where only persons
25 who had at least 10 years of work experience were included in the study, the workers tended not
26 to change job categories over the years. Thus, a job known only at one point in time was a
27 reasonable marker of past diesel exhaust exposure. Unfortunately, the exposure was only
28 qualitatively verified. Quantitative use of this information would have been much more
29 meaningful. Zaebst et al. (1991) conducted an industrial hygiene survey of elemental carbon
30 exposure in the trucking industry by job categories. Using these exposure measurements,
3 1 Steenland et al. ( 1 998) conducted an exposure-response analysis of their earlier lung cancer case-
32 control study (Steenland et ai., 1990). These exposure data are currently being verified and will
33 be used for quantitative risk assessment in the near future.
*3 /\ ^"^ r* r» 11 1^ «a tiri -m c it^xr/'kl^Mnrr nf^t&nti&l &vr\r\ci i»*o +r\ /^t£*o^1 ^•vVtmic'-fr *at-o rv^i*t^»-o ti-»»/-*1y ,-1»-«» r,e»»-r-
S^- I •*_/ ^rV V*fc^WfcfcA^ AAhJ A* A * *_/ i * &**^ LSXS k^AAfe*fe»A w 1 &u \S *J W»d. -V VW X*A WhJW 1. W^t_lAb* tAO V V4*i- V- &1.AJ..LA.W1 •-» J bj. IAW1X VU.L I V W 1. D «
7/25/00 7-52 DRAFT— DO NOT CITE OR QUOTE
-------
metal miners have assessed whether diesel exhaust is associated with lung cancer. Currently,
there are about 385 underground metal mines in the United States. Of these, 250 have been
3 permanently operating and 135 have been intermittently operating (Steenland, 1986).
4 Approximately 20,000 miners are employed, but not all of them are currently working in the
5 mines. Diesel engines were introduced in metal mines in the early to mid-1960s. Although all
Q these mines use diesel equipment, it is difficult to estimate how many of these miners were
7 actually exposed to diesel fumes.
8 Diesel engines were introduced in coal mines at an even later date, and their use is still
9 quite limited. In 1983, approximately 1,000 diesel units were in place in underground coal
10 mines, up from about 200 units in 1977 (Daniel, 1984). The number of units per mine varies
11 greatly; 1 mine may account for more than 100 units.
12 Even if it were possible to estimate how many miners (metal and coal) were exposed to
13 diesel exhaust, it would be very difficult to separate out the confounding effects of other potential
14 pulmonary carcinogens, such as radon decay products or heavy metals (e.g., arsenic, chromium).
15 Furthermore, the relatively short latency period limits the usefulness of these cohorts of miners.
16
17 7.2.4.1. Summary of the Cohort Mortality Studies
«The cohort studies mainly demonstrated an increase in lung cancer. Studies of bus
company workers by Waller (1981), Rushton et al. (1983), and Edling et al. (1987) failed to
20 demonstrate any statistically significant excess risk of lung cancer, but these studies have certain
21 methodological problems, such as small sample sizes, short follow-up periods (just 6 years in the
22 Rushton et al. study), lack of information on confounding variables, and lack of analysis by
23 duration of exposure, duration of employment, or latency that preclude their use in determining
24 the carcinogenicity of diesel exhaust. Although the Waller (1981) study had a 25-year follow-up
25 period, the cohort was restricted to employees (ages 45 to 64) currently in service. Employees
26 who left the job earlier, as well as those who were still employed after age 64 and who may have
27 died from cancer, were excluded.
28 Wong et al. (1985) conducted a mortality study of heavy equipment operators that
29 demonstrated a nonsignificant positive trend for cancer of the lung with length of membership
30 and latency. Analysis of deceased retirees showed a significant excess of lung cancer.
31 Individuals without work histories who started work prior to 1967, when records were not kept,
32 may have been in the same jobs for the longest period of time. Workers without job histories
33 included those who had the same job before and after 1967 and thus may have worked about 12
34 to 14 years longer; these workers exhibited significant excess risks of lung cancer and stomach
^1 cancer. If this assumption about duration of jobs is correct, then these site-specific causes can be
7/25/00 7-53 DRAFT—DO NOT CITE OR QUOTE
-------
1 linked to diesel exhaust exposure. One of the methodologic limitations of this study is that most
2 of these men worked outdoors; thus, this cohort might have had relatively low exposure to diesel
3 exhaust. The authors did not present any environmental measurement data either. Because of
4 the absence of detailed work histories for 30% of the cohort and the availability of only partial
5 work histories for the remaining 70%, jobs were classified and ranked according to presumed
6 diesel exposure. Information is lacking regarding duration of employment in the job categories
7 (used for surrogate of exposure) and other confounding factors (alcohol consumption, cigarette
8 smoking, etc.). Thus, this study cannot be used to support or refute a causal association between
9 exposure to diesel exhaust and lung cancer.
10 A 2-year mortality analysis by Boffetta and Stellman (1988) of the American Cancer
11 Society's prospective study, after controlling for age and smoking, demonstrated an excess risk
12 of lung cancer in certain occupations with potential exposure to diesel exhaust. These excesses
13 were statistically significant among miners (RR = 2.67, 95% CI = 1.63,4.37) and heavy
14 equipment operators (RR = 2.6, 95% CI =1.12, 6.06). Recently Briiske-Hohlfeld et al. (1999)
15 also have observed significantly higher risk for lung cancer, in the range of 2.31 to 4.3, for heavy
16 equipment operators. The elevated risks were nonsignificant in railroad workers (RR = 1.59) and
17 truck drivers (RR = 1.24). A dose response was also observed for truck drivers. With the
18 exception of miners, exposure to diesel exhaust occurred in the three other occupations showing
19 an increase in the risk of lung cancer. Despite methodologic limitations, such as the lack of
20 representiveness of the study population (composed of volunteers only, who were probably
21 healthier than the general population), leading to an underestimation of the risk, and the
22 questionable reliability of exposure data based on self-administered questionnaires that were not
23 validated, this study is suggestive of a causal association between exposure to diesel exhaust and
24 excess risk of lung cancer.
25 Two mortality studies were conducted by Gustavsson et al. (1990) and Hansen (1993)
26 among bus garage workers (Stockholm, Sweden) and truck drivers, respectively. An SMR of
27 122 was found among bus garage workers, based on 17 cases. A nested case-control study was
28 also conducted in this cohort. Detailed exposure matrices based on job tasks were assembled for
29 both diesel exhaust and asbestos exposures. Statistically significant increasing lung cancer
30 relative risks of 1.34, 1.81, and 2.43 were observed for diesel exhaust indices of 10 to 20, 20 to
31 30, and >30, respectively, using 0 to 10 as a comparison group. Adjustment for asbestos
32 exposure did not change the results. The main strength of this study is the detailed exposure
33 matrices; some of the limitations are low power (small cohort) and lack of smoking histories.
34 But smoking is net likely to be different anicng study individuals irrespective of their exposure
7/25/00 7-54 DRAFT—DO NOT CITE OR QUOTE
-------
1 Hansen (1993), on the other hand, found statistically significant SMR of 160 from cancer
fof bronchus and lung. No dose response was observed, although the excesses were observed in
most of the age groups (30 to 39, 45 to 49, 50 to 54, 55 to 59, 60 to 64, and 65 to 74). There are
4 quite a few methodologic limitations to this study. Exposure to diesel exhaust was assumed in
5 truck drivers for diesel-powered trucks, but no validation of exposure was attempted. Follow-up
6 period was short, no latency analysis was done, and smoking data were lacking. However, a
7 population survey carried out in 1988 showed very little difference in smoking habits of residents
8 of rural area and the total Danish male population, thus, smoking is unlikely to confound the
9 finding of excess lung cancer. The findings of both these studies are consistent with the findings
10 of other truck driver studies and are supportive of causal association.
11 Two mortality studies of railroad workers were conducted by Howe et al. (1983) and
12 Garshick et al. (1988). The Howe et al. study, which was conducted in Canada, found relative
13 risks of 1.2 (p<0.01) and 1.35 (pO.OOl) among "possibly" and "probably" exposed groups,
14 respectively. The trend test showed a highly significant dose-response relationship with
15 exposure to diesel exhaust and the risk of lung cancer. The main limitation of the study was the
16 inability to separate overlapping exposures of coal dust/combustion fumes and diesel fumes.
17 Information on jobs was available at retirement only. There also was insufficient detail on the
[8 classification of jobs by diesel exhaust exposure. The exposures could have been nonconcurrent
or concurrent, but because the data are lacking, it is possible that the observed excess could be
20 due to the effect of both coal dust/combustion fumes and diesel fumes and not just one or the
21 other. It should be noted that, so far, coal dust has not been demonstrated to be a pulmonary
22 carcinogen in studies of coal miners. However, lack of data on confounders such as asbestos and
23 smoking (though use of the internal comparison group to compute relative risks minimizes
24 confounding by smoking) makes interpretation of this study difficult. When three diesel exhaust
25 exposure categories were examined for smoking-related diseases such as emphysema, laryngeal
26 cancer, esophageal cancer, and buccal cancer, positive trends were observed, raising a possibility
27 that the dose response demonstrated for diesel exposure may have been due to smoking. The
28 findings of this study are at best suggestive of diesel exhaust being a lung carcinogen.
29 The strong evidence for linking diesel exhaust exposure to lung cancer comes from the
30 Garshick et al. (1988) railroad worker study conducted in the United States. Relative risks of
31 1.57 (95% CI = 1.19, 2.06) and 1.34 (95% CI = 1.02, 1.76) were found for ages 40 to 44 and 45
32 to 49, respectively, after the exclusion of workers exposed to asbestos. The investigators
33 reported that the risk of lung cancer increased with increasing duration of employment. As this
34 was a large cohort study with a lengthy follow-up and adequate analysis, including dose response
(based on duration of employment as a surrogate) as well as adjustment for other confounding
7/25/00 7-55 DRAFT—DO NOT CITE OR QUOTE
-------
1 factors such as asbestos, the observed association between increased lung cancer and exposure to
2 diesel exhaust is more meaningful. Even though the reanalysis of these data by Crump et al.
3 (1991) found that the relative risk could be positively or negatively related to duration of
4 exposure depending on how age was controlled, additional analysis by Garshick et al. (1991)
5 found that the relationship between years exposed when adjusted for the attained age and
6 calendar years was flat to negative, depending on the choice of the model. They also found that
7 deaths were underreported by approximately 20% to 70% between 1977 and 1 980, and their
8 analysis based on job titles, limited to 1 959-1976, showed that the youngest workers still had the
9 highest risk of dying of lung cancer. On the other hand, an analysis of the same data by
1 0 California EPA (CalEPA, 1 998) yielded a positive dose response set using age at 1959 and
1 1 adding an interaction term of age and calendar year in the model. However, Crump (1999)
1 2 reported a negative dose response in his latest analysis. The divergent results of these recent
1 3 analyses do not negate the strong evidence this study provides for the qualitative evaluation.
1 4 The observance of dose response would have strengthened the causal association, but an absence
15 of a dose response does not negate it.
1 6 Suggestive evidence is provided by a recent study of potash miners in Germany.
1 7 The information on the exposure (including elemental carbon and organics), work chronology,
1 8 and work category was used by the investigators to calculate cumulative exposures for each
1 9 worker. Furthermore, information on smoking habits indicated homogeneity in the cohort.
20 A statistically nonsignificant twofold increase in lung cancer was observed in the production
2 1 workers as compared to workshop workers. The lack of significance for this finding could be
22 due to short follow-up, not enough latency, and relatively young age of the cohort.
23
24 7.2.4.2. Summary of the Case-Control Studies of Lung Cancer
25 Among the 1 1 lung cancer case-control studies reviewed in this chapter, only 2 studies
26 did not find any increased risk of lung cancer. Lerchen et al. (1987) did not find any excess risk
27 of lung cancer, after adjusting for age and smoking, for diesel fume exposure. The major
28 limitation of this study was a lack cf adequate exposure data derived from the job titles obtained
29 from occupational histories. Next of kin provided the occupational histories for 50% of the cases
30 that were not validated. The power of the study was small (analysis done on males only, 333
31 cases). Similarly, Boffeta et al. (1990) did not find any excess of lung cancer after adjusting for
32 smoking and education. This study had a few methodological limitations. The lung cancer cases
33 and controls were drawn from the ongoing study of tobacco-related diseases. It is interesting to
34 note that the leading risk factor for lung cancer is cigarette smoking. The exposure was not
35 measured. Instead, occupations Wcic uscu a* surrogates for exposure. Furthermore, tnere were
7/25/00 7-56 DRAFT— DO NOT CITE OR QUOTE
-------
1 very few individuals in the study who were exposed to diesel exhaust. On the other hand,
f statistically nonsignificant excess risks were observed for diesel exhaust exposure by Hall and
Wynder (1984) in workers who were exposed to diesel exhaust versus those who were not (OR =
4 1.4 and 1.7 with two different criteria) and by Damber and Larsson (1987) in professional drivers
5 (OR = 1.2). These rates were adjusted for age and smoking. Hall and Wynder (1984) had a high
6 nonparticipation rate of 36%. Therefore, the positive results found in this study are
7 underestimated at best. In addition, the self-reported exposures used in the study by Hall and
8 Wynder (1984) were not validated. This study also had low power to detect excess risk of lung
9 cancer for specific occupations.
10 The study by Benhamou et al. (1988), after adjusting for smoking, found significantly
11 increased risks of lung cancer among French motor vehicle drivers (RR = 1.42) and transport
12 equipment operators (RR = 1.35). The main limitation of the study was the inability to separate
13 exposures to diesel exhaust from those to gasoline exhaust because both motor vehicle drivers
14 and transport equipment operators probably were exposed to the exhausts of both types of
15 vehicles.
16 Hayes et al. (1989) combined data from three studies (conducted in three different States)
17 to increase the power to detect an association between lung cancer and occupations with a high
potential for exposure to diesel exhaust. They found that truck drivers employed for more than
10 years had a significantly increased risk of lung cancer (OR =1.5, 95% CI = 1.1, 1.9). This
20 study also found a significant trend of increasing risk of lung cancer with increasing duration of
21 employment among truck drivers. The relative odds were computed by adjusting for birth
22 cohort, smoking, and State of residence. The main limitation of this study is again the mixed
23 exposures to diesel and gasoline exhausts, because information on type of engine was lacking.
24 Also, potential bias may have been introduced because the way in which the cause of death was
25 ascertained for the selection of cases varied in the three studies. Furthermore, the methods used
26 in these studies to classify occupational categories were different, probably leading to
27 incompatibility of occupational categories.
28 Emmelin et al. (1993), in their Swedish dockworkers from 15 ports, found increased
29 relative odds of 6.8 (90% CI = 1.3 to 34.9). A strong interaction between smoking and diesel
30 exhaust was observed in this study. Of 50 cases and 154 controls, only 6 individuals were
31 nonsmokers. Although intricate exposure matrices were created using three different variables,
32 no direct exposure measurement was done. Despite the limitations of small number of cases and
33 controls; lack of data on asbestos exposure, which is fairly common in dockworkers; and very
34 few nonsmokers; this study provides consistent support for a real effect of diesel exhaust
exposure and occurrence of lung cancer, at least in smokers.
7/25/00 7-57 DRAFT—DO NOT CITE OR QUOTE
-------
1 In a population-based lung cancer case-control study Swanson et al. (1993) found
2 statistically significant excess risks adjusted for age at diagnosis, smoking, and race, among
3 white male drivers of heavy trucks employed for ;>20 years and railroad workers employed for
4 *10 years (OR = 2.5, 95% CI = 1.1,4.4 and OR = 2.4, 95 % CI = 1.1, 5.1, respectively), and
5 among black farmers employed for *20 years (OR = 10.4, 95% CI = 1.4, 77.1). Although
6 individual ORs were not significant for various occupations with potential exposure to diesel
7 exhaust, statistically significant trends were observed for drivers of heavy trucks, light trucks,
8 farmers, and railroad industry workers among whites, and among black farmers (psO.5). The
9 main strengths of the study are availability of data on lifetime work history and smoking history;
10 the main limitation is absence of actual specific exposure data. This is the first study that found
11 increased lung cancer risk for farmers, who are exposed to diesel exhaust of their farm tractors.
12 The most convincing evidence comes from the case-control studies, among railroad
13 workers by Garshick et al. (1987), among truck drivers of the Teamsters Union by Steenland et
14 al. (1990, 1998), among different professional drivers in Denmark by Hansen et al. (1998), and
15 among male workers occupationally exposed to diesel motor emissions in Germany by Bruske-
16 Hohlfeld et al. (1999). Garshick et al. found that after adjustment for asbestos and smoking, the
17 relative odds for continuous exposure were 1.39 (95% CI = 1.05, 1.83). Among the younger
18 workers with longer diesel exhaust exposure, the risk of lung cancer increased with duration of
19 exposure after adjusting for asbestos and smoking. Even after the exclusion of recent diesel
20 exhaust exposure (5 years before death), the relative odds increased to 1.43 (95% CI = 1.06,
21 1.94). This appears to be a well-conducted and well-analyzed study with reasonably good power.
22 Potential confounders were controlled adequately, and interactions between diesel exhaust and
23 other lung cancer risk factors were tested. Some of the limitations of this study are inadequate
24 latency period, misclassification of exposure because ICC job classification was used as
25 surrogate for exposure, and use of death certificates for identification of cases and controls.
26 Steenland et al. (1990), on the other hand, created two separate work history files, one
27 from Teamsters Union pension files and the other from next-of-kin interviews. Using duration of
28 employment as a categorical variable and considering employment after 1959 (when presumed
29 dieselization occurred) for long-haul drivers, the risk of lung cancer increased with increasing
30 years of exposure. Using 1964 as the cutoff, a similar trend was observed for long-haul drivers.
31 For short-haul drivers, the trend was positive with a 1959 cutoff, but not when 1964 was used as
32 the cutoff. For truck drivers who primarily drove diesel trucks and worked for 35 years, the
33 relative odds were 1.89. The main strengths of the study are availability of detailed records from
34 the Teamsters Union, a relatively large sampl? size, availability of smokine data, and
35 mccisursrnsntc cf sxpcsurs. Ths limitations of this study include possible rnisclassifications of
7/25/00 7-58 DRAFT—DO NOT CITE OR QUOTE
-------
1 exposure and smoking, lack of levels of diesel exposure, a smaller nonexposed group, and an
4 insufficient latency period. Recently Steenland et al. (1998) conducted an exposure-response
analysis on these cases and controls, using the industrial hygiene survey results of Zaebst et al.
4 (1991). The estimates were made for long-haul drivers, short-haul drivers, dockworkers,
5 mechanics, and those outside the trucking industry. The survey found that mechanics had the
6 highest current levels of diesel exhaust exposures and dockworkers who mainly used propane-
7 powered forklifts had the lowest exposure. The finding of the highest lung cancer risk for
8 mechanics and lowest for dock workers is indicative of a causal association between the diesel
9 exhaust exposure and development of lung cancer. However, the risk among mechanics did not
10 increase with increasing duration of employment. The OR for quartile cumulative exposure,
11 computed by using logistic regression adjusted for age, race, smoking, diet, and asbestos
12 exposure, showed a pattern of increasing trends in risk with increasing exposure, between 1.08
13 and 1.72 depending upon exposure level and lag structure used.
14 Hansen et al. (1998), in their study of professional drivers in Denmark, found statistically
15 significant ORs (adjusted for socioeconomic status) of 1.31, 1.64, and 1.39 for lorry/bus drivers,
16 taxi drivers, and unspecified drivers, respectively. The lag time analyses for duration of
17 employment were unchanged for lorry/bus drivers but increased to OR = 3 from 2.2 in taxi
18 drivers with a lag time of 10 years and duration of employment of > 5 years. The authors asserted
that the higher risk seen in the taxi drivers may be due to higher exposure to these drivers
20 because of longer time spent in traffic congestion. Furthermore, the trend tests for increasing
21 risk of lung cancer with increasing duration of employment were statistically significant for both
22 lorry/bus drivers and taxi drivers in both 10-year lag time and no lag time. The main strengths of
23 the study are the large sample size, availability of detailed employment records, and information
24 on socioeconomic status. The main limitations are absence of individual data on smoking habits
25 and asbestos exposure, and information about the type of fuel used for the vehicles driven by
26 these professional drivers. A personal communication with the main investigator revealed that
27 the lorries/buses and taxis have been using diesel fuel since the early 1960s. Moreover, indirect
28 information about smoking and asbestos exposure indicated that these two confounders are
29 unlikely to explain the observed excesses or the trends, resulting in strong support of earlier
30 positive studies.
31 Briiske-Hohlfeld et al. (1999) recently conducted a pooled analysis of two case-control
32 studies among male workers occupationally exposed to DME in Germany. The investigators
33 collected data on demographic information, detailed smoking, and occupational history. Job
34 titles and industries were classified in 33 and 21 categories respectively. Job descriptions were
written and verified to avoid misclassification. Individual cumulative DME exposures and
7/25/00 7-59 DRAFT—DO NOT CITE OR QUOTE
-------
1 smoking pack-years were calculated. Asbestos exposures were estimated by certain job-specific
2 supplementary questions. Analysis of 3,498 lung cancer cases and 3,541 controls yielded
3 statistically significant ORs ranging from 1.25 to 2.31 adjusted for smoking and asbestos
4 exposure. The risk increased with increasing years of exposure for both the first year of exposure
5 and the end year of exposure. These investigators presented analyses by various job categories,
6 by years of exposure, first and end years of exposure and, when possible, separately for West and
7 East Germany. Significantly higher risks were found among all four job categories. For
8 professional drivers (of trucks, buses, and taxis) ORs ranged from 1.25 to 2.53. For other traffic-
9 related jobs (switchmen, diesel locomotive drivers, diesel forklift truck drivers), ORs ranged
10 from 1.53 to 2.88. For heavy equipment operators (bulldozers, graders, and excavators), ORs
11 ranged from 2.31 to 4.3, and for drivers of farming equipment the only significant excess (OR =
12 6.81) was for exposure for < 30 years.
13 This study shows increased risk for all the DME-exposed job categories. The
14 professional drivers and the other traffic-related jobs also have some mixed exposures to gasoline
15 exhaust in general traffic. On the other hand, it should be noted that exposure to DME among
16 heavy equipment and farm tractor drivers is much higher and not as mixed as in professional
17 drivers. The heavy equipment drivers usually drive repeatedly through their own equipment's
18 exhaust. Therefore, the observed highest risk for lung cancer in this job category establishes a
19 strong link with the DME. The only other study that found significantly higher risk for heavy
20 equipment operators (RR = 2.6) was conducted by Boffeta et al. (1988). Although the only
21 significant excess in the group was observed for farming tractor operators with more than 30
22 years of exposure, a steady increase in risk was observed for this job category with increasing
23 exposure. The investigators stated that the working conditions and the DME of tractors remained
24 fairly constant over the years. This increase may be due mainly to exposure to DME and PM10
25 The main strengths of the study are large sample size, resulting in good statistical power;
26 inclusion of incident cases diagnosed not more than 3 months prior to the interview; use of only
27 personal interviews, reducing recall bias; diagnoses ascertained by cytology or histology; and
28 availability of lifelong detailed occupational and smoking history. Exposure estimation done for
29 each individual was based on job codes and industry codes, which were validated by written job
30 descriptions to avoid misclassification.
31 The main limitation of the study is lack of data on actual exposure to DME. The
32 cumulative quantitative exposures were calculated on the basis of time spent in each job with
33 potential exposure to DME and the type of equipment used. Thus, this study provides strong
"34 evidence for caussl association between exposure tc diesel exhaust and occurrence of lung
7/25/00 7-60 DRAFT—DO NOT CITE OR QUOTE
-------
7.2.4.3. Summary of the Reviews and Meta-Analyses of Lung Cancer
Three summaries of studies concerned with the relationship of diesel exhaust exposure
and lung cancer risk are reviewed. The HEI report is a narrative study of more than 3 5
4 epidemiologic studies (16 cohort and 19 case-control) of occupational exposure to diesel
5 emissions published between 1957 and 1993. Control for smoking was identified in 15 studies.
6 Six of the studies (17%) reported relative risk estimates less than 1, whereas 29 (83%) reported at
7 least 1 relative risk, indicating a positive association. Twelve studies indicating a relative risk
8 greater than 1 had 95% confidence intervals that excluded unity. These studies found that the
9 evidence suggests that occupational exposure to diesel exhaust from diverse sources increases the
10 rate of lung cancer by 20% to 40% in exposed workers generally, and to a greater extent among
11 workers with prolonged exposure. They also found that the results are not explicable by
12 confounding due to cigarette smoking or other known sources of bias.
13 Bhatia et al. (1998) identified 23 studies that met criteria for inclusion in the meta-
14 analysis. The observed relative risk estimates were greater than 1 in 21 of these studies. The
15 pooled relative risk weighted by study precision was 1.33 (95% CI= 1.24, 1.44), which indicated
16 increased relative risk for lung cancer from occupational exposure to diesel exhaust.
17 Subanalyses by study design (case-control and cohort studies) and by control for smoking
18 produced results that did not differ from those of the overall pooled analysis. Cohort studies
A using internal comparisons showed higher relative risks than those using external comparisons.
20 Lipsett and Campleman (1999) identify 39 independent estimates of relative risk among
21 30 eligible studies of diesel exhaust and lung cancer published between 1975 and 1995. Pooled
22 relative risks for all studies and for study subsets were estimated using a random effect model.
23 Interstudy heterogeneity was also modeled and evaluated. A pooled smoking-adjusted relative
24 risk was 1.47 (95% CI = 1.29, 1.67). Substantial heterogeneity was found in the pooled-risk
25 estimates. Adjustment for confounding by smoking, having a lower likelihood of selection bias,
26 and increased study power were all found to contribute to lower heterogeneity and increased
27 pooled estimates of relative risk.
28 There is some variability in the conclusions of these summaries of the association of
29 diesel exhaust and lung cancer. The three analyses find that smoking is unlikely to account for
30 the observed effects, and all conclude that the data support a causal association between lung
31 cancer and diesel exhaust exposure. On the other hand, Stober and Abel (1996), Muscat and
32 Wynder (1995), and Cox (1997) call into question the assertions by Cohen and Higgins (1995),
33 Bhatia et al. (1998), and Lipsett and Campleman (1999) that the associations seen for diesel
34 exhaust and lung cancer are unlikely to be due to bias. They argue that methodologic problems
are prevalent among the studies, especially in evaluation of diesel engine exposure and control of
7/25/00 7-61 DRAFT—DO NOT CITE OR QUOTE
-------
1 confounding by cigarette smoking. The conclusions of the two meta-analyses are based on
2 magnitude of pooled relative risk estimates and evaluation of potential sources of heterogeneity
3 in the estimates. Despite the statistical sophistication of the meta-analyses, the statistical models
4 used cannot compensate for deficiencies in the original studies and will remain biased to the
5 extent that bias exists in the original studies.
6
7 1.2 A A. Discussion of Relevant Methodologic Issues
8 A persistent association of risk for lung cancer and diesel exhaust exposure has been
9 observed in more than 30 epidemiologic studies published in the literature over the past 40 years.
1 0 Evaluation of whether this association can be attributed to a causal relation between diesel
1 1 exhaust exposure and lung cancer requires careful consideration of whether chance, bias, or
1 2 confounding might be likely alternative explanations.
1 3 A total of 1 0 cohort and 12 case-control studies are reviewed in this chapter. An
1 4 increased lung cancer risk was observed in 8 cohort and 1 0 case-control studies, even though the
1 5 results were not always statistically significant. There is a consistent tendency for point
1 6 estimates of relative risk to be greater than one in studies that adjusted (either directly or
1 7 indirectly) for smoking, had a long enough follow-up, and sufficient statistical power among
1 8 truck drivers, railroad workers, dock workers, and heavy equipment workers. If this elevated risk
1 9 was due to chance one would expect almost equal distribution of these point estimates to be
20 above and below one. Many of the studies provide confidence intervals for their estimates of
21 excess risk or statistical tests, which indicate that it is unlikely that the individual study findings
22 were due to random variation. The persistence of this association between diesel exhaust and
23 lung cancer risk in so many studies indicates that the possibility is remote that the observed
24 association in aggregate is due to chance. It is unlikely that chance alone accounts for the
25 observed relation between diesel exhaust and lung cancer.
26 The excess risk is observed in both cohort and case-control designs, which contradicts the
27 concern that a methodologic bias specifically characteristic of either design (e.g., recall bias)
28 might account for the observed effect. Selection bias is certainly present in some of the
29 occupational cohort studies that use external population data in estimating relative risks, but this
30 form of selection bias (a healthy worker effect) would only obscure, rather than spuriously
3 1 produce, an association between diesel exhaust and lung cancer. Several occupational
32 epidemiologic studies that use more appropriate data for their estimates are available. Selection
33 biases may be operating in some case-control studies, but it is not obvious how such a bias could
o/i u_ _..rrr-: __ *i-. — :r~ - :_ _.ccu,.* ______ i~_» — J ~* ___ ~ ----- ~u *~ i --- > *-. *u~ — „,.:..*, — *
*j-r uc. OUULI^ICIIUJ uiuiuiiii 111 cucwi, picvaicui, cum ouvsiig tiiuugii IAJ ic.au. IA_> uik> wwiioiaiLui.
3D aSiOCiatiOIi Sceii ili tilt aggregate Uata. oiVcIi tile Variety Oi ueSigHS UScu ill Stiiayilig tiic ui
7/25/00 7-62 DP.AFT — DO NOT CITE OR QUOTE
-------
exhaust and lung cancer association and the number of studies in different populations, it is
unlikely that routinely studying noncomparable groups is an explanation for the consistent
3 association seen. Exposure information bias is certainly a problem for almost all of the studies
4 concerned. Detailed and reliable individual-level data on diesel exhaust exposure for the period
5 of time relevant to the induction of lung cancer are not available and are difficult to obtain.
6 Generally, the only information from which diesel exposure can be inferred is occupational data,
7 which is a poor surrogate for the true underlying exposure distribution. Study endpoints are
8 frequently mortality data taken from death certificate information, which is frequently inaccurate
9 and often does not fully characterize the lung cancer incidence experience of the population in
10 question. Using inaccurate surrogates for lung cancer incidence and for diesel exposure can lead
11 to substantial bias, and these shortcomings are endemic in the field. In most cases these
12 shortcomings will lead to misclassification of exposure and of outcome, which is nondifferential.
13 Nondifferential misclassification of exposure and/or outcome can bias estimates of a diesel
14 exhaust—lung cancer association, if one exists, toward the null; but it is unlikely that such
15 misclassification would produce a spurious estimate in any one study. It is even more unlikely
16 that it would bias a sufficient number of studies in a uniform direction to account for the
17 persistent aggregate association observed.
« Moreover, throughout this chapter, various methodologic limitations of individual studies
have been discussed, such as small sample size, short follow-up period, lack of data on
20 confounding variables, use of death certificates to identify the lung cancer cases, and lack of
21 latency analysis. The studies with small sample sizes (i.e., not enough power) and short follow-
22 up periods (i.e., not enough latent period) have been difficult to interpret due to these limitations.
23 The most important confounding variable is smoking which is a strong risk factor for
24 lung cancer. All the studies considered for this report are either cohort retrospective mortality or
25 case-control studies where history of exposures in the past is elicited. Smoking history is usually
26 difficult to obtain in such instances. The smoking histories obtained from surrogates (next of
27 kin, either spouse or offspring) were found to be accurate by Lerchen and Samet (1986) and
28 McLaughlin et al. (1987). Lerchen and Samet did not detect any consistent bias in the report.of
29 cigarette consumption. In contrast, overreporting of cigarette smoking by surrogates was
30 observed by Rogot and Reid (1975), Kolonel et al. (1977), and Humble et al. (1984). Kolonel et
31 al. found that the age at which an individual started smoking was reported within 4 years of
32 actual age 84% of the time. These studies indicate that surrogates were able to provide fairly
33 credible information on the smoking habits of the study subjects. If the surrogates of the cases
54 were more likely to overreport cigarette smoking compared with the controls, then it might be
7/25/00 7-63 DRAFT—DO NOT CITE OR QUOTE
-------
1 harder to find an effect of diesel exhaust because most of the increase in lung cancer would be
2 attributed to smoking rather than to exposure to diesel exhaust.
3 Some studies do not adjust for tobacco smoke exposure. Even though smoking is a
4 strong risk for lung cancer, it is only a confounder if there are differential smoking habits among
5 individuals exposed to diesel exhaust versus individuals who are not exposed. Most of the
6 occupational cohorts include workers from the same socioeconomic background or used an
7 internal comparison group; hence, it is unlikely that confounding by cigarette smoking is
8 substantial in these studies. Some studies have adjusted for socioeconomic status and some
9 studies have compared the cigarette smoking habits by conducting rural and urban general
1 0 population surveys. Besides, in studies with long enough latency, adjustment for cigarette
1 1 smoking did not alter substantially the observed higher risk.
1 2 Another methodologic concern in these studies is use of death certificates to determine
1 3 cause of death. Death certificates were used by all of the cohort mortality studies and some of
1 4 the case-control studies of lung cancer to determine cause of death. Use of death certificates
1 5 could lead to misclassification bias because of overdiagnosis. Studies of autopsies done between
1 6 1960 and 1 971 demonstrated that lung cancer was overdiagnosed when compared with hospital
1 7 discharge, with no incidental cases found at autopsy (Rosenblatt et al., 1971). Schottenfeld et al.
1 8 (1982) also found an overdiagnosis of lung cancer among autopsies conducted in 1977 and 1978.
1 9 On the other hand, Percy et al. (1981) noted 95% concordance when comparing 10,000 lung
20 cancer deaths observed in the Third National Cancer Survey from 1969 to 1971 (more than 90%
21 were confirmed histologically) to death-certificate-coded cause of death. These more recent
22 findings suggest that the diagnosis of lung cancer on death certificates is better than anticipated.
23 In reality, lung cancer is one cause of death that has been found to be generally reliably
24 diagnosed on the death certificate.
25 Finally, several investigators have not conducted latency analysis in their studies. The
26 latent period for lung cancer development is up to 30 years or more. Considering the fact that
27 dieselization was not complete till almost 1959 for locomotives and the 1970s for the trucking
28 industry in the USA. most of the cohort studies do not have a long enough follow-up period to
29 allow for latency of 30+ years. In addition, the study inclusion criteria for most of the studies are
30 individuals who worked in the industry for at least 6 months /I year from the beginning of the
3 1 follow-up period to the end of the follow-up period. Hence, the later the individual enters the
32 cohort, the shorter the follow-up period; thus, the latent period is insufficient for the occurrence
33 of lung cancer. Therefore, the observed slight to moderate increase in risk of lung cancer could
T •
4 A />*i+-t-fi
Jfc. I>»L
7/25/00 7-64 DRAFT— DO NOT CITE OR QUOTE
-------
enough to allow for the 30+ years latency needed for the development of lung cancer (Hansen et
al., 1998; Bruske-Hohlfeld et al., 1999). These investigators identified lung cancer cases in the
"3 early to mid-1990s and found significant excess risks for lung cancer among the individuals
4 exposed to diesel exhaust. It should be noted that the use of diesel fuel for trucks, buses, and
5 taxis had started in their countries (Denmark and Germany, respectively) in the early 1960s.
6
7 7.2.4.5. Evaluation of Causal Association
8 In most situations, epidemiologic data are used to delineate the causality of certain health
9 effects. Several cancers have been causally associated with exposure to agents for which there is
10 no direct biological evidence. Insufficient knowledge about the biological basis for diseases in
11 humans makes it difficult to identify exposure to an agent as causal, particularly for malignant
12 diseases when the exposure was in the distant past. Consequently, epidemiologists and biologists
13 have provided a set of criteria that define a causal relationship between exposure and the health
14 outcome. A causal interpretation is enhanced for studies that meet these criteria. None of these
15 criteria actually proves causality; actual proof is rarely attainable when dealing with
16 environmental carcinogens. None of these criteria should be considered either necessary (except
17 temporality of exposure) or sufficient in itself. The absence of any one or even several of these
criteria does not prevent a causal interpretation. However, if more criteria apply, this provides
more credible evidence for causality.
20 Thus, applying the Hill criteria (1965) of causal inference, as modified by Rothman
21 (1986), to the studies reviewed here resulted in the following:
22
23 • Strength of association. This phrase refers to the magnitude of the ratio of
24 incidence or mortality (RRs or ORs). Several studies found statistically
25 significant RRs and ORs that ranged from 1.2 to 2.6 (Howe et al., 1983; Rushton
26 et al., 1983; Wong et al., 1985; Gustavsson et al., 1990; Emmlin et al., 1993;
27 Hansen et al., 1993; Hansen et al., 1998) and, after adjustment for smoking and/or
28 asbestos, RRs and ORs remained statistically significant and in the same range in
29 certain studies (Dambar and Larson 1987; Garshick et al., 1987, 1988; Benhamou
30 et al., 1988; Boffetta and Stellman, 1988; Hays et al., 1989; Steenland et al., 1990;
31 Swanson et al., 1993; Briisk-Hohlfeld et al., 1999). In addition, two meta-
32 analyses demonstrated that not only did excess in lung cancer remain the same
33 after stratification/adjustment for smoking and occupation, but in several
instances the pooled RRs showed modest increases, with little evidence of
heterogeneity. Overall, the studies in epidemiologic terms show relatively modest
7/25/00 7-65 DRAFT—DO NOT CITE OR QUOTE
-------
1 to weak association between diesel exhaust and occurrence of lung cancer. Even
2 though strong associations are more likely to be causal than modest-to-weak
3 associations, the fact that association is relatively modest or weak does not rule
4 out the causal link.
5 • Consistency. Increased lung cancer risk has been observed in several cohort and
6 case-control studies, conducted in several industries and occupations in which
7 workers were potentially exposed to diesel exhaust. However, not all the excesses
8 were statistically significant. Statistically significant lung cancer excesses
9 adjusted for smoking were observed in truck drivers (Hayes et al., 1989; Hansen
10 et al., 1993; Swanson et al., 1993; Briiske-Hohlfeld et al., 1999), professional
11 drivers (Benhamou et al., 1988; Briiske-Hohlfeld et al., 1999), railroad workers
12 (Garshick et al., 1987; Swanson et al., 1993), heavy equipment drivers (Boffetta et
13 al., 1988; Briiske-Hohlfeld et al., 1999), and farm tractor drivers (Swanson et al.,
14 1993; Briiske-Hohlfeld et al., 1999). Furthermore, the two recent meta-analyses
15 by Bhatia et al. (1998) and Lipsett and Campleman (1999) found that even though
16 a substantial heterogeneity existed in their initial pooled estimates, stratification
17 on several factors demonstrated a relationship between exposure to DE and excess
18 lung cancer that remained positive throughout various analyses.
19 • Specificity. This criterion requires that a single cause lead to a single effect. With
20 respect to exposure to diesel exhaust, excess for lung cancer is the only effect that
21 is found to be consistently elevated and statistically significant in several studies.
22 Quite a few studies have examined diesel exhaust for other effects such as bladder
23 cancer, leukemia, gastrointestinal cancers, prostate cancer etc. The evidence for
24 these effects is inadequate. Rothman (1986), in his discussion about causality
25 criteria, states "Causes of a given effect, however, cannot be expected to be
26 without other effects on any logical grounds. In fact, everyday experience teaches
27 us repeatedly that single events may have many effects. Hill's discussion of this
28 standard for inference is replete with reservations, but even so, the criterion seems
29 useless and misleading."
30 • Temporality. The only necessary, but not sufficient, criterion described by Hill
31. for causality inference is that exposure to a causal agent precedes the effect in
32 time. This criterion is clearly satisfied in the studies reviewed here. Temporality
33 can be explored further in addressing the latency issue. A certain period is
34 necessary for developrnent of an effect after exposure to a causal agent has
35 occurred. For instance, in cancer-causing agents a latent period can vary from ^
7/25/00 7-66 DRAFT—DO NOT CITE OR QUOTE
-------
1 years (childhood leukemia) to 2 30 years (mesothelioma). Most of the studies
f reviewed here did not conduct the latency analysis. Some studies had a short
follow-up period that did not allow enough time for the latency period (Waller,
4 1981; Howe et al., 1983; Rushton et al., 1983; Wong et al., 1985, Hansen et al.,
5 1993) while several studies clearly allowed for an adequate latency period
6 (Garshick et al., 1987; Gustavsson et al., 1990; Steenland et al., 1990; Swanson et
7 al., 1993; Briiske-Hohlfeld et al., 1999). Both type of studies showed mixed
8 results.
9 Biological gradient. This criterion refers to the dose-response curve. Due to the
10 lack of quantitative data on diesel exhaust exposure in most studies reviewed here,
11 analyzing the dose-response curve directly was not possible. In very few studies
12 was exposure to diesel exhaust addressed specifically. Most investigators have
13 used job titles/categories and duration of employment as surrogates for exposure
14 and thus have presented response in relation to duration of employment.
15 Significant dose-response (using duration of employment as a surrogate) was
16 observed in various studies for railroad workers (Howe et al., 1983; Garshick et
17 al., 1987; Garshick et al., 1988; Swanson et al., 1993), truck drivers (Boffetta and
18 Stellman, 1988; Hayes et al., 1989; Steenland et al., 1990; Swanson et al., 1993;
^P Hansen et al., 1998; Briiske-Hohlfeld et al., 1999), transportation/heavy
20 equipment operators (Wong et al., 1985; Gustavsson et al., 1990; Briiske-Hohlfeld
21 et al., 1999), farmers/farm tractor users (Swanson et al., 1993; Briiske-Hohlfeld et
22 al., 1999), and dockworkers (Emmelin et al., 1993).
23 • Biological plausibility. This criterion refers to the biologic plausibility of the
24 hypothesis, an important concern that may be difficult to judge. The hypothesis
25 considered for this review is that occupational exposure to diesel exhaust is
26 causally associated with the occurrence of lung cancer and is supported by the
27 following: First, diesel exhaust has been shown to cause lung and other cancers in
28 animals (Heinrich et al., 1986b; Iwai et al., 1986b; Mauderly et al., 1987; Pott et
29 al., 1990; Mauderly et al., 1994). Second, it contains highly mutagenic substances
30 such as polycyclic aromatic hydrocarbons as well as nitroaromatic compounds
31 (Claxton, 1983; Ball et al., 1990; Gallagher et al., 1993; Sera et al., 1994; Nielsen
32 et al., 1996a) that are recognized human pulmonary carcinogens (IARC, 1989).
33 Third, diesel exhaust consists of carbon core particles with surface layers of
34 organics and gases; the tumorigenic activity may reside in one, some, or all of
these components. As explained in Chapter 4, there is clear evidence that the
7/25/00 7-67 DRAFT—DO NOT CITE OR QUOTE
-------
1 organic constituents, both in particles and vapor phases, have the capacity to
2 interact with DNA and give rise to mutations, chromosomal aberrations, and cell
3 transformations, all well- established steps in the process of carcinogenesis.
4 Further, increased levels of peripheral blood cell DNA adducts associated with
5 occupational exposure to diesel exhaust have been observed in humans (Nielsen et
6 al., 1996a,b). Thus, the above evidence makes a convincing case that
7 occupational exposures to diesel exhaust are causally associated with the
8 occurrence of lung cancer — highly plausible biologically.
9
10 In conclusion, the epidemiologic studies of exposure to diesel exhaust and occurrence of
1 1 lung cancer furnish evidence that is consistent with a causal association. This association
1 2 observed in several studies is unlikely to be due to chance or bias. Although many studies did
1 3 not have information on smoking, confounding by smoking is unlikely in these studies because
1 4 the comparison population was from the same socioeconomic class. The strength of association
1 5 was weak to modest (RRs/ORs between 1 .2 and 2.6), with dose-response relationship observed
16 in several studies. Last, but not least, there is strong evidence for biological plausibility that
1 7 exposure to diesel exhaust would result in excess risk of lung cancer in humans.
18
1 9 7.3. CARCINOGENICITY OF DIESEL EMISSIONS IN LABORATORY ANIMALS
20 This chapter summarizes studies that assess the carcinogenic potential of diesel exhaust in
21 laboratory animals. The first portion of this chapter summarizes results of inhalation studies.
22 Experimental protocols for the inhalation studies typically consisted of exposure (usually
23 chronic) to diluted exhaust in whole-body exposure chambers using rats, mice, and hamsters as
24 model species. Some of these studies used both filtered (free of particulate matter) diesel exhaust
25 and unfiltered (whole) diesel exhaust to differentiate gaseous-phase effects from effects induced
26 by diesel PM (DPM) and its adsorbed components. Other studies were designed to evaluate the
27 relative importance of the carbon core of the diesel particle versus that of particle-adsorbed
28 compounds. Finally, a number of exposures were carried out to determine the combined effect of
29 inhaled diesel exhaust and tumor initiators, tumor promoters, or cocarcinogens.
30 Particulate matter concentrations in the diesel exhaust used in these studies ranged from
31 0.1 to 12 mg/m3. In this chapter, any indication of statistical significance implies thatp^O.05
32 was reported in the reviewed publications. A summary of the animal inhalation carcinogenicity
33 studies and their results is presented in Table 7-3.
34
35 parucics. cALiacLcu Jicscl
7/25/00 7-68 DRAFT— DO NOT CITE OR QUOTE
ore n e revewe pucaons. summary o e anma naaon carcnogen
and their results is presented in Table 7-3.
Results of lung implantation and intratracheal instillation studies of whole diesel
s. cALiacLcu Jicscl paiticlcS, and pciTticlc cxtrscts Circ reported in Section 7.3.3 2nd in
-------
1 Tables 7-4 and 7-5. Studies destined to assess the carcinogenic effects of DPM as well as solvent
f extracts of DPM following subcutaneous (s.c.) injection, intraperitoneal (i.p.) injection, or
intratracheal (itr.) instillation in rodents are summarized in Section 7.3.5. Individual chemicals
4 present in the gaseous phase or adsorbed to the particle surface were not included in this review
5 because assessments of those of likely concern (i.e., formaldehyde, acetaldehyde, benzene,
6 polycyclic aromatic hydrocarbons [PAHs]) have been published elsewhere (U.S. EPA, 1993).
7
8 7.3.1. Inhalation Studies (Whole Diesel Exhaust)
9 7.3.1.1. Rat Studies
10 The potential carcinogenicity of inhaled diesel exhaust was first evaluated by Karagianes
11 et al. (1981). Male Wistar rats (40 per group) were exposed to room air or diesel engine exhaust
12 diluted to a DPM concentration of 8.3 (± 2.0) mg/m3, 6 hr/day, 5 days/week for up to 20 months.
13 The animals were exposed in 3,000 L plexiglass chambers. Airflow was equal to 50 liters per
14 minute. Chamber temperatures were maintained between 25° and 26.5 °C. Relative humidity
15 ranged from 45% to 80%. Exposures were carried out during the daytime. The connected to an
16 electric generator and operated at varying loads and speeds to simulate operating conditions in an
17 occupational situation. To control the CO concentration at 50 ppm, the exhaust was diluted 35:1
with clean air. Six rats per group were sacrificed after 4, 8, 16, and 20 months exposure for gross
necropsy and histopathological examination.
20 The only tumor detected was a bronchiolar adenoma in the group exposed over 16
21 months to diesel exhaust. No lung tumors were reported in controls. The equivocal response
22 may have been caused by the relatively short exposure durations (20 months) and small numbers
23 of animals examined. In more recent studies, for example, Mauderly et al. (1987), most of the
24 tumors were detected in rats exposed for more than 24 months.
25 General Motors Research Laboratories sponsored chronic inhalation studies at the
26 Southwest Research Institute using male Fischer 344 rats, 30 per group, exposed to DPM
27 concentrations of 0.25, 0.75, or 1.5 mg/m3 (Kaplan et al., 1983; White et al., 1983). The animals
28 were exposed in 12.6 m3 exposure chambers. Airflow was adjusted to provide 13 changes per
29 hour. Temperature was maintained at 22 ± 2 °C. The exposure protocol was 20 hr/day, 7
30 days/week for 9 to 15 months. Exposures were halted during normal working hours for
31 servicing. Some animals were sacrificed following completion of exposure, while others were
32 returned to clean air atmospheres for an additional 8 months. Control animals received clean air.
33 Exhaust was generated by 5.7-L Oldsmobile engines (four different engines used throughout the
34 experiment) operated at a steady speed and load simulating a 40-mph driving speed of a full-size
passenger car.
7/25/00 7-69 DRAFT—DO NOT CITE OR QUOTE
-------
1 Although five instances of bronchoalveolar carcinoma were observed in 90 rats exposed
2 to diesel exhaust for 15 months and held an additional 8 months in clean air, compared with none
3 among controls, statistical significance was not achieved in any of the exposure groups. These
4 included one tumor in the 0.25 mg/m3 group, three in the 0.75 mg/m3 group, and one in the 1.5
5 mg/m3 group. Rats kept in clean-air chambers for 23 months did not exhibit any carcinomas. No
6 tumors were observed in any of the 180 rats exposed to diesel exhaust for 9 or 15 months without
7 a recovery period, or in the respective controls for these groups. Equivocal results may again
8 have been due to less-than-lifetime duration of the study as well as insufficient exposure
9 concentrations. Although the increases in tumor incidences in the groups exposed for 1 5 months
1 0 and held an additional 8 months in clean air were not statistically significant, relative to controls,
1 1 they were slightly greater than the historic background incidence of 3.7% for this specific lesion
1 2 in this strain of rat (Ward, 1983). The first definitive studies linking inhaled diesel exhaust to
1 3 induction of lung cancer in rats were reported by researchers in Germany, Switzerland, Japan,
1 4 and the United States in the mid-to-late 1980s. In a study conducted at the Fraunhofer Institute
1 5 exhaust-generating system and exposure atmosphere characteristics are presented in Appendix A.
1 6 The type of engine used (3-cylinder, 43 bhp diesel) is normally used in mining situations and was
17 of Toxicology and Aerosol Research, female Wistar rats were exposed for 1 9 hr/day, 5
1 8 days/week to both filtered and unfiltered (total) diesel exhaust at an average particulate matter
1 9 concentration of 4.24 mg/m3. Animals were exposed for a maximum of 2.5 years. The exposure
20 system as described by Heinrich et al. (1986a) used a 40 kilowatt 1 .6-L diesel engine operated
21 continuously under the U.S. 72 FTP driving cycle. The engines used European Reference Fuel
22 with a sulfur content of 0.36%. Filtered exhaust was obtained by passing engine exhaust through
23 a Luwa FP-65 HT 610 particle filter heated to 80 °C and a secondary series of filters (Luwa FP-
24 85, Luwa NS-30, and Drager CH 63302) at room temperature. The filtered and unfiltered
25 exhausts were diluted 1:17 with filtered air and passed through respective 12m3 exposure
26 chambers. Mass median aerodynamic diameter of DPM was 0.35 ±0.10 nm (mean ± SD). The
27 gas-phase components of the diesel exhaust atmospheres are presented in Appendix A.
28 The effects of exposure to either filtered or unfiltered exhaust were described by
29 Heinrich et al. (1986b) and Stober (1986). Exposure to unfiltered exhaust resulted in 8
30 bronchoalveolar adenomas and 9 squamous cell tumors in 15 of 95 female Wistar rats examined,
31 for a 1 5.8% tumor incidence. Although statistical analysis was not provided, the increase
32 appears to be highly significant. In addition to the bronchioaiveoiar adenomas and squamous
33 cell tumors, there was a high incidence of bronchioaiveoiar hyperplasia (99%) and metaplasia of
34 th.? bronchioaiveoiar enith
-------
Mohr et al. (1986) provided a more detailed description of the lung lesions and rumors
identified by Heinrich et al. (1986a,b) and Stober (1986). Substantial alveolar deposition of
3 carbonaceous particles was noted for rats exposed to unfiltered diesel exhaust. Squamous
4 metaplasia was observed in 65.3% of the rats breathing unfiltered diesel exhaust, but not in the
5 control rats. Of nine squamous cell tumors, one was characterized as a Grade I carcinoma
6 (borderline atypia, few to moderate mitoses, and slight evidence of stromal invasion), and the
7 remaining eight were classified as benign keratinizing cystic tumors.
8 Iwai et al. (1986) examined the long-term effects of diesel exhaust inhalation on female
9 F344 rats. The exhaust was generated by a 2.4-L displacement truck engine. The exhaust was
1 0 diluted 10:1 with clean air at 20 °C to 25 °C and 50% relative humidity. The engines were
1 1 operated at 1 ,000 rpm with an 80% engine load. These operating conditions were found to
1 2 produce exhaust with the highest particle concentration and lowest NO2 and SO2 content. For
1 3 those chambers using filtered exhaust, proximally installed high-efficiency paniculate air
1 4 (HEP A) filters were used. Three groups of 24 rats each were exposed to unfiltered diesel
1 5 exhaust, filtered diesel exhaust, or filtered room air for 8 hr/day, 7 days/week for 24 months.
1 6 Particle concentration was 4.9 mg/m3 for unfiltered exhaust. Concentrations of gas-phase
1 7 exhaust components were 30.9 ppm NOX, 1 .8 ppm NO2, 13.1 ppm SO2, and 7.0 ppm CO.
No lung tumors were found in the 2-year control (filtered room air) rats, although one
adenoma was noted in a 30-months control rat, providing a spontaneous tumor incidence of
20 4.5%. No lung tumors were observed in rats exposed to filtered diesel exhaust. Nineteen of the
21 24 exposed to unfiltered exhaust survived for 2 years. Of these, 14 were randomly selected for
22 sacrifice at this time. Four of the rats developed lung tumors; two of these were malignant. Five
23 rats of this 2-year exposure group were subsequently placed in clean room air for 3 to 6 months
24 and four eventually (time not specified) exhibited lung tumors (three malignancies). Thus, the
25 lung tumor incidence for total tumors was 42.1% (8/19) and 26.3% (5/19) for malignant tumors
26 in rats exposed to whole diesel exhaust. The tumor types identified were adenoma (3/19),
27 adenocarcinoma (1/19), adenosquamous carcinoma (2/19), squamous carcinoma (1/19), and
28 large-cell carcinoma (1/19). The lung tumor incidence in rats exposed to whole diesel exhaust
29 was significantly greater than that of controls (psO.Ol). Tumor data are summarized in Table
30 7-3. Malignant splenic lymphomas were detected in 37.5% of the rats in the filtered exhaust
31 group and in 25.0% of the rats in the unfiltered exhaust group; these values were significantly
32 (pzQ.05) greater than the 8.2% incidence noted in the control rats. The study demonstrates
33 production of lung cancer in rats following 2-year exposure to unfiltered diesel exhaust. In
addition, splenic malignant lymphomas occurred during exposure to both filtered and unfiltered
34^
7/25/00 7-71 DRAFT— DO NOT CITE OR QUOTE
-------
1 diesel exhaust. This is the only report to date of tumor induction at an extrarespiratory site by
2 inhaled diesel exhaust in animals.
3 A chronic (up to 24 months) inhalation exposure study was conducted by Takemoto et al.
4 (1986), in which female Fischer 344 rats were exposed to diesel exhaust generated by a 269-cc
5 YANMAR-40CE NSA engine operated at an idle state (1,600 rpm). Exposures were 4
6 hours/day, 4 days/week. The animals were exposed in a 376-L exposure chamber. Air flow was
7 maintained at 120 L/min. Exhaust was diluted to produce a particle concentration of 2-4 mg/m3.
8 When not exposed the animals were maintained in an air-conditioned room at a temperature of
9 24 ± 2°C and a relative humidity of 55 ± 5% with 12 hr of light and darkness. Temperature and
10 humidity in the exposure chambers was not noted. The particle concentration of the diesel
11 exhaust in the exposure chamber was 2 to 4 mg/m3. B[a]P and 1-nitropyrene concentrations were
12 0.85 and 93 p.g/g of particles, respectively. No lung tumors were reported in the diesel-exposed
13 animals. It was also noted that the diesel engine employed in this study was originally used as an
14 electrical generator and that its operating characteristics (not specified) were different from those
15 of a diesel-powered automobile. However, the investigators deemed it suitable for assessing the
16 effects of diesel emissions.
17 Mauderly et al. (1987) provided data affirming the carcinogenicity of automotive diesel
18 engine exhaust in F344/Crl rats following chronic inhalation exposure. Male and female rats
19 were exposed to diesel engine exhaust at nominal DPM concentrations of 0.35 (n = 366), 3.5
20 (n = 367), or 7.1 (n = 364) mg/m3 for 7 hr/day, 5 days/week for up to 30 mo. Sham-exposed
21 (n = 365) controls breathed filtered room air. A total of 230,223, 221, and 227 of these rats
22 (sham-exposed, low-, medium-, and high-exposure groups, respectively) were examined for lung
23 tumors. These numbers include those animals that died or were euthanized during exposure and
24 those that were terminated following 30 months of exposure. The exhaust was generated by
25 1980 model 5.7-L Oldsmobile V-8 engines operated through continuously repeating U.S. Federal
26 Test Procedure (FTP) urban certification cycles. The engines were equipped with automatic
27 transmissions connected to eddy-current dynamometers and flywheels simulating resistive and
28 inertial loads of a midsize passenger car. The D-2 diesel control fuel (Phillips Chemical Co.) met
29 U.S. EPA certification standards and contained approximately 30% aromatic hydrocarbons and
30 0.3% sulfur. Following passage through a standard automotive muffler and tailpipe, the exhaust
31 was diluted 10:1 with filtered air in a dilution tunnel and serially diluted to the final
32 concentrations. The primary dilution process was such that particle coagulation was retarded.
33 Mokler et al. (1984) provided a detailed description of the exposure system. No exposure-related
34 changes in body weight or Hfespan were noted for any of the exposed animals, nor were there
35 snv si0ns of overt toxicit*7. Collectiv? !uno turner incidence was greater (7. statisticv
7/25/00 7-72 DRAFT—DO NOT CITE OR QUOTE
-------
1 psQ.05) in the high (7.1 mg/m3) and medium (3.5 mg/m3) exposure groups (12.8% and 3.6%,
f respectively) versus the control and low (0.35 mg/m3) exposure groups (0.9% and 1.3%,
respectively). In the high-dose group the incidences of tumor types reported were adenoma
4 (0.4%), adenocarcinomas plus squamous cell carcinomas (7.5%), and squamous cysts (4.9%). In
5 the medium-dose group adenomas were reported in 2.3% of animals, adenocarcinomas plus
6 squamous cell carcinomas hi 0.5%, and squamous cysts in 0.9%. In the low-exposure group
7 adenocarcinomas plus squamous cell carcinomas were detected in 1.3% of the rats. Using the
8 same statistical analysis of specific tumor types, adenocarcinoma plus squamous cell carcinoma
9 and squamous cyst incidence was significantly greater in the high-exposure group, and the
10 incidence of adenomas was significantly greater in the medium-exposure group. A significant
11 (pO.OOl) exposure-response relationship was obtained for tumor incidence relative to exposure
12 concentration and lung burden of DPM. These data are summarized in Table 7-3. A logistic
13 regression model estimating tumor prevalence as a function of time, dose (lung burden of DPM),
14 and sex indicated a sharp increase in tumor prevalence for the high dose level at about 800 days
15 after the commencement of exposure. A less pronounced, but definite, increase in prevalence
16 with time was predicted for the medium-dose level. Significant effects were not detected at the
17 low concentration. DPM (mg per lung) of rats exposed to 0.35, 3.5, or 7.1 mg of DPM/m3 for 24
months were 0.6, 11.5, and 20.8, respectively, and affirmed the greater-than-predicted
accumulation that was the result of decreased particle clearance following high-exposure
20 conditions.
21 In summary, this study demonstrated the pulmonary carcinogenicity of high
22 concentrations of whole, diluted diesel exhaust in rats following chronic inhalation exposure. In
23 addition, increasing lung particle burden resulting from this high-level exposure and decreased
24 clearance was demonstrated. A logistic regression model presented by Mauderly et al. (1987)
25 indicated that both lung DPM burden and exposure concentration may be useful for expressing
26 exposure-effect relationships.
27 A long-term inhalation study (Ishinishi et al., 1988a; Takaki et al., 1989) examined the
28 effects of emissions from a light-duty (LD) and a heavy-duty (HD) diesel engine on male and
29 female Fischer 344/Jcl rats. The LD engines were 1.8-L, 4-cylinder, swirl-chamber-type power
30 plants, and the HD engines were 11-L, 6-cylinder, direct-injection-type power plants. The
31 engines were connected to eddy-current dynamometers and operated at 1,200 rpm (LD engines)
32 and 1,700 rpm (HD engines). Nippon Oil Co. JIS No. 1 or No. 2 diesel fuel was used. The 30-
33 months whole-body exposure protocol (16 h/day, 6 days/week) used DPM concentrations of 0,
34 0.5, 1, 1.8, or 3.7 mg/m3 from HD engines and 0, 0.1, 0.4, 1.1, or 2.3 mg/m3 from LD engines.
7/25/00 7-73 DRAFT—DO NOT CITE OR QUOTE
-------
1 The animals inhaled the exhaust emissions from 1700 to 0900 h. Sixty-four male rats and 59 to
2 61 female rats from each exposure group were evaluated for carcinogenicity.
3 For the experiments using the LD series engines, the highest incidence of hyperplastic
4 lesions plus tumors (72.6%) was seen in the highest exposure (2.3 mg/m3) group. However, this
5 high value was the result of the 70% incidence of hyperplastic lesions; the incidence of
6 adenomas was only 0.8% and that of carcinomas 1.6%. Hyperplastic lesion incidence was
7 considerably lower for the lower exposure groups (9.7%, 4.8%, 3.3%, and 3.3% for the 1.1, 0.4,
8 and 0.1 mg/m3 and control groups, respectively). The incidence of adenomas and carcinomas,
9 combining males and females, was not significantly different among exposure groups (2.4%,
10 4.0%, 0.8%, 2.4%, and 3.3% for the 2.3, 1.1, 0.4, and 0.1 mg/m3 groups and the controls,
11 respectively).
12 For the experiments using the HD series engines, the total incidence of hyperplastic
13 lesions, adenomas, and carcinomas was highest (26.6%) in the 3.7 mg/m3 exposure group. The
14 incidence of adenomas plus carcinomas for males and females combined equaled 6.5%, 3.3%,
15 0%, 0.8%, and 0.8% at 3.7, 1.8, 1, and 0.4 mg/m3 and for controls, respectively. A statistically
16 significant difference was reported between the 3.7 mg/m3 and the control groups for the HD
17 series engines. The carcinomas were identified as adenomas, adenosquamous carcinomas, and
18 squamous cell carcinomas. Although the number of each was not reported, it was noted that the
19 majority were squamous cell carcinomas. A progressive dose-response relationship was not
20 demonstrated. Tumor incidence data for this experiment are presented in Table 7-3.
•21 The Ishinishi et al. (1988a) study also included recovery tests in which rats exposed to
22 whole diesel exhaust (DPM concentration of 0.1 or 1.1 mg/m3 for the LD engine and 0.5 or
23 1.8 mg/m3 for the HD engine) for 12 months were examined for lung tumors following 6-, 12-, or
24 18-months recovery periods in clean air. The incidences of neoplastic lesions were low, and
25 pulmonary DPM burden was lower than for animals continuously exposed to whole diesel
26 exhaust and not provided a recovery period. The only carcinoma observed was in a rat examined
27 12 months following exposure to exhaust (1.8 mg/m3) from the HD engine.
28 Brightwell et al. (1986,1989) studied the effects of diesel exhaust on male and female
29 F344 rats. The diesel exhaust was generated by a 1.5-L Volkswagen engine that was computer-
30 operated according to the U.S. 72 FTP driving cycle. The engine was replaced after 15 mo. The
31 engine emissions were diluted by conditioned air delivered at 800 m3/h to produce the high-
32 exposure (6.6 mg/m5) diesel exhaust atmosphere. Further dilutions of 1:3 and 1:9 produced the
33 medium- (2.2 mg/m3) and low- (0.7 mg/m3) exposure atmospheres. The CO and NOX
34 concentrations (mean ± SD) were 32 ± 11 pprn and 8 -L 1 ppni in the high-exposure concentration
7/25/00 7-74 DRAFT—DO NOT CITE OR QUOTE
-------
chamber. The inhalation exposures were conducted overnight to provide five 16-h periods per
week for 2 years; surviving animals were maintained for an additional 6 mo.
T For males and females combined, a 1.2% (3/260), 0.7% (1/144), 9.7% (14/144), and
4 38.5% (55/143) incidence of primary lung tumors occurred in F344 rats following exposure to
5 clean air or 0.7,2.2, and 6.6 mg of DPM/m3, respectively (Table 7-3). Diesel exhaust-induced
6 tumor incidence in rats was dose-related and higher in females than in males (Table 7-3). These
7 data included animals sacrificed at the interim periods (6, 12,18, and 24 mo); therefore, the
8 tumor incidence does not accurately reflect the effects of long-term exposure to the diesel
9 exhaust atmospheres. When tumor incidence is expressed relative to the specific intervals, a lung
10 tumor incidence of 96% (24/25), 76% (19/25) of which were malignant, was reported for female
11 rats in the high-dose group exposed for 24 months and held in clean air for the remainder of their
12 lives. For male rats in the same group, the tumor incidence equaled 44% (12/27), of which 37%
13 (10/27) were malignant. It was also noted that many of the animals exhibiting tumors had more
14 than one tumor, often representing multiple histological types. The numbers and types of tumors
15 identified in the rats exposed to diesel exhaust included adenomas (40), squamous cell
16 carcinomas (35), adenocarcinomas (19), mixed adenoma/adenocarcinomas (9), and
17 mesothelioma (1). It should be noted that exposure during darkness (when increased activity
would result in greater respiratory exchange and greater inhaled dose) could account, in part, for
the high response reported for the rats.
20 Lewis et al. (1989) also examined the effects of inhalation exposure of diesel exhaust
21 and/or coal dust on tumorigenesis on F344 rats. Groups of 216 male and 72 female rats were
22 exposed to clean air, whole diesel exhaust (2 mg soot/m3), coal dust (2 mg/m3 respirable
23 concentration; 5 to 6 mg/m3 total concentration), or diesel exhaust plus coal dust (1 mg/m3 of
24 each respirable concentration; 3.2 mg/m3 total concentration) for 7 h/day, 5 days/week during
25 daylight hours for up to 24 mo. Groups of 10 or more males were sacrificed at intermediate
26 intervals (3, 6, and 12 mo). The diesel exhaust was produced by a 7.0-L, 4-cycle, water-cooled
27 Caterpillar Model 3304 engine using No. 2 diesel fuel (<0.5% sulfur by mass). The exhaust was
28 passed through a Wagner water scrubber, which lowered the exhaust temperature and quenched
29 engine backfire. The animals were exposed in 100-cubic-foot chambers. Temperature was
30 controlled at 22±2 °C and relative humidity at 50%±10%. The exhaust was diluted 27-fold with
31 chemically and biologically filtered clean air to achieve the desired particle concentration.
32 Histological examination was performed on 120 to 121 male and 71 to 72 female rats
33 terminated after 24 months of exposure. The exhaust exposure did not significantly affect the
34 rumor incidence beyond what would be expected for aging F344 rats. There was no
7/25/00 7-75 DRAFT—DO NOT CITE OR QUOTE
-------
1 postexposure period, which may explain, in part, the lack of significant tumor induction. The
2 particulate matter concentration was also less than the effective dose in several other studies.
3 In a more recent study reported by Heinrich et al. (1995), female Wistar rats were
4 exposed to whole diesel exhaust (0.8,2.5, or 7.0 mg/m3) 18 h/day, 5 days/week for up to 24 mo,
5 then held in clean air an additional 6 mo. The animals were exposed hi either 6 or 12 m3
6 exposure chambers. Temperature and relative humidity were maintined at 23-25 °C and 50%-
7 70%, repectively. Diesel exhaust was generated by two 40-kw 1.6-L diesel engines
8 (Volkswagen). One of them was operated according to the U.S. 72 cycle. The other was
9 operated under constant load conditions. The first engine did not supply sufficient exhaust,
10 which was filled by the second engine. Cumulative exposures for the rats in the various
11 treatment groups were 61.7, 21.8, and 7.4 g/m3 x h for the high, medium, and low whole-exhaust
12 exposures. Significant increases in tumor incidences were observed in the high (22/100;
13 p<0.001) and mid (11/200; p<0.01) exposure groups relative to clean-air controls (Table 7-3).
14 Only one tumor (1/217), an adenocarcinoma, was observed in clean-air controls. Relative to
15 clean-air controls, significantly increased incidences were observed in the high-exposure rats for
16 benign squamous cell tumors (14/100;/7<0.001), adenomas (4/100; ;?<0.01), and
17 adenocarcinomas (5/100;/><0.05). Only the incidence of benign squamous cell tumors (7/200;
18 joO.Ol) was significantly increased in the mid-exposure group relative to the clean-air controls.
19 Particle lung burden and alveolar clearance also were determined in the Heinrich et al.
20 (1995) study. Relative to clean air controls, alveolar clearance was significantly compromised
21 by exposure to mid and high diesel exhaust. For the high-diesel-exhaust group, 3-mo recovery
22 time in clean air failed to reverse the compromised alveolar clearance.
23 In a study conducted at the Inhalation Toxicology Research Institute (Nikula et al., 1995)
24 F344 rats (114-115 per sex per group) were exposed 16 hr/day, 5 days/week during daylight
25 hours to diesel exhaust diluted to achieve particle concentrations of 2.5 or 6.5 mg/m3 for up to 24
26 mo. Controls (118 males, 114 females) were exposed to clean air. Surviving rats were
27 maintained an additional 6 weeks in clean air, at which time mortality reached 90%. Diesel
28 exhaust was generated with two 1988 Model LH6 General Motors 6.2 L V-8 engines burning D-
29 2 fuel that met EPA certification standards. Chamber air flow was sufficient to provide about 15
30 exchanges per hour. Relative humidity was 40% to 70% and temperature ranged from 23 to 25
31 °C.
32 Following low and high diesel exhaust exposure, the lung burdens were 36.7 and 80.7
33 mg, respectively, for females and 45.1 and 90.1 mg, respectively, for males. The percentages of
34 susceptible rats (males and females combined) with malignant neoplasms were 0.9 (contiol), 3.3
35 (lev/ diesel cxhdust), dnd 12.3 (high chcsc! exhaust). The pcik,cuuigc:> uf iau> (males and females
7/25/00 7-76 DRAFT—DO NOT CITE OR QUOTE
-------
combined) with malignant or benign neoplasms were 1.4 (control), 6.2 (low diesel exhaust), and
17.9 (high diesel exhaust). All primary neoplasms were associated with the parenchyma rather
than the conducting airways of the lungs. The first lung neoplasm was observed at 15 mo.
4 Among 212 males and females examined in the high-dose group, adenomas were detected in 23
5 animals, adenocarcinomas in 22 animals, squamous cell carcinomas in 3 animals, and an
6 adenosquamous carcinoma in 1 animal. For further details see Table 7-3. Analysis of the
7 histopathologic data suggested a progressive process from alveolar epithelial hyperplasia to
8 adenomas and adenocarcinomas.
9 Iwai et al. (1997) carried out a series of exposures to both filtered and whole exhaust
10 using a light-duty (2,369 mL) diesel engine. The protocol for engine operation was not stated.
11 Groups of female SPF F344 Fischer rats were exposed for 2 years for 8 hr/day, 7 days/week, 8
12 hr/day, 6 days/week, or 18 hr/day, 3 days/week to either filtered exhaust or exhaust diluted to a
13 particle concentration of 9.4, 3.2, and 5.1 mg/m3, respectively. Cumulative exposure (mg/m3 x
14 hrs of exposure) equaled 274.4, 153.6, and 258.1 mg/m3. The animals were then held for an
15 additional 6 months in clean air. Lung tumors were reported in 5/121 (4%) of controls, 4/108
16 (4%) of those exposed to filtered exhaust, and 50/153 (35%) among those exposed to whole
17 exhaust. Among rats exposed to whole diesel exhaust the following number of tumors were
« detected; 57 adenomas, 24 adenocarcinomas, 2 benign squamous cell tumors, 7 squamous cell
carcinomas, and 3 adenosquamous carcinomas. The authors stated that benign squamous cell
20 tumors probably corresponded to squamous cysts in another classification.
21
22 7.3.1.2. Mouse Studies
23 A series of inhalation studies using strain A mice was conducted by Orthoefer et al.
24 (1981). Strain A mice are usually given a series of intraperitoneal injections with the test agent;
25 they are then sacrificed at about 9 months and examined for lung tumors. In the present series,
26 inhalation exposure was substituted. Diesel exhaust was provided by one of two Nissan CN6-33
27 diesel engines having a displacement of 3244 cc and run on a Federal Short Cycle. Flow through
28 the exposure chambers was sufficient to provide 15 air changes per hour. Temperature was
29 maintained at 24 °C and relative humidity at 75%. In the first study, groups of 25 male Strong A
30 strain (A/S) mice were exposed to irradiated diesel exhaust (to simulate chemical reactions
31 induced by sunlight) or nonirradiated diesel exhaust (6 mg/m3) for 20 h/day, 7 days/week.
32 Additional groups of 40 Jackson A strain (S/J) mice (20 of each sex) were exposed similarly to
33 either clean air or diesel exhaust, then held in clean air until sacrificed at 9 months of age. No
34 tumorigenic effects were detected at 9 months of age. Further studies were conducted in which
male A/S mice were exposed 8 hr/day, 7 days/week until sacrifice (approximately 300 at 9
7/25/00 7-77 DRAFT—DO NOT CITE OR QUOTE
-------
1 months of age and approximately 100 at 12 months of age). With the exception of those treated
2 with urethan, the number of tumors per mouse did not exceed historical control levels in any of
3 the studies. Exposure to diesel exhaust, however, significantly inhibited the tumorigenic effects
4 of the 5-mg urethan treatment. Results are listed in Table 7-3.
5 Kaplan et al. (1982) also reported the effects of diesel exposure in strain A mice. Groups
6 of male strain A/J mice were exposed for 20 h/day, 7 days/week for 90 days and held until 9
7 months of age. Briefly, the animals were exposed in inhalation chambers to diesel exhaust
8 generated by a 5.7-L Oldsmobile engine operated continuously at 40 mph at DPM concentrations
9 of 0, 0.25, 0.75, or 1.5 mg/m3. Controls were exposed to clean air. Temperature was maintained
10 at 22 ± 2 °C and relative humidity at 50% ± 10% within the chambers. Among 458 controls and
11 485 exposed animals, tumors were detected in 31.4% of those breathing clean air versus 34.2%
12 of those exposed to diesel exhaust. The mean number of tumors per mouse also failed to show
13 significant differences.
14 In a follow-up study, strain A mice were exposed to diesel exhaust for 8 months (Kaplan
15 et al., 1983; White et al., 1983). After exposure to the highest exhaust concentration (1.5
16 mg/m3), the percentage of mice with pulmonary adenomas and the mean number of tumors per
17 mouse were significantly less (p<0.05) than those for controls (25.0% vs. 33.5% and 0.30 ± 0.02
18 [S.E.] vs. 0.42 ± 0.03 [S.E.]) (Table 7-3).
19 Pepelko and Peirano (1983) summarized a series of studies on the health effects of diesel
20 emissions in mice. Exhaust was provided by two Nissan CN 6-33, 6-cylinder, 3.24-L diesel
21 engines coupled to a Chrysler A-272 automatic transmission and Eaton model 758-DG
22 dynamometer. Sixty-day pilot studies were conducted at a 1:14 dilution, providing DPM
23 concentrations of 6 mg/m3 The engines were operated using the Modified California Cycle.
24 These 20-hr/day, 7-days/week pilot studies using rats, cats, guinea pigs, and mice produced
25 decreases in weight gain and food consumption. Therefore, at the beginning of the long-term
26 studies, exposure time was reduced to 8 h/day, 7 days/week at an exhaust DPM concentration of
27 6 mg/m3. During the final 12 months of exposure, however, the DPM concentration was
28 increased to 12 mg/m3. For the chronic studies, the engines were operated using the Federal
29 Short Cycle. Chamber temperature was maintained at 74 °C and relative humidity at 50%.
30 Airflow was sufficient for 15 changes per hour.
31 Pepelko and Peirano (1983) described a two-generation study using Sencar mice exposed
32 to diesel exhaust. Male and female parent-generation mice were exposed to diesel exhaust at a
33 DPM concentration of 6 mg/m3 prior to (from weaning to sexual maturity) and throughout
34 mating. The dams continued exposure through gestation, bnth, oau weaning. Groups 01
3C offspring (13G males cum 130 females) wcic exposed iu cither diesel exhausi or clean air. The
7/25/00 7-78 DRAFT -DO NOT CITE OR QUOTE
-------
exhaust exposure was increased to a DPM concentration of 12 mg/m3 when the offspring were 12
weeks of age and was maintained until termination of the experiment when the mice were 15
3 months old.
4 The incidence of pulmonary adenomas (16.3%) was significantly increased in the mice
5 exposed to diesel exhaust compared with 6.3% in clean-air controls. The incidence in males and
6 females combined was 10.2% in 205 animals examined compared with 5.1% in 205 clean-air
7 controls. This difference was also significant. The incidence of carcinomas was not affected by
8 exhaust exposure in either sex. These results provided the earliest evidence for cancer induction
9 following inhalation exposure to diesel exhaust. The increase in the sensitivity of the study,
10 allowing detection of tumors at 15 mo, may have been the result of exposure from conception. It
11 is likely that Sencar mice are sensitive to induction of lung tumors because they are also sensitive
12 to induction of skin tumors. These data are summarized in Table 7-3.
13 Takemoto et al. (1986) reported the effects of inhaled diesel exhaust (2 to 4 mg/m3,4
14 h/day, 4 days/week, for up to 28 mo) in ICR and C57BL mice exposed from birth. Details of the
15 exposure conditions are presented in Section 7.3.2.1. All numbers reported are for males and
16 females combined. Four adenomas and 1 adenocarcinoma were detected in 34 diesel exhaust-
17 exposed ICR mice autopsied at 13 to 18 mo, compared with 3 adenomas among 38 controls. Six
adenomas and 3 adenocarcinomas were reported in 22 diesel-exposed ICR mice autopsied at 19
to 28 mo, compared with 3 adenomas and 1 adenocarcinoma in 22 controls. Four adenomas and
20 2 adenocarcinomas were detected in 79 C57BL mice autopsied at 13 to 18 mo, compared with
21 none in 19 unexposed animals. Among males and females autopsied at 19 to 28 mo, 8 adenomas
22 and 3 adenocarcinomas were detected in 71 exposed animals, compared with 1 adenoma among
23 32 controls. No significant increases in adenoma or adenocarcinoma were reported for either
24 strain of exposed mice. However, the significance of the increase in the combined incidence of
25 adenomas and carcinomas was not evaluated statistically. A statistical analysis by Pott and
26 Heinrich (1990) indicated that the difference in combined benign and malignant tumors between
27 whole diesel exhaust-exposed C57BL/6N mice and corresponding controls was significant at
28 p<.05. See Table 7-3 for details of tumor incidence.
29 Heinrich et al. (1986b) and Stober (1986), as part of a larger study, also evaluated the
30 effects of diesel exhaust in mice. Details of the exposure conditions reported by Heinrich et al.
31 (1986a) are given in Section 7.3.1.1 and Appendix A. Following lifetime (19 h/day, 5
32 days/week, for a maximum of 120 weeks) exposure to diesel exhaust diluted to achieve a particle
33 concentration of 4.2 mg/m3, 76 female NMRI mice exhibited a total lung tumor incidence of
§ adenomas and adenocarcinomas combined of 32%. Tumor incidences reported for control mice
(n = 84) equaled 11% for adenomas and adenocarcinomas combined. While the incidence of
7/25/00 7-79 DRAFT—DO NOT CITE OR QUOTE
-------
1 adenomas showed little change, adenocarcinomas increased significantly from 2.4% for controls
2 to 17% for exhaust-exposed mice. In a follow-up study, however, Heinrich et al. (1995) reported
3 a lack of tumorigenic response in either female NMRI or C57BL/6N mice exposed 17 h/day, 5
4 days/week for 13.5 to 23 months to whole diesel exhaust diluted to produce a particle
5 concentration of 4.5 mg/m3. These data are summarized in Table 7-3.
6 The lack of a carcinogenic response in mice was reported by Mauderly et al. (1996). In
7 this study, groups of 540 to 600 CD-I male and female mice were exposed to whole diesel
8 exhaust (7.1, 3.5, or 0.35 mg DPM/m3) for 7 hr/day, 5 days/week for up to 24 mo. Controls were
9 exposed to filtered air. Diesel exhaust was provided by 5.7-L Oldsmobile V-8 engines operated
10 continuously on the U.S. Federal Test Procedure urban certification cycle. The chambers were
11 maintained at 25-28 °C, relative humidity at 40%-60%, and a flow rate sufficient for 15 air
12 exchanges per hour. Animals were exposed during the light cycle, which ran from 6:00 AM to
13 6:00 PM. DPM accumulation in the lungs of exposed mice was assessed at 6,12, and 18 months
14 of exposure and was shown to be progressive; DPM burdens were 0.2 ± 0.02, 3.7 ± 0.16, and 5.6
15 ± 0.39 mg for the low-, medium-, and high-exposure groups, respectively. The lung burdens in
16 both the medium- and high-exposure groups exceeded that predicted by exposure concentration
17 ratio for the low-exposure group. Contrary to what was observed in rats (Heinrich et al., 1986b;
18 Stober, 1986; Nikula et al., 1995; Mauderly et al., 1987), an exposure-related increase in primary
19 lung neoplasms was not observed in the CD-I mice, supporting the contention of a species
20 difference in the pulmonary carcinogenic response to poorly soluble particles. The percentage
21 incidence of mice (males and females combined) with one or more malignant or benign
22 neoplasms was 13.4, 14.6, 9.7, and 7.5 for controls and low-, medium-, and high-exposure
23 groups, respectively.
24 Although earlier studies provided some evidence for tumorigenic responses in diesel-
25 exposed mice, no increases were reported in the two most recent studies by Mauderly et al.
26 (1996) and Heinrich et al. (1995), which utilized large group sizes and were well designed and
27 conducted. Overall, the results in mice must therefore be considered to be equivocal.
28
29 7.3.1.3. Hamster Studies
30 Heinrich et al. (1982) examined the effects of diesel exhaust exposure on tumor
31 frequency in female Syrian golden hamsters. Groups of 48 to 72 animals were exposed to clean
32 air or whole diesel exhaust at a mean DPM concentration of 3.9 mg/m3. Inhalation exposures
33 were conducted 7 to 8 hr/day, 5 days/week for 2 years. The exhaust was produced by a 2.4-L
34 Dairnler-Benz engine operated under a constant load and a constant speed of 2,400 rpm. Flow
7/25/00 7-80 DRAFT—DO NOT CITE OR QUOTE
-------
rate was sufficient for about 20 exchanges per hour in the 250-L chambers. No lung tumors were
reported in either exposure group.
3 In a subsequent study, Syrian hamsters were exposed 19 hr/day, 5 days/week for a
4 lifetime to diesel exhaust diluted to a DPM concentration of 4.24 mg/m3 (Heinrich et al., 1986b;
5 Stober, 1986). Details of the exposure conditions are reported in Appendix A. Ninety-six
6 animals per group were exposed to clean air or exhaust. No lung tumors were seen in either the
7 clean-air group or in the diesel exhaust-exposed group.
8 In a third study (Heinrich et al., 1989b), hamsters were exposed to exhaust from a
9 Daimler-Benz 2.4-L engine operated at a constant load of about 15 kW and at a uniform speed of
10 2,000 rpm. The exhaust was diluted to an exhaust-clean air ratio of about 1:13, resulting in a
11 mean particle concentration of 3.75 mg/m3. Exposures were conducted in chambers maintained
12 at 22 to 24 °C and 40% to 60% relative humidity for up to 18 mo. Surviving hamsters were
13 maintained in clean air for up to an additional 6 mo. The animals were exposed 19 hr/day, 5
14 days/week beginning at noon each day, under a 12-hr light cycle starting at 7 AM. Forty animals
15 per group were exposed to whole diesel exhaust or clean air. No lung tumors were detected in
16 either the clean-air or diesel-exposed hamsters.
17 Brightwell et al. (1986, 1989) studied the effects of diesel exhaust on male and female
Syrian golden hamsters. Groups of 52 males and 52 females, 6 to 8 weeks old, were exposed to
diesel exhaust at DPM concentrations of 0.7, 2.2, or 6.6 mg/m3. They were exposed 16 hr/day, 5
20 days/week for a total of 2 years and then sacrificed. Exposure conditions are described in
21 Section 7.3.1.1. No statistically significant (t test) relationship between tumor incidence and
22 exhaust exposure was reported.
23 In summary, diesel exhaust alone did not induce an increase in lung tumors in hamsters
24 of either sex in several studies of chronic duration at high exposure concentrations.
25
26 7.3.1.4. Monkey Studies
27 Fifteen male cynomolgus monkeys were exposed to diesel exhaust (2 mg/m3) for 7
28 hr/day, 5 days/week for 24 months (Lewis et al., 1989). The same numbers of animals were also
29 exposed to coal dust (2 mg/m3 respirable concentration; 5 to 6 mg/m3 total concentration), diesel
30 exhaust plus coal dust (1 mg/m3 respirable concentration for each component; 3.2 mg/m3 total
31 concentration), or filtered air. Details of exposure conditions were listed previously in the
32 description of the Lewis et al. (1989) study with rats (Section 7.3.1.1) and are listed in Appendix
33 A.
34 None of the monkeys exposed to diesel exhaust exhibited a significantly increased
incidence of preneoplastic or neoplastic lesions. It should be noted, however, that the 24-mo
7/25/00 7-81 DRAFT—DO NOT CITE OR QUOTE
-------
1 time frame employed in this study may not have allowed the manifestation of tumors in primates,
2 because this duration is only a small fraction of the monkeys' expected lifespan. In fact, there
3 have been no near-lifetime exposure studies in nonrodent species.
4
5 7.3.2. Inhalation Studies (Filtered Diesel Exhaust)
6 Several studies have been conducted in which animals were exposed to diesel exhaust
7 filtered to remove PM. As these studies also included groups exposed to whole exhaust, details
8 can be found in Sections 7.3.1.1 for rats, 7.3.1.2 for mice, and 7.3.1.3 for hamsters. Heinrich et
9 al. (1986b) and Stober (1986) reported negative results for lung tumor induction in female Wistar
10 rats exposed to filtered exhaust diluted to produce an unfiltered particle concentration of 4.24
11 mg/m3. Negative results were also reported in female Fischer 344 rats exposed to filtered
12 exhaust diluted to produce an unfiltered particle concentration of 4.9 mg/m3 (Iwai et al., 1986), in
13 Fischer 344 rats of either sex exposed to filtered exhaust diluted to produce an unfiltered particle
14 concentration of 6.6 mg/m3 (Brightwell et al., 1989), in female Wistar rats exposed to filtered
15 exhaust diluted to produce an unfiltered particle concentration of 7.0 mg/m3 (Heinrich et al.,
16 1995), and in female Fischer 344 rats exposed to filtered exhaust diluted to produce unfiltered
17 particle concentrations of 5.1, 3.2, or 9.4 mg/m3 (Iwai et al., 1997). In the Iwai et al. (1986)
18 study, splenic lymphomas were detected in 37.5% of the exposed rats compared with 8.2% in
19 controls.
20 In the study reported by Heinrich at al. (1986a) and Stober (1986), primary lung tumors
21 were seen in 29/93 NMRI mice (males and females combined) exposed to filtered exhaust,
22 compared with 11/84 in clean-air controls, a statistically significant increase. In a repeat study
23 by Heinrich et al. (1995), however, significant lung tumor increases were not detected in either
24 female NMRI or C57BL/6N mice exposed to filtered exhaust diluted to produce an unfiltered
25 particle concentration of 4.5 mg/m3.
26 Filtered exhaust also failed to induce lung tumor induction in Syrian Golden hamsters
27 (Heinrich et al., 1986a; Brightwell et al., 1989).
28 Although lung tumor increases were reported in one study and lymphomas in another,
29 these results could not be confirmed in subsequent investigations. It is therefore concluded that
30 little direct evidence exists for carcinogenicity of the vapor phase of diesel exhaust in laboratory
31 animals at concentrations tested.
32
33 7.3.3. Inhalation Studies (Diesel Exhaust Plus Cocarcinogens)
34 Details of the studies reported here have been described earlier and in Table 7-3. Tumor
ethan (1 rng/kg bod> wclglii i.p. al the sian of exposureJ or promotion with
35 initiation with urrth— (\ — ?."
-------
butylated hydroxytolulene (300 mg/kg body weight i.p. week 1, 83 mg/kg week 2, and 150
mg/kg for weeks 3-52) did not influence tumorigenic responses in Sencar mice of both sexes
exposed to concentrations of diesel exhaust up to 12 mg/m3 (Pepelko and Peirano, 1983).
4 Heinrich et al. (1986b) exposed Syrian hamsters of both sexes to diesel exhaust diluted to
5 a particle concentration of 4 mg/m3. See Section 7.3.1.1 for details of the exposure conditions.
6 At the start of exposure the hamsters received either one dose of 4.5 mg diethylnitrosamine
7 (DEN) subcutaneously per kg body weight or 20 weekly intratracheal instillations of 250 \ig BaP.
8 Female NMRI mice received weekly intratracheal instillations of 50 or 100 \igBaP for 10 or 20
9 weeks, respectively, or 50 \ig dibenz[ah]anthracene (DBA) for 10 weeks. Additional groups of
10 96 newborn mice received one s.c. injection of 5 or 10 \ig DBA between 24 and 48 hr after birth.
11 Female Wistar rats received weekly subcutaneous injections of dipentylnitrosamine (DPN) at
12 doses of 500 and 250 mg/kg body weight, respectively, during the first 25 weeks of exhaust
13 inhalation exposure. Neither DEN, DBA, or DPN treatment enhanced any tumorigenic
14 responses to diesel exhaust. Response to BaP did not differ from that of BaP alone in hamsters,
15 but results were inconsistent in mice. Although 20 BaP instillations induced a 71% tumor
16 incidence in mice, concomitant diesel exposure resulted in only a 41% incidence. However,
17 neither 10 BaP instillations nor DBA instillations induced significant effects.
Takemoto et al. (1986) exposed Fischer 344 rats for 2 years to diesel exhaust at particle
concentrations of 2 to 4 mg/m3. One month after start of inhalation exposure one group of rats
20 received di-isopropyl-nitrosamine (DIPN) administered i.p. at 1 mg/kg weekly for 3 weeks.
21 Among injected animals autopsied at 18 to 24 mo, 10 adenomas and 4 adenocarcinomas were
22 reported in 21 animals exposed to clean air, compared with 12 adenomas and 7 adenocarcinomas
23 in 18 diesel-exposed rats. According to the authors, the incidence of adenocarcinomas was not
24 significantly increased by exposure to diesel exhaust.
25 Brightwell et al. (1989) investigated the concomitant effects of diesel exhaust and DEN in
26 Syrian hamsters exposed to diesel exhaust diluted to produce particle concentrations of 0.7, 2.2,
27 or 6.6 mg/m3 for 2 years. The animals received a single dose of 4.5 mg DEN s.c. 3 days prior to
28 start of inhalation exposure. DEN did not affect the lack of responsiveness to diesel exhaust
29 alone. Heinrich et al. (1989b) also exposed Syrian hamsters of both sexes to diesel exhaust
30 diluted to a particle concentration of 3.75 mg/m3 for up to 18 mo. After 2 weeks of exposure,
31 groups were treated with either 3 or 6 mg DEN/kg body weight, respectively. Again, DEN did
32 not significantly influence the lack of tumorigenic responses to diesel exhaust.
33 Heinrich et al. (1989a) investigated the effects of DPN in female Wistar rats exposed to
34 diesel exhaust diluted to achieve a particle concentration of 4.24 mg/m3 for 2-2.5 years. DPN at
doses of 250 and 500 mg/kg body weight was injected subcutaneously once a week for the first
7/25/00 7-83 DRAFT—DO NOT CITE OR QUOTE
-------
1 25 weeks of exposure. The tumorigenic responses to DPN were not affected by exposure to
2 diesel exhaust. For details of exposure conditions of the hamster studies see Section 7.3.1.3.
3 Heinrich et al. (1986a) and Mohr et al. (1986) compared the effects of exposure to
4 particles having only a minimal carbon core but a much greater concentration of PAHs than
5 DPM does. The desired exposure conditions were achieved by mixing coal oven flue gas with
6 pyrolyzed pitch. The concentration of B[a]P and other PAHs per milligram of DPM was about
7 three orders of magnitude greater than that of diesel exhaust. Female rats were exposed to the
8 flue gas-pyrolyzed pitch for 16 hr/day, 5 days/week at particle concentrations of 3 to 7 mg/m3 for
9 22 mo, then held in clean air for up to an additional 12 mo. Among 116 animals exposed, 22
10 tumors were reported in 21 animals, for an incidence of 18.1%. One was a bronchioloalveolar
11 adenoma, one was a bronchioloalveolar carcinoma, and 20 were squamous cell tumors. Among
12 the latter, 16 were classified as benign keratinizing cystic tumors and 4 were classified as
13 carcinomas. No tumors were reported in 115 controls. The tumor incidence in this study was
14 comparable to that reported previously for the diesel exhaust-exposed animals.
15 In analyzing the studies of Heinrich et al. (1986a,b), Heinrich (1990), Mohr et al. (1986),
16 and Stober (1986), it must be noted that the incidence of lung tumors occurring following
17 exposure to whole diesel exhaust, coal oven flue gas, or carbon black (15.8%, 18.1%, and 8% to
18 17%, respectively) was very similar. This occurred despite the fact that the PAH content of the
19 PAH-enriched pyrolyzed pitch was more than three orders of magnitude greater than that of
20 diesel exhaust; carbon black, on the other hand, had only traces of PAHs. Based on these
.21 findings, particle-associated effects appear to be the primary cause of diesel-exhaust-induced
22 lung cancer in rats exposed at high concentrations. This issue is discussed further in Chapter 7.
23
24 7.3.4. Lung Implantation or Intratracheal Instillation Studies
25 7.3.4.1. Rat Studies
26 Grimmer et al. (1987), using female Osbome Mendel rats (35 per treatment group),
27 provided evidence that PAHs in diesel exhaust that consist of four or more rings have
28 carcinogenic potential. Condensate was obtained from the whole exhaust of a 3.0-L passenger-
29 car diesel engine connected to a dynamometer operated under simulated city traffic driving
30 conditions. This condensate was separated by liquid-liquid distribution into hydrophilic and
31 hydrophobic fractions representing 25% and 75% of the total condensate, respectively. The
32 hydrophiiic, hydrophobic, or reconstituted hydrophobic fractions were surgically implanted into
33 the lungs of the rats. Untreated controls, vehicle (beeswax/trioctanoin) controls, and positive
34 (R[a]P) controls were also included in the protocol (Table 7-6). Fraction lib (made up of PAHs
3fi with fnur to seven rings), \vhich accounted foi uul> 0.8% of the loiai weight ol DPM condensate,
7/25/00 7-84 DRAFT—DO NOT CITE OR QUOTE
-------
1 produced the highest incidence of carcinomas following implantation into rat lungs. A
f carcinoma incidence of 17.1% was observed following implantation of 0.21 mg lib/rat, whereas
the nitro-PAH fraction (lid) at 0.18 mg/rat accounted for only a 2.8% carcinoma incidence.
4 Hydrophilic fractions of the DPM extracts, vehicle (beeswax/trioctanoin) controls, and untreated
5 controls failed to exhibit carcinoma formation. Administration of all hydrophobic fractions (Ila-
6 d) produced a carcinoma incidence (20%) similar to the summed incidence of fraction lib
7 (17.1%) and lid (2.8%). The B[a]P positive controls (0.03, 0.1, 0.3 mg/rat) yielded a carcinoma
8 incidence of 8.6%, 31.4%, and 77.1 %, respectively. The study showed that the tumorigenic
9 agents were primarily four- to seven-ring PAHs and, to a lesser extent, nitroaromatics. However,
10 these studies demonstrated that simultaneous administration of various PAH compounds resulted
11 in a varying of the tumorigenic effect, thereby implying that the tumorigenic potency of PAH
12 mixtures may not depend on any one individual PAH. This study did not provide any
13 information regarding the bioavailability of the particle-associated PAHs that might be
14 responsible for carcinogenicity.
15 Kawabata et al. (1986) compared the effects of activated carbon and diesel exhaust on
16 lung tumor formation. One group of 59 F344 rats was intratracheally instilled with DPM (1
17 mg/week for 10 weeks). A second group of 31 rats was instilled with activated carbon using the
same dosing regime. Twenty-seven rats received only the solvent (buffered saline with 0.05%
Tween 80), and 53 rats were uninjected. Rats dying after 18 months were autopsied. All animals
20 surviving 30 months or more postinstillation were sacrificed and evaluated for histopathology.
21 Among 42 animals exposed to DPM surviving 18 months or more, tumors were reported in 31,
22 including 20 malignancies. In the subgroup surviving for 30 mo, tumors were detected in 19 of
23 20 animals, including 10 malignancies. Among the rats exposed to activated carbon, the
24 incidence of lung tumors equaled 11 of 23 autopsied, with 7 cases of malignancy. Data for those
25 dying between 18 and 30 months and those sacrificed at 30 months were not reported separately.
26 Statistical analysis indicated that activated carbon induced a significant increase in lung tumor
27 incidence compared with no tumors in 50 uninjected controls and 1 tumor in 23 solvent-injected
28 controls. The tumor incidence was significantly greater in the DPM-instilled group and was
29 significantly greater than the increase in the carbon-instilled group.
30 A study reported by Rittinghausen et al. (1997) suggested that organic constituents of
31 diesel particles play a role in the induction of lung tumors in rats. An incidence of 16.7%
32 pulmonary cystic keratinizing squamous cell lesions was noted in rats intratracheally instilled
33 with 15 mg whole diesel exhaust particles, compared with 2.1% in rats instilled with 15 mg
34 particles extracted to remove all organic constituents, and none among controls. Instillation of
7/25/00 7-85 DRAFT—DO NOT CITE OR QUOTE
-------
1 30 mg of extracted particles induced a 14.6% incidence of squamous lesions, indicating the
2 greater effectiveness of particles alone as lung particle overload increased.
3 Iwai et al. ( 1 997) instilled 2, 4, 8, and 1 0 mg of whole diesel particles over a 2- to 1 0-
4 week period into female F/344 rats, 50 or more per group. Tumors were reported in 6%, 20%,
5 43%, and 74% of the rats, with incidence of malignant tumors equal to 2%, 13%, 34%, and 48%,
6 respectively. In a second experiment comparing whole with extracted diesel particles, tumor
7 incidence equaled 1/48 (2%) in uninjected controls, 3/55 (5%) in solvent controls, 12/56 (21%)
8 in extracted diesel particles, and 13/106 (12%) in animals injected with unextracted particles.
9 Although the extracted particles appeared to be more potent, when converted to a lung burden
1 0 basis (mg/100 mg dry lung) the incidence was only 14% among those exposed to extracted
1 1 exhaust compared with 3 1% in those exposed to whole particles.
1 2 Dasenbrock et al. (1996) conducted a study to determine the relative importance of the
1 3 organic constituents of diesel particles and particle surface area in the induction of lung cancer in
1 4 rats. Fifty-two female Wistar rats were intratracheally instilled with 16-17 doses of DPM,
1 5 extracted DPM, printex carbon black (PR), lampblack (LB), benzo[a]pyrene (BaP), DPM + BaP,
16 or PR + BaP. The animals were held for a lifetime or sacrificed when moribund. The lungs
1 7 were necropsied and examined for tumors. Diesel particles were collected from a Volkswagen
18 1 .6-L engine operating on a US FTP-72 driving cycle. The mass median aerodynamic diameter
1 9 (MMAD) of the diesel particles was 0.25 (im and the specific surface area was 12 m2/gm.
20 Following extraction with toluene, specific surface area increased to 1 38 m2/gm. The MMAD
21 for extracted PR was equal to 14 nm, while the specific surface area equaled 271 m2/gm. The
22 MMAD for extracted lampblack was equal to 95 nm, with a specific surface area equal to 20
23 m2/gm. The BaP content of the treated particles was 1 1 .3 mg per gm diesel particles and 29.5
24 mg BaP per gm PR. Significant increases in lung tumors were detected in rats instilled with 1 5
25 mg unextracted DPM and 30 mg extracted DPM, but not 1 5 mg extracted DPM. Printex CB was
26 more potent than lampblack CB for induction of lung tumors, whereas BaP was effective only at
27 high doses. Total dose and tumor responses are shown in Table 7-4.
28 A number of conclusions can be drawn from these results. First of aii, particles devoid of
29 organic* are capable of inducing lung tumor formation, as indicated by positive results in the
30 groups treated with high-dose extracted diesel particles and printex. Nevertheless, toluene
3 1 extraction of organics from diesel particles results in a decrease in potency, indicating that the
32 organic fraction does play a role in cancer induction. A relationship between cancer potency and
33 particle surface area was also suggested by the finding that printex with a large specific surface
34 area was more potent than eiihej extracted DPM or iampbiack, which have smaller specific areas.
35 Finally, wliilc vciy Ituge doses of 5aF are very effective in the induction of lung tumors, smaller
7/25/00 7-86 DRAFT— DO NOT CITE OR QUOTE
-------
doses adsorbed to particle surfaces had little detectable effect, suggesting that other organic
components of diesel exhaust may be of greater importance in the induction of lung tumors at
low doses pf BaP (0.2-0.4 mg).
4
5 7.3.4.2. Syrian Hamster Studies
6 Kunitake et al. (1986) and Ishinishi et al. (1988b) conducted a study in which total doses
7 of 1.5, 7.5, or 15 mg of a dichloromethane extract of DPM were instilled intratracheally over 15
8 weeks into male Syrian hamsters that were then held for their lifetimes. The tumor incidences of
9 2.3% (1/44), 0% (0/56), and 1.7% (1/59) for the high-, medium-, and low-dose groups,
10 respectively, did not differ significantly from the 1.7% (1/56) reported for controls. Addition of
11 7.5 mg of B[a]P to a DPM extract dose of 1.5 mg resulted in a total tumor incidence of 91.2%
12 and malignant tumor incidence of 88%. B[a]P (7.5 mg over 15 weeks) alone produced a tumor
13 incidence rate of 88.2% (85% of these being malignant), which was not significantly different
14 from the DPM extract + B[a]P group. Intratracheal administration of 0.03 ng B[a]P, the
15 equivalent content in 15 mg of DPM extract, failed to cause a significant increase in tumors in
16 rats. This study demonstrated a lack of detectable interaction between DPM extract and B[a]P,
17 the failure of DPM extract to induce carcinogenesis, and the propensity for respiratory tract
carcinogenesis following intratracheal instillation of high doses of B[a]P. For studies using the
DPM extract, some concern must be registered regarding the known differences in chemical
20 composition between DPM extract and DPM. As with all intratracheal instillation protocols,
21 DPM extract lacks the complement of volatile chemicals found in whole diesel exhaust.
22 The effects on hamsters of intratracheally instilled DPM suspension, DPM with Fe2O3, or
23 DPM extract with Fe2O3 as the carrier were studied by Shefher et al. (1982). The DPM
24 component in each of the treatments was administered at concentrations of 1.25, 2.5, or 5.0
25 mg/week for 15 weeks to groups of 50 male Syrian golden hamsters. The total volume instilled
26 was 3.0 mL (0.2 mL/week for 15 weeks). The DPM and dichloromethane extracts were
27 suspended in physiological saline with gelatin (0.5% w/v), gum arabic (0.5% w/v), and
28 propylene glycol (10% by volume). The Fe2O3 concentration, when used, was 1.25 mg/0.2 mL
29 of suspension. Controls received vehicle and, where appropriate, carrier particles (Fe2O3)
30 without the DPM component. Two replicates of the experiments were performed. Adenomatous
31 hyperplasia was reported to be most severe in those animals treated with DPM or DPM plus
32 Fe2O3 particles and least severe in those animals receiving DPM plus Fe2O3. Of the two lung
33 adenomas detected microscopically, one was in an animal treated with a high dose of DPM and
the other was in an animal receiving a high dose of DPM extract. Although lung damage was
increased by instillation of DPM, there was no evidence of tumorigenicity.
7/25/00 7-87 DRAFT—DO NOT CITE OR QUOTE
-------
1 7.3.4.3. Mouse Studies
2 Ichinose et al. (1 997a) intratracheally instilled 36 four- week-old male ICR mice per
3 group weekly for 10 weeks with sterile saline or 0.05, 0.1, or 0.2 mg DPM. Particles were
4 collected from a 2.74-L four-cylinder Isuzu engine run at a steady speed of 1 ,500 rpm under a
5 load of 10 torque (kg/m). Twenty-four hours after the last instillation, six animals per group
6 were sacrificed for measurement of lung 8-hydroxydeoxyguanosine(8-OHdG). The remaining
7 animals were sacrificed after 12 months for histopathological analysis. Lung tumor incidence
8 varied from 4/30 (13.3%) for controls to 9/30 (30%), 9/29 (31%), and 7/29 (24.1%) for mice
9 instilled with 0.05, 0.1, and 0.2 mg/week, respectively. The increase in animals with lung tumors
1 0 compared with controls was statistically significant for the 0.1 mg dose group, the only group
1 1 analyzed statistically. Increases in 8-OHdG, an indicator of oxidative DNA damage, correlated
1 2 well with the increase in tumor incidence in the 0.05 mg dose group, although less so with the
1 3 other two. The correlation coefficients r = 0.916, 0.765, and 0.677 for the 0.05, 0.10, and 0.20 mg
1 4 DPM groups, respectively.
15 In a similar study, 33 four- week-old male ICR mice per group were intratracheally
1 6 instilled weekly for 1 0 weeks with sterile saline, 0.1 mg DPM, or 0.1 mg DPM from which the
1 7 organic constituents were extracted with hexane (Ichinose et al., 1997b). Exhaust was collected
1 8 from a 2.74-L four-cylinder Izuzu engine run at a steady speed of 2,000 rpm under a load of 6
1 9 torque (kg/m). Twenty-four hours after the last instillation, six animals per group were sacrificed
20 for measurement of 8-OHdG. Surviving animals were sacrificed after 12 mo. The incidence of
21 lung tumors increased from 3/27 (11.1%) among controls to 7/27 (25.9%) among those instilled
22 with extracted diesel particles and 9/26 (34.6%) among those instilled with unextracted particles.
23 The increase in number of tumor-bearing animals was statistically significant compared with
24 controls (/?<0.05) for the group treated with unextracted particles. The increase in 8-OHdG was
25 highly correlated with lung tumor incidence, r = 0.99.
26
27 7.3.5. Subcutaneous and Intraperitoneal Injection Studies
28 7.3.5.L Mouse Studies
29 In addition to inhalation studies, Ormoefer et al. (1981) also tested the effects of i.p.
30 injections of DPM on male (A/S) strain mice. Three groups of 30 mice were injected with 0. 1
31 mL of a suspension (particles in distilled water) containing 47, 1 17, or 235 ^g of DPM collected
32 from Fluoropore filters in the inhalation exposure chambers. The exposure system and exposure
33 atmosphere are described in Appendix A. Vehicle controls received injections of particle
34 suspension made up cf particulate matter ftum control exposure niters, positive controls received
20 mg
cf urcthan, and negative controls received no injections. Injections were made three times
7/25/00 7-88 DRAFT— DO NOT CITE OR QUOTE
-------
1 weekly for 8 weeks, resulting in a total DPM dose of 1.1,2.8, and 5.6 mg for the low-, medium-,
fand high-dose groups and 20 mg of urethan for the positive control group. These animals were
sacrificed after 26 weeks and examined for lung tumors. For the low-, medium-, and high-dose
4 DPM groups, the tumor incidence was 2/30, 10/30, and 8/30, respectively. The incidence among
5 urethan-treated animals (positive controls) was 100% (29/29), with multiple tumors per animal.
6 The tumor incidence for the DPM-treated animals did not differ significantly from that of vehicle
7 controls (8/30) or negative controls (7/28). The number of tumors per mouse was also unaffected
8 by treatment.
9 In further studies conducted by Orthoefer et al. (1981), an attempt was made to compare
10 the potency of DPM with that of other environmental pollutants. Male and female Strain A mice
11 were injected i.p. three times weekly for 8 weeks with DPM, DPM extracts, or various
12 environmental mixtures of known carcinogenicity, including cigarette smoke condensate, coke
13 oven emissions, and roofing tar emissions. Injection of urethan or dimethylsulfoxide (DMSO)
14 served as positive or vehicle controls, respectively. In addition to DPM from the Nissan diesel
15 previously described, an eight-cylinder Oldsmobile engine operated at the equivalent of 40 mph
16 was also used to compare emission effects from different makes and models of diesel engine.
17 The mice were sacrificed at 9 months of age and their lungs examined for histopathological
changes. The only significant findings, other than for positive controls, were small increases in
numbers of lung adenomas per mouse in male mice injected with Nissan DPM and in female
20 mice injected with coke oven extract. Furthermore, the increase in the extract-treated mice was
21 significant only in comparison with uninjected controls (not injected ones) and did not occur
22 when the experiment was repeated. Despite the use of a strain of mouse known to be sensitive to
23 tumor induction, the overall findings of this study were negative. The authors provided several
24 possible explanations for these findings, the most likely of which were (1) the carcinogens that
25 were present were very weak, or (2) the concentrations of the active components reaching the
26 lungs were insufficient to produce positive results.
27 Kunitake et al. (1986) conducted studies using DPM extract obtained from a 1983 HD
28 MMC—6D22P 1 l-L V-6 engine. Five s.c. injections of DPM extract (500 mg/kg per injection)
29 resulted in a significant (p<0.01) increase in subcutaneous tumors for female C57BL mice (5/22
30 [22.7%] vs. 0/38 among controls). Five s.c. doses of DPM extract of 10, 25, 30, 100, or 200
31 mg/kg failed to produce a significant increase in tumor incidence. One of 12 female ICR mice
32 (8.3%) and 4 of 12 male ICR mice (33.3%) developed malignant lymphomas following neonatal
33 s.c. administration of 10 mg of DPM extract per mouse. The increase in malignant lymphoma
34 incidence for the male mice was statistically significant at/?<0.05 compared with an incidence of
7/25/00 7-89 DRAFT—DO NOT CITE OR QUOTE
-------
1 2/14 (14.3%) among controls. Treatment of either sex with 2.5 or 5 mg of DPM extract per
2 mouse did not result in statistically significant increases in tumor incidence.
3 Additional studies using DPM extract from LD (1.8-L, 4-cylinder) as well as HD engines
4 with female ICR and nude mice (BALB/c/cA/JCL-nu) were also reported (Kunitake et al., 1988).
5 Groups of 30 ICR and nude mice each were given a single s.c. injection of 10 mg HD extract, 10
6 mg HD + 50 ng 12-O-tetradecanoylphorbol 13-acetate (TPA), 10 mg LD extract + 50 jig TPA, or
7 50 ng TPA. No malignant tumors or papillomas were observed. One papillomatous lesion was
8 observed in an ICR mouse receiving LD extract + TPA, and acanthosis was observed in one nude
9 mouse receiving only TPA.
10 In what appears to be an extension of the Kunitake et al. (1986) s.c. injection studies,
11 Takemoto et al. (1988) presented additional data for subcutaneously administered DPM extract
12 from HD and LD diesel engines. In this report, the extracts were administered to 5-week-old and
13 neonatal (<24 hr old) C57BL mice of both sexes. DPM extract from HD or LD engines was
14 administered weekly to the 5-week-old mice for 5 weeks at doses of 10, 25, 50, 100, 200, or 500
15 mg/kg, with group sizes ranging from 15 to 54 animals. After 20 weeks, comparison with a
16 control group indicated a significant increase in the incidence of subcutaneous tumors for the 500
17 mg/kg HD group (5 of 22 mice [22.7%]5jp<0.01), the 100 mg/kg LD group (6 of 32 [18.8%],
18 /KO.01), and the 500 mg/kg LD group (7 of 32 [21.9%], p<0.01) in the adult mouse experiments.
19 The tumors were characterized as malignant fibrous histiocytomas. No tumors were observed in
20 other organs. The neonates were given single doses of 2.5, 5, or 10 mg DPM extract
21 subcutaneously within 24 hr of birth. There was a significantly higher incidence of malignant
22 lymphomas in males receiving 10 mg of HD extract and of lung tumors for males given 2.5 mg
23 HD extract and for males given 5 mg and females given 10 mg LD extract. A dose-related trend
24 that was not significant was observed for the incidences of liver tumors for both the HD extract-
25 and LD extract-treated neonatal mice. The incidence of mammary tumors in female mice and
26 multiple-organ tumors in male mice was also greater for some extract-treated mice, but was not
27 dose related. The report concluded that LD DPM extract showed greater carcinogenicity than did
28 HD DPM extract.
29
30 7.3.6. Dermal Studies
31 7.3.6.1. Mouse Studies
32 In one of the earliest studies of diesel emissions, the effects of dermal application of
33 extract from DPM were examined by Kotin et al. (1955). Acetone extracts were prepared from
34 the DPM of a diesel engine (type and size net provided) operated at \varrnup rncdc and under
35 load. These extracts \vere applied dennally three times •weekly to male and female C57BL and
7/25/00 7-90 DRAFT—DO NOT CITE OP. QUOTE
-------
strain A mice. Results of these experiments are summarized in Table 7-5. In the initial
experiments using 52 (12 male, 40 female) C57BL mice treated with DPM extract from an
3~ engine operated in warmup mode, two papillomas were detected after 13 mo. Four tumors were
4 detected 16 months after the start of treatment in 8 surviving of 50 exposed male strain A mice
5 treated with DPM extract from an engine operated under full load. Among female strain A mice
6 treated with DPM extract from an engine operated under full load, 17 tumors were detected in 20
7 of 25 mice surviving longer than 13 mo. This provided a significantly increased tumor incidence
8 of 85%. Carcinomas as well as papillomas were seen, but the numbers were not reported.
9 Depass et al. (1982) examined the potential of DPM and dichloromethane extracts of
10 DPM to act as complete carcinogens, carcinogen initiators, or carcinogen promoters. In skin-
11 painting studies, the DPM was obtained from an Oldsmobile 5.7-L diesel engine operated under
12 constant load at 65 km/h. The DPM was collected at a temperature of 100°C. Groups of 40
13 C3H/HeJ mice were used because of their low spontaneous tumor incidence. For the complete
14 carcinogenesis experiments, DPM was applied as a 5% or 10% suspension in acetone.
15 Dichloromethane extract was applied as 5%, 10%, 25%, or 50% suspensions. Negative controls
16 received acetone, and positive controls received 0.2% B[a]P. For tumor-promotion experiments,
17 a single application of 1.5% B[a]P was followed by repeated applications of 10% DPM
suspension, 50% DPM extract, acetone only (vehicle control), 0.0001% phorbol 12-myristate 13-
acetate (PMA) as a positive promoter control, or no treatment (negative control). For the tumor-
20 initiation studies, a single initiating dose of 10% diesel particle suspension, 50% diesel particle
21 extract, acetone, or PMA was followed by repeated applications of 0.0001% PMA. Following 8
22 months of treatment, the PMA dose in the initiation and promotion studies was increased to
23 0.01%. Animals were treated three times per week in the complete carcinogenesis and initiation
24 experiments and five times per week in promotion experiments. All test compounds were
25 applied to a shaved area on the back of the mouse.
26 In the complete carcinogenesis experiments, one mouse receiving the high-dose (50%)
27 suspension of extract developed a squamous cell carcinoma after 714 days of treatment. Tumor
28 incidence in the B[a]P group was 100%, and no tumors were observed in any of the other groups.
29 For the promotion studies, squamous cell carcinomas with pulmonary metastases were identified
30 in one mouse of the 50% DPM extract group and in one in the 25% extract group. Another
31 mouse in the 25% extract group developed a grossly diagnosed papilloma. Nineteen positive
32 control mice had tumors (11 papillomas, 8 carcinomas). No tumors were observed for any of the
33 other treatment groups. For the initiation studies, three tumors (two papillomas and one
34 carcinoma) were identified in the group receiving DPM suspension and three tumors (two
7/25/00 7-91 DRAFT—DO NOT CITE OR QUOTE
-------
1 papillomas and one fibrosarcoma) were found in the DPM extract group. These findings were
2 reported to be statistically insignificant using the Breslow and Mantel-Cox tests.
3 Although these findings were not consistent with those of Kotin et al. (1955), the
4 occurrence of a single carcinoma in a strain known to have an extremely low spontaneous tumor
5 incidence may be of importance. Furthermore, a comparison between studies employing
6 different strains of mice with varying spontaneous tumor incidences may result in erroneous
7 assumptions.
8 Nesnow et al. (1982) studied the formation of dermal papillomas and carcinomas
9 following dermal application of dichloromethane extracts from coke oven emissions, roofing tar,
10 DPM, and gasoline engine exhaust. DPM from five different engines, including a preproduction
11 Nissan 220C, a 5.7-L Oldsmobile, a prototype Volkswagen Turbo Rabbit, a Mercedes 300D, and
12 a HD Caterpillar 3304, was used for various phases of the study. Male and female Sencar mice
13 (40 per group) were used for tumor initiation, tumor promotion, and complete carcinogenesis
14 studies. For the tumor-initiation experiments, the DPM extracts were topically applied in single
15 doses of 100, 500, 1,000, or 2,000 (ig/mouse. The high dose (10,000 ^g/mouse) was applied in
16 five daily doses of 2,000 ng. One week later, 2 ^g of the tumor promoter TPA was applied
17 topically twice weekly. The tumor-promotion experiments used mice treated with 50.5 jig of
18 B[a]P followed by weekly (twice weekly for high dose) topical applications (at the
19 aforementioned doses) of the extracts. For the complete carcinogenesis experiments, the test
20 extracts were applied weekly (twice weekly for the high doses) for 50 to 52 weeks. Only extracts
21 from the Nissan, Oldsmobile, and Caterpillar engines were used in the complete carcinogenesis
22 experiments.
23 In the tumor-initiation studies, both B[a]P alone and the Nissan engine DPM extract
24 followed by TPA treatment produced a significant increase in tumor (dermal papillomas)
25 incidence at 7 to 8 weeks postapplication. By 15 weeks, the tumor incidence was greater than
26 90% for both groups. No significant carcinoma formation was noted for mice in the tumor-
27 initiation experiments following exposure to DPM extracts of the other diesel engines, although
28 the Oldsmobile engine DPM extract at 2.0 mg/mouse did produce a 40% papillonia incidence in
29 male mice at 6 mo. This effect, however, was not dose dependent.
30 B[a]P (50.5 ^g/week), coke oven extract (at 1.0, 2.0, or 4.0 mg/week), and the highest
31 dose of roofing tar extract (4.0 mg/week) all tested positive for complete carcinogenesis activity.
32 DPM extracts from only the Nissan, Oldsmobile, and Caterpillar engines were tested for
33 complete carcinogenic potential, and all three proved to be negative using the Sencar mouse
7/25/00 7-92 DRAFT—DO NOT CITE OR QUOTE
-------
The results of the dermal application experiments by Nesnow et al. (1982) are presented
in Table 7-7. The tumor initiation-promotion assay was considered positive if a dose-dependent
response was obtained and if at least two doses provided a papilloma-per-mouse value that was
4 three times or greater than that of the background value. Based on these criteria, only emissions
5 from the Nissan were considered positive. Tumor initiation and complete carcinogenesis assays
6 required that at least one dose produce a tumor incidence of at least 20%. None of the DPM
7 samples yielded positive results based on this criterion.
8 Kunitake et al. (1986,1988) evaluated the effects of a dichloromethane extract of DPM
9 obtained from a 1983 MMC M-6D22P 11-L V-6 engine. An acetone solution was applied in 10
10 doses every other day, followed by promotion with 2.5 \ig of TPA three times weekly for 25
11 weeks. Exposure groups received a total dose of 0.5, 5, 15, or 45 mg of extract. Papillomas
12 were reported in 2 of 50 animals examined in the 45 mg exposure group and in 1 of 48 in the 15
13 mg group compared with 0 of 50 among controls. Differences, however, were not statistically
14 significant.
15
16 7.3.7. Summary and Conclusions of Laboratory Animal Carcinogenicity Studies
17 As early as 1955, Kotin et al. (1955) provided evidence for tumorigenicity and
carcinogenicity of acetone extracts of DPM following dermal application and also provided data
suggesting a difference in this potential depending on engine operating mode. Until the early
20 1980s, no chronic studies assessing inhalation of diesel exhaust, the relevant mode for human
21 exposure, had been reported. Since then long-term inhalation bioassays with diesel exhaust have
22 been carried out in the United States, Germany, Switzerland, and Japan, testing responses of rats,
23 mice, and Syrian hamsters, and to a limited extent cats and monkeys.
24 It can be reasonably concluded that with adequate exposure, inhalation of diesel exhaust
25 is capable of inducing lung cancer in rats. Responses best fit cumulative exposure (concentration
26 x daily exposure duration x days of exposure). Examination of rat data shown in Table 7-8
27 indicates a trend of increasing tumor incidence at exposures exceeding 1 x 104 mg-hr/m3.
28 Exposures greater than approximately this value result in lung particle overload, characterized by
29 slowed particle clearance and lung pathology, as discussed in Chapters 3 and 5, respectively.
30 Tumor induction at high doses may therefore be primarily the result of lung particle overload
31 with associated inflammatory responses. Although tumorigenic responses could not be detected
32 under non-particle-overload conditions, the animal experiments lack sensitivity to determine if a
33 threshold exists. However, studies such as those reported by Driscoll et al. (1996) support the
34 existence of a threshold if it is assumed that inflammation is a prerequisite for lung tumor
7/25/00 7-93 DRAFT—DO NOT CITE OR QUOTE
-------
1 induction. If low-dose effects do occur, it can be hypothesized that the organic constituents are
2 playing a role. See Chapter? for a discussion of this issue.
3 Although rats develop adenomas, adenocarcinomas, and adenosquamous cell carcinomas,
4 they also develop squamous keratinizing lesions. This latter spectrum appears for the most part
5 to be peculiar to the rat. In a recent workshop aimed at classifying these tumors (Boorman et al.,
6 1996), it was concluded that when these lesions occur in rats as part of a carcinogenicity study,
7 they must be evaluated on a case-by-case basis and regarded as a part of the total biologic profile
8 of the test article. If the only evidence of tumorigeniciry is the presence of cystic keratinizing
9 epitheliomas, it may not have relevance to human safety evaluation of a substance or particle.
10 Their use in quantifying cancer potency is even more questionable.
11 The evidence for response of common strains of laboratory mice exposed under standard
12 inhalation protocols is equivocal. Inhalation of diesel exhaust induced significant increases in
13 lung tumors in female NMRI mice (Heinrich et al., 1986b; Stober, 1986) and in female Sencar
14 mice (Pepelko and Peirano, 1983). An apparent increase was also seen in female C57BL mice
15 (Takemoto et al., 1986). However, in a repeat of their earlier study, Heinrich et al. (1995) failed
16 to detect lung tumor induction in either NMRI or C57BL/6N mice. No increases in lung tumor
17 rates were reported in a series of inhalation studies using strain A mice (Orthoefer et al., 1981;
18 Kaplan et al., 1982, 1983; White et al., 1983). Finally, Mauderly et al. (1996) reported no
19 tumorigenic responses in CD-I mice exposed under conditions resulting in positive responses in
20 rats. The successful induction of lung tumors in mice by Ichinose et al. (1997a,b) via
.21 intratracheal instillation may have been the result of focal deposition of larger doses. Positive
22 effects in Sencar mice may be due to use of a strain sensitive to tumor induction in epidermal
23 tissue by organic agents, as well as exposure from conception, although proof for such a
24 hypothesis is lacking.
25 Attempts to induce significant increases in lung tumors in Syrian hamsters by inhalation
26 of whole diesel exhaust were unsuccessful (Heinrich et al., 1982, 1986b, 1989b; Brightwell et al.,
27 1986). However, hamsters are considered to be relatively insensitive to lung tumor induction.
28 For example, while cigarette smoke, a known human carcinogen, was shown to induce laryngeal
29 cancer in hamsters, the lungs were relatively unaffected (Dontenwill et al., 1973).
30 Neither cats (Pepelko and Peirano, 1983 [see Chapter 7]) nor monkeys (Lewis et al.,
31 1986) developed tumors following 2-year exposure to diesel exhaust. The duration of these
32 exposures, however, was likely to be inadequate for these two longer-lived species, and group
33 sizes were quite small. Exposure levels were also below the maximum tolerated dose (MTD) in
34 the monkey studies and, in fact, only borderline for detection of lung turner increases in rats.
7/25/00 7-94 DRAFT—DO NOT CITE OR QUOTE
-------
1 Long-term exposure to diesel exhaust filtered to remove paniculate matter failed to
f induce lung tumors in rats (Heinrich et al., 1986b; Iwai et al., 1986; Brightwell et al., 1989), or in
Syrian hamsters (Heinrich et al., 1986b; Brightwell, 1989). A significant increase in lung
4 carcinomas was reported by Heinrich et al. (1986b) in NMRI mice exposed to filtered exhaust.
5 However, in a more recent study the authors were unable to confirm earlier results in either
6 NMRI or C57BL/6N mice (Heinrich et al., 1995). Although filtered exhaust appeared to
7 potentiate the carcinogenic effects of DEN (Heinrich et al., 1982), because of the lack of positive
8 data in rats and equivocal or negative data in mice it can be concluded that filtered exhaust is
9 either not carcinogenic or has a low cancer potency.
10 Kawabata et al. (1986) demonstrated the induction of lung tumors in Fischer 344 rats
11 following intratracheal instillation of DPM. Rittinghausen et al. (1997) reported an increase in
12 cystic keratinizing epitheliomas following intratracheal instillation of rats with either original
13 DPM or DPM extracted to remove the organic fraction, with the unextracted particles inducing a
14 slightly greater effect. Grimmer et al. (1987) showed not only that an extract of DPM was
15 carcinogenic when instilled in the lungs of rats, but also that most of the carcinogenicity resided
16 in the portion containing PAHs with four to seven rings. Intratracheal instillation did not induce
17 lung tumors in Syrian hamsters (Kunitake et al., 1986; Ishinishi et al., 1988b).
Dermal exposure and s.c. injection in mice provided additional evidence for tumorigenic
effects of DPM. Particle extracts applied dermally to mice have been shown to induce
20 significant skin tumor increases in two studies (Kotin et al., 1955; Nesnow et al., 1982).
21 Kunitake et al. (1986) also reported a marginally significant increase in skin papillomas in ICR
22 mice treated with an organic extract from an HD diesel engine. Negative results were reported
23 by Depass et al. (1982) for skin-painting studies using mice and acetone extracts of DPM
24 suspensions. However, in this study the exhaust particles were collected at temperatures of 100
25 °C, which would minimize the condensation of vapor-phase organics and, therefore, reduce the
26 availability of potentially carcinogenic compounds that might normally be present on diesel
27 exhaust particles. A significant increase in the incidence of sarcomas in female C57B1 mice was
28 reported by Kunitake et al. (1986) following s.c. administration of LD DPM extract at doses of
29 500 mg/kg. Takemoto et al. (1988) provided additional data for this study and reported an
30 increased tumor incidence in the mice following injection of LD engine DPM extract at doses of
31 100 and 500 mg/kg. Results of i.p. injection of DPM or DPM extracts in strain A mice were
32 generally negative (Orthoefer et al., 1981; Pepelko and Peirano, 1983), suggesting that the strain
33 A mouse may not be a good model for testing diesel emissions.
Results of experiments using tumor initiators such as DEN, B[a]P, DPN, or DBA
(Brightwell et al., 1986; Heinrich et al., 1986b; Takemoto et al., 1986) were generally
7/25/00 7-95 DRAFT—DO NOT CITE OR QUOTE
-------
1 inconclusive regarding the tumor-promoting potential of either filtered or whole diesel exhaust.
2 A report by Heinrich et al. (1 982), however, indicated that filtered exhaust may promote the
3 tumor-initiating effects of DEN in hamsters.
4 Several reports (Wong et al., 1985; Bond et al., 1990) affirm observations of the potential
5 carcinogenicity of diesel exhaust by providing evidence for DNA damage in rats. These findings
6 are discussed in more detail in Section 3.6. Evidence for the mutagenicity of organic agents
7 present in diesel engine emissions is also provided in Chapter 4.
8 Evidence for the importance of the carbon core was initially provided by studies of
9 Kawabata et al. (1986), which showed induction of lung tumors following intratracheal
1 0 instillation of carbon black that contained no more than traces of organics, and studies of
1 1 Heinrich (1990) that indicated that exposure via inhalation to carbon black (Printex 90) particles
1 2 induced lung tumors at concentrations similar to those effective in DPM studies. Additional
1 3 studies by Heinrich et al. (1995) and Nikula et al. (1995) confirmed the capability of carbon
1 4 particles to induce lung tumors. Induction of lung tumors by other particles of low solubility,
1 5 such as titanium dioxide (Lee et al., 1986), confirmed the capability of particles to induce lung
1 6 tumors. Pyrolyzed pitch, on the other hand, essentially lacking a carbon core but having much
1 7 higher PAH concentrations than DPM, also was effective in tumor induction (Heinrich et al.,
18 1986a, 1994).
1 9 The relative importance of the adsorbed organics, however, remains to be elucidated and
20 is of some concern because of the known carcinogenic capacity of some of these chemicals.
21 These include polycyclic aromatics as well as nitroaromatics, as described in Chapter 2. Organic
22 extracts of particles also have been shown to induce tumors in a variety of injection, intratracheal
23 instillation, and skin-painting studies, and Grimmer et al. (1987) have, in fact, shown that the
24 great majority of the carcinogenic potential following instillation resided in the fraction
25 containing four- to seven-ring PAHs.
26 In summary, based on positive inhalation studies in rats exposed to high concentrations,
27 intratracheal instillation studies in rats and mice exposed to high doses, and supported by positive
28 mutagenicity studies, the evidence for carcinogenicity of diesel exhaust is considered to be
29 adequate in animals. The contribution of the various fractions of diesel exhaust to the
30 carcinogenic response is less certain. Exposure to filtered exhaust generally failed to induce lung
3 1 tumors. The presence of known carcinogens adsorbed to diesel particles and the demonstrated
32 tumorigenicity of particle extracts in a variety of injection, instillation, and skin-painting studies
33 indicate a carcinogenic potential for the organic fraction. Studies showing that long-term
O A ov»-»/-»ovM%a ot Kirrk *"**"* £TvtT-otir\Tic of r»r*r\rliw cr*1viV*1o rvdrti/^liac i (* rr
•« r* A**A AaCi-m+w TC* £»Tri /•?£»»•»/-*£» +V\ o + t V* » /» o t* « /•* TI /•
7/25/00 7-96 DRAFT— DO NOT CITE OR QUOTE
-------
1 diesel particle is primarily instrumental in the carcinogenic response observed in rats under
f sufficient exposure conditions. The ability of diesel exhaust to induce lung rumors at non-
particle-overload conditions, and the relative contribution of the particles' core versus the
4 particle-associated organics (if effects do occur at low doses) remains to be determined.
5
6 7.4. MODE OF ACTION OF DIESEL EMISSION-INDUCED CARCINOGENESIS
7 As noted in Chapter 2, diesel exhaust is a complex mixture that includes a vapor phase
8 and a particle phase. The particle phase consists of an insoluble carbon core with a large number
9 of organic compounds, as well as inorganic compounds such as sulfates, adsorbed to the particle
10 surface. Some of the semivolatile and particle-associated compounds, in particular PAHs, nitro-
11 PAHs, oxy-PAHs, and oxy-nitro-PAHs (Scheepers and Bos, 1992), are considered likely to be
12 carcinogenic in humans. The vapor phase also contains a large number of organic compounds,
13 including several known or probable carcinogens such as benzene and 1,3-butadiene. Because
14 exposure to the vapor phase alone, even at high concentrations, failed to induce lung cancer in
15 laboratory animals (Heinrich et al., 1986), discussion will focus on the particulate matter phase.
16 Additive or synergistic effects of vapor-phase components, however, cannot be totally
17 discounted, as chronic inhalation bioassays involving exposure to diesel particles alone have not
been carried out.
Several hypotheses regarding the primary mode of action of diesel exhaust have been
20 proposed. Initially it was generally believed that cancer was induced by particle-associated
21 organics acting via a genotoxic mechanism. By the late 1980s, however, studies indicated that
22 carbon particles virtually devoid of organics could also induce lung cancer at sufficient inhaled
23 concentrations (Heinrich, 1990). This finding provided support for a hypothesis originally
24 proposed by Vostal (1986) that induction of lung tumors arising in rats exposed to high
25 concentrations of diesel exhaust is related to overloading of normal lung clearance mechanisms,
26 accumulation of particles, and cell damage followed by regenerative cell proliferation. The
27 action of particles is therefore mediated by epigenetic mechanisms that can be characterized
28 more by promotional than initiation stages of the carcinogenic process. More recently several
29 studies have focused upon the production of reactive oxygen species generated from particle-
30 associated organics, which may induce oxidative DNA damage at exposure concentrations lower
31 than those required to produce lung particle overload. Because it is likely that more than one of
32 these factors is involved in the carcinogenic process, a key consideration is their likely relative
33 contribution at different exposure levels. The following discussion will therefore consider the
34^ possible relationship of the organic components of exhaust, inflammatory responses associated
with lung particle overload, reactive oxygen species, and physical characteristics of diesel
7/25/00 7-97 DRAFT—DO NOT CITE OR QUOTE
-------
1 particles to cancer induction, followed by a hypothesized mode of action, taking into account the
2 likely contribution of the factors discussed.
3
4 7.4.1. Potential Role of Organic Exhaust Components in Lung Cancer Induction
5 More than 100 carcinogenic or potentially carcinogenic components have been
6 specifically identified in diesel emissions, including various PAHs and nitroarenes such as
7 1-nitropyrene (1-NP) and dinitropyrenes (DNPs). The majority of these compounds are adsorbed
8 to the carbon core of the particulate phase of the exhaust and, if desorbed, may become available
9 for biological processes such as metabolic activation to mutagens. Among such compounds
10 identified from diesel exhaust are benzo(a)pyrene (B[a]P), dibenz[a,/z]anthracene, pyrene,
11 chrysene, and nitroarenes such as 1-NP, 1,3-DNP, 1,6-DNP, and 1,8-DNP, all of which are
12 mutagenic, carcinogenic, or implicated as procarcinogens or cocarcinogens (Stenback et al.,
13 1976; Weinstein and Troll, 1977; Thyssen et al., 1981; Pott and Stober, 1983; Howard et al.,
14 1983; Hirose et al., 1984; Nesnow et al., 1984; El-Bayoumy et al., 1988). More recently Enya et
15 al. (1997) reported isolation of 3-nitrobenzanthrone, one of the most powerful direct-acting
16 mutagens known to date, from the organic extracts of diesel exhaust.
17 Grimmer et al. (1987) separated diesel exhaust particle extract into a water- and a lipid-
18 soluble fraction, and the latter was further separated into a PAH-free, a PAH-containing, and a
19 polar fraction by column chromatography. These fractions were then tested in Osborne-Mendel
20 rats by pulmonary implantation at doses corresponding to the composition of the original diesel
21 exhaust. The water-soluble fraction did not induce tumors; the incidences induced by the lipid-
22 soluble fractions were 0% with the PAH-free fraction, 25% with the PAH and nitro-PAH-
23 containing fractions, and 0% with the polar fraction. The PAH and nitro-PAH-containing
24 fraction, comprising only 1% by weight of the total extract, was thus shown to be responsible for
25 most, if riot all, of the carcinogenic activity.
26 Exposure of rats by inhalation to 2.6 mg/m3 of an aerosol of tar-pitch condensate with no
27 carbon core but containing 50 |ag/m3 benzo[a]pyrene along with other PAHs for 10 months
28 induced lung tumors in 39% of the animals. The same amount of tar-pitch vapor condensed onto
29 the surface of carbon black particles at 2 and 6 mg/m3 resulted in tumor rates that were roughly
30 two times higher (89% and 72%). Because exposure to 6 mg/m3 carbon black almost devoid of
31 extractable organic material induced a lung tumor rate of 18%, the combination of PAHs and
32 particles increases their effectiveness (Heinrich et al., 1994). Although this study shows the
33 tumor- inducing capability of PAHs resulting from combustion, it should be noted that the
34 benzoffl]pyrene content in the coal-tar pitch was about three orders of magnitude greater than in
35 diesel soot. Moreover, because organics are present on diesel particles in a thinner layer and the
7/25/00 7-98 DRAFT—DO NOT CITE OR QUOTE
-------
particles are quite convoluted, they may be more tightly bound and less bioavailable.
Nevertheless, these studies provide evidence supporting the involvement of organic constituents
of diesel particles hi the carcinogenic process.
4 Exposure of humans to related combustion emissions provides some evidence for the
5 involvement of organic components. Mumford et al. (1989) reported greatly increased human
6 lung cancer mortality in Chinese communes burning so-called smoky coal, but not wood, in
7 unvented open-pit fires used for heating and cooking. Although particle concentrations were
8 similar, PAH levels were five to six times greater in the air of communes burning smoky coal.
9 Coke oven emissions, containing high concentrations of PAHs but lacking an insoluble carbon
10 core, have also been shown to be carcinogenic in humans (Lloyd, 1971).
11 Adsorption of PAHs to a carrier particle such as hematite, CB, aluminum, or titanium
12 dioxide enhances their carcinogenic potency (Farrell and Davis, 1974). As already noted,
13 adsorption to carbon particles greatly enhanced the tumorigenicity of pyrolyzed pitch condensate
14 containing B[a]P and other aromatic carcinogens (Heinrich et al., 1995). The increased
15 effectiveness can be partly explained by more efficient transport to the deep lung. Slow release
16 also enhances residence time in the lungs and prevents overwhelming of activating pathways. As
17 discussed in Chapter 3, free organics are likely to be rapidly absorbed into the bloodstream,
18 which may explain why the vapor-phase component of exhaust is relatively ineffective in the
~^P induction of pathologic or carcinogenic effects.
20 Even though the organic constituents may be tightly bound to the particle surface,
21 significant elution is still likely because particle clearance half-times are nearly 1 year in humans
22 (Bohning et al., 1982). Furthermore, Gerde et al. (199la) presented a model demonstrating that
23 large aggregates of inert dust containing crystalline PAHs are unlikely to form at doses typical of
24 human exposure. This allows the particles to deposit and react with the surrounding lung
25 medium, without interference from other particles. Particle-associated PAHs can then be
26 expected to be released more rapidly from the particles. Bond et al. (1984) provided evidence
27 that alveolar macrophages from beagle dogs metabolized B[a]P coated on diesel particles to
28 proximate carcinogenic forms. Unless present on the particle surface, B[a]P is more likely to
29 pass directly into the bloodstream and escape activation by phagocytic cells.
30 The importance of DE-associated PAHs in the induction of lung cancer in humans may
31 be enhanced because of the possibility that the human lung is more sensitive to these compounds
32 than are rat lungs. Rosenkranz (1996) summarized information indicating that in humans and
33 mice, large proportions of lung cancers contain both mutated p5 3 suppressor genes and K-ras
34 genes. Induction of mutations in these genes by genotoxins, however, is much lower in rats than
in humans or mice.
7/25/00 7-99 DRAFT—DO NOT CITE OR QUOTE
-------
1 B[a]P, although only one of many PAHs present in diesel exhaust, is the one most
2 extensively studied. Bond et al. (1983, 1984) demonstrated metabolism of particle-associated
3 B[a]P and free B[a]P by alveolar macrophages (AM) and by type II alveolar cells. The
4 respiratory tract cytochrome P-450 systems have an even greater concentration in the nonciliated
5 bronchiolar cells (Boyd, 1984). It is worth noting that bronchiolar adenomas that develop
6 following diesel exposure have been found to resemble both Type II and nonciliated bronchiolar
7 cells. It should also be noted that any metabolism of procarcinogens by these latter two cell types
8 probably involves the preextraction of carcinogens in the extracellular lining fluid and/or other
9 endocytotic cells, as they are not especially important in phagocytosis of particles. Thus,
10 bioavailability is an important issue in assessing the relative importance of PAHs.
11 Additionally, a report by Borm et al. (1997) indicates that incubating rat lung epithelial-
12 derived cells with human polymorphonucleocytes (PMNs) (either unactivated or activated by
13 preexposure to phorbol myristate acetate) increases DNA adduct formation caused by exposure
14 to benzofa] pyrene; at 0.05 to 0.5 micromolar concentration, addition of more activated PMN in
15 relation to the number of lung cells further increased adduct formation in a dose-dependent
16 manner. The authors suggest that "an inflammatory response in the lung may increase the
17 biologically effective dose of PAHs, and may be relevant to data interpretation and risk
18 assessment of PAH-containing particles." These data raise the possibility that diesel exhaust
19 exposure at low concentrations may result in levels of neutrophil influx that would not
20 necessarily be detectable via histopathological examination as acute inflammation, but that might
21 be effective at amplifying any potential diesel exhaust genotoxic effect.
22 Nitro-PAHs have also been implicated as potentially involved in diesel-exhaust-induced
23 lung cancer. Although the nitro-PAH fraction of diesel was less effective than PAHs in the
24 induction of lung cancer when implanted into the lungs of rats (Grimmer et al., 1987), in a study
25 of various extracts of diesel exhaust particles, 30%-40% of the total mutagenicity could be
26 attributed to a group of six nitroarenes (Salmeen et al., 1984). Moreover, Gallagher et al. (1994)
27 reported results suggesting that DNA adducts are formed from nitro-PAHs present in DNA and
28 may play a role in the carcinogenic process. Nitroarenes, however, quantitatively represent a
29 very small percentage of diesel particle extract (Grimmer et al., 1987), making their role in the
^^ W V V«s«.AA^^ A. *.^*f*. A»* i. WhJ^-S **S •
31 The induction of DNA adducts in humans occupationally exposed to diesel exhaust
32 indicates the likelihood that PAHs are participating in the rumorigenic response, and that these
33 effects can occur at exposure levels less than those required to induce lung particle overload.
34 Distinct adduct patterns were found among garage workers occupationally exposed to diesel
35 exhaust when compared with nonexposed controls (Nielsen and Autrup, 1994). Furthermore, the
36 findings were concordant with the adduct patterns observed in groups exposed to low
7/25/00 7-100 DRAFT—DO NOT CITE OR QUOTE
-------
concentrations of PAHs from combustion processes. Hemminki et al. (1994) also reported
significantly elevated levels of DNA adducts in lymphocytes from garage workers with known
3 diesel exhaust exposure compared with unexposed mechanics. Hou et al. (1995) found elevated
4 adduct levels in bus maintenance workers exposed to diesel exhaust. Although no difference in
5 mutant frequency was observed between the groups, the adduct levels were significantly different
6 (3.2 vs. 2.3 x 10'8). Nielsen et al. (1996) measured three biomarkers in DE-exposed bus garage
7 workers: lymphocyte DNA adducts, hydroxyethylvaline adducts in hemoglobin, and
8 1-hydroxypyrene in urine. Significantly increased levels were reported for all three. Quetal.
9 (1996) detected increased adduct levels, as well as increases in some individual adducts, in the
10 blood of underground coal miners exposed to DE.
11
12 7.4.2. Role of Inflammatory Cytokines and Proteoiytic Enzymes in the Induction of Lung
13 Cancer in Rats by Diesel Exhaust
14 It is well recognized that the deposition of particles in the lung can result in the efflux of
15 PMNs from the vascular compartment into the alveolar space compartment in addition to
16 expanding the AM population size. Following acute exposures, the influx of the PMNs is
17 transient, lasting only a few days (Adamson and Bowden, 1978; Bowden and Adamson, 1978;
18 Lehnert et al., 1988). During chronic exposure the numbers of PMNs lavaged from the lungs of
I^F diesel-exposed rats generally increased with increasing exposure duration and inhaled DPM
20 concentration (Strom, 1984). Strom (1984) also found that PMNs in diesel-exposed lungs
21 remained persistently elevated for at least 4 months after cessation of exposure, a potential
22 mechanism that may be related to an ongoing release of phagocytized particles. Evidence in.
23 support of this possibility was reported by Lehnert et al. (1989) in a study in which rats were
24 intratracheally instilled with 0.85, 1.06, or 3.6 mg of polystyrene particles. The PMNs were not
25 found to be abnormally abundant during the clearance of the two lower lung burdens, but they
26 became progressively elevated in the lungs of the animals in which alveolar-phase clearance was
27 inhibited. Moreover, the particle burdens in the PMNs became progressively greater over time.
28 Such findings are consistent with an ongoing particle relapse process, in which particles released
29 by dying phagocytes are ingested by new ones.
30 The inflammatory response, characterized by efflux of PMNs from the vascular
31 compartment, is mediated by inflammatory chemokines. Driscoll et al. (1996) reported that
32 inhalation of high concentrations of carbon black stimulated the release of macrophage
33 inflammatory protein 2 (MIP-2) and monocyte chemotactic protein 1 (MCP-1). They also
34 reported a concomitant increase in hprt mutants. In a following study it was shown that particle
f exposure stimulates production of tumor necrosis factor TNF-a, an agent capable of activating
expression of several proteins that promote both adhesion of leucocytes and chemotaxis (Driscoll
7/25/00 7-101 DRAFT—DO NOT CITE OR QUOTE
-------
1 et al., 1997a). In addition, alveolar macrophages also have the ability to release several other
2 effector molecules or cytokines that can regulate numerous functions of other lung cells,
3 including their rates of proliferation (Bitterman et al., 1983; Jordana et al., 1988; Driscoll et al.,
4 1996).
5 Another characteristic of AMs and PMNs under particle overload conditions is the release
6 of a variety of potentially destructive hydrolytic enzymes, a process known to occur
7 simultaneously with the phagocytosis of particles (Sandusky et al., 1977). The essentially
8 continual release of such enzymes during chronic particle deposition and phagocytosis in the lung
9 may be detrimental to the alveolar epithelium, especially to Type I cells. Evans et al. (1986)
10 showed that injury to Type I cells is followed shortly thereafter by a proliferation of Type II cells.
11 Type II cell hyperplasia is a common feature observed in animals that have received high lung
12 burdens of various types of particles, including unreactive polystyrene microspheres.
13 Exaggerated proliferation as a repair or defensive response to DPM deposition may have the
14 effect of amplifying the likelihood of neoplastic transformation in the presence of carcinogens
15 beyond that which would normally occur with lower rates of proliferation, assuming an increase
16 in the cycling of target cells and the probability of a neoplastic-associated genomic disturbance.
17
18 7.4.3. Role of Reactive Oxygen Species in Lung Cancer Induction by Diesel Exhaust
19 Phagocytes from a variety of rodent species produce elevated levels of oxidant reactants
20 in response to challenges, with the physiochemical characteristics of a phagocytized particle
21 being a major factor in determining the magnitude of the oxidant-producing response. Active
22 oxygen species released by the macrophages and lymphatic cells can cause lipid peroxidation in
23 the membrane of lung epithelial cells. These lipid peroxidation products can initiate a cascade of
24 oxygen free radicals that progress through the cell to the nucleus, where they damage DNA. If
25 this damage occurs during the epithelial cell's period of DNA synthesis, there is some probability
26 that the DNA will be replicated unrepaired (Lechner and Mauderly, 1994). The generation of
27 reactive oxygen species by both AMs and PMNs should therefore be considered as one potential
28 factor of what probably is a multistep process that culminates in the development of lung tumors
29 in response to chronic deposition of DPM.
30 Even though products of phagocytic oxidative metabolism, including superoxidc anions,
31 hydrogen peroxide, and hydroxyl radicals, can kill tumor cells (Klebanoff and Clark, 1978), and
32 the reactive oxygen species can peroxidize lipids to produce cytotoxic metabolites such as
33 malonyldialdehyde, some products of oxidative metabolism apparently can also interact with
34 DNA to produce mutations. Cellular DNA is damaged by oxygen free radicals generated from a
35 variety of sources (Ames, 1983; Trotter, 1980). Along this line, Weitzman and Stossel (1981)
36 found that human peripheral leukocytes are mutagenic in the Ames assay. This mutagenic
7/25/00 7-102 DRAFT—DO NOT CITE OR QUOTE
-------
activity was related to PMNs and blood monocytes; blood lymphocytes alone were not
mutagenic. These investigators speculated that the mutagenic activity of the phagocytes was a
result of their ability to produce reactive oxygen metabolites, inasmuch as blood leukocytes from
4 a patient with chronic granulomatous diseases, in which neutrophils have a defect in the
5 NADPH oxidase generating system (Klebanoff and Clark, 1978), were less effective in
6 producing mutations than were normal leukocytes. Of related significance, Phillips et al. (1984)
7 demonstrated that the incubation of Chinese hamster ovary cells with xanthine plus xanthine
8 oxidase (a system for enzymatically generating active oxygen species) resulted in genetic
9 damage hallmarked by extensive chromosomal breakage and sister chromatid exchange and
10 produced an increase in the frequency of thioguanidine-resistant cells (HGPRT test). Aside from
11 interactions of oxygen species with DNA, increasing evidence also points to an important role of
12 phagocyte-derived oxidants and/or oxidant products in the metabolic activation of
13 procarcinogens to their ultimate carcinogenic form (Kensler et al., 1987).
14 Driscoll et al. (1997b) have demonstrated that exposure to doses of particles producing
15 significant neutrophilic inflammation are associated with increased mutation in rat alveolar type
16 II cells. The ability of particle-elicited macrophages and neutrophils to exert a mutagenic effect
17 on epithelial cells in vitro supports a role for these inflammatory cells for the in vivo mutagenic
« effects of particle exposure. The inhibition of bronchoalveolar lavage cell-induced mutations by
catalase implies a role for cell-derived oxidants in this response.
20 Hatch and co-workers (1980) have demonstrated that interactions of guinea pig AMs with
21 a wide variety of particles, such as silica, metal oxide-coated fly ash, polymethylmethacrylate
22 beads, chrysotile asbestos, fugitive dusts, polybead carboxylate microspheres, glass and latex
23 beads, uncoated fly ash, and fiberglass increase the production of reactive oxygen species.
24 Similar findings have been reported by numerous investigators for human, rabbit, mouse, and
25 guinea pig AMs (Drath and Karnovsky, 1975; Allen and Loose, 1976; Beall et al., 1977; Lowrie
26 and Aber, 1977; Miles et al., 1977; Rister and Baehner, 1977; Hoidal et al., 1978). PMNs are
27 also known to increase production of superoxide radicals, hydrogen peroxide, and hydroxyl
28 radicals in response to membrane-reactive agents and particles (Goldstein et al., 1975; Weiss et
29 al., 1978; Root and Metcalf, 1977). Although these responses may occur at any concentration,
30 they are likely to be greatly enhanced at high exposure concentrations with slowed clearance and
31 lung particle overload.
32 Reactive oxygen species can also be generated from particle-associated organics. Sagai
33 et al. (1993) reported that DPM can nonenzymatically generate active oxygen species (e.g.,
34 superoxide [O2~] and hydroxyl radical [OH] in vitro without any biologically activating systems)
such as microsomes, macrophages, hydrogen peroxide, or cysteine. Because DPM washed with
7/25/00 7-103 DRAFT—DO NOT CITE OR QUOTE
-------
1 methanol could no longer produce these radicals, it was concluded that the active components
2 were compounds extractable with organic solvents. However, the nonenzymatic contribution to
3 the DPM-promoted active oxygen production was negligible compared with that generated via an
4 enzymatic route (Ichinose et al., 1997a). They reported that O2~ and OH can be enzymatically
5 generated from DPM by the following process. Soot-associated quinone-like compounds are
6 reduced to the semiquinone radical by cytochrome P-450 reductase. These semiquinone radicals
7 then reduce O2 to O2~, and the produced superoxide reduces ferric ions to ferrous ions, which
8 catalyzes the homobiotic cleavage of H2O2 dismutated from O2 by superoxide dismutase or
9 spontaneous reactions to produce OH. According to Kumagai et al. (1997), while quinones are
10 likely to be the favored substrates for this reaction, the participation of nitroaromatics cannot be
11 ruled out.
12 One of the critical lesions to DNA bases generated by oxygen free radicals is 8-
13 hydroxydeoxyguanosine (8-OHdG). The accumulation of 8-OhdG as a marker of oxidative
14 DNA damage could be an important factor in enhancing the mutation rate leading to lung cancer
15 (Ichinose et al., 1997a). For example, formation of 8-OHdG adducts leads to G:C to T: A
16 transversions unless repaired prior to replication. Nagashima et al. (1995) demonstrated that the
17 production of (8-OHdG) is induced in mouse lungs by intratracheal instillation of DPM.
18 Ichinose et al. (1997b) reported further that although intratracheal instillation of DPM in mice
19 induced a significant increase in lung tumor incidence, comparable increases were not reported
20 when mice were instilled with extracted DPM (to remove organics). Lung injury was also less in
21 the mice instilled with extracted DPM. Moreover, increases in 8-OHdG in the mice instilled
22 with unextracted DPM correlated very well with increases in tumor rates. In a related study,
23 Ichinose et al. (1997a) intratracheally instilled small doses of DPM, 0.05, 0.1, or 0.2 mg weekly
24 for 3 weeks, in mice fed standard or high-fat diets either with or without p-carotene. High
25 dietary fat enhanced DPM-induced lung tumor incidence, whereas p-carotene, which may act as a
26 free radical scavenger, partially reduced the tumorigenic response. Formation of 8-OHdG was
27 again significantly correlated with lung tumor incidence in these studies, except at the highest
28 dose. Dasenbrock et al. (1996) reported that extracted DPM, intratracheally instilled into rats (15
29 mg total dose) induced only marginal increases in lung tumor induction, while unextracted DPM
30 was considerably more effective. Although adducts were not measured in this study, it
31 nevertheless provides support for the likelihood that activation of organic metabolites and/or
32 generation of oxygen free radicals from organics are involved in the carcinogenic process.
33 Additional support for the involvement of particle-associated radicals in tissue damage
34 was provided by the finding mat preireatment with superoxide dismutase (SOD), an antioxidant,
35 markedly ieduced lung injury and death due to instillation of DPM. Similarly, Hirafuji et al.
7/25/00 7-104 DRAFT—DO NOT CITE OR QUOTE
-------
(1995) found that the antioxidants catalase, deferoxamine, and MK-447 inhibited the toxic
effects of DPM on guinea pig tracheal cells and tissues in vitro.
Although the data presented supported the hypothesis that generation of reactive oxygen
4 species resulting from exposure to DPM is involved in the carcinogenic process, it should be
5 noted that 8-OHdG is efficiently repaired and that definitive proof of a causal relationship in
6 humans is still lacking. It is also uncertain whether superoxide or hydroxyl radicals chemically
7 generated by DPM alone promote 8-OHdG production in vivo and induce lung toxicity, because
8 SOD is extensively located in mammalian tissues. Nevertheless, demonstration that oxygen free
9 radicals can be generated from particle-associated organics, that their presence will induce adduct
10 formation and DNA damage unless repaired, that tumor induction in experimental animals
11 correlates with OhdG adducts, and that treatment with antioxidant limits lung damage, provides
12 strong support for the involvement of oxygen free radicals in the toxicologic and carcinogenic
13 response to diesel exhaust.
14
15 7.4.4. Relationship of Physical Characteristics of Particles to Cancer Induction
16 The biological potential of inhaled particles is strongly influenced by surface chemistry
17 and character. For example, the presence of trace metal compounds such as aluminum and iron,
as well as ionized or protonated sites, is important in this regard (Langer and Nolan, 1994). A
major factor is specific surface area (surface area/mg). PMNs characteristically are increased
20 abnormally in the lung by diesel exhaust exposure, but their presence in the lungs does not
21 appear to be excessive following the pulmonary deposition of even high lung burdens of
22 spherical TiO2 particles in the 1-2 p.m diameter range (Strom, 1984). In these studies lung
23 tumors were detected only at an inhaled concentration of 250 M-g/m3. In a more recent study in
24 which rats were exposed to TiO2 in the 15-40 nm size range, inhibition of particle clearance and
25 tumorigenesis were induced at concentrations of 10 mg/m3 (Heinrich et al., 1995). Comparison
26 of several chronic inhalation studies correlating particle mass and particle surface area retained in
27 the lung with tumor incidence indicated that particle surface area is a much better dosimeter than
28 paticle mass (Oberdorster and Yu, 1990; Driscoll et al., 1996). Heinrich et al. (1995) also found
29 that lung tumor rates increased with specific particle surface area following exposure to diesel
30 exhaust, carbon black, or titanium dioxide, irrespective of particle type. Langer and Nolan
31 (1994) reported that the hemolytic potential of Min-U-Sill 5, a silica flour, increased in direct
32 relationship to specific surface area at nominal particle diameters ranging from 0.5 to 20 \im.
33 Ultrafine particles appear to be more likely to be taken up by lung epithelial cells. Riebe-
34 Imre et al. (1994) reported that CB is taken up by lung epithelial cells in vitro, inducing
chromosomal damage and disruption of the cytoskeleton, lesions that closely resemble those
7/25/00 7-105 DRAFT—DO NOT CITE OR QUOTE
-------
1 present in tumor cells. Johnson et al. (1993) reported that 20-nm polytetrafluoroethylene
2 particles are taken up by pulmonary epithelial cells as well as polymorphonuclear leucocytes,
3 inducing an approximate 4-, 8-, and 40-fold increase in the release of interleukin-1 alpha and
4 beta, inducible nitric oxide synthetase, and macrophage inflammatory protein, respectively.
5 The carcinogenic potency of diesel particles, therefore, appears to be related, at least to
6 some extent, to their small size and convoluted shape, which results in a large specific particle
7 surface area. Toxicity and carcinogenicity increased with increasing particle size into the
8 submicron range. For example, Heinrich et al. (1995) have shown that ultrafine titanium dioxide
9 (approximately 0.2 p.m diameter) is much more toxic than particles with a 10-fold greater
10 diameter of the same composition used in an earlier study by Lee et al. (1986). This increase in
11 toxicity has been noted with even smaller particles. For example, carbon black particles 20 nm
12 in diameter were shown to be significantly more toxic than 50 nm particles (Murphy et al.,
13 1999). The relationship between particle size and toxicity is of concern because, as noted in
14 Chapter 2, approximately 50%-90% of the number of particles in diesel exhaust are in the size
15 range from 5 to 50 nm. Other than disruption of the cytoskeleton of epithelial cells, there is little
16 information regarding the means by which particle size influences carcinogenicity as well as
17 noncancer toxicity.
18
19 7.4.5. Integrative Hypothesis for Diesel-Induced Lung Cancer
20 The induction of lung cancer by large doses of carbon black via inhalation (Heinrich et
21 al., 1995; Mauderly et al., 1991; Nikula et al., 1995) or intratracheal instillation (Kawabata et al.,
22 1994; Pott et al., 1994; Dasenbrock et al., 1996) led to the development of the lung particle
23 overload hypothesis. According to this hypothesis the induction of neoplasia by insoluble low-
24 toxicity particles is associated with an inhibition of lung particle clearance and the involvement
25 of persistent alveolar epithelial hyperplasia. Driscoll (1995), Driscoll et al. (1996), and
26 Oberdorster and Yu (1990) outlined a proposed mechanism for the carcinogenicity of diesel
27 exhaust at high doses that emphasizes the role of phagocytic cells. Following exposure,
28 phagocytosis of particles acts as a stimulant for oxidant production and inflammatory cytokine
29 release by lung phagocytes. It was hypothesized that at high particle exposure concentrations the
30 quantity of mediators released by particle-stimulated phagocytes exceeds the inflammatory
31 defenses of the lung (e.g., antioxidants, oxidant-metabolizing enzymes, protease inhibitors,
32 cytokine inhibitors), resulting in tissue injury and inflammation. With continued particle
33 exposure and/or the persistence of excessive particle burdens, there then develops an
34 environment cf phagccytic activation, excessive nicuiatur release-Lissue injury and,
35 consequently, more tissue injury, infiaiumatiun, ami tissue reiease. This is accompanied by cell
7/2 5/00 7-106 DRAFT—DO NOT CITE OR QUOTE
-------
1 proliferation. As discussed in a review by Cohen and Ellwein (1991), conceptually, cell
f proliferation can increase the likelihood that any oxidant-induced or spontaneously occurring
genetic damage becomes fixed in a dividing cell and is clonally expanded. The net result of
4 chronic particle exposures sufficient to elicit inflammation and cell proliferation in the rat lung is
5 an increased probability that the genetic changes necessary for neoplastic transformation will
6 occur. A schematic of this hypothesis has been outlined by McClellan (1997) (see Figure 7-3).
7 In support of this hypothesis, it was reported that concentrations of inhaled CB resulted in
8 increased cytokine expression and inflammatory influx of neutrophils (Oberdorster et al., 1995),
9 increased formation of 8-OhdG (Ichinose et al., 1997b), and increase in the yield of hprt mutants,
10 an effect ameliorated by treatment with antioxidants (Driscoll, 1995; Driscoll et al., 1996).
11 Metabolism of carcinogenic organics to active forms as well as the generation of reactive oxygen
12 species from certain organic species are likely to contribute to the toxic and carcinogenic process.
13 At low concentrations, inflammatory effects associated with lung particle overload are
14 generally absent. However, activation of organic carcinogens and generation of oxidants from
15 the organic fraction can still be expected. Actual contribution depends upon elution and the
16 effectiveness of antioxidants. Direct effects of ultrafine diesel particles taken up by epithelial
17 cells are also likely to play a role.
Although high-dose induction of cancer is logically explained by this hypothesis, particle
overload has not been clearly shown to induce lung cancer in other species. As noted in the
20 quantitative chapter, the relevance of the rat pulmonary response is therefore problematic. The
21 rat pulmonary noncancer responses to DPM, however, have fairly clear interspecies and human
22 parallels. In response to poorly soluble particles such as DPM, humans and rats both develop an
23 alveolar macrophage response, accumulate particles in the interstitium, and show mild interstitial
24 fibrosis (ILSI, 2000). Other species (mice, hamsters) also have shown similar noncancer
25 pulmonary responses to DPM, but without accompanying cancer response. The rat response for
26 noncancer pulmonary histopathology, however, seems to be more pronounced compared with
27 humans or other species, i.e., rats appear to be more sensitive. Although many critical elements
28 of interspecies comparison, such as the role of airway geometry and patterns of particle
29 deposition, need further elucidation, this basic interspecies similarity and greater sensitivity of
30 pulmonary response seen after longer exposures at high doses make pulmonary histopathology in
31 rats a valid basis for noncancer dose-response assessment.
32
33 7.4.6. Summary
34 Recent studies have shown tumor rates resulting from exposures to nearly organic-free
CB particles at high concentrations to be similar to those observed for diesel exhaust exposures,
7/25/00 7-107 DRAFT—DO NOT CITE OR QUOTE
-------
1 thus providing strong evidence for a particle overload mechanism for DE-induced pulmonary
2 carcinogenesis in rats. Such a mechanism is also supported by the fact that carbon particles per
3 se cause inflammatory responses and increased epithelial cell proliferation and that AM function
4 may be compromised under conditions of particle overload.
5 The particle overload hypothesis appears sufficient to account for DE-induced lung
6 cancer in rats. However, there is increasing evidence for lung cancer induction in humans at
7 concentrations insufficient to induce lung particle overload as seen in rats (Section 3.7 and ILSI
8 2000). Uptake of particles by epithelial cells at ambient or occupational exposure levels, DNA
9 damage resulting from oxygen-free radicals generated from organic molecules, and the gradual in
10 situ extraction and activation of procarcinogens associated with the diesel particles are likely to
11 play a role in this response. The slower particle clearance rates in humans (up to a year or more)
12 may result in greater extraction of organics. This is supported by reports of increased DNA
13 adducts in humans occupationally exposed to diesel exhaust at concentrations unlikely to induce
14 lung particle overload. Although these modes of action can be expected to function at lung
15 overload conditions also, they are likely to be overwhelmed by inflammatory associated effects.
16 The evidence to date indicates that caution must be exercised in extrapolating
17 observations made in animal models to humans when assessing the potential for DE-induced
18 pulmonary carcinogenesis. The carcinogenic response and the formation of DNA adducts in rats
19 exposed to diesel exhaust and other particles at high exposure concentrations may be species-
20 specific and not particle-specific. The likelihood that different modes of action predominate at
21 high and low doses also renders low-dose extrapolation to ambient concentrations uncertain.
22
23 7.5. WEIGHT-OF-EVIDENCE EVALUATION FOR POTENTIAL HUMAN
24 CARCINOGENICITY
25 A weight-of-evidence evaluation is a synthesis of all pertinent information addressing the
26 question of how likely an agent is to be a human carcinogen. EPA's 1986 Guidelines for
27 Carcinogen Risk Assessment (U.S. EPA, 1986) provide a classification system for the
28 characterization of the overall weight of evidence for potential human carcinoge'riicity based on
29 human evidence, animal evidence, and other supportive data. This system includes Group A:
30 Human Carcinogen'., Group B: Probable Human Carcinogen; Group C: Possible Human
31 Carcinogen; Group D: Not Classifiable as to Human Carcinogenicity; and Group E: Evidence
32 for Noncarcinogenicity to Humans.
33 As part of the guidelines development and updating process, the Agency has developed
34 revisions tc the 1986 guidelines to take into account knowledge gained in recent years about the
35 carcinogenic processes. \Vith rcgaiu to llic wcighi-of-evidence evaluation for potential human
7/25/00 7-108 DRAFT—DO NOT CITE OR QUOTE
-------
carcinogenicity, EPA's 1996 Proposed Guidelines for Carcinogen Risk Assessment (U.S. EPA,
1996b) and the subsequent revised external review draft (U.S. EPA, 1999) emphasize the need
for characterizing cancer hazard, in addition to hazard identification. Accordingly, the question
4 to be addressed in hazard characterization is expanded to how likely an agent is to be a human
5 carcinogen, and under what exposure conditions a cancer hazard may be expressed. The revised
6 guidelines also stress the importance of considering the mode(s) of action information for
7 making an inference about potential cancer hazard beyond the range of observation, typically
8 encountered at levels of exposure in the general environment. "Mode of action" refers to a series
9 of key biological events and processes that are critical to the development of cancer. This is
10 contrasted with "mechanisms of action," which is defined as a more detailed description of the
11 complete sequence of biological events at the molecular level that must occur to produce a
12 carcinogenic response.
13 To express the weight of evidence for potential human carcinogenicity, EPA's proposed
14 guidelines utilize a hazard narrative in place of the classification system. However, in order to
15 provide some measure of consistency, standard hazard descriptors are used as part of the hazard
16 narrative to express the conclusion regarding the weight of evidence for potential human
17 carcinogenicity.
The sections to follow evaluate and weigh the individual lines of evidence and combine
all evidence to make an informed judgement about the potential human carcinogenicity of DE. A
20 conclusion in accordance with EPA's 1986 classification system (U.S. EPA, 1986) is provided,
21 as well as a hazard narrative along with appropriate hazard descriptors according to EPA's
22 Proposed Revised Guidelines (U.S. EPA, 1996b, 1999). These sections draw on information
23 reviewed in Chapters 2, 3, 4, and 7.
24
25 7.5.1. Human Evidence
26 Twenty-two epidemiologic studies about the carcinogenicity of workers exposed to DE
27 in various occupations are reviewed in Section 7.2. Exposure to DE has typically been inferred
28 based on job classification within an industry. Increased lung cancer risk, although not always
29 statistically significant, has been observed in 8 out of 10 cohort and 10 of 12 case-control studies
30 within several industries, including railroad workers, truck drivers, heavy equipment operators,
31 and professional drivers. The increased lung cancer relative risks generally range from 1.2 to
32 1.5, though a few studies show relative risks as high as 2.6. Statistically significant increases in
33 pooled relative risk estimates (1.33 to 1.47) from two independent meta-analyses further support
34^ a positive relationship between DE exposure and lung cancer in a variety of DE-exposed
occupations.
7/25/00 7-109 DRAFT—DO NOT CITE OR QUOTE
-------
1 The generally small increased lung cancer relative risk (less than 2) observed in these
2 analyses potentially weakens the evidence of causality. When a relative risk is less than 2, if
3 confounders (e.g.;, smoking, asbestos exposure) are having an effect on the observed risk
4 increases, it could be enough to account for the increased risk. With the strongest risk factor for
5 lung cancer being smoking, there is a concern that smoking effects may be influencing the
6 magnitude of the observed increased relative risks. However, in studies for which the effects of
7 smoking were accounted for, increased relative risks for lung cancer prevailed. Though some
8 studies did not have information on smoking, confounding by smoking is unlikely in these
9 studies because the comparison population was from the same socioeconomic class. Moreover,
10 when the meta-analysis focused only on the smoking-controlled studies, the relative risks tended
11 to increase.
12 As evaluated in Chapter 7 (Section 7.2.4.5), application of the criteria for causality
13 provides evidence that the increased risks observed in available epidemiologic studies are
14 consistent with a causal association between exposure to DE and occurrence of lung cancer.
15 ' Overall, the human evidence for potential carcinogenicity for DE is judged to be strong, but less
16 than sufficient for DE to be considered as a human carcinogen, because of exposure uncertainties
17 (lack of historical exposure of workers to DE) and an inability to satisfactorily account for all
18 confounders.
19
20 7.5.2. Animal Evidence
21 DE and its organic constituents, both in the gaseous and particle phase, have been
22 extensively tested for carcinogenicity in many experimental studies using several animal species
23 and with different modes of administration. Several well-conducted studies have consistently
24 demonstrated that chronic inhalation exposure to sufficiently high concentrations of DE
25 produced dose-related increases in lung tumors (benign and malignant) in rats. In contrast,
26 chronic inhalation studies of DE in mice showed mixed results whereas negative findings were
27 consistently seen in hamsters. The gaseous phase of DE (filtered exhaust without particulate
28 fraction), however, was found not to be carcinogenic in rats, mice, or hamsters.
29 In several intratracheal instillation studies, diesel particulate matter (DPM), DPM
30 extracts, and carbon black, which was virtually devoid of PAHs, have been found to produce
31 increased lung tumors in rats. When directly implanted into the rat lung, DPM condensate
32 containing mainly four- to seven-ring PAHs induced increases in lung tumors. In several dermal
33 studies in mice, DPM extracts have also been shown to cause skin tumors and sarcomas in mice
34 renewing SuL/CutuiicGus injection.
7/25/00 7-110 DPvAFT—DO NOT CITE OR QUOTE
-------
Overall, there is sufficient evidence for the potential carcinogenicity of whole DE in the
rat at high exposure concentration or administered dose, both by inhalation and intratracheal
instillation. Available data indicate that both the carbon core and the adsorbed organics have
4 potential roles in inducing lung tumors hi the rat, although their relative contribution to the
5 carcinogenic response remains to be determined. The gaseous phase of DE, however, does not
6 appear to have any significant role in DE-induced lung cancer response hi the rat.
7 Available data also indicate that among the traditional animal test species, the rat is the
8 most sensitive species to DE. As reviewed in Section 7.4, the lung cancer responses in rats from
9 high-concentration exposures to DE appear to be mediated by impairment of lung clearance
10 mechanisms through particle overload, resulting in persistent chronic inflammation and
11 subsequent pathologic and neoplastic changes in the lung. Overload conditions are not expected
12 to occur in humans as a result of environmental or most occupational exposures to DE. Thus, the
13 animal evidence (i.e., increased lung tumors in the rat) provides additional support for identifying
14 potential cancer hazard to humans, but is not considered suitable for dose-response analysis and
15 estimation of human risk to DE.
16 The consistent findings of carcinogenic activity by the organic extracts of DPM in
17 noninhalation studies (intratracheal instillation, lung implantation, skin painting) further
contribute to the overall animal evidence for a human hazard potential for DE.
20 7.5.3. Other Key Data
21 Other key data, although not as extensive as the human and animal carcinogenicity data,
22 are judged to be supportive of potential carcinogenicity of DE. As discussed in Chapter 2, DE is
23 a complex mixture of hundreds of constituents in either gaseous phase or particle phase.
24 Although present in small amounts, several organic compounds in the gaseous phase (e.g. PAHs,
25 formaldehyde, acetaldehyde, benzene, 1,3-butadiene) are known to exhibit mutagenic and/or
26 carcinogenic activities. PAHs and PAH derivatives, including nitro-PAHs, present on the diesel
27 particle are also known to be mutagenic and carcinogenic. As reviewed in Chapter 4, DPM and
28 DPM organic extracts have been shown to induce gene mutations in a variety of bacteria and
29 mammalian cell test systems. In addition, DE, DPM and DPM extracts have been found to cause
30 chromosomal aberrations, aneuploidy, and sister chromatid exchange in both in vivo and in vitro
31 tests.
32 There is also suggestive evidence for the bioavailability of the organics from DE (Chapter
33 3). Elevated levels of DNA adducts in lymphocytes have been reported in workers exposed to
34^ DE. In addition, animal studies showed that some of the radiolabeled organic compounds are
eluted from DE particles following deposition in the lungs (Section 3.6).
7/25/00 7-111 DRAFT—DO NOT CITE OR QUOTE
-------
1 7.5.4. Mode of Action
2 As discussed in Section 7.4, the modes of action of DE-induced carcinogenicity in
3 humans are not well understood. It is likely that multiple modes of action are involved. These
4 may include: (a) mutagenic and genotoxic events (e.g., direct and indirect effects on DNA and
5 effects on chromosomes) by organic compounds in the gaseous and particle phases; (b) indirect
6 DNA damage via the production of reactive oxygen species (ROS) induced by particle-
7 associated organics; and (c) particle-induced chronic inflammatory response leading to oxidative
8 DNA damage through the release of cytokines, ROS, etc., and an increase in cell proliferation.
9 The particulate phase appears to have the greatest contribution to the carcinogenic effects,
10 and both the particle core and the associated organic compounds have demonstrated carcinogenic
11 properties, although a role for the gas-phase components cannot be ruled out. The carcinogenic
12 activity of DE also appears to be related to the small size of the particles. Moreover, the relative
13 contribution of the various modes of action may be different at different exposure levels.
14 Available evidence from animal studies indicates the importance of the role of the DE particles
15 in mediating lung tumor response at high exposure levels. Thus, the role of the adsorbed organic
16 compounds may take on increasing importance at lower exposure levels.
17
18 7.5.5. Characterization of Overall Weight of Evidence: EPA's 1986 Carcinogen Risk
19 Assessment Guidelines
20 The totality of evidence supports the conclusion that DE is a. probable hitman carcinogen
21 (Group Bl). This conclusion is based on:
22
23 • Limited human evidence (less than sufficient) for a causal association between DE
24 exposure and increased lung cancer risk among workers of different occupations;
25 • Sufficient animal evidence for the induction of lung cancer in the rat from inhalation
26 exposure to high concentrations of DE, DPM, and the carbon core; and supporting
27 evidence of carcinogenicity of DPM and the associated organics in rats and mice by
28 noninhalation route of exposure; and
29 • Extensive supporting data including the demonstrated mutagenic and/oi chromosom?"
30 effects of DE and its organic constituents, suggestive evidence for the bioavailability
31 of the organics from DE, and the known mutagenic and/or carcinogenic activity of a
32 number of individual organic compounds present on the particles and in the gaseous
33 phase.
7/25/00 7-112 DRAFT—DO NOT CITE OR QUOTH
-------
1 7.5.6. Weight-of-Evidence Hazard Narrative: EPA's Proposed Revised Carcinogen Risk
f Assessment Guidelines (1996b, 1999)
The combined evidence supports the conclusion that DE is likely to be carcinogenic to
4 humans by inhalation exposure at any exposure condition. In comparison with other agents
5 designated as likely to be carcinogenic to humans, the weight of evidence for DE is at the upper
6 end of the spectrum. The weight of evidence of human carcinogenicity is based on:
7
8 • Strong but less than sufficient epidemiologic evidence for a causal association between
9 occupational exposure and elevated risk of lung cancer;
10 • Consistent evidence of increases of lung tumors in rats from chronic inhalation
11 exposure to high concentration of whole DE, DPM, or the particle elemental carbon
12 core;
13 • Supportive evidence of carcinogenicity in rats for the diesel particle (DPM) via
14 intratracheal instillation, and for DPM organic extracts in rats and mice in
15 noninhalation studies (intratracheal instillation, lung implantation, skin painting,
16 subcutaneous injection);
17 • Extensive evidence of mutagenic and chromosomal effects of DE and its organic
^L constituents;
^^ • Suggestive evidence of the bioavailability of the DPM organics in studies of humans
20 and animals; and
21 • The presence of a number of individual organic compounds on the diesel particles
22 (e.g., PAHs and derivatives) and in the gaseous phase (e.g., benzene, acetaldehydes)
23 that are known to exhibit mutagenic and/or carcinogenic properties.
24
25 A major uncertainty in characterizing the potential cancer hazard for DE at low levels of
26 environmental exposure is the incomplete understanding of its mode of action for the induction
27 of lung cancer in humans. Nonetheless, available data indicate that DE-induced lung
28 carcinogenicity seems to be mediated by mutagenic and nonmutagenic events by both the
29 particles and the associated organic compounds, although a role for the organics in the gaseous
30 phase cannot be ruled out. Given that there is some evidence for a mutagenic mode of action, a
31 cancer hazard is presumed at any exposure level. This is consistent with EPA's science policy
32 position, which assumes a nonthreshold effect for carcinogens in the absence of definitive data
33 demonstrating a nonlinear or threshold mechanism. Accordingly, linear low-dose extrapolation
34^ should be assumed in dose-response assessment. Because of insufficient information, the human
carcinogenic potential of DE by oral and dermal exposures cannot be determined.
7/25/00 7-113 DRAFT—DO NOT CITE OR QUOTE
-------
1 7.6. EVALUATIONS BY OTHER ORGANIZATIONS
2 Several organizations have reviewed the relevant data and evaluated the potential human
3 carcinogenicity of DE or its paniculate component. The conclusions reached by these
4 organizations are generally comparable to the evaluation made in this assessment using EPA's
5 Carcinogen Risk Assessment Guidelines. A summary of available evaluations conducted by
6 other organizations is provided in Table 7-9.
7
8 7.7. CONCLUSION
9 It is concluded that environmental exposure to DE may present a cancer hazard to
10 humans. The particulate phase appears to have the greatest contribution to the carcinogenic
11 effects, and both the particle core and the associated organic compounds have demonstrated
12 carcinogenic properties, although a role for the gas-phase components cannot be ruled out.
13 Using either EPA's 1986 Carcinogen Risk Assessment Guidelines (U.S. EPA, 1986) or the
14 proposed revisions (U.S. EPA, 1996b, 1999), DE is judged to be a probable human carcinogen,
15 or likely to be carcinogenic to humans by inhalation, respectively. The weight of evidence for
16 potential human carcinogenicity for DE is considered strong, even though inferences are
17 involved in the overall assessment. Major uncertainties of the hazard assessment include the
18 following unresolved issues:
19 • There has been a considerable scientific debate about the significance of the available
20 human evidence for a causal association between occupational exposure and increased
21 lung cancer risk. Many experts view the evidence as weak while many others consider
22 the evidence as strong. This is due to a lack of consensus about whether the effects of
23 smoking have been adequately accounted for in key studies, and the lack of historical
24 DE exposure data for the available studies.
25 • Although the mode of action for DE-induced lung tumors in rats from high exposure is
26 sufficiently understood, the mode of action for lung cancer risk in humans is not fully
27 known. To date, available evidence for the role of both the adsorbed organics and the
28 carbon core particle has been shown to be associated with high exposure conditions.
29 There is virtually no information about the relative role of DE constituents in
30 mediating carcinogenic effects at the low exposure levels. Furthermore, there is only a
31 limited understanding regarding the relationship between particle size and
32 • carcinogenicity.
33 • DE is present in ambient PM (e.g., PM2 5 or PMIO); however, a cancer hazard for
34 ambient FM has not been clearly identified.
7/2 5/00 7-114 DRAFT—DO NOT CITE OR QUOTE
-------
1 Additional research is needed to address these issues to reduce the uncertainty associated with
the potential cancer hazard of exposure to DE.
7/25/00 7-115 DRAFT—DO NOT CITE OR QUOTE
-------
NJ
[J\
o
o
Table 7-1. Epidcmiologic studies of the health effects of exposure to diesel exhaust: cohort mortality studies
Authors
Population studied
Diesel exhaust exposure
assessment
Results
Limitations.
Waller Approximately 20,000 male
Five job categories used to SMR = 79 for lung cancer for the Exposure measurement of
London transportation workers define exposure
Aged 45 to 64 years
25 years follow-up (1950-1974)
total cohort benzo[a]pyrene showed very little
difference between inside and outside
SMRs for all five job categories the garage
Environmental
benzo[a]pyrene were less than 100 for lung
concentrations measured in cancer
1957 and 1979
Incomplete information on cohort
members
No adjustment for confounding such
as other exposures, cigarette smoking,
etc.
No latency analysis
H
a
o
z
o
H
O
KH
H
W
O
Howe ct al. 43,826 male pensioners of the
(1983) Canadian National Railway
Company
Mortality between 1965 and
1977 among these pensioners
was compared with mortality
of general Canadian population
Exposure groups
classified by a group
of experts based on
occupation at the time
of retirement
Three exposure groups:
Nonexposed
Possibly exposed
Probably exposed
RR=1.2(p=0.013)and
RR= 1.3(p=0.001)forlung
cancer for possible and probable
exposure, respectively
A highly significant
dose-response relationship
demonstrated by trend
test (p<0.001)
Incomplete exposure assessment due
to lack of lifetime occupational
history
Mixed exposures to coal
dust/combustion products and diesel
exhaust
No validation of method was used to
categorize exposure
Lack of data on smoking but use of
internal comparison group to compute
RRs minimizes the potential
confounding by smoking
No latency analysis
a
o
H
r-i
-------
K)
U<
O
o
B'ablc 7-1. Epidcmiologic studies of the health effects o
(continued)
osure to diesel exhaust: cohort mortality studies
Authors
Population studied
Diesel exhaust exposure
assessment
Results
Limitations
Rushton 8,490 male London transport
et al. (1983) maintenance workers
100 different job titles were
grouped in
20 broad categories
Mortality of workers employed for
1 continuous year between January The categories were not
1, 1967, and December 31, 1975, ranked for diesel exhaust
was compared with mortality of exposure
general population of England and
Wales
SMR=133(/><0.03)forlung
cancer in the general hand job
group
Several other job
categories showed SS increased
SMRs for several other sites
based on fewer than five cases
Ill-defined diesel exhaust exposure
without any ranking
Average 6-year follow-up i.e., not
enough time for lung cancer latency
No adjustment for confounders
Wong et al. 34,156 male heavy construction
(1985) equipment operators
Members of the local union for
at least 1 year between
January 1, 1964, and December 1,
1978
20 functional job titles
grouped into three job
categories for potential
exposure
SMR=166 (p<0.05) for liver
cancer for total cohort
No validation of exposure categories,
which were based on surrogate
information
SMR = 343 (observed = 5,
/?<0.05) for lung cancer for high- Incomplete employment records
Exposure groups (high, low, exposure bulldozer operators with
and unknown) based on job 15-19 years of membership, 20+ Employment history other than from
description and proximity to years of follow-up the union not available
source of diesel exhaust
emissions SMR = 119 (observed = 141, 15 year follow-up may not provide
/j<0.01) for workers with no work sufficient time for lung cancer latency
histories
No data on confounders such as other
exposures, alcohol, smoking, etc.
Edling et al. 694 male bus garage employees Three exposure groups
(1987)
Follow-up from 1951 through
1983
Mortality of these men was
compared with mortality of
general population of Sweden
based on job titles:
High exposure, bus
garage workers
Intermediate exposure,
bus drivers
Low exposure, clerks
No SS differences were observed Small sample size
between observed and expected
for any cancers by different No validation of exposure
exposure groups
No data on confounders such as other
exposures, smoking, etc.
-------
;J
KJ
O
O
Table 7-1. Epiih miologic studies of the health effects of exposure to diesel exhaust: cohort mortality studies
(continued)
Authors
Population studied
Diesel exhaust exposure
assessment
Results
Limitations
oo
Bcffetta and 46,981 male volunteers enrolled in Self-reported occupations
were coded into 70 job
categories
Stdline.il the American Cincer Society's
(1(.I88) Prospective Mortality Study of
Cancer in 1982
Employment in high diesel truck drivers (RR= 1.19)
Aged 40 to 79 years at enrollment exhaust exposure jobs were
compared with nonexposed
Total mortality (SS) elevated for Exposure information based on self-
railroad workers (RR= 1.43), reported occupation for which no
heavy equipment operators validation was done
(RR=1.7), miners (RR=1.34), and
Volunteer population, probably
healthy population
First 2-year foi low-up
jobs
Lung cancer mortality (SS)
adjusted for age & smoking,
elevated for total cohort
(RR=1.31), miners (RR=2.67),
and heavy equipment operators
(RR=2.6)
Lung cancer mortality (SNS)
elevated among railroad workers
and truck drivers
Truck drivers also showed a
dose-response
Ga-shick 55,407 white nu.le railroad
et;,l. (1988) workers
Aged 40 to 64 years Li 1959
Started work 10-20 years earlier
than 1959
Industrial hygiene data
correlated with job titles to
dichotomize the jobs as
"exposed" or "not exposed"
Ga snick
(1991)
RR = 1.45 (40-44 year age group) Years of exposure used as surrogate
RR = 1.33 (45-49 year age group) for dose
Both SS
Not possible to separate the effect of
After exclusion of workers time since first exposure and duration
exposed to asbestos of exposure
RR = 1.57 (40-44 year age group)
RR = 1.34 (45-49 year age group) Lack of smoking data but case-control
Both SS study showed very little difference
between those exposed to diesel
Dose response indicated by exhaust versus those who were not
increasing lung cancer risk with
increasing cumulative exposure
Further analysis using attained
age, limited through 1976 showed
youngest workers still had the
highest risk
-------
-J
to
o
o
Table 7-1. Epidemiologic studies of the health effects
(continued)
osure to diesel exhaust: cohort mortality studies
Authors
Population studied
Diesel exhaust exposure
assessment
Results
Limitations
Crump et al. Reanalysis of Garshick et al., 1988
(1991) data
Crump et al.
(1999)
Dose response found to be
positive or negative depending
upon how the age was controlled
in the model
Negative dose-response upheld in
the latest analysis
California Reanalysis of Garshick et al., 1988
EPA (1998)
Positive dose response using age
at 1959 and interaction term of
age & calendar year
Gustavsson 695 male workers from 5 bus
et al. (1990) garages in Stockholm, Sweden,
who had worked for 6 months
between 1945 and 1970
34 years follow-up (1952-1986)
Nested case-control study
17 cases, six controls for each case
matched on age ± 2 years
duration of work
Four diesel exhaust indices SNS SMRs of 122 and 115 (OA
were created: and GP), respectively
Oto 10, 10 to 20, 20-30, and
>30 based on job tasks and Case-control study results
showed dose response:
RR= 1.34 (10 to 20)
RR= 1.81(20 to 30)
RR = 2.43 (>30)
All SS with 0-10 as comparison
group
Exposure matrix based on job tasks
(not on actual measurements)
Small cohort, hence low power
Lack of smoking data is unlikely to
confound the results since it is a
nested case-control study
O
O
2
O
H
n
O
G
O
H
W
Hansen Cohort of 57,249 unskilled Diesel exhaust exposure
(1993) laborers, ages 15 to 74, in assumed based on diesel-
Denmark (nationwide census file) powered trucks
November 9, 1970
Follow-up through November 9,
1980
SS SMRs for lung cancer :
SMR = 160 for total population
SMR = 229 for age 55-59 years
SMR = 227 for age 60-64 years
No actual exposure data available
Lack of smoking data but population
survey showed very little difference
between rural and urban smoking
habits
Job changes may have occurred from
laborer to driver
Short follow-up period
-------
^ Table 7-1. Epidoiniologic studies of the health effects of exposure to diesel exhaust: cohort mortality studies
|^J (continued)
o Authors Population studied Diesel exhaust exposure Results Limitations
assessment
Saverin et Cohort of 5,536 potash miners Diesel exhaust exposure SNS increased RRs adjusted for Small, young cohort
al. (1999) who had worked underground for categories defined as: smoking: 1.68 and 2.7 for total
at least 1 year after 1969 production (high) cohort & subcohort, respectively Few deaths
maintenance (medium)
Subcohort of 3,258 who had workshop (low) No latency analysis
worked for at least 10 years
underground 225 air samples obtained:
for total carbon, organics, &
Follow-up from 1970 to 1994 fine dust in 1992
Abbreviations: RR = relative risk; SMR = standardized mortality ratio; SNS = statistically nonsignificant; SS = statistically significant;
0 = occupationally active; GP = general population.
-4
o
O
O
2
O
H
n
/O
o
3
-------
Table 7-2. Epidemiologic studies of the health effects of exposure to diesel exhaust: case-control studies of lung cancer
to
LTt
Authors
Population studied
Diesel exhaust exposure
Results
Limitations
K)
Hall and 502 histologically confirmed
Wynder lung cancers
(1984) Cases diagnosed 12 mo prior to
interviews
502 matched hospital controls
without tobacco-related diseases,
matched for age, sex, race, and
geographical area
Population from 18 hospitals in
controls
Based on previous
Industrial Hygiene
Standards for a
particular occupation,
usual lifetime occupation
coded as "probably high
exposure" and "no
exposure"
N1OSH standards used
to classify exposures:
High
Moderate
Low
SNS excess risk after adjustment for Complete lifetime employment
j
smoking for lung cancer:
RR= 1.4 (1st criteria)
and
RR=1.7 (NIOSH criteria)
history not available
Self-reported occupation history not
validated
No analysis by dose, latency, or
duration of exposure
No information on nonoccupational
diesel exposure
O
O
z
O
H
O
HH
a
Damber and 589 lung cancer cases who had
Larsson died prior to 1979 reported to
(1987) Swedish registry between 1972
and 1977
582 matched dead controls (sex,
age, year of death, municipality)
drawn from National Registry
of Cause of Death
453 matched living controls
(sex, year of birth, municipality)
drawn from National
Population Registry
Occupations held for at
least I year or more
For underground miners: SS OR = Uncertain diesel exhaust exposure
2.7 (a 1 year of employment)
No validation of exposure done
A 5-digit code was used to SS OR = 9.8 (i20 years of
classify the occupations employment)
according to Nordic
Classification of
= 1.2 (a 20 years of employment)
with dead controls
Occupations
Underground miners data not adjusted
for other confounders such as radon,
For professional drivers: SNS OR etc.
All ORs adjusted for smoking
O
G
O
H
CD
-------
^J
N>
O
O
Tnblc 7-2. Epidcmiologic studies of the health effects of exposure to diesel exhaust:
case-coitrol studies of lung cancer (continued)
Authors
Population studied
Diesel exhaust exposure
Results
Limitations
NJ
NJ
Ler ;hen 506 bng cancel' cases from
et a.. (1987) New Mexico tuiror registry
(333 males and 173 females)
Aged 25-84 yem
Diagnosed between January 1,
1980, and December 31, 1982
771 (499 males and 272 females)
frequency matched wr!h cases,
selected from telephone directory
Lifetime occupational
history and self-reported
exposure history were
obtained
Coded according to
Standard Industrial
Classification Scheme
No excess of relative odds were
observed for diesel exhaust
exposure
Exposure based on occupational
history and self-report, which was not
validated
50% occupational history provided by
next of kin
Absence of lung cancer association
with asbestos suggests
misclassification of exposure
Gar ihick 1,319 lung cance: cases who died
etal.(1987) between March 1, 1981,
and February 28, 1982
2,385 matched controls (two each,
age and date of death)
Both cases and controls drawn
from railroad worker cohort
who had worked for 10 or
more years
Personal exposure assessed SS OR =1.41 (s64 year age group)
for 39 job categories
SS OR = 1.64 (s64 year age group)
for i20 years diesel exhaust
exposure group when compared to
0- to 4-year exposure group
This was corrected with
job titles to dichotomize
the exposure into:
Exposed
Not exposed
Industrial hygiene
sampling done
All ORs adjusted for lifetime
smoking and asbestos exposure
Probable misclassification of diesel
exhaust exposure jobs
Years of exposure used as surrogate
for dose
13% of death certificates not
ascertained
Overestimation of smoking history
r
o
o
2
o
H
O
i— i
H
m
O
c
o
a
Ben lamou 1,260 histologica ly confirmed
etal.(1988) lung cancer cases
2,084 non-tobacco-related
disease matched controls
(sex, age at diagnosis,
hospital admissioi, and
interviewer)
Occurring between 1976 and
1980 in France
Based on exposures
determined by panel of
experts
The occupations were
recorded blindly using
International Standard
Classification of
Occupations as chemical
or physical exposures
Significant excess risks were found
in motor vehicle drivers
(RR= 1.42) and
transport equipment operators
(RR = 1.35) (smoking adjusted)
Exposure based on occupational
histories not validated
Exposures classified as chemical and
physical exposures, not specific to
diesel exhaust
-------
O
o
Table 7-2. Epidemiologic studies of the health effects of exposure to diesel exhaust:
case-control studies of lung cancer (continued)
Authors
Population studied
Diesel exhaust exposure
Results
Limitations
Hayes et al. Pooled data from three different
(1989) studies consisting of 2,291 male
lung cancer cases
2,570 controls
SS OR = 1.5 for truck drivers (>10
years of employment)
Occupational information
from next of kin for all
jobs held
SS positive trend with increasing
Jobs classified with respect employment as truck driver
to potential exposure to
known and suspected Adjusted for age, smoking, & study
pulmonary carcinogens area
Exposure data based on job
description given by next of kin,
which was not validated
Could have been mixed exposure to
both diesel and gasoline exhausts
Job description could have led to
misclassification
-J
to
Steenland 1,058 male lung cancer deaths
et al. (1990) between 1982 and 1983
1,160, every sixth death from
entire mortality file, sorted by
Social Security number
(excluding lung cancer,
bladder cancer, and motor
vehicle accidents)
Cases and controls were from
Central State Teamsters who
had filed claims (requiring 20-year
tenure)
Longest job held: diesel As 1964 cut-off point:
truck driver, gasoline truck
driver, both types
of trucks, truck
mechanic, and
SS OR = 1.64 for long-haul drivers
with 13+ years of employment
dockworkers
Positive trend test for long-haul
drivers (p=0.04)
SS OR = 1.89 for diesel truck
drivers of 35+ years of employment
Adjusted for age, smoking, &
asbestos
Exposure based on job titles not
validated
Possible misclassification of exposure
and smoking, based on next-of-kin
information
Lack of sufficient latency
O
O
2:
o
H
n
i— t
a
o
G
O
m
Steenland et Exposure-response analyses of
al. (1998) their 1990 case-control study
Industrial hygiene data of For mechanics: OR = 1.69 (had the
elemental carbon in highest diesel exhaust exposure)
trucking industry collected
by Zaebst et al. (1991) Lowest diesel exhaust exposure and
used to estimate individual lowest OR = 0.93 observed for
exposures dockworkers
Cumulative exposures
calculated based on
estimated lifetime
exposures
Increasing risk of lung cancer with
increasing exposure
Adjusted for age & smoking
-------
O
Table 7-2. Epidemiologic studies of the health effects of exposure to diesel exhaust:
case-control studies of lung cancer (continued)
Authors
Population studied
Diesel exhaust exposure
Results
Limitations
Bolfetta et From 18 hospitals (since 1969),
al. (1990) 2,584 male lung cancer cases
matched to either one control (69)
or two controls' (2,515) were
drawn. Matched on age, hospital,
and year of interview
A priori aggregation of OR slightly below unity SNS
occupations categorized
into low probability, Adjusted for smoking
possible exposure (19
occupations), and probable
exposure (13 occupations)
to diesel exhaust
No verification of exposure
Duration of employment used as
surrogate for dose
Number of individuals exposed to
diesel exhaust was small
Emmelin et 50 male lung cancer cases from
al. (1993) 15 ports (worked for at least
Indirect diesel exhaust SS OR for high-exposure group =
exposure assessment done 6.8
6 months between 1950 and 1974), based on (1) exposure
154 controls matched on age and intensity, (2)
port characteristics of
Positive trend for diesel exhaust
observed (trend much steeper for
ventilation, (3) measure of smokers than nonsmokers)
proportion of time in
higher exposure jobs Adjusted for smoking
Numbers of cases and controls are
small
Very few nonsmokers
Lack of exposure information on
asbestos
No latency analysis
O
D
O
Z
O
H
O
»—H
H
m
o
&
/o
G
O
H
W
Swiinson et Population based case-control
al. (199:i) study in metropolitan Detroit
Telephone interviews with
the individual or surrogate
about lifetime work history
3,792 lung cancer cases and 1,966
colon cancer (cases) controls, Occupation and industry
diagnosed between 1984 and 1987 data coded per 1980 U.S.
in white and black males (aged Census Bureau
between 40-84) classification codes
Certain occupations and
industries were selected as
unexposed to carcinogens
SS excess ORs observed for
- black farmers OR= 10.4 for 20+
years employment
- white railroad industry workers
OR= 2.4 for 10+ years employment
Among white trend tests were SS
for
-drivers of heavy duty trucks
- drivers of light duty trucks
- farmers
- railroad workers
Among blacks trend test was SS for
farmers only
All the ORs were adjusted for age at
diagnosis, pack-years of cigarette
smoking and race
Lack of direct information on specific
exposures
No latency analysis
-------
K)
L/l
O
o
Table 7-2. Epidemiologic studies of the health effects of exposure to diesel exhaust:
case-control studies of lung cancer (continued)
Authors
Population studied
Diesel exhaust exposure
Results
Limitations
^J
K)
Hansen et Population-based case-control
al. (1998) study of professional drivers in
Denmark
Male lung cancer cases diagnosed
between 1970-1989, controls
matched by year of birth and sex
Information about past
employment obtained by
linkage with nationwide
pension fund
Employment as lorry/bus
drivers (n= 1,640) and taxi
drivers (n=426) was used
as surrogate for exposure
to diesel exhaust
For lorry/bus drivers: SS OR = 1.31
For taxi drivers: SS OR = 1.64,
which increased to 2.2 in > 5-year
employment with no lag time & 3.0
in > 5 year employment with 10-
year lag time
SS trend test for increasing risk
with increasing employment for
both lorry/bus drivers & taxi drivers
(p<0.001)
All ORs adjusted for socioeconomic
status
Lack of information on the type of
fuel (personal communication with
the principal investigator confirmed
that diesel fuel is used for the
lorry/buses and taxis since early
1960s)
Even though direct adjustment was
not done for smoking/asbestos,
indirect methods indicate that the
results are not likely to be confounded
by these factors
O
O
2
O
H
O
HH
H
m
o
&
o
c
o
H
-------
-J
K5
O
O
Table 7-2. Epidemiologic studies of the health effects of exposure to diesel exhaust:
case-control studies of lung cancer (continued)
Authors
Population studied
Diesel exhaust exposure
Results
Limitations
to
o\
Briiske- Pooled analysis of two case-
Hohlfeld et control studies (3,498 cases &
al. (1999) 3,541 controls)
Controls frequency matched on
sex, age, & region, randomly
selected from the compulsory
population registry
Inclusion criteria: (1) born in or
after 1913/less than 75 years old,
(2) German nationality/resident of
the region - lived in Germany for
more than 25 years, & (3) lung
cancer diagnosis should be 3
months prior to the study
Information obtained by personal
interview on:
Lifetime detailed
occupational & smoking
histories obtained from
each individual in a
personal interview
SS higher risk adjusted for smoking Lack of data on actual exposure to
observed for all 4 categories:
diesel exaust
A- ORs ranged from 1.25 to 2.53
B- ORs ranged from 1.53 to 2.88
C- ORs ranged from 2.31 to 4.3
Based on job codes (33 job D- 6.81 (exposure < 30 years)
titles & 21 industries)
potential diesel exhaust Risk increased with increasing
exposure classified in 4 exposure
categories: A- professional
drivers of trucks, buses, &
taxis; B- other traffic
related i.e., switchman,
locomotive, & forklift
drivers; C- bulldozer
operators, graders,&
excavators; D- farm tractor
drivers
Cumulative diesel exhaust
exposures and pack-years
(smoking) calculated for
each individual
Abbreviations: OR = odds ratio; RR = relative risk; SNS = statistically nonsignificant; SS = statistically significant.
-------
-J
I
H--*
to
D
O
O
z
o
H
n
h-H
H
W
O
&
O
c
o
H
Species/
Study strain
Karagianes et Ral/Wistar
al. (1981)
Kaplan et al. Rat/F344
(1983)
White etal.
(1983)
Heinrich et al. Rat/ Wistar
(1986a,b)
Mohretal.
(1986)
Iwai et al. Rat/F344
(1986)
Takemoto et Rat/F344
al. (1986)
Mauderly et al. Rat/F344
(1987)
Sex/total
number
M, 40
M, 40
M.30
M, 30
M, 30
M, 30
F.96
F.92
F, 95
F.24
F,24
F.24
F, 12
F,2I
F, 15
F, 18
M + F, 230b
M + F, 223
M + F, 221
M + F, 227
Exposure
atmosphere
Clean air
Whole exhaust
Clean air
Whole exhaust
Whole exhaust
Whole exhaust
Clean air
Filtered
exhaust
Whole exhaust
Clean air
Filtered
exhaust
Whole exhaust
Clean air
Clean air
Whole exhaust
Whole exhaust
Clean air
Whole exhaust
Whole exhaust
Whole exhaust
Particle
concentration (nig/ Other
m3) treatment
8.3 None
None
0 None
0.25 None
0.75 None
1.5 None
4 None
None
None
4.9 None
None
None
0 None
0 DIPN"
2-4 None
2-4 DIPN"
0 None
0.35 None
3.5 None
7.1 None
Post-
Exposure exposure
protocol observation
6 hr/day, NA
5 days/week,
for up to
20 mo
20 hr/day, 8 mo
7 days/week, 8 mo
for up to 8 mo
15 mo 8 mo
19 hr/day, NA
5 days/week
for up to
35 mo
8 hr/day, NA
7 days/week,
for 24 mo
4 hr/day, NA
4 days/week,
18-24 mo
7 hr/day, NA
5 days/week
up to 30 mo
Tumor type and incidence (%)"
Adenomas
0/6 (0)
1/6(16.6)
Bronchoalveolar carcinoma
0/30 (0)
1/30(3.3)
3/30(10.0)
1/30(3.3)
Squamous
Adenomas Carcinomas cell tumors
0/96 (0) 0/96 (0) 0/96 (0)
0/92 (0) 0/92 (0) 0/92 (0)
8/95 (8.4) 0/95 (0) 9/95 (9.4)
Large cell
Adenocarcinoma and
and squamous
adenosquamous cell
Adenomas carcinoma carcinomas
1/22(4.5) 0/22(0) 0/22(0)
0/16(0) 0/16(0) 0/16(0)
3/19(0) 3/19(15.8) 2/19(10.5)
Adenoma Carcinoma
0/12(0) 0/12(0)
10/21 (47.6) 4/21 (19)
0/15(0) 0/15(0)
12/18(66.7) 7/18(38.9)
Adenocarcinoma
+ squamous cell Squamous
Adenomas carcinoma cysts
(0) (0.9) (0)
(0) (1-3) (0)
(2.3) (0.5) (0.9)
(0.4) (7.5) (4.9)
Comments
All tumors
0/96 (0)
0/92 (0)
17/95 (17.8)c
All tumors
1/22 (4.5)r
0/16(0)
8/19
All
tumors
(0.9)
(1.3)
(3.6)'
(12.8)'
-------
Table 7-3. Summary of animal inhalation carcinogenicity studies (continued)
^
/)
3
•3
.j
«4
O
x>
3
^
>
T-l
*•!
H
J
2
H
">
-H
TJ
D
a
D
-H
3
Tl
Species/
Study strain
Ishinish clal. Rat/F344
(I988a)
Heavy-duty
engine
Ishinish etal. Rat/F344
(I988a)
Light dity
Heavy duly
Brightw.:!! et Rat/344
al.(!98(.)
Henrich ;t al. Rat/Wistar
(I989a)
Lewis el al. Rat/F344
(1989)
Sex/total
number
M + F, ;:3
M + F, ;;3
M + F, ::5
M + F, ;:3
M + F, :'4
NS, 5
NS,8
NS, II
NS, 5
NS, 9
NS, II
NS, 5
NS.9
NS 11
NS, 5
NS,6
NS, 13
M + F, 260
M + F, 144
M + F, 143
M + F, 143
M + F, 144
M + F, 143
F, NS
F, NS
F, NS
F, NS
F,NS
F, NS
M + F, 288"
Particle
Exposure concentration (ing/
atmosphere m1)
Clean air
Whole exhaust
Whole exhaust
Whole exhaust
Whole exhaust
Whole exhaust
Whole exhaust
Whole exhaust
Whole exhaust
Whole exhaust
Whole exhaust
Whole exhaust
Whole exhaust
Whole exhaust
Whole exhaust
Whole exhaust
Whole exhaust
Clean air
Filtered
exhaust
(medium
exposure)
Filtered
exhaust (high
exposure)
Whole exhaust
Whole exhaust
Whole exhaust
Clean air
Whole exhaust
Filtered
exhaust
Clean air
Whole exhaust
Filtered
exhaust
Clean air
Whole exhaust
0
0.5
1.0
1.8
3.7
0.
0.
0.
1.
1.
1.
0.5
0.5
0.5
1.8
1.8
18
0
0
0
0.7
2.2
6.6
0
4.2
0
0
4.2
0
2
Other
treatment
None
None
None
None
None
None
None
None
None
None
None
None
None
None
None
None
None
None
None
None
None
None
None
DPNJ
DPNJ
DPNd
DPNC
DPN'
DPN'
None
None
Post-
Exposure exposure
protocol observation
16hr/day,
6 days/week,
for up to
30 mo
I6hr/day,
6 days/week,
for 12 mo
16hr/day,
6 days/week,
for 12 mo
I6hr/day,
5 days/week,
for 24 mo
19hr/day,
5 days/week
for 24 to
30 mo
7 hr/day,
5 days/week,
24 mo
NA
6 mo
12 mo
18 mo
6 mo
12 mo
18 mo
6 mo
12 mo
18 mo
6 mo
12 mo
18 mo
NA
NA
NA
Tumor type and incidence (%)'
Adenomas
0/123(0)
0/123(0)
0/125(0)
0/123(0)
0/124(0)
Adenomas
0/5 (0)
0/8 (0)
0/1 1 (0)
0/5 (0)
0/9 (0)
0/1 1 (0)
0/5 (0)
0/9 (0)
0/1 1 (0)
0/5 (0)
0/6 (0)
0/13(0)
No tumors
Adenosquamous Squamous
carcinomas cell
1/123(0.8) carcinomas
0/123(0) 0/123(0)
0/125(0) 1/123(0.8)
4/123(3.3) 0/125(0)
6/124(4.8) 0/123(0)
2/124(1.6)
Carcinomas All tumors
0/5 (0) 0/5 (0)
0/8 (0) 0/8 (0)
0/11(0) 0/11(0)
0/5 (0) 0/5 (0)
0/9 (0) 0/9 (0)
0/1 1 (0) 0/1 1 (0)
0/5 (0) 0/5 (0)
0/9 (0) 0/9 (0)
0/1 1 (0) 0/1 1 (0)
0/5 (0) 0/1 1 (0)
0/6 (0) 0/6 (0)
1/13(0) 1/13(0)
Primary lung tumors
3/260(1.2)
0/144 (0)
0/143 (0)
1/143(0.7)
14/144(9.7)'
55/143(38.5)'
Squamous
cell
carcinoma
(4.4)
(46.8)'
(4.4)
(16.7)
(31.3)'
(14.6)
All tumors
1/123(0.8)
1/123(0.8)
0/125(0)
4/123(3.3)
8/124(6.5)'
All lung
tumors
(84.8)
(83.0)
(67.4)
(93.8)
(89.6)
(89.6)
0/192(0)
0/192(0)
Comments
Tumor
incidence for
all rats dying
or sacrificed
? 24/25 (96%)
after 24 mo
tf 12/27 (44%)
after 24 mo
-------
Table 7-3. Summary of animal inhalation carciri^lnicity studies (continued)
K)
Species/
Study strain
Takakietal. Rat/F344
(1989)
Light-duty
engine
lleinrichetal. Rat/Wistar
(1995)
Nikulaetal. Ral/F344
(1995)
Iwaietal. F/344
(1997)
Sex/total
number
M + F
M + F
M + F
M + F
M + F
F,
F,
F,
F,
F,
F,
M + F
M + F
M + F
M + F
M + F
, 123
,123
,125
, 123
, 124
220
200
200
100
100
100
,214"
,210
,212
,213
,211
121, F
108, F
153, F
Particle
Exposure concentration (ing/
atmosphere m1)
Clean air
Whole exhaust
Whole exhaust
Whole exhaust
Whole exhaust
Clean air
Whole exhaust
Whole exhaust
Whole exhaust
Carbon black
Ti02
Clean air
Whole exhaust
Whole exhaust
Carbon black
Carbon black
Clean air
Filtered air
Whole exhaust
0
0.1
0.4
II
2.3
0
0.8
2.5
7.0
11.6
10.0
0
2.5
6.5
2.5
6.5
0
0
3.2-9.4
Other
treatment
None
None
None
None
None
None
None
None
None
None
None
None
None
None
None
None
None
None
None
Exposure
protocol
16 hr/day,
6 days/week,
for up to
30 mo
1 8 hr/day,
5 days/week,
for up to
24 mo
16 hr/day,
5 days/week
for up to
24 mo
NA
48-56 hr/day
48-56 hr/day
Post-
exposure
observation
Adenosquainous
carcinomas
NA 1/23 (0.8)
1/23 (0.8)
1/25 (0.8)
0/23 (0)
1/24(8.1)
Adenomas
6 mo 0/217(0)
0/198(0)
2/200(1)
4/100(4)
13/100(13)
4/100(4)
Adenomas
6 weeks 1/214 (<1)
7/210(3)
23/212(11)
3/213(1)
13/211 (6)
NA
6 mo
Tumor type and incidence (%)'
Squamous cell
carcinomas All tumors
2/123(1.6) 1/23(0.8)
1/23 (0.8) 1/23(0.8)
0/125(0) 0/125(0)
5/123(4.1) 0/123(0)
2/124(1.6) . 0/124(0)
Squamous
cell
Adenocarcinomas carcinomas
1/217 (
-------
Table 7-3. Summary of animal inhalation carcinogenicity studies (continued)
o
o
U)
o
O
O
2
3
n
H H
Si
/O
G
O
H
tn
Species/
Study strain
Mouse/
Jackson A
Mouse/
Jackson A
Kaplan el al. Mouse
(1982) A/J
Kaplan etal. Mouse/
(1983) A/J
White etal.
(1983)
Pepelko and Mouse/
Peirano(l98.'i) Sencar
Sex/total
number
M + 1:, 40
M + l:, 40
F.60
F,60
F.60
F,60
M, 429
M.430
M, 458
M, 18
M, 485
M.388
M, 388
M.399
M.396
M + F, 260
Exposure
atmosphere
Clean air
Whole exhaust
Clean air
Clean air
Whole exhaust
Whole exhaust
Clean air
Whole exhaust
Clean air
Clean air
Whole exhaust
Clean air
Whole exhaust
Whole exhaust
Whole exhaust
Cle;m air
Clean air
Clean air
Whole exhaust
Whole exhaust
Whole exhaust
Particle
concentration (nig/
m>)
0
6.4
0
0
6.4
6.4
0
6.4
1.5
0
0.25
0.75
1.5
121212
Other
treatment
None
None
None
Urethan1
None
Urethan1
None
None
None
Urethan"
None
None
None
None
None
None
BHT1
Urethan1
None
BHT1
Urethan1
Post-
Exposure exposure
protocol observation
20 hr/day, 8 weeks
7 days/week,
for 8 weeks
8 weeks
20 hr/day,
7 days/week,
for approx.
7 mo.
20 hr/day, 6 mo
7 days/week,
for 3 mo
20 hr/day, NA
7 days/week,
for up to
8 mo
Continuous NA
for 1 5 mo
Tumor type and incidence (%)'
16/36(44.4)
1 1/34 (32.3)
4/58 (6.9)
9/52(17.3)
14/56(25.0)
22/59 (37.3)
73/403(18.0)
66/368(17.9)
Pulmonary adenomas
144/458(31.4)
18/18(100)
165/485(34.2)
Pulmonary adenoma
130/388(33.5)
131/388(33.8)
109/399 (27.3)
99/396 (25.0)
Adenomas Carcinomas All tumors
(5.1) (0.5) (5.6)
(12.2) (1.7) (2.8)
(8.1) (0.9) (9.0)
(10.2)c (1.0) (11.2)c
(5.4) (2.7) (8.1)
(8.7) (2.6) (11.2)
Comments
0.5 tumors/
mouse
0.4 tumors/
mouse
0.09 tumors/
mouse
0.25 tumors/
mouse
0.32 tumors/
mouse
0.39 tumors/
mouse
0.23 tumors/
mouse
0.20 tumors/
mouse
-------
Table 7-3. Summary of animal inhalation carciriBfEnicity studies (continued)
J\
-^
-••^
O Species/
Study strain
I'epelko and Mouse/
Peirano(l983) Strain A
_j Meinrich et al. Mouse/
L. (I986a,b) NMRI
-»J
Takcinolo el Mouse/IRC
al.
(1986)
Mouse/
C57BL
;g Hcinrich et al. Mouse/
> (1995) C57BL/6N
n
-j
J
. Mouse/
^ NMRI
H
->
-H
H
^ Mouse/
D NMRI
50
D
3
H
T)
Sex/total
number
M + F, 90
M + F, 84
M + F, 93
M + F, 76
M + F, 45
M + F, 69
M + F, 12
M + F, 38
F, 120
F, 120
F, 120
F, 120
F, 120
F, 120
F, 120
F, 120
Exposure
atmosphere
Clean air
Clean air
Whole exhaust
Whole exhaust
Clean air
Whole exhaust
Clean air
Filtered
exhaust
Whole exhaust
Clean air
Whole exhaust
Clean air
Whole exhaust
Clean air
Whole exhaust
Particle-free
exhaust
Clean air
Whole exhaust
Carbon black
TiOj
Clean air
Whole exhaust
Particle-free
exhaust
Particle
concentration (nig/ Other
m3) treatment
1212012 None
Exposure
(darkness)
Exposure
(darkness)
Urcthan"1
Urethan"1
4 None
None
None
0 None
2-4 None
0 None
2-4 None
4.5 None
None
None
0 None
4.5 None
11.6 None
10 None
4.5 None
None
None
Post-
Exposure exposure
protocol observation
NA
19 hr/day, NA
5 days/week
for up to
30 mo
4 hr/day, NA
4 days/week,
for 19-28 mo
4 hr/day, NA
4 days/week
for 19-28 mo
18 hr/day, 6 mo
5 days/week,
for up to 2 1
mo
18 hr/day, 9.5 mo
5 days/week
for up to
13.5 mo
18 hr/day, None
5 days/week,
23 mo
Tumor type and incidence (%)"
All tumors
21/87(24)
59/237 (24.9)
10/80(12.5)
22/250(0.10)
66/75 (88)
42/75 (0.95)
Squamous
cell
Adenomas Adenocarcinoma tumors All tumors
9/84(11) 2/84(2) — 11/84(13)
11/93(12) 18/93(19)' — 29/93(31)'
11/76(15) 13/76(17)' — 24/76(32)'
Adenoma Adenocarcinoma
3/45(6.7) 1/45(2.2)
6/69 (8.7) 3/69 (4.3)
1/12(8.3) 0/12(0)
8/38(21.1) 3/38(7.9)
Adenomas Adenocarcinomas
(25) (15.4)
(21.8) (15.4)
(11.3) (10)
(U.3) (2.5)
(25) (8.8)
(18.3) (5.0)
(31.7) (15)
Comments
0.29 tumors/
mouse
0.27 tumors/
mouse
0.14
0.10
2.80
0.95
5.1% tumor
rate
8.5% tumor
rate
3.5% tumor
rate
-------
^*
J\
5
— '
-J
*J
o
3
0
>
r)
H
3
5
z;
:>
H
"5
•H
•-J
H
Species/ Sex/total
Study strain number
Maudenyetal. Mouse/CD-I M + F, I5',b
(I996) M + F, I7l
M + F, 1 5J
M + F, I8(
Heinriclietal. Hamster/ M + F, 96
(I986a,b) Syrian M + F, 96
M + F, 96
I3riglitw;ll el Hamster/ M + F,
al. Syrian M + F, 202
(1 989) Golden M + F, 1 04
M + F, I04
M + F, I0l
M + F, I02
M + F, lOI
M + F, 204
M + F, 203
Particle
Exposure concentration (mg/
atmosphere
Clean air
Whole exhaust
Whole exhaust
Whole exhaust
Clean air
Filtered exhaust
Whole exhaust
Clean air
Clean air
Filtered
exhaust
(medium
dose)
Filtered
exhaust
(high dose)
Whole exhaust
Whole exhaust
Whole exhaust
Filtered
exhaust
(high dose)
Whole exhaust
mj)
0
0.35
3.5
7.0
4
0
0
0
0
0.7
2.2
6.6
0
6.6
Other
treatment
None
None
None
None
None
None
None
None
DEN'
DEN'
DEN'
DEN'
DEN'
DENi
None
None
'Table values indicate number with tumors/number examined (% animals with tumors).
'Number of animals examined for tumors.
'Significantly different from clean air controls
Post-
Exposure exposure
protocol observation
7 hr/day, 5 None
days/week,
for up to 24
mo
19 hr/day
5 days/week
for up to
30 mo NA
16 hr/day, NA
5 days/week,
for 24 mo
Tumor type and incidence (%)'
Multiple
adenomas
1/157(0.6)
2/171(1.2)
0/155(0)
0/186(0)
Adenomas
0/96(0)
0/96(0)
0/96(0)
Multiple
carcinomas
2/157(1.3)
1/171 (0.6)
1/155(0.6)
0/186(0)
Adenomas/
carcinoma
1/157(0.6)
1/171 (0.6)
0/155(0)
0/186(0)
Squamous
cell
Adenocarcinoma tumors
0/96(0)
0/96(0)
0/96(0)
Primary lung
tumors
7/202 (3.5)
4/104(3.8)
9/104(8.7)
2/101 (2.0)
6/102(5.9)
4/101 (3.9)
1/204(0.5)
0/203 (0)
'Butylated hydroxytoluene 300 mg/kg, i.p. for week
weeks 3 to 52.
0/96
0/96
0/96
Alveolar/
bronchiolar
adenoma
10/157(6.4)
16/171(9.4)
8/155(5.2)
10/186(5.4)
All tumors
0/96(0)
0/96(0)
0/96(0)
Comments
Alveolar/
bronchiolar
carcinoma
7/157(4.5)
5/171 (2.9)
6/155(3.9)
4/186(2.2)
Respiratory
tract tumors
not related to
exhaust
exposure for
any of the
groups
1 , 83 mg/kg for week 2, and 1 50 mg/kg for
"12 mg/mjfrom 12 weeks of age to termination of exposure. Prior exposure (in utero) and of parents
dDipcntylnitrosainine; 6.25 mg/kg/weck s.c. during first 25 weeks of exposure.
'Dipentylnitrasamine; 1 2.5 mg/kg/week s.t. during first 25 weeks of exposure.
~v 'Splenic lymplumas also delected in controls (8.3%), filtered exhaust group (37.5%) and whole
•^
V
"}
.^
-H
:5
•^
n
exhaust group (25%).
"5.3% incidence oflarge cell carcinomas.
hl g/kg, i p. I/week for 3 weeks starting I mo into exposure.
'Includes adenomas, squamous cell carcinomas, adenocarcinomas,
and mesolheliomas.
was 6 mg/m5.
"120-121 males and 71-72 females examined histologically.
"Not all animals were exposed for full term, at least
exposure.
NS = Not specified.
NA = Not applicable
10 males were killed at 3, 6, and
12 mo of
adenosquamous cell carcinoma,
'4.5 mg/diethylnitiosamine (DEN)/kg, s.c., 3 days prior to start of inhalation exposure.
'Single i.o. dos; I mg/kg at start of exposure.
-------
Table 7-4. Tumor incidences in rats following intratracheal instillation of diesel
exhaust particles (DPM), extracted DPM, carbon black (CB), benzo[a]pyrene (BaP),
or particles plus BaP
Experimental
group
Control
DPM (original)
DPM (extracted)
DPM (extracted)
CB (printex)
CB (lampblack)
BaP
BaP
DEP + BaP
CB (printex) + BaP
Number
of
animals
47
48
48
48
48
48
47
48
48
48
Total dose
4.5 mL
15 mg
30 mg
15 mg
15 mg
14 mg
30 mg
15 mg
15 nig + 170 ^g
BaP
1 5 mg + 443 ng
BaP
Animals with
tumors
(percent)
0 (0)
8 (17)
10 (21)
2 (4)
10 (21)
4 (8)
43 (90)
12 (25)
4 (8)
13 (27)
Statistical
significance8
-
<0.01
< 0.001
NS
< 0.001
NS
< 0.001
< 0.001 .
NS
< 0.001
7/25/00
7-133
DRAFT—DO NOT CITE OR QUOTE
-------
7/25/00
-pj
UJ
Table 7-5. Tumorigenic
Number of
animals Strain/sex
:>2 C57BL/40
F
C57BL/12
M
.50 Strain A/M
25 Strain A/F
effects of dermal application
Sample material
Extract of DPM obtained
during warmup
Extract of DPM obtained
during full load
Extract of DPM obtained
during full load
of acetone extracts of DPM
Time to first
tumor (mo)
13
15
13
Survivors at
time of first
tumor Total tumors
33 2
8 4
20 17
Duration of
experiment
(mo)
22
23
17
Source: Kotineta;., 1955.
O
O
2
O
H
O
o
JO
/o
G
O
H
m
-------
-J
N>
O
O
U)
O
O
2
O
H
O
O
to
O
c
3
W
7-6. Tumor incidence and survival time of rats treateTrT)y surgical lung implantation with fractions from diesel
exhaust condensate (35 rats/group)
Material portion by weight (%)
Hydrophilic fraction (I) (25)
Hydrophobia fraction (II) (75)
Nonaromatics +
PACC 2 + 3 rings (Ha) (72)
PAHd 4 to 7 rings (lib) (0.8)
Polar PAC (lie) (1.1)
Nitro-PAH (lid) (0.7)
Reconstituted hydrophobics
(la, b, c, d) (74.5)
Control, unrelated
Control (beeswax/trioctanoin)
Benzo[a]pyrene
Dose (mg)
6.7
20.00
19.22
0.21
0.29
0.19
19.91
0.3
0.1
0.03
Median
survival time
in weeks Number of
(range) carcinomas3
97(24-139) 0
99(50-139) 50601
103(25-140)
102(50-140)
97 (44-138)
106(32-135)
93(46-136) 70027113
110(23-138)
103(51-136)
69(41-135)
98(22-134)
97(32-135)
Number of Carcinoma
adenomas'* incidence (%)
1 0
1000 14.2
0
17.1
0
2.8
101000 20.0
0
0
77.1
31.4
8.6
aSquamous cell carcinoma.
bBronchiolar/alveolar adenoma.
CPAC = polycyclic aromatic compounds.
dPAH = polycyclic aromatic hydrocarbons.
Source: Adapted from Grimmer et al., 1987.
-------
-J
to
u<
o
o
Table 7-7. Dermal tumorigenic and carcinogenic effects of various emission extracts
-J
U)
N.X
£
O
O
2!
O
H
O
Tumor initiation
Sample Papillomas" Carcinomas'1
B ;nzo[« jpyrene +/+c +/+
Topside coke oven ' +/+ -/+
C ake oven main +/+ +/+
Raofingtar +/+ +/+
Nissan +/+ +/+
Oldsrnobile +/+ -/-
VW Rabbit +/+ -/-
I^iercedes +/- -/-
Caterpillar -/- -/-
Rssidential furnace -/- -/-
Mustang +/+ -/+
Complete
carcinogenesis
Carcinomas'1
+/+
NDd
+/+
+/+
-/-
-/-
Ie
ND
-/-
ND
ND
Tumor promotion
Papillomas"
+/+
ND
+/+
+/+
ND
ND
ND
ND
ND
ND
ND
"Scored at 6 mo.
bCumulative score at 1 year.
cMale/f3male.
''ND = Not determined.
el = Incomplete.
Scarce: Nesnowetal., 1982.
o
e:
o
H
m
-------
NJ
O
o
Table 7-8. Cumulative (concentration x time) exposure dafa for rats exposed to whole diesel exhaust
Cumulative exposure
-J
_a
+)
-4
3
£
n
-3
3
D
t
D
^
Tl
3
D
— <
Study
Mauderly et al.
(1987)
Nikula et al.
(1995)
Heinrich et al.
(1986a)
Heinrich et al.
(1995)
Ishinishi et al.
(1988a)
(Light-duty
engine)
(Heavy-duty
engine)
Exposure
rate/duration Total Particle
(hr/vveek, exposure concentration
mo) time (hr) (mg/m3)
35, 30 4.20042004e
35,30 +15
35,30
35,30
80, 23 73607360736
80, 23 0
80,23
95, 35 1330013300
95,35
90, 24 8.64086409e
90,24 +15
90, 24
90,24
96, 30 1. 15201 152e
96, 30 +49
96,30
96,30
96,30
96,30
96,30
96,30
96,30
96,30
0
0.35
3.5
7.1
0
2.5
6.5
4.24
0
0.8
2.5
7.0
0
0.1
0.4
1.1
2.3
0
0.5
1.0
1.8
3.7
(mg-hr/m3)
Per week
0
12.25
122.5
248.5
0
200.0
520.0
402.8
0
72.0
225.0
630.0
0
9.6
38.4
105.6
220.8
0
48.0
96.0
172.8
355.2
Tumor
Total incidence (%)a
147014700298 0.9
20 1.3
3.6
12.8
1840047840 1.0
7.0
18.0
56392 17.8
740021800617 0
00 0
5.5
22.0
1.1524 3.3
60813e+37 2.4
0.8
4.1
2.4
0.8
0.8
0
3.3
6.5
-------
to
Table 7-8. Cumulative (concentration x time) exposure data for rats exposed to whole diesei exhaust (continued)
o
o
Cumulative exposure
Exposure
rate/duration Total Particle
(hr/week, exposure concentration
-4
i
UJ
CO
o
•n
I
O
O
H
o
H
tn
O
O
c
o
H
tn
Study
Brightwell et al.
(1989)
Kaplan et al.
(1983)
Iwaietal. (1986)
Takemoto et al.
(1986)
Karagianes et al.
(1981)
Iwaietal.(1997)
mo) time (hr)
80,24 7.680768 le+1
80, 24 5
80,24
80,24
140, 15 8.4008401e+l
140, 15 5
140, 15
140, 15
56, 24 53765376
56,24
16,18-24 1,152-1,536
16,18-24 1,152-1,536
30, 20 24002400
30,20
56, 24 53764992561
48, 24 6
54,24
(mg/m3)
0
0.7
2.2
6.6
0
0.25
0.75
1.5
4.9
0
2-4
8.3
9.4
3.2
5.1
(mg-hr/m3)
i
Tumor
Per week Total incidence (%)a
0 537616896506
56.0 88
176.0
528.0
0 210063001260
35.0 0
105.0
210.0
274.4 26342
0 0
32-64 3,456-4,608
249 19920
526154275 5.47041597e+l
4
1.2
0.7
9.7
38.5
0
3.3
10.0
3.3
36.8
0
16.6
421242
-------
Table 7-9. Evaluations of diesel exhaust as to human carcinogenic potential
Organization
NIOSH(1988)
IARC(1989)
IPCS (1996)
California EPA
(1998)
U.S. DHHS (2000)
Human data
Limited
Limited
N/Aa
"Consistent evidence
for a causal
association"
"Elevated lung
cancer in
occupationally
exposed groups"
Animal data
Confirmatory
Sufficient
N/A
"Demonstrated
carcinogenicity"
"Supporting animal
and mechanistic
data"
Overall
evaluation
Potential
occupational
carcinogen
Probably
carcinogenic to
humans
Probably
carcinogenic to
humans
DPM as a "toxic air
contaminant"
(California Air
Resources Board)
Reasonably
anticipated to be a
carcinogen
"Not applicable.
7/25/00
7-139
DRAFT—DO NOT CITE OR QUOTE
-------
0.5
RR •attmata* & 95* Cf
1 1.5
All Studies
CaseOontrof Studtes
Cohort Studios
Internal Comparison
Population
External Comparison
Population
Smoking Adjusted
Smoking Not Adjusted
Sub-analysis by
Occupation
Railroad Workers
Eauipmorrt Operators
Truck Drivers
Bus Workers
1 n
i a—
— .
i a
In •
LJ I
i a—
1 C
i i
. — i
^
1
Figure 7-1. Pooled relative risk estimates and heterogeneity-adjusted 95% confidence
intervals for all studies and subgroups of studies included in the meta-analysis.
Source: Bhatia et al., 1998.
7/25/00
7-140 UKAFi—UO NU1 Cl 11 UK U/UO1 Jb
-------
1.8
oi 1.6
•o
c
a
n 1.4
•o
e
"5
o
Q.
1 •
0.8
Categorias of Epidemiological Studies Included
Note. Cl = confidence interval; HWE = healthy worker effect.
Figure 1-2. Pooled estimates of relative risk of lung cancer in epidemiological studies
involving occupational exposure to diesel exhaust (random-effects models).
Source: Lipsett and Campleman, 1999.
7/25/00
7-141
DRAFT—DO NOT CITE OR QUOTE
-------
Diesel exhaust
Particulate matter
Carbon black
Exposure
Clearance
Macrophage
I Exposure
I Deposition
Desorption
Unique
to
diesel
Deposition
Organic chemicals
I
Reactive
oxygen
species
I
Cytokines
Growth factors
Proteases
Activation of
protooncogenes
Inflammation
Cell injury
Cell proliferation
Hyperplasia
Inactivation of tumor
suppressor genes
Fibrosis
Initiated cell
Preneoplastic
lesion
Malignant
tumor
Figure 1-5. Fathegenesis of lung disease in rats with chronic, high-level exposures to
particles.
Source: Modified from McClellan, 1997.
7/25/00
•7 1 /1
/ - i-rz
)RAFT—:
-------
7.8. REFERENCES
Adamson, IYR; Bowden, DH. (1978) Adaptive response of the pulmonary macrophagic system to carbon. II.
4 Morphologic studies. Lab Invest 38:430-438.
5
6 Ahlberg, J; Ahlbom, A; Lipping, H; et al. (1981) [Cancer among professional drivers—a problem-oriented register-
7 based study]. LSkartidningen 78:1545-1546.
8
9 Allen, RC; Loose, LD. (1976) Phagocytic activation of a luminol-dependent chemiluminescence in rabbit alveolar
10 and peritoneal macrophages. Biochem Biophys Commun 69:245-252.
11
12 Ames, BN. (1983) Dietary carcinogens and anticarcinogens. Science 221:1256-1264.
13
14 Aronsen, KJ; Siemiatycki, J; Dewar, R; et al. (1996) Occupational risk factors for prostate cancer: results from a
15 case-control study in Montreal, Quebec, Canada. Am J Epidemiol 143(4):363-373.
16
17 Balarajan, R; McDowell, ME. (1983) Malignant lymphomas and road transport workers. J Epidemiol Community
18 Health 37:316-317.
19
20 Balarajan, R; McDowall, ME. (1988) Professional drivers in London: a mortality study. Br J Ind Med 45:483-486.
21
22 Ball, JC; Green, B; Young, WC; et al., (1990) S9-activated Ames assays of diesel particle extracts: detecting
23 indirect-acting mutagens in samples that are direct-acting. Environ Sci Technol 24:890-894.
24
25 Beall, GD; Repine, JE; Hoidal, JR; et al. (1977) Chemiluminscence by human alveolar macrophages: stimulation
26 with heat killed bacteria or phorbol myristate acetate. Infect Immunol 17:117-120.
Benhamou, S; Benhamou, E; Flamant, R. (1988) Occupational risk factors of lung cancer in a French case-control
29 study. BrJ Ind Med 45:231-23 3.
30
31 Bender, AP; Parker, DL; Johnson, RA; et al. (1989) Minnesota Highway Maintenance Workers Study: cancer
32 mortality. Am J Ind Med 15:545-556.
33
34 Bhatia, R; Lopipero, P; Smith, A. (1998) Diesel exhaust exposure and lung cancer. Epidemiology 9(1):84-91.
35
36 Bitterman, PB; Aselberg, S; Crystal, RG. (1983) Mechanism of pulmonary fibrosis: spontaneous release of the
37 alveolar macrophage-derived growth factor in interstitial lung disorders. J Clin Invest 72:1801 -1813.
38
39 Boffetta, P; Stellman, SD. (1988) Association between diesel exhaust exposure and multiple myeloma: an example
40 of confounding. Prev Med 17:236-237.
41
42 Boffetta, P; Harris, RE; Wynder, EL. (1990) Case-control study on occupational exposure to diesel exhaust and lung
43 cancer risk. Am J Ind Med 17:577-591.
44
45 Bohning, DE; Atkins, HL; Conn, SH. (1982) Long-term particle clearance in man: normal and impaired. Ann
46 OccupHyg 26:259-271.
47
48 Bond, JA; Mitchell, CE; Li, AP. (1983) Metabolism and macromolecular covalent binding of benzo[a]pyrene in
49 cultured Fischer-344 rat lung type II epithelial cells. Biochem Pharmacol 32:3771-3776.
50
51 Bond, JA; Butler, MM; Medinsky, MA; et al. (1984) Dog pulmonary macrophage metabolism of free and particle-
52 associated ['"C]benzo[a]pyrene. J Toxicol Environ Health 14:181-189.
7/25/00 7-143 DRAFT—DO NOT CITE OR QUOTE
-------
1 Bond, JA; Johnson, NF; Snipes, MB; et al. (1990) DNA adduct formation in rat alveolar type II cells: cells
2 potentially at risk for inhaled diesel exhaust. Environ Mol Mutagen 16:64-69.
3
4 Boorman, GA; Brockman, M; Carlton, WW; et al. (1996) Classification of cystic keratinizing squamous lesions of
5 the rat lung:report of a workshop. Toxicol Pathol 24:564-572.
6
7 Borm, PJA; Knaapen, AM; Schins, RFP; et al. (1997) Neutrophils amplify the formation of DNA adducts by
8 benzo[a]pyrene in lung target cells. Environ Health Perspect 105 (Suppl 5): 1089-1093.
9
10 Bowden, DH; Adamson, IYR. (1978) Adaptive responses of the pulmonary macrophagic system to carbon. I.
11 Kinetic studies. Lab Invest 38:422-438.
12
13 Boyd, MR. (1984) Metabolic activation and lung toxicity: a basis for cell-selective pulmonary damage by foreign
14 chemicals. Environ Health Perspect 55:47-51.
15
16 Brightwell, J; Fouillet, X; Cassano-Zoppi, AL; et al. (1986) Neoplastic and functional changes in rodents after
17 chronic inhalation of engine exhaust emissions. In: Carcinogenic and mutagenic effects of diesel engine exhaust.
18 Ishinishi, N; Koizumi, A; McClellan, R; et al., eds. Amsterdam: Elsevier; pp. 471-485.
19
20 Brightwell, J; Fouillet, X; Cassano-Zoppi, AL; et al. (1989) Tumors of the respiratory tract in rats and hamsters
21 following chronic inhalation of engine exhaust emissions. J Appl Toxicol 9:23-31.
22
23 Briiske-Hohlfeld, I; Mohner, M; Ahrens, W; et al. (1999) Lung cancer risk in male workers occupationally exposed
24 to diesel motor emissions in Germany. Am J Ind Med 36:405-414.
25
26 Buiatti, E; Kriebel, D; Geddes, M; et al. (1985) A case control study of lung cancer in Florence, Italy. I.
27 Occupational risk factors. J Epidemiol Commun Health 39:244-250.
28
29 California Environmental Protection Agency. (Cal-EPA, OEHHA) (April, 1998) Health risk assessment for diesel
30 exhaust. Public and Scientific Review Draft.
31
32 Claxton, LD. (1983) Characterization of automotive emissions by bacterial mutagenesis bioassay: a review.
33 Environ Mutagen 5:609-631.
34
35 Coggon, D; Pannett, B; Acheson, ED. (1984) Use of job-exposure matrix in an occupational analysis of lung and
36 bladder cancers on the basis of death certificates. J Natl Cancer Inst 72:61-65.
37
38 Cohen, AJ; Higgins, MWP. (1995) Health effects of diesel exhaust: epidemiology. In: Diesel exhaust: a critical
39 analysis of emissions, exposure, and health effects. Cambridge, MA: Related HEI Publications; pp. 251-292.
40
41 Cohen, SM; Ellwein, LB. (1991) Genetic error, cell proliferation and carcinogenesis. Cancer Res 51:6493-6505.
42
43 Commins, BT; Walier, RE; Lawther, PJ. (1957) Air poiiution in diesel bus garages. Br J Ind Med 14:232-239.
44
45 Cuiciler, S; Clave1., J; Lirnassct, JC; at al. (1993) Occupational risks cf bladder car.cer in Francs: z. mult-center case-
46 control study. Int J Epidemiol 22:403-411.
47
48 Cox. LA. (1997) Does diesel exhaust cause human lung cancer? J Risk Anal 17(6):807-829.
49
50 Crump, KS. (1999) Lung cancer mortality and diesel exhaust: reanalysis of a retrospective cohort study of U.S.
51 railroad workers. Inhal Toxicol 11:1-17.
52
53 Crump, KS; Lambert, T; Chen, C; et al. (1991) Assessment of risk from exposure to diesel engine emissions.
54 Clement International Corporation. Prepared for U.S. EPA under contract no. 68-02-4601; 56.
55
7/25/00 7-144 DRAFT—DO NOT CITE OR QUO!
TT:
-------
Damber, LA; Larsson, LG. (1987) Occupation and male lung cancer: a case-control study in northern Sweden. Br J
Ind Med 44:446-453.
Daniel, JH. (1984) The use of diesel-powered equipment in U.S. underground coal operations. Presented at:
5 American Mining Congress International Coal Show; May, 1984; Chicago, IL. Washington, DC: Department of the
6 Interior, Bureau of Mines.
7
8 Dasenbrock-, C; Peters, L; Creutzenberg, O; et al. (1996) The carcinogenic potency of carbon particles with and
9 without PAH after repeated intratracheal administration in the rat. Toxicol Lett 88:15-21.
10
11 Depass, LR; Chen, KC; Peterson, LG. (1982) Dermal carcinogenesis bioassays of diesel particulates and
12 dichloromethane extract of diesel particulates in C3H mice. In: Toxicological effects of emissions from diesel
13 engines. Lewtas, J, ed. Amsterdam: Elsevier; pp. 321-326.
14
15 Dontenwill, W; Chevalier, H-J; Harke, H-P; et al. (1973) Investigations on the effects of chronic cigarette smoke
16 inhalation in Syrian golden hamsters. J Natl Cancer Inst 51:1681-1832.
17
18 Drath, DB; Kamovsky, ML. (1975) Superoxide production by phagocytic leukocytes. J Exp Med 141:257-262.
19
20 Driscoll, K. (1995) Role of inflammation in the development of rat lung tumors in response to chronic particle
21 exposure. In: Proceedings of the MIT toxicology symposium: particle overload in the rat lung and lung cancer:
22 relevance for human risk assessment. Cambridge, MA: Massachusetts Institute of Technology; 29 pp.
23
24 Driscoll, KE. (1996) Mechanisms of rat lung tumors: Role of persistent inflammation. Inhal Toxicol 8:139-153.
25
26 Driscoll, KE; Carter, JM; Howard, BW; et al. (1996) Pulmonary inflammatory, chemokine, and mutagenic
27 responses in rats after subchronic inhalation of carbon black. Toxicol Appl Pharmacol 136:372-380.
Driscoll, KE; Carter, JM; Hassenbein, DG; et al. (1997a) Cytokines and particle-induced inflammatory cell
30 recruitment. Environ Health Perspect 105 (Suppl 5): 1159-1164.
31
32 Driscoll, KE; Deyo, LC; Carter, JM; et al. (1997b) Effects of particle exposure and particle-elicited inflammatory
33 cells on mutation in rat alveolar epithelial cells. Carcinogenesis 18:423-430.
34
35 Edling, C; Anjou, C-G; Axelson, O; et al. (1987) Mortality among personnel exposed to diesel exhaust. Int Arch
36 Occup Environ Health 59:559-565.
37
38 El-Bayoumy, K; Rivenson, A; Johnson, B; et al. (1988) Comparative tumorigenicity of 1-nitropyrene, 1-
39 nitrosopyrene, and 1-aminopyrene administered by gavage to Sprague-Dawley rats. Cancer Res 48:4256-4260.
40
41 Emmelin, A; Nystrom, L; Wall, S. (1993) Diesel exhaust exposure and smoking: a case-referent study of lung
42 cancer among Swedish dock workers. Epidemiology 4:237-244.
43
44 Enya, T; Suzuki, H; Watanabe, T; et al. (1997) 3-Nitrobenzanthrone, a powerful bacterial mutagen and suspected
45 human carcinogen found in diesel exhaust and airborne particulates. Environ Sci Technol 31:2772-2776.
46
47 Evans, M; Shami, SG; Martinez, LA. (1986) Enhanced proliferation of pulmonary alveolar macrophages after
48 carbon instillation in mice depleted of blood monocytes by strontium-89. Lab Invest 54:154-159.
49
50 Farrell, RL; Davis, GW. (1974) Effect of particulate benzo(a)pyrene carrier on carcinogenesis in the respiratory
51 tract of hamsters. In: Experimental lung cancer. Karbe, E; Davis, GW, eds. Berlin: Springer; pp. 186-198.
52
• Flodin, U; Fredriksson, M; Persson, B. (1987) Multiple myeloma and engine exhausts, fresh wood, and creosote: a
case-referent study. Am J Ind Med 12:519-529.
55
7/25/00 7-145 DRAFT—DO NOT CITE OR QUOTE
-------
1 Gallagher, J; George, M; Kohan, M; et al. (1993) Detection and comparison of DNA adducts after in vitro and in
2 vivo diesel emission exposures. Environ Health Perspect 99:225-228.
3
4 Gallagher, J; Heinrich, U; George, M; et al. (1994) Formation of DNA adducts in rat lung following chronic
5 inhalation of diesel emissions, carbon black, and titanium dioxide particles. Carcinogenesis 15:1291-1299.
6
7 Garland, FC; Gorham, ED; Garland, CF; et al. (1988) Testicular cancer in US Navy personnel. Am J Epidemiol
8 127:411-414.
9
10 Garshick, E; Schenker, MB; Munoz, A; et al. (1987) A case-control study of lung cancer and diesel exhaust
11 exposure in railroad workers. Am Rev Respir Dis 135:1242-1248.
12
13 Garshick, E; Schenker, MB; Munoz, A; et al. (1988) A retrospective cohort study of lung cancer and diesel exhaust
14 exposure in railroad workers. Am Rev Respir Dis 137:820-825.
15
16 Gerde, P; Medinsky, MA; Bond, JA. (1991a) Particle-associated polycyclic aromatic hydrocarbons—a reappraisal of
17 their possible role in pulmonary carcinogenesis. Toxicol Appl Pharmacol 188:1-13.
18
19 Gerde, P; Medinsky, MA; Bond, JA. (1991b) The retention of polycyclic aromatic hydrocarbons in the bronchial
20 airways—a theoretical comparison. Toxicol Appl Pharmacol 108:1-13.
21
22 Goldstein, IM; Roos, D; Kaplan, HB; et al. (1975) Complement and immunoglobulins stimulate superoxide
23 production by human leukocytes independently of phagocytosis. J Clin Invest 56:1155-1163.
24
25 Grimmer, G; Brune, H; Deutsch-Wenzel, R; et al. (1987) Contribution of polycyclic aromatic hydrocarbons and
26 nitro-derivatives to the carcinogenic impact of diesel engine exhaust condensate evaluated by implantation into the
27 lungs of rats. Cancer Lett 37:173-180.
28
29 Guberan, E; Usel, M; Raymond, L; et al. (1992) Increased risk for lung cancer and for cancer of the gastrointestinal
30 tract among Geneva professional drivers. BrJ IndMed 49:337-344.
31
32 Gustafsson, L; Wall, S; Larsson, LG; et al. (1986) Mortality and cancer incidence among Swedish dock workers—a
33 retrospective cohort study. Scand J Work Environ Health 12:22-26.
34
35 Gustavsson, P; Plato, N; Lidstrom, EB; et al. (1990) Lung cancer and exposure to diesel exhaust among bus garage
36 workers. Scand J Work Environ Health 16:348-354.
37
38 Hall, NEL; Wynder, EL. (1984) Diesel exhaust exposure and lung cancer: a case-control study. Environ Res 34:77-
39 86.
40
41 Hansen, J; Raaschou-Nielsen, O; Olsen, JH; et al. (1998) Increased risk of lung cancer among different types of
42 professional drivers in Denmark. Occup Environ Med 55:115-118.
43
44 Hansen, ES. (1993) A follow-up study on the mortality of truck drivers. Am J Ind Med 23:811-821.
45
46 Hatch, GE; Gardner, DE; Menzel, DB. (1980) Stimulation of oxidant production in alveolar macrophages by
47 pollutant and latex particles. Environ Res 32:121-136.
48
49 Hayes, RB; Thomas, T; Silverman, DT; et al. (1989) Lung cancer in motor exhaust-related occupations. Am J Ind
50 Med 16:685-695.
51
52 Health Effects Institute. (nEI) (1995) Diesel exhaust: a critical analysis of emissions, exposure, and health ettects.
53 Cambridge, MA: Related HEI Publications; D. 27.
54
7/25/00 7-146 DRAFT—DO NOT CITE OR QUOTE
-------
Heinrich, U. (1990) Carcinogenicity of carbon black "Printex 90" in female Wistar rats. Presented at: U.S.
Environmental Protection Agency Diesel Emissions Peer Review Workshop; July 18-19; Chapel Hill, NC.
Heinrich, U. (1990) Presentation at workshop on health effects of diesel emissions. U.S. EPA, Research Triangle
5 Park, July 1990.
6
7 Heinrich, U; Peters, L; Funcke, W; et al. (1982) Investigation of toxic and carcinogenic effects of diesel exhaust in
8 long-term inhalation exposure of rodents. In: Lewtas, J, ed. Toxicological effects of emissions from diesel
9 emissions. Amsterdam: Elsevier; pp. 225-242.
10
11 Heinrich, U; Pott, F; Rittinghausen, S. (1986a) Comparison of chronic inhalation effects in rodents after long-term
12 exposure to either coal oven flue gas mixed with pyrolized pitch or diesel engine exhaust. In:. Carcinogenic and
13 mutagenic effects of diesel engine exhaust. Ishinishi, N; Koizumi, A; McClellan, R; St8ber, W, eds. Amsterdam:
14 Elsevier; pp. 441-457.
15
16 Heinrich, U; Muhle, H; Takenaka, S; et al. (1986b) Chronic effects on the respiratory tract of hamsters, mice and
17 rats after long-term inhalation of high concentrations of filtered and unfiltered diesel engine emissions. J Appl
18 Toxicol 6:383-395.
19
20 Heinrich, U; Peters, L; Fuhst, R; et al. (1989a) The effect of automotive exhaust exposure on the carcinogenicity of
21 diphenylnitrosamine (DPN) in the respiratory tract of rats. Exp Pathol 37:51-55.
22
23 Heinrich, U; Mohr, U; Fuhst, R; et al. (1989b) Investigation of a potential cotumorigenic effect of the dioxides of
24 nitrogen and sulfur, and of diesel-engine exhaust, on the respiratory tract of Syrian golden hamsters. Research
25 Report No. 26. Cambridge, MA: Health Effects Institute; pp. 27.
26
27 Heinrich, U; Dungworth, DL; Pott, F; et al. (1994) The carcinogenic effects of carbon black particles and tar-pitch
condensation aerosol after inhalation exposure of rats. Ann Occup Hyg 48 (suppl. 4):351-356.
30 Heinrich, U; Fuhst, R; Rittinghausen, S; et al. (1995) Chronic inhalation exposure of Wistar rats and two different
31 strains of mice to diesel engine exhaust, carbon black, and titanium dioxide. Inhal Toxicol 7:533-556.
32
33 Hemminki, K; Pershagen, G. (1994) Cancer risk of air pollution: epidemiological evidence. Environ Health Perspect
34 102(Suppl4):187-192.
35
36 Hemminki, K; Soderling, J; Ericson, P; et al. (1994) DNA adducts among personnel servicing and loading diesel
37 vehicles. Carcinogenesis 15:767-769.
38
39 Hill, AG. (1965) The environment and disease: association or causation? Proc R Soc Med 58:295-300.
40
41 Hirafuji, M; Sakakibara, M; Endo, T; et al. (1995) Biological effects of diesel exhaust particles (DEP) on tissues and
42 cells isolated from respiratory tracts of guinea pigs. Res Commun Mol Pathol Pharmacol 90:221-233.
43
44 Hirose, M; Lee, M-S; Wang, CY; et al. (1984) Induction of rat mammary gland tumors by 1-nitropyrene, a recently
45 recognized environmental mutagen. Cancer Res 44:1158-1162.
46
47 Hoar, SK; Hoover, R. (1985) Truck driving and bladder cancer mortality in rural New England. J Natl Cancer Inst
48 74:771-774.
49
50 Hoidal, JR; Repine, JE; Beall, GD; et al. (1978) The effect of phorbol myristate acetate on metabolism and
51 ultrastructure of human alveolar macrophages. Am J Pathol 91:469-482.
52
Hou, S-M; Lambert, B; Hemminki, K. (1995) Relationship between hprt mutant frequency, aromatic DNA adducts
and genotypes for GSTM1 and NAT2 in bus maintenance workers. Carcinogenesis 16:1913-1917.
55
7/25/00 7-147 DRAFT—DO NOT CITE OR QUOTE
-------
1 Howard, PC; Gerrard, JA; Milo, GE; et al. (1983) Transformation of normal human skin fibroblasts by 1-
2 nitropyrene and 6-nitrobenzo[a]pyrene. Carcinogenesis 4:353-355.
3
4 Howe, GR; Lindsay, JP. (1983) A follow-up study of a ten percent sample of the Canadian labour force: I. Cancer
5 mortality in males 1965 -1973.
6
7 Howe, GR; Burch, JD; Miller, AB; et al. (1980) Tobacco use, occupation, coffee, various nutrients, and bladder
8 cancer. J Natl Cancer Inst 64:701 -713.
9
10 Howe, GR; Fraser, D; Lindsay, J; et al. (1983) Cancer mortality (1965-77) in relation to diesel fume and coal
11 exposure in a cohort of retired railway workers. J Natl Cancer Inst 70:1015-1019.
12
13 Huisingh, J; Bradow, R; lungers, R; et al. (1978) Application of bioassay to the characterization of diesel particle
14 emissions. In: Application of short-term bioassay in the fractionation and analysis of complex mixtures. Waters,
15 MD; Nesnow, S; Huisingh, JL; et al., eds. New York: Plenum; pp. 381-418.
16
17 Humble, CG; Samet, JM; Skipper, BE. (1984) Comparison of self- and surrogate-reported dietary information. Am
18 JEpidemiol 119:86-98.
19
20 IARC. (International Agency for Research on Cancer) (1989) Monographs on the evaluation of carcinogenic risks to
21 humans. Vol. 46. Diesel and gasoline engine exhaust and some nitroarenes. Lyon, France: IARC.
22
23 Ichinose, T; Yajima, Y; Nagashima, M; et al. (1997a) Lung carcinogenesis and formation of 8-hydroxy-
24 deoxyguanosine in mice by diesel exhaust particles. Carcinogenesis 18:185-192.
25
26 Ichinose, T; Yamanushi, T; Seto, H; et al. (1997b) Oxygen radicals in lung carcinogenesis accompanying
27 phagocytosis of diesel exhaust particles. Int J Oncol 11:571-575.
28
29 International Life Sciences Institute (ILSI). (2000) ILSI Risk Science Institute workshop: the relevance of the rat
30 lung response to particle overload for human risk assessment. Gardner, DE, ed. Inhal Toxicol 12(1-2).
31
32 IPCS. (1996) International Program on Chemical Safety Environmental Health Criteria 171. Diesel fuel and exhaust
33 emissions. Geneva: World Health Organization, pp. 172-176.
34
35 Iscovich, J; Castelleto, R; Esteve, J; et al. (1987) Tobacco smoking, occupational exposure and bladder cancer in
36 Argentina. Int J Cancer 40:734-740.
37
38 Ishinishi, N; Kuwabara, N; Takaki, Y; et al. (1988a) Long-term inhalation experiments on diesel exhaust. In: Diesel
39 exhaust and health risks. Results of the HERP studies. Ibaraki, Japan: Research Committee for HERP Studies; pp.
40 11-84.
41
42 Ishinishi, N; Inamasu, T; Hisanaga, A; et al. (1988b) Intratracheal instillation study of diesel paniculate extracts in
43 hamsters. In: Diesel exhaust and health risk. Ibaraki, Japan: Research Committee for HERP Studies; pp. 209-216.
44
45 hvai, K.; Udagawa, T; Yarr.agishi, M; et a!. (1986) Long-term inhalation studies of diese! exhaust on F344 SPF rats.
46 Incidence of lung cancer and lymphoma. In: Ishinishi, N; Koizumi, A; McClellan, R; Stober, W, eds. Carcinogenic
47 and mutagenic effects of diesel engine exhaust. Amsterdam: Elsevier; pp. 349-360.
48
49 Iwai, K; Udwaga, T; Yamagishi, M; et al. (1986) Long-term inhalation studies of diesel exhaust on F344 SPF rats.
50 Dev Toxicol Environ Sci 13:349-360.
51
52 Iwai, K.; Higuchi, K; Udagawa, T; st al. (1997) Lung turner induced by long-term inhalation cr ir.tratrachea!
53 instillation of diesel exhaust particles. Exp Pathol Toxicol 49:393-401.
54
55 Iyer, V; Harris, RE; Wynder, EL. (1990) Diesel exhaust exposure and bladder cancer risk. Eur J Epidemiol 6:49-54.
7/25/00 7-148 DRAFT—DO NOT CITE OR QUOT
"C
-------
Jensen, OM; Wahrendorf, J; Knudsen, JB; et al. (1987) The Copenhagen case-referent study on bladder cancer:
risks among drivers, painters, and certain other occupations. Scand J Work Environ Health 13:129-134.
Jockel, KH; Ahrens, W; Jahn, I; et al. (1995) Untersuchungen zu Lungenkrebs und Risiken am Arbeitsplatz
5 (Schlussbericht). Berlin: Wirtschaftsverlag NW. FbOl HK546.
6
7 Jockel, KH; Ahrens, W; Jahn, I; et al. (1998) Occupational risk factors for lung cancer—a case-control study in
8 West Germany. Int J Epidemiol 27:549-560.
9
10 Johnson, C; Finkelstein, J; Gelein, R; et ai. (1993) Early alterations in the mRNA abundance of IL-la, IL-lp, iNOS,
11 MIP-2, TGFp, and VEGF associated with ultrafine particle exposure. Fundam Appl Toxicol 50:138 (abstr).
12
13 Jordana, M; Richards, C; Irving, LB; et al. (1988) Spontaneous in vitro release of alveolar macrophage cytokines
14 after the intratracheal instillation of bleomycin in rats: characterization and kinetic studies. Am Rev Respir Dis
15 137:1135-1140.
16
17 Kaplan, HL; Mackenzie, WF; Springer, KJ; et al. (1982) A subchronic study of the effects of exposure of three
18 species of rodents to diesel exhaust. In: Toxicological effects of emissions from diesel engines. Lewtas, J, ed. New
19 York: Elsevier Biomedical; pp. 161 -220.
20
21 Kaplan, HL; Springer, KJ; MacKenzie, WF. (1983) Studies of potential health effects of long-term exposure to
22 diesel exhaust emissions. Final Report No. 01-0750-103 (SWRI) and No. 1239 (SFRE). San Antonio, TX:
23 Southwest Research Institute.
24
25 Kaplan, I. (1959) Relationship of noxious gases to carcinoma of the lung in railroad workers. J Am Med Assoc
26 171:2039-2043.
27
Karagianes, RF; Palmer, RF; Busch, RH. (1981) Effects of inhaled diesel emissions and coal dust in rats. Am Ind
Hyg Assoc J. 42:382-391.
30
31 Kawabata, Y; Iwai, K; Udagawa, T; et al. (1986) Effects of diesel soot on unscheduled DNA synthesis of tracheal
32 epithelium and lung tumor formation. In: Carcinogenic and mutagenic effects of diesel engine exhaust. Ishinishi, N;
33 Koizumi, A; McClellan, R; et al., eds. Amsterdam: Elsevier; pp. 213-232.
34
35 Kawabata, Y; Udagawa, T; Higuchi, K; et al. (1994) Early one year exposure to diesel engine exhaust causes lung
36 tumors. Toxic and carcinogenic effects of solid particles on the respiratory tract. In: Toxic and carcinogenic effects
37 of solid particles in the respiratory tract. Mohr, U; Dungworth, DL; Mauderly, JL; et al., eds. Washington, DC:
38 International Life Sciences Institute Press, pp. 429-431.
39
40 Kensler, TW; Egner, PA; Moore, KG; et al. (1987) Role of inflammatory cells in the metabolic activity of
41 polycyclic aromatic hydrocarbons in mouse skin. Toxicol Appl Pharmacol 90:337-346.
42
43 Klebanoff, SJ; Clark, RA. (1978) The metabolic burst. In: The neutrophil: function and clinical disorders.
44 Amsterdam: North-Holland; pp. 283-368.
45
46 Kolonel, LN; Hirohata, T; Nomura, AMY. (1977) Adequacy of survey data collected from substitute respondents.
47 Am J Epidemiol 106:476-484.
48
49 Kotin, P; Falk, H; Thomas, M. (1955) Aromatic hydrocarbons. III. Presence in the paniculate phase of diesel engine
50 exhausts and the carcinogenicity of exhaust extracts. Arch Ind Health 11:113-120.
51
52 Kumagai, Y; Arimoto, T; Shinyashiki, M; et al. (1997) Generation of reactive oxygen species during interaction of
diesel exhaust particle components with NADPH cytochrome P450 reductase and involvement of the bioactivation
in the DNA damage. Free Radical Biol Med 22:479-487.
55
7/25/00 7-149 DRAFT—DO NOT CITE OR QUOTE
-------
1 Kunitake, E; Shimamura, K; Katayama, H; et al. (1986) Studies concerning carcinogenesis of diesel particulate
2 extracts following intratracheal instillation, subcutaneous injection, or skin application. In: Carcinogenic and
3 mutagenic effects of diesel engine exhaust. Ishinishi, N; Koizumi, A; McClellan, R; et al., eds. Amsterdam:
4 Elsevier; pp. 235-252.
5
6 Kunitake, E; Imase, A; Shimamura, K; et al. (1988) Skin application and subcutaneous injection experiments of
7 diesel particulate extracts using ICR mice and nude mice. In: Diesel exhaust and health risks. Results of the HERP
8 studies. Ibaraki, Japan: Research Committee for HERP Studies; pp. 217-225.
9
10 Langer, AM; Nolan, RP. (1994) Factors controlling the biological potential of inorganic dusts: surface area,
11 chemistry and character. In: Toxic and carcinogenic effects of solid particles in the respiratory tract. Mohr, U;
12 Dungworth, DL; Mauderly, JL; et al., eds. Washington, DC: International Life Sciences Institute Press, pp. 147-156.
13
14 Lechner, JF; Mauderly, JL. (1994) Sequence of events in lung carcinogenesis: initiation and promotion,
15 protooncogenes and tumor suppressor genes and particulates. In: Toxic and carcinogenic effects of solid particles
16 in the respiratory tract. Mohr, U; Dungworth, DL; Mauderly, JL; et al., eds. Washington, DC: International Life
17 Sciences Institute Press, pp. 235-251.
18
19 Lee, KP; Henry, NW, III; Trochimowicz, HJ; et al. (1986) Pulmonary response to impaired lung clearance in rats
20 following excessive TiO, dust deposition. Environ Res 41:144-167.
21
22 Lehnert, BE; Valdez, YE; Bomalaski, SH. (1988) Analyses of particles in the lung free cell, tracheobronchial lymph
23 nodal, and pleural space compartments following their deposition in the lung as related to lung clearance
24 mechanisms. Ann Occup Hyg 32:125-140.
25
26 Lehnert, BE; Ortiz, JB; London, JE; et al. (1989) Migratory behavior of alveolar macrophages during clearance of
27 light to heavy lung burdens of particles. 1989 Annual Meeting of the American Thoracic Society, Cincinnati, OH,
28 May 14-17.
29
30 Lerchen, ML; Samet, JM. (1986) An assessment of the validity of questionnaire responses provided by a surviving
31 spouse. Am J Epidemiol 123:481-489.
32
33 Lerchen, ML; Wiggins, CL; Samet, JM. (1987) Lung cancer and occupation in New Mexico. J Natl Cancer Inst
34 79:639-645.
35
36 Lewis, TR; Green, FHY; Moorman, WJ; et al. (1989) A chronic inhalation toxicity study of diesel engine emissions
37 and coal dust, alone and combined. J Am Coll Toxicol 8:345-375.
38
39 Lipsett, M; Campleman, S. (1999) Occupational exposure to diesel exhaust and lung cancer: a meta-analysis. Am
40 J Public Health 80(7): 1009-1017.
41
42 Lloyd, JW. (1971) Long-term mortality of steelworkers. V. Respiratory cancer in coke plant workers. J Occup Med
43 13:53-68.
44
45 LUWHC, DB, Abfer, VR.. (1977) Supcioxide production by iabbil aivcoloi iiidciophdges. Life Sci 21.1575-1584.
46
47 Mauderly, JL. (1994) Toxicological and epidemiological evidence for health risks from inhaled engine emissions.
48 Environ Health Perspect 102(suppl 4): 165-171.
49
50 Mauderly, JL; Jones, RK; Griffith, WC; et al. (1987) Diesel exhaust is a pulmonary carcinogen in rats exposed
51 chronically by inhalation. Fundam Appl Toxicol 9:208-221.
rr ^
•J4.
53 Mauderly, JL; Snipes, MB; Barr, EB; et al. (1991) Influence of particle-associated organic compounds on
54 carcinogenicity of diesel exhaust. Eighth Health Effects Institute Annual Conference, April 21-24, Colorado
55 Springs, CO.
7/25/00 7-150 DRAFT—DO NOT CITE OR QUOTE
-------
Mauderly, JL. (1994) Toxicological and epidemiological evidence for health risks from inhaled engine emissions.
Environ Health Perspec 102(suppl 4): 165-171.
Mauderly, JL; Banas, DA; Griffith, WC; et al. (1996) Diesel exhaust is not a pulmonary carcinogen in CD-I mice
5 exposed under conditions carcinogenic to F344 rats. Fundam Appl Toxicol 30:233-242.
6
7 McClellan, R. (1997) Use of mechanistic data in assessing human risks from exposure to particles. Environ Health
8 Perspect 105(Suppl 5):1-17.
9
10 McLaughlin, JK; Dietz, MS; Mehl, ES; et al. (1987) Reliability of surrogate information on cigarette smoking by
11 type of informant. Am J Epidemiol 126:144-146.
12
13 Miles, PR; Lee, P; Thrush, MA; et al. (1977) Chemiluminescence associated with phagocytosis of foreign particles
14 in rabbit alveolar macrophages. Life Sci 20:165-170.
15
16 Mohr, U; Takenaka, S; Dungworth, DL. (1986) Morphologic effects of inhaled diesel engine exhaust on lungs of
17 rats: comparison with effects of coal oven flue gas mixed with pyrolized pitch. In: Carcinogenic and mutagenic
18 effects of diesel engine exhaust. Ishinishi, N; Koizumi, A; McClellan, R; et al., eds. Amsterdam: Elsevier; pp. 459-
19 470.
20
21 Mokler, BV; Archibeque, FA; Beethe, RL; et al. (1984) Diesel exhaust exposure system for animal studies. Fundam
22 Appl Toxicol 4:270-277.
23
24 Mumford, JL; Chapman, RS; Harris, DB. (1989) Indoor air exposure to coal and wood combustion emissions
25 associated with a high lung cancer rate in Xuan Wei, China. Environ Int 15:315-320.
26
Murphy, SAM; Berube, KA; Richards, RJ. (1999) Bioreactivity of carbon black and diesel-exhaust particles to
primary Clara and type II epithelial cell cultures. Occup Environ Med 56:813-819.
30 Muscat, JE. (1996) Carcinogenic effects of diesel emissions and lung cancer: the epidemic logic evidence is not
31 causal. J Clin Epidemiol 49(8):891-892.
32
33 Muscat, JE; Wynder, EL. (1995) Diesel engine exhaust and lung cancer: an unproven association. Environ Health
34 Perspect 103:812-818.
35
36 Nagashima, M; Kasai, H; Yokota, J; et al. (1995) Formation of an oxidative DNA damage, 8-
37 hydroxydeoxyguanosine, in mouse lung DNA after intrarracheal instillation of diesel exhaust particles and effects of
38 high-dietary fat and beta-carotene on this process. Carcinogenesis 16:1441-1445.
39
40 National Institute for Occupational Safety and Health. (1988) Carcinogenic effects of exposure to diesel exhaust.
41 NIOSH Current Intelligence Bulletin 50. DHHS (NIOSH) Publication No. 88-116. Atlanta, GA: Centers for Disease
42 Control.
43
44 National Toxicology Program. (2000) National Institute for Environmental Health Science- National Institutes of
45 Health. Report on Carcinogens, 9th Edition. Research Triangle Park, NC.
46
47 Nesnow, S; Evans, C; Stead, A; et al. (1982) Skin carcinogenesis studies of emission extracts. In: Toxicological
48 effects of emissions from diesel engines. Lewtas, J, ed. Amsterdam: Elsevier; pp. 295-320.
49
50 Nesnow, S; Triplett, LL; Slaga, TJ. (1984) Tumor initiating activities of 1-nitropyrene and its nitrated products in
51 Sencar mice. Cancer Lett 23:1-8.
52
Nielsen, PS; Autrup, H. (1994) Diesel exhaust-related DNA adducts in garage workers. Clin Chem 40:1456-1458.
7/25/00 7-151 DRAFT—DO NOT CITE OR QUOTE
-------
1 Nielsen, PS; Andreassen, A; Farmer, PB; et al. (1996a) Biomonitoring of diesel exhaust-exposed workers. DNA
2 and hemoglobin adducts and urinary I-hydroxypyrene as markers of exposure. Toxicol Lett 86:27-37.
3
4 Nielsen, PS; de Pater, N; Okkels, H; et al. (1996b) Environmental air pollution and DNA adducts in Copenhagen
5 bus drivers—effect of (incomplete entry)
6
7 Nikula, KJ; Snipes, MB; Barr, EB; et al. (1995) Comparative pulmonary toxicities and carcinogenicities of
8 chronically inhaled diesel exhaust and carbon black in F344 rats. Fundam Appl Toxicol 25:80-94.
9
10 Nikula, KJ; Avila, KJ; Griffith, WC. (1997) Sites of particle retention and lung tissue responses to chronically
11 inhaled diesel exhaust and coal dust in rats and Cynomolgus monkeys. Environ Health Perspect 105(5):1231-1234.
12
13 Notani, PN; Shah, P; Jayant, K; et al. (1993) Occupation and cancers of the lung and bladder: a case-control study in
14 Bombay. Int J Epidemiol22:185-191.
15
16 Oberdorster, G; Yu, CP. (1990) The carcinogenic potential of inhaled diesel exhaust: a particle effect? J Aerosol Sci
17 21 (suppl. 1):S397-S401.
18
19 Oberdorster, G; Baggs, R; Gelein, R; et al. (1995) Pulmonary effects of inhaled carbon black in rats. Toxicologist
20 15:46.
21
22 Orthoefer, JG; Moore, W; Kraemer, D. (1981) Carcinogenicity of diesel exhaust as tested in strain A mice. Environ
23 Int 5:461-471.
24
25 Pepelko, WE; Peirano, WB. (1983) Health effects of exposure to diesel engine emissions. J Am Coll Toxicol 2:253-
26 306.
27
28 Percy, C; Stanek, E, III; Gloeckler, L. (1981) Accuracy of cancer death certificates and its effect on cancer mortality
29 statistics. Am J Public Health 71:242-250.
30
31 Phillips, BJ; James, TEB; Anderson, D. (1984) Genetic damage in CHO cells exposed to enzymatically generated
32 active oxygen species. Mutat Res 126:265-271.
33
34 Pott, F; Stober, W. (1983) Carcinogenicity of airborne combustion products observed in subcutaneous tissue and
35 lungs of laboratory rodents. Environ Health Perspect 47:293-303.
36
37 Pott, F; Heinrich, U. (1990) Relative significance of different hydrocarbons for the carcinogenic potency of
38 emissions from various incomplete combustion processes. In: complex mixtures and scientific risk. Scientific
39 publication 104. Vaino, A; Sorsa, M; McMichaei, AJ; eds. Lyon, France: IARC; pp. 288-297.
40
41 Pott, F; Heinrich, U; Vainio, H; et al. (1990) Complex mixtures and cancer risk. In: Relative significance of
42 hydrocarbons for the carcinogenic potency of emissions from various incomplete combustion processes. IARC,
43 288-297.
44
A^ D*-\-H- P- O»"1<"">T»f*'*'l^ T\l - U^inf^r-K IT* ot «al (1 QQ/t."\ T urtrr tiimruii-c in rote aft^r intToiT5>'*^>^^>^ inctiliatirm r»frinoto Ann
I *•> *. W Vk.) •. } L^> M**^ « > V>*. «») ~-" ^- •, »^W,*.»»•»*, V} *•"• —». ^ • S * I j .L-.^..*^ ..U... W i». — i-.. . — i._- —it —. lii— _W — —-»W_. *-.__--,!_—-i— --. W -. —-ij-*. . . ^*^I.
46 Occup Hyg 38:357-363, Suppl. 1.
47
48 Qu, SX; Leigh, J; Koelmeyer, H; et al. (1996) Detection of DNA adducts in coal miners exposed to diesel emissions
49 in Australia. Fundam Appl Toxicol 50:237 (abstr).
50
51 Raffle, PAB. (1957) The health of the worker. Br J Ind Med 14:73-80.
52
53 Riebe-Imre, M; Aufderheide, M; Gartner-Hubsch, S; et al. (1994) Cytotoxic and genotoxic effects of insoluble
54 poarticles in vitro. In: Toxic and carcinogenic effects of solid particles in the respiratory tract. Mohr, U;
-------
1 Dungworth, DL; Mauderly, DL; et al., eds. Washington, DC: International Life Sciences Institute Press, pp. 519-523.
Risch, HA; Burch, JD; Miller, AB; et al. (1988) Occupational factors and the incidence of cancer of the bladder in
Canada. Br J Ind Med 45:361-367.
5
6 Rister, M; Baehner, RL. (1977) Effect of hyperoxia on superoxide anion and hydrogen peroxide production of
7 polymorphonuclear leukocytes and alveolar macrophages. Br J Haematol 36:241-248.
8
9 Rittinghausen, S; Mohr, U; Dungworth, DL. (1997) Pulmonary cystic keratinizing squamous cell lesions after
10 inhalation/instillation of different particles. Exper Toxicol Pathol 49:433-446.
11
12 Rogot, E; Reid, DD. (1975) The validity of data from next-of-kin in studies of mortality among migrants. Int J
13 Epidemiol 4:51-54.
14
15 Root, RK; Metcalf, JA. (1977) H2O2 release from human granulocytes during phagocytosis. J Clin Invest 60:1266-
16 1279.
17
18 Rosenblatt, MB; Teng, PK; Kerpe, S; et al. (1971) Causes of death in 1,000 consecutive autopsies. NY State J Med
19 71:2189-2193.
20
21 Rosenkranz, HS. (1996) Mutagenic nitroarenes, diesel emissions, particle-induced mutations and cancer: an essay
22 on cancer causation by a moving target. Mutat Res 367:65-72.
23
24 Rosenkranz, HS; Mermelstein, R. (1983) Mutagenicity and genotoxicity of nitroarenes: all nitro-containing
25 chemicals were not created equal. Mutat Res 114:217-267.
26
27 Rothman, KJ. (1986) Modem epidemiology. Boston/Toronto: Little, Brown and Co.; pp. 16-21.
Rushton, L; Alderson, MR. (1983) Epidemiological survey of oil distribution centres in Britain. Br J Ind Med
30 40:330-339.
31
32 Rushton, L; Alderson, MR; Nagarajah, CR. (1983) Epidemiological survey of maintenance workers in London
33 Transport Executive bus garages at Chiswick Works. Br J Ind Med 40:340-345.
34
35 Sagai, M; Saito, H; Ichinose, T; et al. (1993) Biological effects of diesel exhaust particles. I. In vitro production of
36 superoxide and in vivo toxicity in mouse. Free Radical Biol Med 14:37-47.
37
38 Salmeen, IT; Pero, AM; Zator, R; et al. (1984) Ames assay chromatograms and the identification of mutagens in
39 diesel exhaust particle extracts. Environ Sci Technol 18:375-382.
40
41 Sandusky, CB; Cowden, MW; Schwartz, SL. (1977) Effect of particle size on regurgitive exocytosis by rabbit
42 alveolar macrophages. In: Pulmonary macrophages and epithelial cells. CONF-760927, National Technical
43 Information Service, U.S. Dept. of Commerce, pp. 85-105.
44
45 Saverin. R; Braunlich, A; Dahman, D; et al. (1999) Diesel exhaust and lung cancer mortality in potash mining. Am
46 J Ind Med 36:415-422.
47
48 Scheepers, PJT; Bos, RP. (1992) Combustion of diesel fuel from a toxicological perspective. II. Toxicity. Int Arch
49 Occup Environ Health 64:163-177.
50
51 Schenker, MB; Smith, T; Munoz, A; et al. (1984) Diesel exposure and mortality among railway workers: results of a
52 pilot study. Br J Ind Med 41:320-327.
Schottenfeld, D; Eaton, M; Sommers, SC; et al. (1982) The autopsy as a measure of accuracy of the death
55 certificate. Bull NY Acad Med 58:778-794.
7/25/00 7-153 DRAFT—DO NOT CITE OR QUOTE
-------
1 Sera, N; Fukuhara, K; Miyata, N; et al. (1994) Detection of nitro-azabenzol[a]pyrene derivatives in the semivolatile
2 phase originating from aibome paniculate matter, diesel and gasoline vehicles. Mutagenesis 9:47-52.
3
4 Shefrier, AM; Collins, SC; Dooley, L; et al. (1982) Respiratory carcinogenicity of diesel fuel emissions interim
5 results. In: Toxicological effects of emissions from diesel engines. Lewtas, J, ed. Amsterdam: Elsevier; pp. 329-350.
6
7 Siemiatycki, J; Gerin, M; Stewart, P; et al. (1988) Associations between several sites of cancer and ten types of
8 exhaust and combustion products: results from a case-referent study in Montreal. Scand J Work Environ Health
9 14:79-90.
10
11 Silverman, DT; Hoover, RN; Albert, S; et al. (1983) Occupation and cancer of the lower urinary tract in Detroit. J
12 Natl Cancer Inst 70:237-245.
13
14 Silverman, DT; Hoover, RN; Mason, TJ; et al. (1986) Motor exhaust-related occupations and bladder cancer.
15 Cancer Res 46:2113-2116.
16
17 Steeniand, K. (1986) Lung cancer and diesel exhaust: a review. Am J Ind Med 10:177-189.
18
19 Steeniand, K; Burnett, C; Osoria, AM. (1987) A case-control study of bladder cancer using city directories as a
20 source of occupational data. Am J Epidemiol 126:247-257.
21
22 Steeniand, NK; Silverman, DT; Homung, RW. (1990) Case-control study of lung cancer and truck driving in the
23 Teamsters Union. Am J Public Health 80:670-674.
24
25 Steeniand, K; Deddens, J; Stayner, L. (1998) Diesel exhaust and lung cancer in the trucking industry: exposure-
26 response analyses and risk assessment. Am J Ind Med 34:220-228.
27
28 Steineck, G; Plato, N; Gerhardsson, M; et al. (1990) Increased risk of urothelial cancer in Stockholm during 1985-
29 87 after exposure to benzene and exhausts. Int J Cancer 45:1012-1017.
30
31 Stenback, F; Rowland, J; Sellakumar, A. (1976) Carcinogenicity of benzo[a]pyrene and dusts in the hamster lung
32 (instilled intratracheally with titanium oxide, aluminum oxide, carbon and ferric oxide). Oncology 33:29-34.
33
34 Stenberg, U; Alsberg, T; Westerholm, R. (1983) Emission of carcinogenic components with automobile exhausts.
35 Environ Health Perspect 47:53-63.
36
37 Stem, FB; Lemen, RA; Curtis, RA. (1981) Exposure of motor vehicle examiners to carbon monoxide: a historical
38 prospective mortality study. Arch Environ Health 36:59-66.
39
40 Stober, W. (1986) Experimental induction of tumors in hamsters, mice and rats after long-term inhalation of filtered
41 and unfiltered diesel engine exhaust. In: Carcinogenic and mutagenic effects of diesel engine exhaust. Ishinishi, N;
42 Koizumi, A; McClellan, R; et al., eds. Amsterdam: Elsevier; pp. 421-429.
43
44 Stober, W; Abel, UR. (1996) Lung cancer due to diesel soot particles in ambient air? A critical appraisal of
45 epidcmiological studies addressing this question. Int Arch Occup Environ Health 6S(Suppl).S3-S61.
46
47 Strom, KA. (1984) Response of pulmonary cellular defenses to the inhalation of high concentrations of diesel
48 exhaust. J Toxicol Environ Health 13:919-944.
49
50 Swanson, GM; Lin, CS; Burns, PB. (1993) Diversity in the association between occupation and lung cancer among
51 black and white men. Cancer Epidemiol Biomarkers Prev 2:313-320.
r ^
•Jf.
53 Takaki, Y; Kitamura, S; Kuwabara, N; et al. (1989) Long-term inhalation studies of exhaust from diesel engine in F-
54 344 rats: the quantitative relationship between pulmonary hyperplasia and anthracosis. Exp Pathol 37:56-61.
55
7/25/00 7-154 DRAFT—DO NOT CITE OR QUOTE
-------
Takemoto, K; Yoshimura, H; Katayama, H. (1986) Effects of chronic inhalation exposure to diesel exhaust on the
development of lung tumors in di-isopropanol-nitrosamine-treated F344 rats and newborn C57BL and ICR mice. In:
Carcinogenic and mutagenic effects of diesel engine exhaust. Ishinishi, N; Koizumi, A; McClellan, R; et al., eds.
Amsterdam: Elsevier; pp. 311-327.
5
6 Takemoto, K; Katayama, H; Kuwabara, T; et al. (1988) Carcinogenicity by subcutaneous administration of diesel
7 paniculate extracts in mice. In: Diesel exhaust and health risks. Ibaraki, Japan: Research Committee for HERP
8 . Studies; pp. 227-234.
9
10 Thyssen, J; Althoff, J; Kimmerle, G; et al. (1981) Inhalation studies with benzo[a]pyrene in Syrian golden hamsters.
11 J Natl Cancer Inst 66:575-577.
12
13 Trotter, JR. (1980) Spontaneous cancer and its possible relationship to oxygen metabolism. Proc Natl Acad Sci
14 77:1763-1767.
15
16 U.S. Environmental Protection Agency (U.S. EPA). (1993) Motor vehicle-related air toxics study. Ann Arbor, MI:
17 Office of Mobile Sources. EPA/420/R-93/005. Available from National Technical Information Service, Springfield,
18 VA, PB93-182590/XAB.
19
20 U.S. EPA. (1996a) Air quality criteria for paniculate matter. Prepared by the National Center for Environmental
21 Assessment, U.S. Environmental Protection Agency, Research Triangle Park, NC. EPA/600/P-95/001bF.
22
23 U.S. EPA. (1996b) Proposed guidelines for carcinogen risk assessment. Office of Research and Development.
24 Federal Register 61 (79): 17960-18011. EPA/600/P-92/003C.
25
26 U.S. EPA. (1999) Guidelines for carcinogen risk assessment-review draft. NCEA-F-0644 (July 1999).
27
»Vineis, P; Magnani, C. (1985) Occupation and bladder cancer in males: a case-control study. Int J Cancer 35:599-
606.
30
31 Vostal, JJ. (1986) Factors limiting the evidence for chemical carcinogenicity of diesel emissions in long-term
32 inhalation experiments. In: Carcinogenic and mutagenic effects of diesel engine exhaust: proceedings of the
33 international satellite symposium on toxicological effects of emissions from diesel engines; July 1986; Tsukuba
34 Science City, Japan. Ishinishi, N; Koizumi, A; McClellan, RO; et al., eds. (Developments in toxicology and
35 environmental science: v. 13.) Amsterdam: Elsevier Science Publishers BV; pp. 381-396.
36
37 Waller, RE. (1981) Trends in lung cancer in London in relation to exposure to diesel fumes. Environ Int 5:479-483.
38
39 Ward, JM. (1983) Background data and variations in tumor rates of control rats and mice. Prog Exper Tumor Res
40 26:241-258.
41
42 Waxweiler, RJ; Wagoner, JK; Archer, VE. (1973) Mortality of potash workers. J Occup Med 15:486-489.
43
44 Weinstein, IB; Troll, W. (1977) National Cancer Institute workshop on tumor promotion and cofactors in
45 carcinogenesis. Cancer Res 3 7:3461 -3463.
46
47 Weiss, SJ; Rustagei, PK; LoBuglio, AF. (1978) Human granulocyte generation of hydroxyl radical. J Exp Med
48 147:316-323.
49
50 Weitzman, SA; Stossel, TP. (1981) Mutation caused by human phagocytosis. Science 212:546-547.
51
52 White, HJ; Vostal, JJ; Kaplan, HL; et al. (1983) A long-term inhalation study evaluates the pulmonary effects of
diesel emissions. J Appl Toxicol 1:332.
7/25/00 7-155 DRAFT—DO NOT CITE OR QUOTE
-------
1 Wichmann, HE; Kreienbrock, L; Kreuzer, M; et al. (1998) Lungenkrebrisiko durch Radon in der Bundesrepublik
2 Deutschland (West). In: Fortschritte in der Umwelmedizin, ecomed Verlagsgesellschaft
3 Landsberg/Lech. Wichmann, HE; SchliepkSter, HW; Fulgraff, G; et al., eds.
4
5 Williams, RR; Stegens, NL; Goldsmith, JR. (1977) Associations of cancer site and type with occupation and
6 industry from the Third National Cancer Survey interview. J Natl Cancer Inst 59:1147-1185.
7
8 Wong, D; Mitchell, CE; Wolff, RK; et al. (1986) Identification of DNA damage as a result of exposure of rats to
9 diesel engine exhaust. Carcinogenesis 7:1595-1597.
10
11 Wong, O; Morgan, RW; Kheifets, L; et al. (1985) Mortality among members of a heavy construction equipment
12 operators union with potential exposure to diesel exhaust emissions. Br J Ind Med 42:435-448.
13
14 Woskie, SR; Smith, TJ; Hammond, SK; et al. (1988a) Estimation of the diesel exhaust exposures of railroad
15 workers: I. Current exposures. Am J Ind Med 13:381-394.
16
17 Woskie, SR; Smith, TJ; Hammond, SK; et al. (1988b) Estimation of the diesel exhaust exposures of railroad
18 workers: II. National and historical exposures. Am J Ind Med 13:395-404.
19
20 Wynder, LE; Dieck, G; Hall, ME. (1985) A case control study of diesel exhaust exposure and bladder cancer.
21 Environ Res 34:475-489.
22
23 Zaebst, DD; Clapp, DE; Blade, LM; et al. (1991) Quantitative determination of trucking industry workers'
24 exposures to diesel exhaust particles. Am Ind Hyg Assoc J 52:529-541.
25
26 Zamora, PO; Gregory, RE; Brooks, AL. (1983) In vitro evaluation of the tumor-promoting potential of diesel-
27 exhaust-particle extracts. J Toxicol Environ Health 11:188-197.
7/25/00 7-156 DRAFT—DO NOT CITE OR QUOTE
-------
8. DOSE-RESPONSE ASSESSMENT: CARCINOGENIC EFFECTS
t
8.1. INTRODUCTION
2 Dose-response assessment defines the relationship between the exposure/dose of an agent
3 and the degree of carcinogenic response, and evaluates potential cancer risks to humans at
4 exposure/dose levels of interest. Most often, the exposure-dose-response of interest is well
5 below the range of observation. As a result, dose-response assessment usually entails an
6 extrapolation from the generally high exposures in studies in humans or laboratory animals to the
7 exposure levels expected from human contact with the agent in the environment. It also includes
8 considerations of the scientific validity of these extrapolations based on available knowledge
9 about the underlying mechanisms or modes of carcinogenic action. The complete sequence of
10 biological events that must occur to produce an adverse effect is defined as "mechanism of
11 action." In cases where only partial information is available, the term "mode of action" is used to
12 refer to the mechanisms for key events that are judged to be sufficient to inform about the shape
13 of the dose-response curve beyond the range of observation.
14 This chapter evaluates the available exposure-dose-response data, discusses extrapolation
15 issues in estimating the cancer risk of environmental exposure to diesel exhaust (DE). It is
16 concluded that available data are inadequate to confidently derive a cancer unit risk estimate for
^P DE or its component, diesel paniculate matter (DPM). Unit risk is one possible output from a
18 dose-response assessment and is defined as the estimated upper-bound cancer risk at a specific
19 exposure or dose from a continuous average lifetime exposure of 70 years (in this case, cancer
20 risk per ng/m3 of DPM).. In lieu of unit-risk-based quantitative risk estimates, this chapter
21 provides some perspective about potential risk at environmental levels. Approaches to dose-
22 response assessment for DE follow EPA's guidelines for carcinogen risk assessment (U.S. EPA,
23 1986, 1996).
24 Subsequent sections of this chapter discuss issues related to dose-response evaluation of
25 human cancer risk to DE. including the target tumor site and underlying mode of action, suitable
26 measures of dose, approaches to low-dose extrapolation, and appropriate data to be used in the
27 dose-response analysis. This is followed by a simple analysis of the possible degree and extent
28 of risk from environmental exposure to DE. Appendix D provides a summary review of dose-
29 response assessments conducted to date by other organizations and investigators.
30
31 8.2. MODE OF ACTION AND DOSE-RESPONSE APPROACH
32 According to EPA's 1996 Proposed Guidelines for Carcinogen Risk Assessment, dose-
t response assessment is performed in two steps: assessment of observed data to derive a point of
departure, followed by extrapolation to lower exposures to the extent necessary. Human data are
7/25/00 8-1 DRAFT—DO NOT CITE OR QUOTE
-------
1 always preferred over animal data, if available, as their use obviates the need for extrapolation
2 across species. Mode of action information is critical to dose-response evaluation, as it informs
3 about the relevance of animal data to assessment of human hazard and risk, the shape of the
4 dose-response curve at low doses, and the most appropriate measure(s) of dose and response.
5 If there are sufficient quantitative data (humans and/or animals) and adequate
6 understanding of the carcinogenic process, the preferred approach is to use a biologically based
7 model for both the range of observation and extrapolation below that range. Otherwise, as a
8 default procedure, a standard mathematical model is used to curve-fit the observed dose-response
9 data to obtain a point of departure, which is the lower 95% confidence limit of the lowest
10 exposure/dose that is associated with a selected magnitude of excesses of cancer risk in human or
11 animal studies. Default approaches for low-dose extrapolation should be consistent with current
12 understanding of the mode(s) of action. These include approaches that assume linearity or
13 nonlinearity, or both. Linear extrapolation is used when there is insufficient understanding of the
14 modes of action, or the mode of action information indicates that the dose-response curve at low-
15 dose is, or is expected to be, linear. Linear extrapolation involves the calculation of the slope of
16 the line drawn from the point of departure to zero exposure or dose (i.e., above background).
17 When there is sufficient evidence for a nonlinear mode of action but not enough data to construct
18 a biologically based model for the relationship, a margin of exposure is used as a default
19 approach. A margin-of-exposure analysis compares the point of departure (i.e., the lowest
20 exposure associated with some cancer risk) with the dose associated with the environmental
21 exposure(s) of interest and determines whether or not the exposure margins are adequate. Both
22 default approaches may be used for a tumor response, if it is mediated by linear and nonlinear
23 modes of action.
24 As reviewed in Chapter 7, there is substantial evidence from combined human and
25 experimental evidence that DE likely poses a cancer hazard to humans at anticipated levels of
26 environmental exposure. The critical target organ is the lung. Limited evidence exists for a
27 casual relationship between risk for lung cancer and occupational exposure to DE in certain
28 occupational workers such as railroad workers, truck drivers, heavy equipment operators, transit
29 workers, etc. In addition, it has been shown unequivocally in several studies that DE can cause
30 benign and malignant lung tumors in rats in a dose-related manner following chronic inhalation
31 exposure to sufficiently high concentrations.
32 The mechanism(s) by which DE induces lung cancer in humans has not been established.
33 As discussed in Section 7.4, several modes of action have been postulated based on available
34 mechanistic studies, including direct DNA effects (gene mutations) by the adsorbed organic
35 compounds and the gaseous fractions, indirect DNA effects (e.g., chromosomal aberrations,
36 sister chromatid exchange [SCE], micronuclei) by DE and DPM, oxidative DNA damage by
7/25/00 8-2 DRAFT—DO NOT CITE OR QUOTE
-------
DPM via release of reactive oxygen species (ROS), and particle-induced chronic inflammatory
response leading to epithelial cell cytotoxicity and regenerative cell proliferation via release of
3 cytokines, growth factors, and ROS. It is likely that a combination of modes of action contribute
4 to the overall carcinogenic activity of DE, and that the relative contribution of the various modes
5 of action may vary with different exposure levels.
6 In the absence of a full understanding of the relative roles of DE constituents in inducing
7 lung cancer in humans, and because there is some evidence for a mutagenic mode of action, this
8 assessment takes the position that linear low-dose extrapolation is most appropriate and prudent
9 (U.S. EPA, 1986, 1996). It should be noted that other individuals and organizations have used
10 either linear risk extrapolation models and or mechanistically based models to estimate cancer
11 risk from environmental exposure to DE (e.g., IPCS, 1996; Cal EPA, 1998; also see Appendix
12 D).
13 On the other hand, there is an adequate understanding of how DE causes lung tumors in
14 the rat under experimental exposure conditions. Prolonged exposure to high concentrations of a
15 variety of insoluble particles including DPM (and its carbon core, devoid of organics) causes
16 lung tumors in rats through a mode of action that involves impairment of lung clearance
17 mechanisms (referred to as "lung overload response"), leading to persistent chronic
^^ inflammation, cell proliferation, metaplasia, and ultimately the development of lung tumors
19 (ILSI, 2000). Because this mode of action is not expected to be operative at environmental
20 exposure conditions, the rat lung tumor dose-response data are not considered suitable for
21 predicting human risk at low environmental exposure concentrations.
22
23 8.3. USE OF EPIDEMIOLOGIC STUDIES FOR QUANTITATIVE RISK ASSESSMENT
24 As discussed above, human data are considered more appropriate than animal data in
25 estimating environmental cancer risk for DE. Still, there are many uncertainties in using the
26 available epidemiologic studies that have quantitative exposure data to extrapolate the risk to the
27 general population for ambient-level DE exposure.
28
29 8.3.1. Sources of Uncertainty
30 The greatest uncertainty in estimating DE-induced cancer risk from epidemiologic studies
31 is the lack of knowledge of actual historical exposures for individual workers, particularly for the
32 early years. Reconstruction of historic exposures are based on job exposure categories, industrial
33 hygiene measurements, and assumptions made about exposure patterns.
Another related uncertainty is the choice of markers of exposure to DE. As discussed
above, the modes of action for DE-induced lung cancer in humans are not fully understood, and
7/25/00 8-3 DRAFT—DO NOT CITE OR QUOTE
-------
1 thus the best measure of DE exposure is unknown. Various markers of DPM (e.g., respirable-
2 sized particles, elemental carbon [EC]) have been used as dosimeters for DE. Though EC is
3 more sensitive and more specific than respirable-sized particles, both are considered appropriate
4 dosimeters. Related to the choice of dosimeter, having a relatively constant relationship between
5 the organics (on the particle) and the particle mass would be consistent with a possible mode of
6 action role for both the particle and organic components. However, evidence of such a constant
7 historic relationship remains unclear. As discussed in Chapter 2 (Section 2.5.2), it appears that
8 newer model on-road engine exhaust may have somewhat less organics adsorbed onto the
9 particle compared with older model engines. On the other hand, with regard to DE in the
10 ambient air, there is significant variation of the amounts of DPM organic emitted because of
11 aged vehicles in the on-road fleet, driving patterns, and the additional presence of nonroad DE
12 (e.g., marine vessels and locomotives, which generally use older technology than on-road
13 engines).
14 Another major uncertainty associated with many of the DE epidemiologic studies was the
15 inability to fully control for smoking effects, resulting in possible errors in estimating relative
16 risk increases. Changes in adjustments for smoking could result in considerable changes in
17 relative risk because smoking has a much larger effect on relative lung cancer risk than is likely
18 for DE. It is difficult to effectively control for a smoking effect in a statistical analysis because
19 cigarette smoke contains an array of biologically active compounds and affects multiple steps of
20 carcinogenesis, thus probably making smokers more susceptible to DE-induced lung cancer than
21 are nonsmokers. A traditional statistical analysis (e.g., logistic regression) would not be able to
22 adjust for such an effect. Although both case-control and cohort studies are subjected to the
23 same difficulty, controlling for smoking effects is more problematic in case-control studies than
24 in cohort studies because a majority of the lung cancer cases (about 85%; U.S. Surgeon General,
25 1982) are usually also smokers.
26 Another uncertainty is the use of occupational worker data to extrapolate cancer hazard
27 risk to the general population and sensitive subgroups. By sex, age, and general health status,
28 workers are not fully representative of the general population. There is virtually no information
23 to determine whether infants and children or people in poor health respond differently to DE
30 exposure than do workers. Finally, the use of linear low-dose extrapolation may contribute
31 significantly to uncertainly in estimating environmental risks.
32
7/25/00 8-4 DRAFT—DO NOT CITE OR QUOTE
-------
f 8.3.2. Evaluation of Key Epidemiologic Studies for Potential Use in Quantitative Risk
Estimates
3 Among the available epidemiologic studies, only the railroad worker studies and the
4 Teamster truck driver studies have quantitative exposure data for possible use in deriving a unit
5 risk estimate for DE-induced lung cancer. This section evaluates the strengths and limitations of
6 these data and their suitability for dose-response analysis.
7
8 8.3.2.1. Railroad Worker Studies
9 Garshick and colleagues conducted both cohort and case-control studies of lung cancer
10 mortalities among U.S. railroad workers registered with the U.S. Railroad Retirement Board
11 (RRB).
12 In the cohort study (Garshick et al., 1988), lung cancer mortality was ascertained through
13 1980 in 55,407 railroad workers, age 40 through 64 in 1959, with at least 10 years of work in
14 selected railroad jobs (39 job titles). The cohort was selected on the basis of job titles in 1959.
15 Industrial hygiene evaluations and descriptions of job activities were used to classify jobs as
16 exposed or unexposed to diesel emissions. Workers with recognized asbestos exposure were
17 excluded from the job categories selected for study. However, a few jobs with some potential for
^P asbestos exposure were included in the cohort. Each subject's work history was determined from
19 a yearly job report filed by his employer with the RRB from 1959 until death or retirement. The
20 year 1959 was chosen as the effective start of DE exposure for this study because by this time
21 95% of the locomotives in the United States were diesel powered. The author reported
22 statistically significant relative risk increases of 1.57 for the 40-44 year age group and 1.34 for
23 the 45-49 year age group, after exclusion of workers exposed to asbestos and controls for
24 smoking. Age groups were determined by their ages in 1959.
25 A main strength of the cohort study is the large sample size (55,407), which allowed
26 sufficient power to detect small risks. This study also permitted the exclusion of workers with
27 potential past exposure to asbestos. The stability of job career paths in the cohort ensured that of
28 the workers 40 to 64 years of age in 1959 classified as DE-exposed, 94% of the cases were still in
29 DE-exposed jobs 20 years later.
30 The main limitation of the cohort study is the lack of quantitative data on exposure to DE.
31 The number of years exposed to DE was used as a surrogate for dose. The dose, based on
32 duration of employment, has inaccuracies because individuals were working on both steam and
33 diesel locomotives during the transition period. It should be noted that the investigators included
Ponly exposures after 1959; the duration of exposure prior to 1959 was not known. Other
limitations of this study include its inability to examine the effect of years of exposure prior to
7/25/00 8-5 DRAFT—DO NOT CITE OR QUOTE
-------
1 1959 and the less-than-optimal latency period for lung cancer expression. No adjustment for
2 smoking was made in this study. For a detailed description of this study please refer to Section
3 7.2.1.7.
4 Garshick and colleagues also conducted a case-control study of railroad workers who
5 died of lung cancer between 1981 and 1982 (Garshick et al., 1987). The author reported
6 statistically significant increased odds ratios (with asbestos exposure accounted for) of 1.41 for
7 the <64 year age group and 1.64 for the <64 year age group with s;20 years of exposure when
8 compared to the 0-4 year exposure group. The population base for this case-control study was
9 approximately 650,000 active and retired male U.S. railroad workers with 10 years or more of
10 railroad service who were born in 1900 or later. The cases were selected from deaths with
11 primary lung cancer, which was the underlying cause of death in most cases. Each case was
12 matched to two deceased controls whose dates of birth were within 2.5 years of the date of birth
13 of the case and whose dates of death were within 31 days of the date of death noted in the case.
14 Controls were selected randomly from workers who did not have cancer noted anywhere on their
15 death certificates and who did not die of suicide or of accidental or unknown causes. A total of
16 1,256 cases and 2,385 controls were selected for the study. Among younger workers,
17 approximately 60% had exposure to DE, whereas among older workers, only 47% were exposed
18 to DE. DE exposure surrogates for workers were similar to those in the cohort study. Asbestos
19 exposure was categorized on the basis of jobs held in 1959, or on the last job held if the subject
20 retired before 1959. Smoking history information was obtained from the next of kin.
21 The strengths of the case control study are consideration of confounding factors such as
22 asbestos exposure and smoking; classification of DE exposures by job titles and industrial
23 hygiene sampling; and exploration of interactions between smoking, asbestos exposure, and DE
24 exposure. Major limitations of this study include: (a) possible overestimation of cigarette
25 consumption by surrogate respondents; (b) use of the Interstate Commerce Commission (ICC)
26 job classification as a surrogate for exposure, which may have led to misclassification of DE
27 exposure jobs with low intensity and intermittent exposure, such as railroad police and bus
28 drivers, as unexposed; (c) lack of data on the contribution of unknown occupational or
29 eaviujmnental exposures and passive srncking; and (d') a suboptimal latency period of 22 veais.
30 which may not be long enough to observe a full expression of lung cancer. For a detailed
31 description of this study, please see Section 7.2.2.4.
32 As a part of these epidemiologic studies Woskie et al. (1988a) conducted an industrial
33 hygiene survey in the early 1990s for selected jobs in four small northern railroads. DE exposure
34 was considered as a yes/no variable based on job in 1959 and estimated years of work in a diesei-
3 5 exposed job as an index of exposure. Thirty-nine job titles were originally identified and were
7/25/00 8-6 DRAFT—DO NOT CITE OR QUOTH
-------
then collapsed into 13 job categories and, for some statistical analyses, into 5 categories (clerks,
signal maintainers, engineers/firers, brakers/conductors/hostlers, and shop workers) (Woskie et
3 al., 1988b; Hammond et al., 1988). As discussed below, these exposure estimations were used
4 by Crump et al. (1991) and by Cal EPA (1998) for their dose-response analyses.
5
6 8.3.2.1.1. Potential for the data to be used for dose-response modeling. Usually dose-response
7 analyses are performed on data from cohort studies. Case-control studies can also be used for
8 dose-response analysis if exposure for each case and control is available. Control of a smoking
9 effect is important when lung cancer is the disease of interest. However, as discussed previously
10 (see Section 8.3.1), one may not be able to control smoking completely in a dose-response
11 analysis.
12 Garshick et al. (1988) reported a positive relationship of relative risk and duration of
13 exposure by modeling age in 1959 as a covariate in an exposure-response model. The positive
14 relationship disappeared when attained age was used instead of age in 1959 and a negative dose-
15 response was observed (Crump et al., 1991). This negative dose-response continued to be upheld
16 in a subsequent reanalysis (Crump, 1999). Garshick (letter to Chao Chen, U.S. EPA, dated
17 August 15, 1991) performed further analysis and reported that the relationship between years of
^fc exposure, when adjusted for attained age and calendar year, was flat to negative depending upon
19 which model was used. In contrast, California EPA (Cal EPA, 1998) found a positive dose-
20 response by using age in 1959 but allowing for an interaction term of age and calendar year in the
21 model.
22 Crump et al. (1991) also found, and Garshick (letter to Chao Chen, U.S. EPA, dated
23 August 15, 1991) confirmed, that in the years 1977-1980 the death ascertainment was not
24 complete. About 20% to 70% of deaths were missing, depending upon the calendar year.
25 Further analysis, based on job titles in 1959 and limited to deaths occurring through 1976,
26 showed that the youngest workers still had the highest risk of dying of lung cancer.
27 Extensive statistical analyses were conducted by a panel convened by HEI (1999) to
28 investigate the utility of the railroad worker cohort for use in dose-response based quantitative
29 risk assessment. Seven models were used to test the data, and the models were formed by
30 varying a number of covariates in different combinations. The covariates included employment
31 duration, cumulative exposure with and without correction for background exposure, and three
32 job categories: clerks and signalmen, train workers (which include engineers/firers/brakers/
33 conductors), and shop workers. The coefficient for each covariate in a model is used to calculate
relative risk for the associated covariate. In summary, the panel found that effects of exposure as
defined by an exposure-response curve were either flat or negative in all of the models. In these
7/25/00 8-7 DRAFT—DO NOT CITE OR QUOTE
-------
1 analyses, relative risk for each job category was assumed to be constant with respect to age.
2 Further exploration of the data showed that the relative risk for train workers was not constant.
3 The panel's statistical analyses also revealed the complexity of the data and difficulties of
4 providing an adequate summary measure of effect, probably because calendar year and
5 cumulative exposure are highly correlated, which makes it especially difficult to sort out their
6 separate effects. The difficulty of providing an adequate measure of DE effect was further
7 demonstrated in Table C.3 of the HEI report, in which negative or positive effects for cumulative
8 exposure (with background exposure adjustment) were obtained depending on whether or not job
9 category was included in the model.
10 The diverging results about the presence or absence of exposure-response for the railroad
11 worker data have become a source of continuing debate about the suitability of these data for
12 estimating DE risk. Although it is difficult to identify the exact reason for the diverging findings,
13 the "age effect" appears to be a main source of uncertainty because age, calendar year, and
14 cumulative exposure are not mutually independent. An ideal dose-response analysis would
15 account for the ages when exposure to DE began and terminated, along with the attained age and
16 other covariates for each person, using exposure intensity over age rather than cumulative
17 exposure as a dosimeter. This analysis would be possible for the railroad workers if information
18 were available on the ages when exposure began and terminated.
19 Given the equivocal evidence for positive exposure-response, EPA has not derived a unit
20 risk on the basis of the available railroad worker data. This determination should not be
21 construed, however, to imply that the railroad worker studies contain no useful information on
22 lung cancer risk from exposure to DE.
23
24 8.3.2.2. Teamsters Union Trucking Industry Studies
25 Steenland et al. (1990) conducted a case-control study of lung cancer deaths in the
26 Central States Teamsters Union to determine the risk of lung cancer among different trucking
27 industry occupations. The study found statistically significant increased odds ratios for lung
28 cancer of 1.89 and 1.64, depending on years of employment. Cases comprised all deaths from
29 lung cancer (1,288). The 1,452 controls comprised every sixth death from the entire file,
30 excluding deaths from lung cancer, bladder cancer, and motor vehicle accidents. Individuals
31 were required to have 20 years tenure in the union to be eligible to claim benefits.
32 Detailed information en work history and potential confounders such as smoking, diet,
33 and asbestos exposure was obtained by questionnaire. On the basis of interview data and the
34 1980 census occupation and industry codes, subjects were classified either as nonexposed or as
35 having held other jobs with potential DE exposure. The Teamsters Union work history file did
7/25/00 8-8 DRAFT—DO NOT CITE OR QUOTE
-------
not have information on whether men drove diesel or gasoline trucks, and the four principal
occupations were long-haul drivers, short-haul or city drivers, truck mechanics, and
3 dockworkers. Subjects were assigned the job category in which they had worked the longest.
4 The main strengths of the study are the availability of detailed records from the Teamsters
5 Union, a relatively large sample size, availability of smoking data, and measurement of possible
6 asbestos exposures. Some limitations of this study include possible misclassifications of
7 exposure and smoking habits, as information was provided by next of kin; lack of sufficient
8 latency to observe lung cancer excess; and a small nonexposed group (n = 120).
9 Steenland et al. (1998) conducted an exposure-response analysis by supplementing the
10 data from their earlier case-control study of lung cancer and truck drivers in the Teamsters Union
11 with exposure estimates based on a 1990 industrial hygiene survey of elemental carbon (EC)
12 exposure (Zaebst, 1991), a surrogate for DE in the trucking industry. Available data indicate that
13 exposure to workers in the trucking industry in 1990 averaged 2-27 ng/m3 of EC. The 1990
14 exposure information was used by Steenland as a baseline exposure measurement to reconstruct
15 past exposure (in the period of 1949 to 1983) by assuming that the exposure for workers in
16 different job categories is a function of highway mileages traveled by heavy-duty vehicles, and
17 efficiency of the engine over the years.
^B The industrial hygiene survey by Zaebst et al. (1991) of EC exposures in the trucking
19 industry provided exposure estimates for each job category in 1990. The EC measurements were
20 generally consistent with the epidemiologic results, in that mechanics were found to have the
21 highest exposures and relative risk, followed by long-haul and short-haul drivers. Dockworkers
22 who had the lowest exposures also had the lowest relative risks.
23 Past exposures were estimated assuming that they were a function of (1) the number of
24 heavy-duty trucks on the road, (2) the particulate emissions (grams/mile) of diesel engines over
25 time, and (3) leaks from truck exhaust systems for long-haul drivers. Estimates of past exposure
26 to EC (as a marker for DE exposure) were made based on the assumption that average 1990
27 levels for a particular job category could be assigned to all subjects in that category, and that
28 levels prior to 1990 were directly proportional to vehicle miles traveled by heavy-duty trucks and
29 the estimated emission levels of diesel engines. For example, a 1975 exposure level was
30 estimated by the following equation: 1975 level = 1990 level * (vehicle miles 1975/vehicle miles
31 1990) * (emissions 1975/emissions 1990). Once estimates of exposure for each year of work
32 history were derived for each subject, analyses were conducted by cumulative level of estimated
33 carbon exposure.
7/25/00 8-9 DRAFT—DO NOT CITE OR QUOTE
-------
1 8.3.2.2.1. Potential for the data to be used for dose-response modeling. Steenland et al. (1998)
2 analyzed their case-control data and showed a significant positive trend in lung cancer risk with
3 increasing cumulative exposure to DE. The study by Steenland et al. (1998) provides a
4 potentially valuable database for calculating unit risk for DE emissions. The strength of this data
5 set is that the smoking histories of workers were obtained to the extent possible. Smoking is
6 especially important in assessing the lung cancer risk due to DE exposure because smoking has
7 much higher relative risk (or odds ratio) of lung cancer than does DE. In the Steenland et al.
8 (1998) study, the overall (ever-smokers vs. nonsmokers) odds ratio for smoking is about 7.2,
9 which is about five-fold larger than the 1.4 relative risk increase from a large synthesis of many
10 DE epidemiologic studies. It is possible that a modest change of information on smoking and
11 diesel exposure might alter the conclusion and risk estimate.
12 Another strength of the Teamster data for use in environmental risk assessment for the
13 general population is that exposures of Teamsters are closer to ambient exposures than are those
14 of railroad workers. The Teamsters Union truck driver case control workers had cumulative
15 exposure ranging from 19 to 2,440 ng/m3-years of EC, with the median and 95th percentile,
16 respectively, of 358 and 754 u.g/m3-years of EC. The median and 95th percentile of an
17 environmentally equivalent exposure would be 3 and 6 ng/m3, respectively.1 These
18 environmental equivalent exposures for the Teamsters Union truck drivers are close to the
19 estimated ambient exposures of <1.0 ng/m3 to 4.0 ng/m3 (see Table 2-30). It should be noted that
20 Steenland's study is a case-control study in which both case and control could be exposed to DE.
21 Therefore, it is not informative to merely observe that environmental and occupational exposures
22 overlap, thus the 95th percentile exposure of 6 ng/m3 for the truck drivers should be used for
23 comparison to ensure that the exposure is likely to be associated with the observed increment of
24 cancer mortality.
25 Steenland et al. (1998) stated that their risk assessment is exploratory because it depends
26 on estimates about unknown past exposures. Reanalysis of DE exposure for this study is
27 underway. In a recent review, HEI (1999) concluded that the Teamsters studies may be useful
28 for quantitative risk assessment, but significant further evaluation and development are needed.
29 Given the ongoing reanaiysis of exposure, EPA will not, at this time, use the Steeniand (199S)
30 occupational risk assessment findings to derive equivalent environmental parameters and cancer
31 unit risk estimates.
'The conversion assumes (1) DPM = 40% EC as reported by Steenland et al. (1998), (2) environmental equivalent
exposure is approximately = 0.21 x occupational exposure, and (3) 70 jig/nV -years is equivalent to a lifetime of exposure
at 1
7/25/00 8- 1 0 DRAFT — DO NOT CITE OR QUOTE
-------
8.3.3. Conclusion
Because of uncertainties associated with the key epidemiologic data and related exposure
3 information, this health assessment is not deriving a cancer unit risk or cancer unit risk range that
4 can be confidently used to estimate population risk. Two significant activities are underway to
5 improve the epidemiologic database for dose-response assessment: (1) to correct the
6 undercounting of mortality in the Garshick et al. (1988) railroad worker study, and (2) to improve
7 exposure estimates for Teamsters Union truck drivers (Steenland et al., 1998). These activities
8 are being pursued by EPA, NIOSH, and the investigators of these studies. EPA will monitor
9 ongoing research, including the longer term work by NCI-NIOSH regarding a new study of
10 miners and the shorter term work reanalysis of epidemiology-exposure studies, and at a later date
11 determine the merit of conducting additional dose-response analysis and unit risk derivation.
12
13 8.4. PERSPECTIVES ON CANCER RISK
14 Although the available data are considered inadequate to confidently establish a cancer
15 unit risk, this does not mean there is no information about the possible cancer risk of DE. To
16 examine the significance of the potential cancer hazard from environmental exposure to DE, all
17 relevant epidemiologic and exposure data as well as simple risk assessment tools can be used.
^B Such an approach does not produce confident estimates of cancer unit risk. Rather, these
19 approaches provide a perspective on the possible magnitude of cancer risk and thus insight about
20 the significance of the hazard. This section describes approaches and methods that are used to
21 gauge the magnitude of potential cancer risk from ambient exposure to DE.
22 The first approach involves examining the differences between the levels of occupational
23 and ambient environmental exposures, and assuming that cancer risk to DE is proportional
24 linearly with cumulative lifetime exposure. Risks to the general public would be low in
25 comparison with occupational risk, if the differences in exposure are large (i.e, about three orders
26 of magnitude or more). On the other hand, if the differences are smaller (i.e., within one to two
27 orders of magnitude), the environmental risks are of concern, as they would approach workers'
28 risk as observed in epidemiologic studies of past occupational exposures.
29 Table 8-1 shows occupational exposure estimates representative of some of the
30 occupational groups where increased relative risks of lung cancer have been observed. Given the
31 limited availability of exposure data, a broad estimate of DPM concentrations in the workplace is
32 also included as a surrogate for high and low bounding of the exposures, recognizing that actual
33 exposures from such concentration ranges would probably be less. These exposure or
7/25/00 8-11 DRAFT—DO NOT CITE OR QUOTE
-------
1 concentration estimates2 are not intended to be precise, or to match with specific epidemiologic
2 data, but rather to provide a broad range of probable exposures. Environmental exposure data
3 from on-road vehicle emissions are based on the 1990 nationwide exposure estimates from the
4 HAPEM model (see Section 2.4.3.3.1). Both average (0.8 ng/rn3) and high-end exposure (4
5 Hg/m3) are used.
6 In order to compare differences between occupational and environmental exposures, it is
7 necessary to convert occupational exposure to continuous exposure (i.e., environmental
8 equivalent exposure = 0.21 * occupational exposure, see Section 2.4.3.1). Accordingly, Table 8-
9 1 shows equivalent environmental levels and the ratios of occupational to environmental
10 exposures, referred to as exposure margins (EMs). An EM of 1 or less indicates that
11 environmental exposure is comparable to occupational exposure. An EM >1 means that the
12 occupational equivalent exposure is greater than the environmental exposure.
13 Table 8-1 shows that the EMs based on the average nationwide environmental exposure
14 (0.8 M-g/m3) approach three orders of magnitude. However, the EMs based on a high-end
15 environmental exposure (i.e., 4 u.g/m3) range from within an order of magnitude to less than two
16 orders of magnitude. This analysis, therefore, indicates that cancer risks from environmental
17 exposure to DE are of potential public health concern. This exposure analysis, however, only
18 addresses on-road sources for DE exposure. With additional DE exposures from non-road
19 sources, which cannot be quantified at this time, there is a potential for greater concern for DE-
20 induced cancer risk.
21 To further characterize possible cancer risk to the general population from environmental
22 exposure to DE, one can begin by examining the risk observed in DE exposed workers. As
23 reviewed in Section 7.2, numerous epidemiologic studies have shown increased lung cancer risks
24 (i.e., some are deaths, some are cases) among workers in certain occupations. The relative risks
25 or odds ratios range from 1.2 to 2.6. Two independent meta-analyses show smoking adjusted
26 relative risk increase of 1.35 (Bhatia et al., 1997) and 1.47 (Lipsett and Campleman, 1999). For
27 the purpose of this analysis, a relative risk of 1.4 is selected as a reasonable estimate. The
28 relative risk of 1.4 means that the workers faced an extra risk that is 40 % higher than the 5%
23 background lifetime luiig cancer risk in Ihe U.S. population/ Thus, using the relationship
Concentration is defined as the amount of DPM in the air; exposure takes into account human exposure patterns
3The background rate of 0.05 is an approximated lifetime risk calculated by the method of lifetable analysis using
—„ :c.~ i.. .„_„_ _„_»„!;»,. ,!„*„ „.,,) __«u^u;i:>,. „<• J,,~»u ;_ *u« „ ,~.._ »-i J= *u- -XT-,*: 1 TT — uu
U£,w~^pfwll IV IWllg WUI1VV1 lllvfl bulllj' lAUbU M11W pi Vl/Ul_rjllkj wl \AVUlll 111 111W ClgW ^1VIU^> 1CUVVI1 Ill/Ill HIV 1>IUL1U11U1 llCCllllI
Statistics (MRS) monographs of Vital Statistics of the U.S. (Vol. 2, Part A, 1992). Similar values based on two rather
crude approaches can also be obtained: (1) 59.8 * 10E-5 / 8.8 x 10E-3 = 6.8 * 10E-2 where 59.8 * 10E-5 and 8.8 *
10E-3 are respectively the crude estimates of lung cancer deaths (including intrathoracic organs, estimated to be less
7/25/00 8-12 DRAFT—DO NOT CITE OR QUOTE
-------
[excess risk = (relative risk-1) x background risk], these DE-exposed workers would have an
excess risk of 2% (10'2) (i.e., to develop lung cancer) due to occupational exposure to DE [(1.4 -
3 l)x 0.05)= 0.02].
4 Next, one would consider the exposure margin (i.e., the EM ratio) between the
5 occupational exposures and general-population environmental exposures. The DPM
6 concentrations in the workplace, used as a surrogate for worker exposure, have been reported to
7 range from 4 to 1,740 |ig/m3 (or an equivalent continuous exposure of 1-365 |ig/m3). Table 8-1
8 shows that the DPM exposure margin ratio between occupational and environmental exposure,
9 using the nationwide average exposure value of 0.8 jig/m3, may range from 1 to 457. Risks from
10 environmental exposure depend on the shape of the dose-response curve in the range between
11 occupational and environmental exposures. If lifetime risks in this range were to fall
12 proportionately with reduced exposure, and if one assumes that past occupational exposures were
13 at the high end, then the risk from average environmental exposure could be between 10"5 and
14 10"4 (0.02 +• 450 = 4 x 10'5). On the other hand, if occupational exposures for different groups
15 were lower, risks from environmental exposure would be higher than 10"4 - 10"5. For example, if
16 occupational concentrations or exposures were closer to 100 ng/m3, a value that is represented in
17 several data sets shown in Table 8-1 (with an equivalent environmental exposure of 20 jig/m3 and
^fe a corresponding EM of 25), then risks from environmental exposure would approach 10"3 (0.02 +•
19 25 = 8 x 10"4). If lifetime risks were to fall more than proportionately, then risks from
20 environmental exposure would be lower. The latter two sources of dose-response uncertainty
21 (i.e., the actual occupational exposures and the shape of the dose-response curve at low
22 exposures) cannot be defined with currently available information, but they affect the
23 environmental risk estimates in opposite directions.
24 The magnitude of the estimated lifetime cancer risk (between 10"5 and 10"4), derived from
25 using a high-end occupational to environmental exposure difference, establishes a reasonable
26 basis for concern that the general population faces possible risks higher than 10"6. Adding to this
27 concern are two other areas where this analysis does not directly address the segments of the
28 population that may be at highest risk: those who are additionally exposed to nonroad sources of
29 DE, and children who may be more sensitive to early life DE exposure.
30 The analyses presented above are not intended to be precise but are useful in gauging the
31 possible range of risk based on applying scientific judgment and simple risk exploration methods
than 105 of the total cases) and total deaths for 1996 reported in Statistical Abstract of the U.S. (Bureau of the
Census, 1998, 118th Edition), and (2) 156,900/270,000,000 * 76 = 0.045, where 156,900 is the projected lung cancer
deaths for the year 2000 as reported in Cancer Statistics 9J of American Cancer Society, Jan/Feb 2000), 270,000,000
is the current U.S. population, and 76 is the expected lifespan.
7/25/00 8-13 DRAFT—DO NOT CITE OR QUOTE
-------
1 to the relative risk findings from available epidemiologic studies. These analyses provide a sense
2 of where an upper limit (or "upper bound") of the cancer risk may be. The simple methodologies
3 used are generic in that they are valid for any increased relative risk data and thus are not unique
4 to the DE data. These analyses are subject to considerable uncertainties, particularly the lack of
5 actual exposure information and the underlying assumption that cancer risk is linearly
6 proportional to cumulative exposure. Nevertheless, these analyses indicate that environmental
7 exposure to DE may pose a lifetime cancer risk ranging from 10~5 to 10'3. These findings are
8 general indicators of the potential significance of the lung cancer hazard, and should not be
9 viewed as a definitive quantitative characterization of risk. Further research is needed to more
10 accurately assess and characterize environmental cancer risks to DE.
11
12 8.5. SUMMARY
13 As concluded in Section 7.5, DE is considered likely to be a carcinogen to humans at
14 environmental levels of exposure. There have been many quantitative dose-response
15 assessments in the peer-reviewed literature using epidemiologic and or experimental data to
16 estimate human cancer risk from environmental exposure to DE (see Appendix D). In light of
17 increased mechanistic understanding in recent years about how DE causes lung tumors in the rat,
18 the present scientific consensus is that the rat lung tumor dose-response data are not suitable for
19 predicting human risk at low exposure concentrations. Therefore, EPA has focused on the use of
20 epidemiological data in characterizing the exposure-response relationship in the observed range
21 of occupational exposure and extrapolating to the presumably lower levels of environmental
22 exposure to derive a dose/exposure-specific unit risk. As discussed in the section, in the absence
23 of a complete understanding of the modes of action for DE-induced lung cancer in humans
24 coupled with the consideration that DE contains many mutagenic and carcinogenic constituents,
25 this assessment takes the position that linear low-dose extrapolation is appropriate (i.e., risk is
26 proportional to total lifetime exposure).
27 This chapter evaluates the railroad worker studies (Garshick et al., 1987., 1988) and the
28 Teamster Union truck driver studies (Steenland et al., 1990, 1998), which have the best available
29 exposure data for possible use in establishing an exposure-response relationship and deriving a
30 cancer unit risk. Because of the uncertainties about the exposure-response for the railroad
31 workers and exposure uncertainties for the truck drivers. EPA is not developing a cancer unit risk
32 estimate for DE from these data sets at this time.
33 In the absence of a cancer unit risk to assess environmental cancer risk, this assessment
34 provides perspectives about the possible magnitude of risk from environmental exposure to DE.
7/25/00 8-14 DRAFT—DO NOT CITE OR QUOTE
-------
The small exposure margins between some occupational and environmental levels indicates a
likelihood of cancer risk from environmental exposure to DE. Furthermore, based on the
3 observed lung cancer from occupational exposures, and conservative assumptions discussed
4 previously, the environmental cancer risks from DE may range from 10~5 to 10~3. These findings
5 are general indicators of the potential significance of the lung cancer hazard and should not be
6 viewed as a definitive quantitative characterization of risk. A major assumption used in these
7 analyses is that cancer risk is linearly proportional to total lifetime exposure. Further research is
8 needed to more accurately assess and characterize environmental cancer risks to DE.
7/25/00 8-15 DRAFT—DO NOT CITE OR QUOTE
-------
Table 8-1. DPM exposure margins for occupational vs. environmental exposures
Occupational
group
Non-coal
miners
Public transit
workers
U.S. railroad
workers
Broad
concentration
range
Estimated occupational
exposure/concentration
Gig/m3)
Environmental
equivalent
10-1,280
2-269
15-98
3-21
39-191
8-40
4-l,740d
1-365
Exoosure margin
ratio
for 0.8 ng/m3 of
environmental
exposure1*
3-336
4-26
10-50
1-457
Exposure margin
ratio
for 4.0 ng/m3 of
environmental
exposure1*
0.5-67
0.8-5
2-10
0.21-91
Reference0
SSverin et al.,
1999
Birch and
Gary, 1996
Woskie et al.,
1988b
HEI, 1995
1 Occupational exposure * 0.21 = equivalent environmental exposure, see Chapter 2, Section 2.4.3.1.
b 0.8 Mg/m3 = average 1990 nationwide exposure estimate from HAPEM model; the companion rural estimate is 0.5 (ig/m3, and 4 ng/m3 is
a high-end estimate. The 1996 nationwide average is 0.7 ug/m3. The companion rural estimate is 0.2 ug/m3; however, a high-end estimate is not
available for 1996. See Chapter 2, Sections 2.4.3.2.1 and 2.4.3.2.2.
' See Table 2-27 for more details about Saverin, Birch and Clay, and Woskie.
d Broadest range of average concentrations across many occupational groups. Use of concentration as a surrogate for high and low boundary for
exposure, may overstate exposure.
7/25/00
8-16
DRAFT—DO NOT CITE OR QUOTE
-------
8.6. REFERENCES
3 Bhatia, R; Lopipero, P; Smith, A. (1997) Diesel exhaust exposure and lung cancer. Epidemiology 9(1):84-91.
4
5 Birch, ME; Gary, RA. (1996) Elemental carbon-based method for monitoring occupational exposures to particulate
6 diesel exhaust. Aerosol SciTechnol 25:221-241.
7
8 California Environmental Protection Agency (Cal EPA). (1998) Health risk assessment for DE. Public and
9 Scientific Review Draft.
10
11 Crump, KS; Lambert, T; Chen, C. (1991) Assessment of risk from exposure to diesel engine emissions. Clement
12 International Corporation. Prepared for U.S. EPA under contract no. 68-02-4601; 56 pp.
13
14 Crump, KS. (1999) Lung cancer mortality and DE: reanalysis of a retrospective cohort study of U.S. railroad
15 workers. Inhal Toxicol 11:1-17.
16
17 Garshick, E; Schenker, MB; Munoz, A; et al. (1987) A case-control study of lung cancer and DE exposure in
18 railroad workers. Am Rev Respir Dis 135:1242-1248.
19
20 Garshick, E; Schenker, MB; Munoz, A; et al. (1988) A retrospective cohort study of lung cancer and DE exposure in
21 railroad workers. Am Rev Respir Dis 137:820-825.
22
23 Hammond, SK; Smith, TJ; Woskie, SR; et al. (1988) Markers of exposure to diesel exhaust and cigarette smoke in
24 railroad workers. Am Ind Hyg Assoc J 49:516-522.
25
Health Effects Institute (HEI). (1995) DE: A critical analysis of emissions, exposure, and health effects. A special
report of the Institute's Diesel Working Group. Cambridge, MA: Health Effects Institute.
29 HEI. (1999) Diesel emissions and lung cancer: epidemiology and quantitative risk assessment. A special report of
30 the Institute's Diesel Epidemiology Expert Panel. Cambridge, MA: Health Effects Institute.
31
32 International Life Sciences Institute (ILSI). (2000) ILSI Risk Science Institute workshop: The relevance of the rat
33 lung response to particle overload for human risk assessment. Gardner, DE, ed. Inhal Toxicol: 12(1-2).
34
35 International Programme on Chemical Safety: World Health Organization (IPCS). (1996) Diesel fuel and exhaust
36 emissions. Environmental Health Criteria 171. Geneva: World Health Organization.
37
38 Lipsett, M; Campleman, S. (1999) Occupational exposure to DE and lung cancer: a meta-analysis. Am J Publ Health
39 89(7): 1009-1017.
40
41 Saverin. R; Braunlich, A; Dahman, D; et al. (1999) Diesel exhaust and lung cancer mortality in potash mining. Am
42 J Ind Med 36:415-422.
43
44 Steenland, NK; Silverman, DT; Hornung, RW. (1990) Case-control study of lung cancer and truck driving in the
45 Teamsters Union. Am J Publ Health 80:670-674.
46
47 Steenland, K; Deddens, J; Stayner, L. (1998) DE and lung cancer in the trucking industry: exposure-response
48 analysis and risk assessment. Am J Ind Med 34:220-228.
49
50 U.S. Environmental Protection Agency (U.S. EPA). (1986) Guidelines for carcinogen risk assessment. Federal
51 Register 51(185):33992-34003.
7/25/00 8-17 DRAFT—DO NOT CITE OR QUOTE
-------
1 U.S. Environmental Protection Agency. (1996) Proposed guidelines for carcinogen risk assessment. Federal
2 Register 61 (79): 17960-18011.
3
4 U.S. Surgeon General. (1982) The health consequences of smoking: cancer. NIH Publication 82-50179, Washington,
5 DC: U.S. DHHS.
6
7 Woskie, SR; Smith, TJ; Hammond, SK; et al. (1988a) Estimation of the DE exposures of railroad workers: II.
8 National and historical exposures. Am J Ind Med 13:395-404.
9
10 Woskie, SR; Smith, TJ; Hammond, SK; et al. (1988b) Estimation of the DE exposures of railroad workers: I. Current
11 exposures. Am J Ind Med 13:381-394.
12
13 Zaebst, D; Clapp, D; Blade, L; et al. (1991) Quantitative determination of trucking industry workers' exposures to
14 diesel particles. Am Ind Hyg Assoc J 52:529-541.
7/25/00 8-18 DRAFT—DO NOT CITE OR QUOTE
-------
9. CHARACTERIZATION OF POTENTIAL HUMAN HEALTH EFFECTS OF
DIESEL EXHAUST: HAZARD AND DOSE-RESPONSE ASSESSMENTS
1 9.1. INTRODUCTION
2 Environmental human health risk assessment entails the evaluation of all pertinent
3 information on the hazardous nature of environmental agents, on the extent of human exposure to
4 them, and on the characterization of the potential risk to the exposed population. Risk
5 assessment consists of four components: hazard assessment, dose-response assessment,
6 exposure assessment, and risk characterization". This document focuses only on hazard and dose-
7 response assessment. The overall objectives of this assessment are:
8
9 • to identify and characterize the human health effects that may result from
10 environmental exposure to diesel exhaust (DE); and
11 • to determine whether there is a quantitative exposure- (or dose-) response relationship
12 for DE exposure and health effects in the range of observation and, if sufficient data
13 are available, to derive toxicity values, estimates of exposure, or dose-specific unit
14 risk for subsequent use in the characterization of potential risk to the general human
15 population and vulnerable subgroups.
V
17 This chapter integrates the key findings about the nature and characteristics of
18 environmental exposure to DE (Chapter 2), health hazard information (Chapters 3,4, 5, and 7),
19 and dose-response analyses (Chapters 6 and 8) that are relevant to the characterization of
20 potential human health effects associated with current-day environmental exposure to DE. It also
21 discusses major uncertainties of this assessment, including critical data and knowledge gaps, key
22 assumptions, and EPA's science policy choices to bridge the data and knowledge gaps.
23
24 9.2. PHYSICAL AND CHEMICAL COMPOSITION OF DIESEL EXHAUST
25 As reviewed in Chapter 2, DE is a complex mixture of hundreds of constituents in gas or
26 particle phases. Gaseous components of DE include carbon dioxide, oxygen, nitrogen, water
27 vapor, carbon monoxide, nitrogen compounds, sulfur compounds, and low-molecular- weight
28 hydrocarbons and their derivatives. The particulate matter of DE, diesel particulate matter
29 (DPM), is composed of elemental carbon, adsorbed organic compounds, and small amounts of
30 sulfate, nitrate, metals, trace elements, water, and unidentified compounds. DPM is either
31 directly emitted from diesel-powered engines (primary particulate matter) or is formed from the
gaseous compounds emitted by a diesel engine (secondary particulate matter). Incomplete
7/25/00 9-1 DRAFT—DO NOT CITE OR QUOTE
-------
1 combustion of fuel hydrocarbons as well as engine oil and other'fuel components such as sulfur
2 leads to the formation of DPM.
3 After emission from the tailpipe, DE undergoes dilution, chemical and physical
4 transformations, and dispersion and transport in the atmosphere. The atmospheric lifetime for
5 some compounds present in DE ranges from hours to days. In general, secondary pollutants
6 formed in an aged aerosol mass are more oxidized, and therefore have increased polarity and
7 water solubility.
8 DE emissions vary significantly in chemical composition and particle sizes among
9 different engine types, fuel formulations, and age of emissions. There have been both qualitative
10 and quantitative changes in DE emissions over time as a result of changes in engine technology
11 and fuel reformulation. The following sections identify and characterize the key components of
12 DE that are of special concern in possible health outcomes, and discuss the changes in the
13 composition of DE over time. The latter information is critical for making a scientific judgment
14 about the appropriateness of using epidemiologic and toxicological findings from past DE
15 exposures to assess hazard and risk from current-day environmental exposures. It should be
16 noted that available animal studies are based on exhaust exposures from various model year on-
17 road diesel engines since 1980, whereas many of the epidemiologic studies refer to exposures
18 from on-road and non-road diesel engines in use from the 1950s through the mid-1990s.
19
20 9.2.1. Diesel Exhaust Components of Possible Health Concern
21 The components of DE that are of health concern for this assessment are the particles
22 (elemental carbon core), the organic compounds adsorbed to the particles, and the organic
23 compounds present in the gas phase.
24
25 9.2.1.1. Diesel Particles
26 Approximately 80%-95% of DPM mass is in the fine particle size range (0.05-1.0
27 microns), with a mean particle diameter of about 0.2 microns. Ultrafine particles (0.005-0.05
28 microns), averaging about 0.02 microns in diameter, account for about 1%-20% of the DPM
29 mass and 50%-90% of the total number of particles in DPM (Section 2.2.8.3).
30 Particle size is important for a number of reasons. Particles with aerodynamic diameters
31 larger than 2.5 microns (i.e., >PM25) tend to be retained in the upper portions of the respiratory
32 tract, whereas particles with diameters smaller than 2.5 microns (i.e.,
-------
DPM is part of ambient particulate matter (PM). The major characteristics that
distinguish DPM from ambient PM are (1) a high portion of elemental carbon, (2) the large
3 surface area associated with carbonaceous particles in the 0.2 micron range; (3) enrichment of
4 certain polycyclic aromatic hydrocarbons (PAHs), and (4) a large percentage of ultrafine
5 particles. The EPA Emissions Trends Report (U.S. EPA, 2000) indicates that annual emissions
6 of diesel PM25 nationwide in 1998 were 6% of the total PM2 5 inventory. Some geographic areas
7 are expected to have a higher percentage of DPM in PM2S because of variations in the number
8 and types of diesel engines present in the area. For instance, DPM contributions to total PM2 5
9 mass were reported to be about 13%-36% in several urban California regions in 1982. More
10 recent studies in the Phoenix and Denver areas showed diesel PM2 5 to be 10%-15% of total
11 PM25 mass, and in Manhattan, diesel PM was reported to contribute about 50% of ambient PM10
12 (Chapter 2, Section 2.4.2.1).
13 DPM generally contains a high percentage of elemental carbon per unit mass, which can
14 be used as a distinguishing feature from other combustion and noncombustion sources of PM25.
15 The DPM elemental carbon content can range from more than 50% to approximately 75% of the
16 DPM mass depending on age of engine, type of engine (heavy-duty versus light-duty), fuel
17 characteristics, and driving conditions. The organic carbon portion of DPM can range
18 approximately from 19% to 43%, although some DPM organic constituents can be higher or
lower than these numbers. In comparison, gasoline engine exhaust generally has a lower
20 elemental carbon content and a higher percentage of organics in the particle mass (Table 2-13).
21
22 9.2.1.2. Organic Compounds
23 The organic compounds present in the gases and adsorbed onto the particles cover a wide
24 spectrum of compounds related to unburned diesel fuel, lube oil, low levels of partial
25 combustion, and pyrolysis products (Table 2-19). The organic compounds present in the gaseous
26 phase include alkanes, alkenes, aldehydes, monocyclic aromatic compounds, and PAHs. Among
27 the gaseous components of DE, the aldehydes are particularly important because of their
28 potential carcinogenic effects and because they make up an important fraction of the gaseous
29 emissions. Formaldehyde accounts for a majority of the aldehyde emissions (65%-80%) from
30 diesel engines. Acetaldehyde and acrolein are the next most abundant aldehydes. Other gaseous
31 components of DE that are notable for their carcinogenic effects include benzene, 1,3-butadiene,
32 PAHs, and nitro-PAHs (including those with ^4 rings and nitro-PAHs with 2 and 3 rings). A
33 number of the gaseous compounds (e.g., aldehydes, alkanes, alkenes, NOX, SOX) are also known
34 to induce respiratory tract irritation given sufficient exposure (see Table 2-21). Very small
amounts of dioxins have been measured in diesel truck exhaust. Dioxin emissions from heavy -
7/25/00 9-3 DRAFT—DO NOT CITE OR QUOTE
-------
1 duty engine truck exhausts are estimated to represent about 1.2% of the national dioxin
2 inventory; dioxin emissions from non-road exhausts have not been estimated (Section 2.2.7.2).
3 Organic substances adsorbed onto DPM include C14_35 hydrocarbon compounds, PAHs
4 with >4 rings, and nitro-PAHs. PAHs and their derivatives comprise <1% of the DPM mass
5 (Section 2.2.8). Many of these hydrocarbons are known to have mutagenic and carcinogenic
6 properties. California EPA (Cal EPA, 1998) identified at least 19 hydrocarbons present in DE
7 that are known or suspected carcinogens, according to evaluations by the International Agency
8 for Research on Cancer (IARC).
9
10 9.2.2. "Fresh" Versus "Aged" Diesel Exhaust
11 Newly emitted exhaust is termed "fresh" whereas exhaust that is more than 1 or 2 days
12 old is referred to as "aged" because of alterations caused by sunlight and other chemical-physical
13 conditions of the ambient atmosphere. It is not clear what the overall toxicological consequence
14 of DE aging is because some compounds in the DE mixture are altered during aging to more
15 toxic forms while others are made less toxic. For example, PAHs present in fresh emissions may
16 be nitrated by atmospheric NO3 to form nitro-PAHs, thus adding to the existing burden of nitro-
17 PAHs present in fresh exhaust. On the other hand, PAHs present in the gas phase can react with
18 hydroxyl radicals present in the ambient air, leading to reduced atmospheric lifetime of the
19 original PAH. Alkanes and alkenes may be converted to aldehydes, and oxides of nitrogen to
20 nitric acid (Section 2.3).
21
22 9.2.3. Changes of DE Emissions and Composition Over Time
23 Chapter 2, with its summary in Section 2.5, provides a full review of emissions trends and
24 a complete characterization of the physical and chemical changes in DE over the years, taking
25 into consideration the lack of consistent analytical and measurement techniques, and the
26 variability in emissions based on vehicle mix, driving cycles, engine deterioration, and other
27 factors. Key findings relevant to the potential health effects of DE are discussed below.
28 As discussed in Chapter 2, Section 2.2.3, the EPA Emissions Trend Report estimates that
29 DPM10 on-road emissions decreased 27% between 1980 and 1998. DPM emission factors
30 (g/miie by model year) from new on-road diesel vehicles decreased on average by a tactor of six
31 in the period from the mid-1970s to the mid-1990s. These significant reductions are largely
32 attributable to reductions in three PM components: elemental carbon, organic carbon, and
33 sulfate. Limited data are available to assess the changes in emission rates from locomotive,
34 marine, or other non-road diesel sources over time. It is estimated that DPM,n (< 10 (am)
35 emissions from non-road diesel engines increased 17% between 1980 and 1998. Despite
36 significant reductions in DPM from diesel vehicles, combined non-road and on-road diesel
7/25/00 9-4 DPxAFT—-DO NOT CITE OR QUOTE
-------
engines still contributed approximately 23% of DPM25 (^2.5 um) emissions to the 1998
inventory (not including the contribution of natural and miscellaneous sources) (Section 2.2.5).
"3 Because of changes in engine technology and fuel composition, the chemical composition
4 of DPM from on-road vehicles has also changed over time. The percentage of soluble organic
5 material associated with DPM from new on-road vehicles decreased by model year from the
6 1980s to the 1990s, and the proportion of elemental carbon is correspondingly higher. PAHs and
7 nitro-PAHs are present in DPM from both new and older diesel engine exhaust. There are
8 insufficient data to provide insight into the potential for changes in total PAH emissions over
9 time or specific organic constituents such as benzo[a]pyrene and 1-nitropyrene. It should be
10 noted that the chemical composition of DPM to which people are currently exposed is
11 determined by a combination of older and newer technology on-road and non-road engines.
12 Consequently, the decrease in the soluble organic fraction of DPM by model year does not
13 directly translate into a proportional decrease in DPM-associated organic material to which
14 people are currently exposed. In addition, the impact from high-emitting and/or smoking diesel
15 engines has not been quantified (Section 2.5.2).
16 Because of these uncertainties, changes in DPM composition over time cannot be
17 confidently quantified. Available data clearly indicate that lexicologically significant organic
18 components of DE (e.g., PAHs, PAH derivatives, nitro-PAHs) were present in DPM and DE in
^P the 1970s and are still present. Even though a significant fraction of ambient DPM (possibly
20 more than 50%) is also emitted by non-road equipment, there are no data available to
21 characterize changes in the chemical composition of DPM from non-road equipment over time.
22 Given the variation in fuel, engine technology, and in-use operational factors over the years,
23 caution should be exercised in presuming that a decrease in the amount of emissions or emission
24 constituents will result in a decrease in risk.
25
26 9.3. AMBIENT CONCENTRATIONS AND EXPOSURE TO DIESEL EXHAUST
27 Section 2.4 provides some information on ambient concentrations of DE, and on
28 occupational and environmental exposures to DE, in order to provide a context for hazard
29 assessment and dose-response analysis. Highlights of available information are discussed below.
30 DE is emitted from a variety of sources, both on-road (e.g., motor vehicles, construction
31 equipment) and non-road (e.g., farm equipment, railway locomotives, marine diesel engines).
32 Environmental exposure to DE is generally higher in urban areas than in rural areas. The
33 concentration of DE constituents in the air is also expected to vary within any geographic area
34 depending on the number and types of diesel engines in the area, the atmospheric patterns of
dispersal, and the proximity of the exposed individuals to the DE source. Certain occupational
7/25/00 9-5 DRAFT—DO NOT CITE OR QUOTE
-------
1 populations (e.g., transportation and garage workers, heavy equipment operators) can be exposed
2 to much higher levels of DE than is the general population.
3 As DE is a complex mixture of a great variety of compounds, "exposure levels" are
4 difficult to define. Even though the environmental levels of a number of individual constituents
5 are generally known, it is difficult to quantify the portion that directly or indirectly comes from
6 diesel engine emissions. Moreover, there is still incomplete knowledge about the relative roles
7 of the relevant DE constituents in mediating the potential health effects of DE. Accordingly,
8 exposure levels to DPM have historically been measured using surrogate markers for whole DE.
9 Although considerable uncertainty exists as to whether DPM mass (expressed as ug/m3 of DPM )
10 is the most appropriate dosimeter, it is considered to be a reasonable choice on the basis of
11 available data until more definitive information about the mechanisms or mode(s) of action of
12 DE becomes available.
13 Several techniques exist for estimating ambient concentrations of DPM, including
14 chemical mass balance (CMB) source apportionment, dispersion modeling, and using elemental
15 carbon as a surrogate for DPM. DPM concentrations reported from CMB and dispersion
16 modeling studies in the 1980s suggest that in urban and suburban areas (Phoenix and Southern
17 California), the annual average DPM concentration ranged from 2 to 13 (Jg/m3. In the 1990s,
18 annual or seasonal average DPM concentrations in suburban or urban locations have ranged
19 from 1.2 to 4.5 ug/m3. DPM concentrations at a major bus stop in downtown Manhattan ranged
20 from 13.2 to 46.7 ug/m3 over a 3-day period in 1993. In nonurban and rural areas in the 1980s,
21 DPM concentrations were reported to range from 1.4 to 5 ug/m3. In the 1990s, nonurban air
22 basins in California were reported to have DPM concentrations ranging from 0.2 to 2.6 ug/m3
23 (Section 2.4.2).
24 A comprehensive exposure assessment cannot be currently conducted because of lack of
25 data. Interim exposure estimation based on EPA's Hazardous Air Pollutant Exposure Model
26 (HAPEM-MS3 model), for on-road sources only, suggests that in 1996 annual average DPM
27 exposure in urban areas from only on-road engines was 0.7 ng/m3, while in rural areas exposure
28 was 0.3 ug/m3. A high-end exposure estimate for 1996 is not yet available. Among 10 urban
29 areas, the 1996 annual average estimated exposure ranged from 0.5 to 1.2 ng/m3. Comparable
30 1990 exposure estimates for uri-road suuices ranged from 0.9 ug/'m for urban areas and from 0.5
31 ug/m3 for rural areas. Exposure estimates for the most highly exposed individuals (e.g., outdoor
32 workers and children who spend large amounts of time outdoors) for 1990 had DPM exposures
33 up to 4.0 ug/m3 (Section 2.4.3.2, Table 2-29). Based on the national inventory, DPM exposure
34 that includes non-road emission sources could at least double the on-road exposure.
35 Estimates for occupational exposures to DE as DPM mass have been generally higher
36 than environmental exposures. The Health Effects Institute (HEI, 1995) reported that mean air
7/25/00 9-6 DRAFT—DO NOT CITE OR QUOTE
-------
concentrations of DPM in the workplace as shown in the available literature ranged from 4 to
l,740^g/m3. Tables 2-27 and 2-28 provide some exposure estimates for specific worker
3 categories. Available information indicates that DPM exposure estimates range up to 1,280
4 ug/m3 for miners, with lower exposures for railroad workers (39-191 ug/m3), firefighters (4-748
5 ug/m3), public transit workers who work with diesel equipment (7-98 ug/m3), mechanics and
6 dock workers (5-65 ug/m3), truck drivers (2-7 ug/m3), and bus drivers (1-3 ug/m3).
7 For direct comparison of lifetime exposures between an occupational setting (8 hours per
8 day, 5 days per week, for 45 years) and environmental exposure (continuous exposure for 70
9 years), the occupational estimates are converted to an equivalent environmental lifetime
10 estimate,1 which is also shown in Table 2-28. A conversion of EC-based measurements to total
11 DPM may also be needed for some estimates. The estimated 70-year lifetime exposures
12 equivalent to those for the occupational groups discussed above range from 0.4 to 2 ug/m3 on the
13 low end to 2 to 269 on the high end. These data indicate that some lower-end occupational
14 estimates of DPM, when converted to environmental equivalents, overlap the range of estimated
15 environmental exposures to DPM (national average in 1990 of 0.8 ug/m3, with high-end
16 exposures up to 4 ug/m3).
17
18 9.4. HAZARD CHARACTERIZATION
With DE being a component of ambient particles in the general environment, it may
20 partly contribute to the range of health effects associated with ambient PM. However, the
21 spectrum of health effects associated with DE exposure are somewhat different, though not
22 entirely inconsistent, with those reported for ambient PM. The primary health effects of concern
23 from environmental exposure to DE, on the basis of combined human and experimental
24 evidence, are lung cancer and noncancer respiratory effects resulting from chronic exposure, and
25 possibly immunologic and allergenic effects from acute and repeated exposures. On the other
26 hand, a wide range of noncancer health effects has been associated with acute, short-term, and
27 long-term exposure to ambient PM. Community epidemiologic studies have shown that ambient
28 PM exposure is statistically associated with increased mortality (especially among people over 65
29 years of age with preexisting cardiopulmonary conditions) and morbidity as measured by
30 increases in hospital admissions, respiratory symptom rates, and decrements in lung function. A
31 cancer hazard has not been characterized for ambient PM, although there is some indication of a
32 possible association between particle air pollution and increased lung cancer risk (U.S. EPA,
33 1996a,b; also see Chapter 7, Section 7.1.2).
'Environmental equivalent occupational exposure = 0.21 * occupational exposure.
7/25/00 9-7 DRAFT—DO NOT CITE OR QUOTE
-------
1 9.4.1. Acute and Short-Term Exposures
2 The combined human and animal evidence indicates that DE can induce irritation to the
3 eye, nose; and throat, as well as inflammatory responses in the airways and the lung following
4 acute and/or short-term exposure to high concentrations. There is also suggestive evidence for
5 possible immunological and allergenic effects of DE.
6
7 9.4.1.1. Acute Irritation
8 DE contains various respiratory irritants in the gas phase and in the particulate phase
9 (e.g., SOX, NOX, aldehydes). Acute exposure to DE has been associated with irritation of the eye,
10 nose, and throat, respiratory symptoms (cough and phlegm), and neurophysiological symptoms
11 such as headache, lightheadedness, nausea, vomiting, and numbness or tingling of the
12 extremities. Such symptoms have been described mainly in reports of individuals exposed to DE
13 in the workplace, or in clinical studies in humans exposed acutely to high concentrations of DE.
14 Because of the general lack of exposure information in available reports, the exact role of DE in
15 causing these effects is not known. An exposure-response relationship for these acute irritation
16 and respiratory symptoms has not been demonstrated (Chapter 5, Section 5.1.1.1).
17
18 9.4.1.2. Respiratory Effects
19 Available studies of occupational exposure to DE have not provided evidence for
20 significant decrements of lung function in workers over a work shift or after a short-term
21 exposure period. Short-term and subchronic inhalation studies of DE in animals (rats, mice,
22 hamsters, cats, guinea pigs) showed inflammation of the airways and minimal or no lung
23 function changes. These effects were associated with high DE exposures (up to 6 mg/m3).
24 Exposure-response relationships have not been established for these responses (Chapter 5,
25 Sections 5.1.2.2 and 5.1.1.1).
26
27 9.4.1.3. Immunological Effects
28 Recent human and animal studies show that acute DE exposure episodes may exacerbate
29 immunological reactions to other allergens or initiate a DE-specific allergenic reaction. The
30 effects seem to be associated with both the organic and carbon core fraction of DPM. In human
31 subjects, intranasal administration of DPM has resulted in measurable increases of IgE antibody
32 production and increased nasal mRNA for the proinflammatory cytokines. The ability of DPM to
33 act as an adjuvant to other allergens has been demonstrated in human subjects. For example, co-
34 exposure to DPM and ragweed pollen was reported to significantly enhance the IgE antibody
3 5 response and cytokine expression relative to ragweed pollen alone. Available animal studies also
36 demonstrate the potential adjuvant effects of DPM with model allergens. For instance, DPM has
7/25/00 9-8 DRAFT—DO NOT CITE OR QUOTE
-------
been shown to enhance IgE antibody production and cytokine production response to several
model allergens (ovalbumin, Japanese cedar pollen) in mice (Chapter 5, Sections 5.1.1.1.3,
3 5.1.1.1.4, 5.1.2.3.5, and 5.1.2.3.6). Additional research is needed to further characterize possible
4 immunological effects of DE and to determine whether or not the immunological effects
5 constitute a low-exposure hazard. This health endpoint is of considerable public health concern,
6 given the increases in allergic hypersensitivity in the U.S. population (Section 5.6.2.6).
7
8 9.4.2. Chronic Exposure
9 9.4.2.1. Noncancer Effects
10 Available long-term and cross-sectional studies have provided evidence for an association
11 between respiratory symptoms (cough and phlegm) and DE exposure, but there was no consistent
12 effect on lung function. DE has been shown in many animal studies of several species to induce
13 lung injury (chronic inflammation and histopathologic changes) following long-term inhalation
14 exposure. DE has also been tested in laboratory animals for other health effects, and no
15 significant effects have been found. Overall, available data support the conclusion of a potential
16 chronic respiratory hazard to humans from long-term exposure to DE.
17
9.4.2.1.1. Respiratory effects. A few human studies in various diesel occupational settings
suggest that DE exposure may impair pulmonary function, as evidenced by increases in
20 respiratory symptoms and some reductions in baseline pulmonary function consistent with
21 restrictive airway disease. Other studies found no particular effects. The methodologic
22 limitations in available human studies limit their usefulness in drawing any firm conclusions
23 about DE exposure and noncancer respiratory effects (Chapter 5, Section 5.1.1.2).
24 Available studies in animals, however, provide a considerable body of evidence
25 demonstrating that prolonged inhalation exposure to DE can result in pulmonary injury. A
26 number of long-term laboratory studies in rats, mice, hamsters, cats, and monkeys found varying
27 degrees of adverse lung pathology including focal thickening of the alveolar walls, replacement
28 of Type I alveolar cells by type II cells, and fibrosis. The rat is the most sensitive animal species
29 to DE-induced pulmonary toxicity (Chapter 5, Sections 5.1.2.3 and 5.4).
30 Available'mechanistic data, mainly in rats, indicate that the DPM fraction of DE is
31 primarily involved in the etiology of pulmonary toxicity, although a role for the adsorbed organic
32 compounds on the particles and in the gaseous phase cannot be ruled out. The lung injury
33 appears to be mediated by an invasion of alveolar macrophages that release chemotactic factors
34 that attract neutrophils and additional alveolar macrophages, which in rum release mediators
B (e.g., cytokines, growth factors) and oxygen radicals. These mediators result in persistent
36 inflammation, cytotoxicity, impaired phagocytosis and clearance of particles, and eventually
7/25/00 9-9 DRAFT—DO NOT CITE OR QUOTE
-------
1 deposition of collagen by activated fibroblasts. This postulated mode of action seems to be
2 operative for a variety of poorly soluble particles in addition to DPM (ILSI, 2000). Because
3 long-term exposure to DE has been shown to induce exposure-dependent chronic respiratory
4 effects in a wide range of animal species, and the postulated mode of action is deemed relevant to
5 humans, there is a sufficient scientific basis to support a conclusion that humans could also be at
6 hazard for these effects under a chronic exposure condition. This inference is deemed reasonable
7 in the absence of information to the contrary.
8
9 9.4.2.1.2. Other noncancer effects. The negative results from available studies in several
10 animal species (rats, mice, hamsters, rabbits, monkeys) indicate that DE is not likely to pose a
11 reproductive or developmental hazard to humans. There has been some evidence from animal
12 studies indicating possible neurological and behavioral effects, as well as liver effects. These
13 effects, however, are seen at exposures higher than the respiratory effects. Overall, there is
14 inadequate evidence for a low-exposure human hazard for these health endpoints (Chapter 5,
15 Sections 5.1.2.3.7, 5.1.2.3.11, and 5.1.2.3.12).
16
17 9.4.2.2. Carcinogenic Effects
18 Many epidemiologic and toxicologic studies have been conducted to examine the
19 potential for DE to cause or contribute to the development of cancer in humans and animals,
20 respectively. In addition, there have been extensive mechanistic studies that provide an
21 improved understanding about the underlying carcinogenic process and the likelihood of hazard
22 to humans. The available evidence indicates that chronic inhalation of DE has the potential to
23 induce lung cancer in humans. There is insufficient information for an evaluation of the potential
24 cancer hazard of DE by oral and dermal routes of exposure.
25
26 9.4.2.2.1. Epidemiologic studies. Twenty-two epidemiologic studies about the carcinogenicity
27 of workers exposed to DE in various occupations are reviewed in Chapter 7, Section 7.2.
28 Exposure to DE has typically been inferred on the basis of job classification within an industry,
29 with cumulative exposure based on duration of employment or age. Increased lung cancer risk,
30 although not always statistically significant, has been observed in 8 out of 10 cohort studies and
31 10 of 12 case-control studies within several industries, including railroad workers, truck drivers,
32 heavy equipment operators, and professional drivers. The increased lung cancer relative risks
33 generally range from 1.2 to 1.5, although a few studies show relative risks as high as 2.6.
34 Statistically significant increases in pooled relative risk estimates (1.33 to 1.47) from two
35 independent meta-analyses further support a positive relationship between DE exposure and lung
36 cancer in a variety of DE-exposed occupations.
7/25/00 9-10 DRAFT—DO NOT CITE OR QUOTE
-------
The generally small increased lung cancer relative risk (less than 2) observed in the
epidemiologic studies potentially weakens the evidence of causality. This is because with a
"3 relative risk of less than 2, if confounders (e.g., smoking, asbestos exposure) were having an
4 effect on the observed risk increases, it could be enough to account for the increased risk. With
5 the strongest risk factor for lung cancer being smoking, there is a concern that smoking effects
6 may be influencing the magnitude of the observed increased relative risks. However, in studies
7 in which the effects of smoking were accounted for, increased relative risks for lung cancer
8 prevailed. Although some studies did not have information on smoking, confounding by
9 smoking is unlikely because the comparison populations were from the same socioeconomic
10 class. Moreover, when the meta-analysis focused only on the smoking-controlled studies, the
11 relative risks tended to increase.
12 As evaluated in Chapter 7 (Section 7.2.4.5), application of the criteria for causality
13 provides evidence that the increased risks observed in available epidemiologic studies are
14 consistent with a causal association between exposure to DE and occurrence of lung cancer.
15 Overall, the human evidence for potential carcinogenicity for DE is judged to be strong but less
16 than sufficient to be considered as a human carcinogen because of exposure uncertainties (lack of
17 historical exposure of workers to DE) and uncertainty as to whether all confounders have been
18 satisfactorily accounted for. The epidemiologic evidence for DE being associated with other
^P forms of cancer is inconclusive.
20
21 9.4.2.2.2. Animal studies. DE and its organic constituents, both in the gaseous and particle
22 phase, have been extensively tested for carcinogenicity in many experimental studies using
23 several animal species and with different modes of administration. Several well-conducted
24 studies have consistently demonstrated that chronic inhalation exposure to sufficiently high
25 concentrations of DE produced dose-related increases in lung tumors (benign and malignant) in
26 rats. In contrast, chronic inhalation studies of DE in mice showed mixed results, whereas
27 negative findings were consistently seen in hamsters. The gaseous phase of DE (filtered exhaust
28 without particulate fraction), however, was found not to be carcinogenic in rats, mice, or
29 hamsters.
30 In several intratracheal instillation studies, DPM, DPM organic extracts, and carbon
31 black, which is virtually devoid of PAHs, have been found to produce increased lung tumors in
32 rats. When directly implanted into the rat lung, DPM condensate containing mainly four- to
33 seven-ring PAHs induced increases in lung tumors. DPM extracts have also been shown to cause
34 skin tumors in several dermal studies in mice, and sarcomas in mice following subcutaneous
injection.
7/25/00 9-11 DRAFT—DO NOT CITE OR QUOTE
-------
1 Overall, there is sufficient evidence for the potential carcinogenicity of whole DE in the
2 rat at high exposure concentration or administered dose, both by inhalation and intratracheal
3 instillation. Available data indicate that both the carbon core and the adsorbed organics have
4 potential roles in inducing lung tumors in the rat, although their relative contribution to the
5 carcinogenic response remains to be determined. The gaseous phase of DE, however, does not
6 have any observable role in the DE-induced lung cancer response in the rat.
7 Available data also indicate that among the traditional animal test species, the rat is the
8 most sensitive species to DE. As reviewed in Chapter 7, Section 7.4, the lung cancer responses
9 in rats from high-concentration exposures to DE appear to be mediated by impairment of lung
1 0 clearance mechanisms owing to particle overload, resulting in persistent chronic inflammation
1 1 and subsequent pathologic and neoplastic changes in the lung. Overload conditions are not
1 2 expected to occur in humans as a result of environmental or most occupational exposures to DE.
1 3 Thus, the animal evidence (i.e., increased lung tumors in the rat) provides additional support for
1 4 identifying a potential cancer hazard to humans, but is considered not suitable for subsequent
1 5 dose-response analysis and estimation of human risk with DE.
1 6 The consistent findings of carcinogenic activity by the organic extracts of DPM in
1 7 noninhalation studies (intratracheal instillation, lung implantation, skin painting) further
1 8 contribute to the overall animal evidence for a human hazard potential for DE.
19
20 9.4.2.2.3. Other key data. Other key data, while not as extensive as the human and animal
2 1 carcinogenicity data, are judged to be supportive of potential carcinogenicity of DE. As
22 discussed above, DE is a complex mixture of hundreds of constituents in either gaseous phase or
23 particle phase. Although present in small amounts, several organic compounds in the gaseous
24 phase (e.g., PAHs, formaldehyde, acetaldehyde, benzene, 1,3 -butadiene) are known to exhibit
25 mutagenic and/or carcinogenic activities. PAHs and PAH derivatives, including nitro-PAHs
26 present on the diesel particle, are also known to be mutagenic and carcinogenic. As reviewed in
27 Chapter 4, DPM and DPM organic extracts have been shown to induce gene mutations in a
28 variety of bacteria and mammalian cell test systems. DPM and DPM organic extracts have also
29 been shown to induce chromosomal aberrations, aneuploidy, and sister chromatid exchange in
30 vitro tests using rodent cells as well as human cells.
3 1 There is also suggestive evidence for the bioavailability of the organic compounds from
32 DE. Elevated levels of DNA adducts in lymphocytes have been reported in workers exposed to
33 DE. In addition, inhalation studies of animals using radio-labeled materials indicate some
"34 (=>ii_itir»n r\f Qrcr3Tiic corpr>ounds from DE after denosition in the lun", ?.s measured bv their
36
7/25/00 9-12 DRAFT — DO NOT CITE OR QUOTE
-------
9.4.2.2.4. Modes of carcinogenic action. As discussed above, there is an adequate
understanding of the modes of action of DE-induced lung tumors in the rat. However, the modes
3 of action by which DE increases lung cancer risks in humans are not fully known. The term
4 "mode of action" refers to a series of key biological events and processes that are critical to the
5 development of cancer. This is contrasted with "mechanisms of action," which is defined as a
6 more detailed description of the complete sequence of biological events at the molecular level
7 that must occur to produce a carcinogenic response.
8 As discussed in Section 7.4, it is likely that multiple modes of action are involved in
9 mediating the carcinogenic effect of DE. These may include (a) mutagenic and genotoxic events
10 (e.g., direct and indirect effects on DNA and effects on chromosomes) by organic compounds in
11 the gas and particle phase, (b) indirect DNA damage via the production of reactive oxygen
12 species (ROS) induced by particle-associated organics, and (c) particle-induced chronic
13 inflammatory response leading to oxidative DNA damage through the release of cytokines, ROS,
14 etc., and an increase in cell proliferation.
15 The particulate phase appears to have the greatest contribution to the carcinogenic effects,
16 and both the particle core and the associated organic compounds have demonstrated carcinogenic
17 properties, although a role for the gas-phase components cannot be ruled out. The carcinogenic
activity of DE also appears to be related to the small size of the particles. Moreover, the relative
contribution of the various modes of action may be different at different exposure levels.
20 Available evidence from animal studies indicates the importance of the role of DE particles in
21 mediating lung tumor response at high exposure levels. Thus, the role of the adsorbed organic
22 compounds may take on increasing importance at lower exposure levels.
23
24 9.4.2.2.5. Weight-of-evidence evaluation. Section 7.5 provides an evaluation of the overall
25 weight of evidence for potential human carcinogenicity in accordance with EPA's Carcinogen
26 Risk Assessment Guidelines (U.S. EPA, 1986, 1996a). The totality of evidence supports the
27 conclusion that DE is a probable human carcinogen (Group Bl) using the criteria as laid out in
28 the 1986 guidelines. A cancer hazard narrative for DE is also provided in accordance with the
29 proposed revised guidelines, which concludes that DE is likely to be carcinogenic to humans by
30 inhalation at any exposure condition. The common bases for either conclusion include the
31 following lines of evidence:
32
33 • strong but less than sufficient evidence for a causal association between DE exposure
34 and increased lung cancer risk among workers of different occupations;
^B • sufficient animal evidence for the induction of lung cancer in the rat from inhalation
36 exposure to high concentrations of DE, DPM, and the elemental carbon core;
7/25/00 9-13 DRAFT—DO NOT CITE OR QUOTE
-------
1 • supporting evidence of carcinogenicity of DPM and the associated organic
2 compounds in rats and mice by noninhalation routes of exposure;
3 • extensive evidence for mutagenic effects of the organic constituents in both
4 particulate matter and gaseous phase, and chromosomal effects of DE, DPM and
5 DPM organics;
6 • suggestive evidence for the bioavailability of DE organics from DE in humans and
7 animals; and
8 • the known mutagenic and carcinogenic activity of a number of individual organic
9 compounds present on the particles (PAHs and their derivatives) and in the gaseous
10 phase (e.g., formaldehyde, acetaldehyde, benzene, 1,3-butadiene, PAHs).
11
12 A major uncertainty in the characterization of the potential cancer hazard of DE at low
13 levels of environmental exposure is the incomplete understanding of its mode of action for the
14 induction of lung cancer in humans. Available data indicate that DE-induced lung
15 carcinogenicity appears to be mediated by mutagenic and nonmutagenic events by both the
16 particles and the associated organic compounds, and that a role for the organics in the gaseous
17 phase cannot be ruled out. Given that there is some evidence for a mutagenic mode of action, a
18 cancer hazard is presumed at any exposure level. This is consistent with EPA's science policy
19 position that assumes a nonthreshold effect for carcinogens in the absence of definitive data
20 demonstrating a nonlinear or threshold mechanism. Because of insufficient information, the
21 human carcinogenic potential of DE by oral and dermal exposures cannot be determined.
22 Several organizations have previously reviewed available relevant data and evaluated the
23 potential human carcinogenicity of DE or the particulate component of DE. Similar conclusions
24 were reached by various organizations (see Table 7-9). For example, some organizations have
25 concluded that DE is probably carcinogenic to humans (IARC, 1989; IPCS, 1996), or reasonably
26 anticipated to be a carcinogen (U.S. DHHS, 2000).
27 Overall, the weight of evidence for potential human carcinogenicity for DE is considered
28 strong, even though inferences are involved in the overall assessment. Major uncertainties of the
29 cancer hazard assessment include the following unresolved issues.
30 First, there has been a considerable scientific debate about the significance of the
31 available human evidence for a causal association between occupational exposure and increased
32 lung cancer risk. Many experts view the evidence as weak, while others consider the evidence as
33 strong. This is due to a lack of consensus about whether the effects of smoking have been
34 adequately accounted for in key studies, and the lack of historical DE exposure data for the
35 available studies.
7/25/00 9-14 DRAFT—DO NOT CITE OR QUOTE
-------
Second, while the mode of action for DE-induced lung tumors in rats from high exposure
is sufficiently understood, the mode of action for lung cancer risk in humans is not fully known.
3 To date, available evidence for the role of both the adsorbed organics and the carbon core particle
4 has been shown to be associated with high-exposure conditions. There is virtually no
5 information about the relative role of DE constituents in mediating carcinogenic effects at the
6 low-exposure levels. Furthermore, there is only a limited understanding regarding the
7 relationship between particle size and carcinogenicity.
8 Third, DE is present in ambient PM (e.g., PM2 5 or PMIO); however, examination of the
9 available PM data has not resulted in the identification of a cancer hazard for ambient PM,
10 although there is some evidence indicating a possible association between ambient PM and lung
11 cancer. Additional research is needed to address these issues to reduce the uncertainty associated
12 with the potential cancer hazard of exposure to DE.
13
14 9.5. DOSE-RESPONSE ASSESSMENT
15 For agents that are known to cause adverse health effects to humans at the exposure of
16 interest, such as the general environment (e.g., air pollutants regulated under the National
17 Ambient Air Quality Standards [ambient PM, ozone, carbon monoxide, sulfur dioxide, nitrogen
oxide, lead, environmental tobacco smoke, etc.]), estimates of human health risks are based on
exposure-/dose-response data of the affected populations. However, for most environmental
20 agents, available health effects information is generally limited to high exposures in studies of
21 humans (e.g., workers) or laboratory animals. For these agents, dose-response assessment is
22 performed in two steps: assessment of observed data to derive a point of departure (which
23 usually is the lowest exposure or dose that induces some, minimal, or no apparent effects),
24 followed by extrapolation to lower exposures to the extent necessary. Human data are always
25 preferred over animal data, if available, as their use obviates the need for extrapolation across
26 species. Extrapolation to low dose is based on the understanding of mode of toxic action of the
27 agent. In the absence of sufficient data that would allow the development of biologically based
28 dose-response models, default methods are generally used to derive toxicity values for estimation
29 of human risks at low doses.
30 For DE, there is sufficient evidence to conclude that acute or short-term inhalation
31 exposure at relatively high levels can cause irritant effects to the eye and upper respiratory tract
32 and inflammation of the lung; however, no quantitative data are available to derive an estimate of
33 human exposure that is not likely to elicit irritant and inflammatory effects in humans.
34 There is also adequate evidence to support the conclusion that DE has the potential to
H cause cancer and noncancer effects of the lung from long-term inhalation exposure. Chapters 6
36 and 8 provide dose-response information and analyses related to the noncancer and cancer
7/25/00 9-15 DRAFT—DO NOT CITE OR QUOTE
-------
1 hazards to humans, respectively, from lifetime exposure to DE. The results of the analyses are
2 discussed below.
3
4 9.5.1. Evaluation of Risk for Noncancer Health Effects
5 As discussed above (Section 9.4), the evidence for potential chronic noncancer health
6 effects of DE is based primarily on findings from chronic animal inhalation studies showing a
7 spectrum of dose-dependent chronic inflammation and histopathological changes in the lung in
8 several animal species including rats, mice, hamsters, and monkeys. On the other hand, available
9 epidemiologic studies of workers exposed to DE, although considered limited because of the lack
10 of exposure information and short exposure duration, have not provided evidence of significant
11 chronic health effects associated with DE exposure, and respiratory symptoms were the only
12 effects reported in a few studies.
13 One approach to derive an estimation of an exposure air level of DE to which humans
14 may be exposed throughout their lifetime without experiencing any untoward or adverse
15 noncancer health effects is to derive a reference concentration (RfC) for DE based on available
16 animal studies. This approach assumes that humans would respond to DE similarly to the tested
17 animals under similar exposure conditions. A major uncertainty of this approach is that animal
18 studies have generally used high DE exposures, and the potential chronic health effects of DE in
19 humans at environmental exposure levels could not be ascertained with available human data. In
20 addition, as DPM is a component of ambient PM, it is conceivable that DPM may partly
.21 contribute to the adverse health effects of ambient PM. Ambient PM has been shown to be
22 statistically associated with increased mortality (especially among people over 65 years of age
23 with preexisting cardiopulmonary conditions) and morbidity, as measured by increases in
24 hospital admissions, respiratory symptoms rates, and decrements in lung function.
25 To address these uncertainties, this assessment also provides two additional approaches
26 for estimating noncancer risk from environmental exposure to DE as bounding estimates. The
27 first approach is to assume that quantitative estimates of risk derived for ambient fine particles
28 (PM2 5) would represent a plausible upper bound for persons potentially exposed to DPM as one
29 of the numerous constituents of ambient PM2 5. Another alternative approach would be to
30 assume equal potency of DPM with other constituents comprising ambient PM25. The support
31 for this approach is that DPM has been shown to have comparable capacity in inducing lung
32 injury in a variety of animal species, as do other poorly soluble particles (ILSI, 2000). Thus,
33 estimation of DE noncancer risks could be based on apportionment of DPM contributions in
34 relationship to the ambient PM-,
3R
36 9.5.1.1. Chronic Reference Concentrations for Diesel Exhaust
7/25/00 9-16 DRAFT—DO NOT CITE OR QUOTE
-------
EPA's Inhalation Reference Concentration Methodology (U.S. EPA, 1994) for the
evaluation of human risks for health effects other than cancer assumes that there is an exposure
3 threshold below which effects will not occur. The RfC can be derived on the basis of either
4 human or animal data. A chronic RfC is defined as "an estimate of a continuous inhalation
5 exposure to the human population, including sensitive subgroups, with uncertainty spanning
6 perhaps an order of magnitude, that is likely to be without appreciable risks of deleterious
7 noncancer effects during a lifetime. " The RfC is not a bright line; rather, as the human exposure
8 increases above the RfC, the margin of protection decreases.
B In the absence of exposure-response data in humans, this assessment derives an RfC for
10 DE based on dose-response data from four chronic inhalation studies in rats (Mauderly et al.,
11 1987; Ishinishi et al., 1988; Heinrich et al., 1995; Nikula et al., 1995). All of these four studies
12 used DPM (expressed as ng/m3) as a measure of DE exposure. The pulmonary effects, including
13 inflamation and histopathologic lesions, were considered to be the critical noncancer effects. As
14 shown in Table 6-2, the no-observable-adverse-effects levels (NOAELs), the lowest-observable-
15 adverse-effects levels (LOAELs), and the adverse effects levels (AELs) for lung inflammation
16 and histopathologic changes were identified for the first three studies. Lower 95% confidence
17 estimates of the concentrations of DPM associated with a 10% incidence (BMCL,0) of chronic
pulmonary inflammation and fibrosis were derived for the Nikula et al. study. Human equivalent
concentrations (HECs) corresponding to the animal exposure levels (NOAEL, LOAEL, AEL,
20 BMCL10) were then computed by using a dosimetry model developed by Yu et al. (1991) as
21 described in Chapter 6, Section 6.5.2, and Appendix A. The dosimetry model accounts for
22 species differences (rat to human) in respiratory exchange rates, particle deposition efficiency,
23 differences in particle clearance rates at high and low doses, and transport of particles to lymph
24 nodes.
25 The highest HEC value associated with no apparent effects, i.e., a NOAEL of 0.14 ug/m3
26 was selected as the point of departure for deriving an RfC. To obtain the RfC, this point of
27 departure was then divided by an uncertainty factor (UF) of 10 to account for inter-individual
28 variation. In the absence of mechanistic or specific data, a default value of 10 is considered
29 appropriate to account for possible human variability in sensitivity, particularly for children and
30 people with preexisting respiratory conditions. The resulting RfC for DE is 14 ng/m3 of DPM.
31 Overall, the confidence level of the RfC assessment for DE is considered medium. A
32 principal uncertainty of the assessment is the reliance on animal data to predict human risk. The
33 critical effects, chronic inflammation and pathologic changes, which are well characterized in
34 four animal species, are considered relevant to humans. Collective evidence for all poorly
^fc soluble particles indicates that the rat is the most sensitive laboratory animal species tested to
36 date and appears to be more sensitive to lung injury induced by any solid particles (including DE)
7/25/00 9-17 DRAFT—DO NOT CITE OR QUOTE
-------
1 than the human (ILSI, 2000). In addition, differences in particle deposition, retention, and
2 clearance mechanisms have been addressed to some extent by the use of the rat-to-human
3 dosimetry model. Thus, the use of rat data is not likely to underestimate human risk for
4 noncancer health effects. In addition, available toxicologic information for DE is relatively
5 complete, as it has been extensively tested in standard toxicologic studies. Still, some
6 uncertainties remain given that there is growing evidence suggesting the potential for DE to
7 cause immunological effects and/or to exacerbate allergenic effects to known sensitizers. The
8 potential relevance for these health endpoints to public health is significant because of increases
9 in the number of individuals with preexisting respiratory conditions and possible interactions
10 with other air pollutants.
11
12 9.5.1.2. Risks Based on Ambient PM2.5
13 As discussed in Chapter 6 (Section 6.3), the EPA has promulgated a long-term PM2 5
14 NAAQS of 15 ug/m3 as an acceptable level for annual-average fine particles to protect against
15 effects from chronic exposure. The standard is based on combined findings of excess daily
16 mortality and morbidity from short-term exposures and findings from long-term fine PM studies
17 (e.g., Harvard Six City and ACS studies) showing increases in mortality around or above the
18 annual average level of 15 ug/m3. If one assumes that the adverse health effects of ambient fine
19 particles are due entirely to DPM, i.e., that DPM is exceptionally toxic, then any characterization
20 of health effects attributable to ambient fine particles could therefore represent an upper-limit
21 estimate for DPM. Accordingly, the upper-limit for DE would be 15 ug/m3.
22
23 9.5.1.3. Apportionment Method Based on Ambient PM2,S
24 As discussed in Chapters 2 and 6, DPM is a component of ambient PM. In some urban
25 areas, the fraction of PM2 5 attributable to DPM from DE sources may exceed 30%, although the
26 proportion appears to be more typically in the range of 10%. If one assumes that DPM is as toxic
27 as other constituents of ambient PM2 5, then ambient concentration to DPM needs to be below the
28 range of 1.5 to 5.0 ug/m3 (i.e., 10% x 15 ug/m3 to 30% x 15 ug/m3) to achieve the same
29 protection for the annual average standard for ambient fine particles of 15 ug/m3.
30
31 9.5.1.4. Conclusions
32 Three approaches are used to estimate an exposure air level of DE (as measured by DPM)
33 to which humans may be exposed throughout their lifetime without experiencing any untoward
34 or adverse noncancer health effects. The RfC method produces an RfC of 14 ug/m3 of DPM on
3R the basis of four chronic inhalation studies of DE in rats. This value is almost the same as the
36 long-term PM25 NAAQS of 15 ug/m3, and close to the 1.5 to 5.0 ug/m3 derived from the
7/25/00 9-18 DRAFT— DO NOT CITE OR QUOTE
-------
apportionment of the PM2 5 standard. As the accuracy of the RfC is part of the definition
("within an order of magnitude "), this dose-response estimate could be considered not to be
substantially different from the other two approaches. This congruence of estimates attests to the
4 reasonableness of the data used and the judgments made in the RfC process, as well as tending to
5 support the accuracy of the estimates of DPM within ambient PM2 5. This congruence of
6 independent methods should also increase overall confidence in these estimates regarding
7 toxicity of DE and its potential health risks for the human population.
8
9 9.5.2. Evaluation of Cancer Risks
10 As discussed above (Section 9.4.3), the combined weight of evidence indicates that DE
11 has the potential to pose a cancer hazard to humans at anticipated levels of environmental
12 exposure. The critical target organ of DE-induced carcinogenicity is the lung. Strong but less
13 than sufficient evidence exists for a causal relationship between risk for lung cancer and
14 occupational exposure to DE in certain occupational workers such as railroad workers, truck
15 drivers, heavy equipment operators, transit workers, etc. In addition, it has been shown
16 unequivocally in several studies that DE can cause benign and malignant lung tumors in rats in a
17 dose-related manner following chronic inhalation exposure to high concentrations. The
18 mechanism(s) by which DE induces lung cancer in humans has not been established, but
^B available data indicate that mutagenic and nonmutagenic modes of action are possible. Hence,
20 for estimating DE cancer risk at low environmental exposures, linear low-dose extrapolation is
21 considered most appropriate, which is consistent with EPA's science policy position that in the
22 absence of an understanding of modes of carcinogenic action, a nonthreshold effect is to be
23 presumed (U.S. EPA, 1986, 1996a). This approach is consistent with the approaches taken by
24 other organizations or individuals who have previously used either linear risk extrapolation
25 models or mechanistically based models to estimate cancer risk from environmental exposure to
26 DE (e.g., IPCS, 1996; Cal EPA, 1998; also see Appendix D).
27 Dose-response assessment is generally based on either human or animal data, although
28 human data are always preferred if available. Many quantitative assessments have been
29 conducted by several organizations and investigators on the basis of both occupational data and
30 rat data (see Appendix D). However, more recent cumulative evidence indicates that DE causes
31 tumors in the rat via a mode of action that involves impairment of lung clearance mechanisms
32 (referred to as "lung overload response") associated with high exposures. Although the dose-
33 response for increases in lung tumors in rats is supportive for identifying a cancer hazard in
34 humans, the mode of action in the rat is not expected to be operative at environmental exposure
P conditions. Therefore, the rat lung tumor dose-response data are not considered suitable for
predicting human risk at low environmental exposures. Given that the rat data are not
7/25/00 9-19 DRAFT—DO NOT CITE OR QUOTE
-------
1 appropriate for estimating cancer risk to humans, this assessment focuses on the use of
2 occupational data for estimating environmental risk of DE to humans.
3 Even though occupational data are considered most relevant for use hi dose-response
4 assessment, considerable uncertainties exist, including the following issues:
5
6 • the use of DPM (expressed as |ag/m3) as a surrogate dosimeter for DE exposure, given
7 that the relative roles of various constituents in mediating carcinogenic effects and the
8 mode of carcinogenic action are still not fully known;
9 • the representativeness of occupational populations for the general population and
1 0 vulnerable subgroups, including infants and children and individuals with preexisting
1 1 diseases, particularly respiratory conditions;
12 • the lack of actual DE workers' exposure data in available epidemiologic studies;
13 • possible confounders (smoking and asbestos exposure) that could contribute to the
1 4 observed lung cancer risk in occupational studies of DE; and
1 5 • whether or not exposure-response relationships for lung cancer risks have been
1 6 demonstrated for available occupational studies of DE.
17
1 8 Chapter 8, Section 8.3 provides a discussion of these uncertainties, along with an evaluation of
1 9 the suitability of available occupational studies for a derivation of a cancer unit risk estimate for
20 DE. Unit risk is defined as the estimated upper-bound cancer risk at a specific exposure or dose
2 1 from a continuous average lifetime exposure of 70 years (in this case, cancer risk per u.g/m3 of
22 DPM).
23 Among the occupational studies, the railroad worker studies (Garshick et al., 1987, 1988)
24 and the Teamsters Union truck driver studies (Steenland et al., 1990, 1998) are considered to
25 have the best available exposure data for possible use in establishing exposure-response
26 relationships and deriving a cancer unit risk. There have been different views on the suitability
27 of either set of studies for estimating environmental cancer risks (e.g., Cal EPA, 1998; HEI,
28 1995, 1999). Given the equivocal evidence for the presence or absence of an exposure-response
29 relationship for the studies of railroad workers, and exposure uncertainties for the studies of truck
30 drivers, it is judged that available data are too uncertain at this time for a confident quantitative
3 1 dose-response analysis and subsequent derivation of cancer unit risk for DE.
32 In the absence of a cancer unit risk to assess environmental cancer risk, this assessment
33 provides some perspective about the possible magnitude of risk from environmental exposure to
34 DE. One approach involves examining the differences between the levels of occupational and
35 ambient environmental exposures bv d) using a nationwide average and upper limit
36 environmental exposures of 0.8 u.g/m3 and 4 ng/m3, respectively, and (2) assuming that cancer
o
~
-------
1 risk to DE is linearly proportional with cumulative lifetime exposure. Risks to the general public
f would be low in comparison with occupational risk, if the differences in exposure are large. On
the other hand, if the differences are small, the environmental risks would approach the workers'
4 risk observed in studies of past occupational exposures. The comparative exposure analysis
5 indicates that for certain occupations, there is a potential for overlap between environmental
6 exposure and environmental equivalent of occupational exposure, having exposure margins of
7 less than 1 to about 460 (see Table 8-1). When environmental exposure is at the high end, the
8 resultant cancer risk may approach that of workers in certain occupations.
9 A second approach is to derive a rough estimate of lung cancer risks from occupational
10 exposures to DE, and then take into account the exposure margins between occupational and
11 environmental exposures to derive an upper limit range of possible lung cancer risks from
12 lifetime environmental exposure to DE. Given the range of observed relative risks or odds ratios
13 of lung cancer in a number of occupational studies (1.2 to 2.6) and the pooled relative risk
14 estimates from two independent meta-analyses (1.35 and 1.47), a relative risk of 1.4 is selected as
15 a reasonable estimate for the purpose of this analysis. The relative risk of 1.4 means that the
16 workers faced an extra risk that is 40% higher than the approximately 5% background lifetime
17 lung cancer risk in the U.S. population.2 Thus, using the relationship [excess risk = (relative
18 risk-1) x background risk], 2% (10~2) of these DE-exposed workers would have been at risk (and
!^P developed lung cancer) attributable to occupational exposure to DE [(1.4-1) * 0.05) = 0.02].
20 Using a nationwide average environmental exposure (0.8 ug/m3 DPM), and assuming (a)
21 the excess lung cancer risk from occupational exposure is about 10"2; (b) the risks fall
22 proportionally with reduced exposure; and (c) the past occupational exposures were at the high
23 end of the range (about 1740 ug/m3 which corresponds to an environmental equivalent exposure
24 of 365 ug/m3, resulting in an exposure margin of 457), then the environmental cancer risk could
25 be between 10"4 to 10"5. On the other hand, if occupational exposures for some groups were
26 lower, e.g., closer to 100 ug/m3, (i.e., an equivalent environmental exposure of 21 ug/m3 with an
27 exposure margin of 25), the environmental risk would approach 10"3.
2 The background rate of 0.05 is an approximated lifetime risk calculated by the method of lifetable
analysis using age-specific lung cancer mortality data and probability of death in the age group taken from the
National Health Statistics (HRS) monographs of Vital Statistics of the U.S. (Vol. 2, Part A, 1992). Similar values
based on two rather crude approaches can also be obtained: (1) 59.8 x 10E-5/8.8 x 10E-3 = 6.8 x 10E-2, where
59.8 x 10E-5 and 8.8 x 10E-3 are, respectively, the crude estimates of lung cancer deaths (including intrathoracic
organs, estimated to be fewer than 105 of the total cases) and total deaths for 1996 reported in the Statistical
Abstract of the U.S. Bureau of the Census (1998, 118th Edition), and (2) 156,900/270,000,000 x 76 = 0.045, where
156,900 is the projected number of lung cancer deaths for the year 2000 as reported in Cancer Statistics 9J of the
American Cancer Society, Jan/Feb 2000; 270,000,000 is the current U.S. population; and age 76 is the expected
lifespan.
7/25/00 9-21 DRAFT—DO NOT CITE OR QUOTE
-------
1 The analyses presented above are not intended to be precise, but are useful in gauging the
2 possible range of risk based on applying scientific judgment and simple risk exploration methods
3 to the relative risk findings from the epidemiologic studies. The analyses provide a sense of
4 where an upper limit (or "upper bound") of the risk may be. The simple methodologies used are
5 generic hi that they are valid for any increased relative risk data, and thus are not unique to the
6 DE data. It should be pointed out that these analyses are subject to considerable uncertainties,
7 particularly the lack of actual exposure information and the underlying assumption that cancer
8 risk is linearly proportional to cumulative exposure. Nevertheless, these analyses, which include
9 the use of public health conservative assumptions, indicate that environmental exposure to DE
10 may pose a lifetime cancer risk that could range from 10"5 to 10~3. These findings are general
11 indicators of the potential significance of the lung cancer hazard, and should not be viewed as a
12 definitive quantitative characterization of risk. Further research is needed to more accurately
13 assess and characterize environmental cancer risks from DE.
14
15 9.6. SUMMARY AND CONCLUSIONS
16 Adverse human health effects may result from current-day environmental exposure to
17 DE. DE may cause acute and chronic respiratory effects and has the potential to cause lung
18 cancer in humans.
19 DE may cause acute irritation to the eye and upper respiratory airways, and mild
20 respiratory symptoms at relatively high exposures. DE may also have immunological properties
21 and may induce allergic responses and/or exacerbate existing respiratory allergies. Quantitative
22 dose-response estimates for these effects could not be developed because of the lack of exposure-
23 response information for these acute and short-term effects.
24 Long-term exposure to low levels of DE may cause chronic inflammation and
25 pathological changes in the lung. The RfC for chronic respiratory effects is estimated to be 14
26 Mg/m3 of DPM. This value is almost the same as the long-term PM25 NAAQS of 15 ng/m3, and
27 close to the 1.5 to 5.0 |ig/m3 derived from an apportionment of DPM from the PM25 standard.
28 The congruence of these estimates supports the reasonableness of the data used and the accuracy
29 of the risk estimates of DPM within ambient PM, v This congruence should also increase the
30 overall confidence that these estimates identify a protective exposure level for the chronic
31 toxicity of DE and its potential health risks for the human population.
32 DE is considered to be a probable human carcinogen, or is likely to be carcinogenic in
33 humans, by inhalation under any exposure condition. Because of considerable uncertainty in the
34. available exposure-response data, a cancer unit risk for DE has not been derived at this time.
.TR Simple analyses using conservative assumptions provide a perspective of the possible range of
36 lung cancer risk from environmental exposure to DE. These analyses indicate that lifetime
7/25/00 9-22 DRAFT—DO NOT CITE OR QUOTE
-------
cancer risk could range from 10~5 to 10"3. These analyses are subject to considerable
uncertainties, particularly the lack of actual exposure information and the underlying assumption
3 that cancer risk is linearly proportional to cumulative exposure. Nevertheless, these findings are
4 general indicators of the potential significance of the lung cancer hazard, although they should
5 not be viewed as a definitive quantitative characterization of risk.
6 Even though the evidence for potential human health effects of DE is convincing and
7 persuasive, uncertainties exist because of the use of many assumptions to bridge data and
8 knowledge gaps about human exposures to DE and the underlying mechanisms by which DE
9 causes observed toxicities in humans and animals. As discussed in Section 9.2, a major
10 uncertainty of this assessment is how the physical and chemical nature of past exposures to DE
11 compares with present-day exposures, and how the DE exposure-response data from
12 occupational and toxicological studies can be used for the characterization of possible hazard and
13 risk from present-day environmental exposures. Available data are not sufficient to provide
14 definitive answers to these questions, as the modes of action for DE toxicity and carcinogenicity
15 are still not known. Clearly, there have been qualitative and quantitative differences in DE
16 emissions and their physical and chemical composition. Given that the changes in DE (e.g.,
17 DPM) over time cannot be quantified, and that the mode of action for DE toxicity is unknown,
this assessment assumes that prior-year toxicologic and epidemiologic findings can be applied to
more current exposures, both of which use DPM mass as the dosimeter.
20 Other uncertainties include the assumptions that health effects observed at high doses
21 may be applicable to low doses, and that toxicologic findings in laboratory animals are predictive
22 of human responses. Available data are not sufficient to demonstrate the presence or absence of
23 an exposure-dose-response threshold for DE toxicity and carcinogenicity. This is due to the lack
24 of complete understanding of how DE may cause adverse health effects in exposed humans and
25 laboratory animals. Although there are hypotheses about the specific mechanisms by which DE
26 might cause cancer and other toxicities, no specific biological pathways or specific constituents
27 of DE have been firmly established as the responsible agents for low-dose effects. The
28 assumptions used in this assessment, i.e., a biological threshold for chronic respiratory effects
29 and the absence of a threshold for lung cancer, are considered prudent and reasonable.
30 The assessment assumes that the potential DE health hazards are for average healthy
31 . adults. There is no DE-specific information that provides direct insight into the question of
32 variable susceptibility within the general human population and vulnerable subgroups. Although
33 default approaches to account for uncertainty in interindividual variation have been included in
34 the derivation of the RfC (i.e., use of an uncertainty factor of 10), they may not be adequately
^P protective for certain vulnerable subgroups. For example, adults who predispose their lungs to
36 increased particle retention (e.g., smoking, high particulate burdens from nondiesel sources),
7/25/00 9-23 DRAFT—DO NOT CITE OR QUOTE
-------
1 have existing respiratory or lung inflammation or repeated respiratory infections, or have chronic
2 bronchitis or asthma could be more susceptible to adverse impacts from DE exposure. Infants
3 and children could also have a greater susceptibility to the acute/chronic toxicity of PM2 5, of
4 which DPM is a part, because of a greater breathing frequency, resulting in greater respiratory
5 tract particle deposition. Increased respiratory symptoms and decreased lung function in children
6 have been associated with ambient PM levels (U.S. EPA, 1996b). Despite these uncertainties,
7 the default approach for using a UF of 10 to account for possible interindividual variation in
8 reaction to DE is appropriate and reasonable given the lack of DE-specific data.
9 Variation in DE exposure is another source of uncertainty. Because of variation in
10 activity patterns, different population subgroups could potentially receive higher or lower
11 exposure to DE depending on their proximity to DE sources. The highest exposed are clearly
12 occupational subgroups whose job brings them very close to diesel emission sources, such as
13 trucking industry workers, engine mechanics, some types of transit operators, railroad workers,
14 diesel powered machinery operators, underground miners, etc. High exposures in the general
15 population would be to those living very near or having time outdoors in proximity to diesel
16 engine exhaust sources. For example, children with outdoor playtime adjacent to roadways
17 where diesel-engine vehicles are in use are likely to have higher DE exposures. Accordingly, DE
18 exposure estimates used in this assessment have included possible high-end exposures as
19 bounding estimates.
20 Lastly, this assessment considers only potential heath effects from exposures to DE alone.
21 DE exposure could be additive or synergistic to concurrent exposures to many other air
22 pollutants. For example, there is suggestive evidence that DPM that has been altered by being in
23 the presence of ambient ozone may significantly increase the rat lung inflammatory effect
24 compared to DPM that was not subjected to ozone (Madden et ah, 2000). It would follow then
25 that DPM in areas with ambient ozone present could be more potent in causing noncancer
26 inflammatory effects. Other concerns include the possible impacts for children and adults on the
27 potentiation of allergenicity from DE exposure. However, in the absence of more definitive data
28 demonstrating interactive effects from combined exposures to DE and other pollutants, it is not
29 possible to further address these issues at this time.
30
31 9.7. REFERENCES
32
33 Cal EPA (California Environmental Protection Agency-OEHHA). (1998) Part B: Health risk assessment for diesel
34 exhaust, Public and Scientific Review Draft. February 1998.
35
36 Garshick, E; Schenker, MB; Munoz, A; et al. (1987) A case-control study of lung cancer and diesel exhaust exposure
37 in railroad workers. Am Rev Kespir Uis 135:1242-1248.
5/00 9-24 DRAFT—DO NOT CITE OR QUOTE
-------
Garshick, E; Schenker, MB; Munoz, A; et al. (1988) A retrospective cohort study of lung cancer and diesel exhaust
exposure in railroad workers. Am Rev Respir Dis 137:820-825.
4 Health Effects Institute (HEI). (1995) Diesel exhaust: a critical analysis of emissions, exposure, and health effects.
5 Cambridge, MA: Health Effects Institute.
6
7 HEI. (1999) Diesel emissions and lung cancer: epidemiology and quantitative risk assessment. A special report of
8 the Institute's Diesel Epidemiology Expert Panel. Cambridge, MA: Health Effects Institute.
9
10 Heinrich, U; Fuhst, R; Rittinghausen, S; et al. (1995) Chronic inhalation exposure of Wistar rats and two different
11 strains of mice to diesel engine exhaust, carbon black, and titanium dioxide. Inhal Toxicol 7:553-556.
12
13 International Agency for Research on Cancer (IARC). (1989) Diesel and gasoline engine exhausts and some
14 nitroarenes. IARC monographs on the evaluation of carcinogenic risks to humans: v. 46. Lyon, France: World
15 Health Organization; pp. 41-185.
16
1 7 International Life Sciences Institute (ILSI). (2000) ILSI Risk Science Institute workshop: the relevance of the rat
1 8 lung response to particle overload for human risk assessment. Gardner, DE, ed. Inhal Toxicol 12(1-2).
19
20 International Programme on Chemical Safety (IPCS), World Health Organization. (1996) Diesel fuel and exhaust
21 emissions. Environmental Health Criteria 171. Geneva: World Health Organization.
22
23 Ishinishi, N; Kuwabara, N; Takaki, Y; et al. (1988) Long-term inhalation experiments on diesel exhaust. In: Diesel
24 exhaust and health risks. Results of the HERP studies. Ibaraki, Japan: Research Committee for HERP Studies; pp.
25 11-84.
Madden, M; Richards, J; Dailey, L; et al. (2000) Effect of ozone on diesel exhaust toxicity in rat lung. Toxicol Appl
28 Pharmacol: in press.
29
30 Mauderly, JL; Jones, RK; Griffith, WC; et al. (1987) Diesel exhaust is a pulmonary carcinogen in rats exposed
31 chronically by inhalation. Fundam Appl Toxicol 9:208-221.
32
33 Nikula, KJ; Snipes, MB; Barr, EB; et al. (1995) Comparative pulmonary toxicities and carcinogenicities of
34 chronically inhaled diesel exhaust and carbon black in F344 rats. Fundam Appl Toxicol 25:80-94.
35
36 Steenland, K; Silverman, DT; Homung, RW. (1990) Case-control study of lung cancer and truck driving in the
37 Teamsters Union. Am J Public Health 80:670-674.
38
39 Steenland, K; Deddens, J; Stayner, L. (1998) Diesel exhaust and lung cancer in the trucking industry: exposure-
40 response analyses and risk assessment. Am J Ind Med 34:220-228.
41
42 U.S. Department of Health and Human Services (DHHS). (2000) 9th report on carcinogens. National Toxicology
43 Program, Research Triangle Park, NC. http://ntp-server.niehs.nih.gov.res.
44
45 U.S. Environmental Protection Agency (U.S. EPA). (1986) Guidelines for carcinogen risk assessment. Federal
46 Register 51(185):33992-34003..
47
48 U.S. EPA. (1994) Methods for derivation of inhalation reference concentrations and application of inhalation
dosimetry. Research Triangle Park, NC: Office of Research and Development, National Center for Environmental
Assessment; EPA/600/8-90/066F.
51
7/25/00 9-25 DRAFT—DO NOT CITE OR QUOTE
-------
1 U.S. EPA. (1996a) Proposed guidelines for carcinogen risk assessment. Office of Research and Development.
2 Federal Register 61(79):17960-18011. EPA/600/P-92/003C.
3
4 U.S. EPA. (1996b) Air quality criteria for paniculate matter. National Center for Environmental Assessment;
5 Research Triangle Park, NC: EPA/600/P-95/001 aF.
6
7 Yu, CP; Yoon, KJ; Chen, YK. (1991) Retention modeling of DE particles in rats and humans. J Aerosol Med
8 4:79-115.
9
7/25/00 9-26 DRAFT—DO NOT CITE OR QUOTE
-------
Appendix A
Calculation of Human Equivalent Continuous
Exposure Concentrations (HECs)
7/25/00 A-1 DRAFT—DO NOT CITE OR QUOTE
-------
A.l. INTRODUCTION
As discussed in Chapter 3, the lung burden of diesel participate matter (DPM) during
3 exposure is determined by both the amount and site of particle deposition in the lung and,
4 subsequently, by rates of translocation and clearance from the deposition sites. Mathematical
5 models have often been used to complement experimental studies hi estimating the lung burdens
6 of inhaled particles in different species under different exposure conditions. This appendix
7 presents a mathematical model that simulates the deposition and clearance of DPM in the lungs
8 of rats and humans of Yu et al.(1991) also published as Yu and Yoon (1990).
9 Diesel particles are aggregates formed from primary spheres 15-30 nrn in diameter. The
10 aggregates are irregularly shaped and range in size from a few molecular diameters to tens of
11 microns. The mass median aerodynamic diameter (MMAD) of the aggregates is typically 0.2
12 [im and is poly disperse with a geometric standard deviation of around 2.3. The organics
13 adsorbed onto the aggregates normally account for 10% to 30% of the particle mass. However,
14 the exact size distribution of DPM and the specific composition of the adsorbed organics depend
15 upon many factors, including engine design, fuels used, engine operating conditions, and the
16 thermodynamic process of exhaust. The physical and chemical characteristics of DPM have been
17 reviewed extensively by Amann and Siegla (1982) and Schuetzle (1983).
»Four mechanisms deposit DPM within the respiratory tract during exposure: impaction,
sedimentation, interception, and diffusion. The contribution from each mechanism to deposition,
20 however, depends upon lung structure and size, the breathing condition of the subject, and
21 particle size distribution. Under normal breathing conditions, diffusion is the most dominant
22 mechanism and the other three mechanisms play minor roles.
23 Once DPM is deposited hi the respiratory tract, both the carbonaceous core and the
24 adsorbed organics will be removed from the deposition sites by mechanical clearance, provided
25 by mucociliary transport in the ciliated conducting airways as well as macrophage phagocytosis
26 and migration hi the nonciliated airways, and dissolution. As the carbonaceous core or soot of
27 DPM is insoluble, it is removed from the lung primarily by mechanical clearance, whereas the
28 adsorbed organics are removed principally by dissolution (Chapter 3).
29
30 A.2. PARTICLE MODEL
31 To develop a mathematical model that simulates the deposition and clearance of DPM in
32 the lung, an appropriate model for diesel particles must be introduced. For the deposition study,
33 an equivalent sphere model developed by Yu and Xu (1987) was used to simulate the dynamics
§and deposition of DPM in the respiratory tract by various mechanisms. For the clearance study, a
diesel particle is assumed to be composed of three different material components according to
7/25/00 A-2 DRAFT—DO NOT CITE OR QUOTE
-------
1 their characteristic clearance rates: (1) a carbonaceous core of approximately 80% of the particle
2 mass; (2) absorbed organics of about 10% of particle mass, which are slowly cleared from the
3 lung; and (3) adsorbed organics quickly cleared from the lung, accounting for the remaining 10%
4 of particle mass. The presence of two discrete organic phases in the particle model is suggested
5 by observations that the removal of particle-associated organics from the lung exhibits a biphasic
6 clearance curve (Sun et al., 1984; Bond et al., 1986), as discussed in Chapter 3. This curve
7 represents two major kinetic clearance phenomena: a fast- phase organic washout with a half-
8 time of a few hours, and a slow phase with a half-time that is a few hundred times longer. The
9 detailed components involved hi each phase are not known. It is possible that the fast phase
10 consists of organics that are leached out primarily by diffusion mechanisms while the slow phase
11 might include any or all of the following components: (a) organics that are "loosened" before
12 they are released, (b) organics that have become intercalated in the carbon core and whose release
13 is thus impeded, (c) organics that are associated for longer periods of time because of
14 hydrophobic interaction with other organic-phase materials, (d) organics that have been ingested
15 by macrophages and as a result effectively remain in the lung for a longer period of time because
16 of metabolism by the macrophage (metabolites formed may interact with other cellular
17 components), and (e) organics that have directly acted on cellular components, such as the
18 formation of covalent bonds with DNA and other biological macromolecules to form adducts.
19 The above distinction of the organic components is general and made to account for the
20 biphasic clearance of DPM; it does not specifically imply the actual nature of the adsorbed
21 organics. For aerosols made of pure organics, such as benzo(a)pyrene (BaP) and nitropyrene
22 (NP) in the same size range of DPM, Sun et al. (1984) and Bond et al. (1986) observed a nearly
23 monophasic clearance curve. This might be explained by the absence of intercalative phenomena
24 (a) and of hydrophobic interaction imposed by a heterogeneous mixture of organics (b). The
25 measurement of a pure organic might also neglect that quantity which has become intracellularly
26 (c) or covalently bound (d).
27
28 A3. CQMPARTMENTAL LUNG MODEL
29 The model of Yu et al. (1991) comprises three principal compartments involved in
30 deposition and clearance: tracheobronchial (T or TB), alveolar (A), and lung-associated lymph
31 node (L), as shown in Figure A-l. The outside compartments blood (B) and GI tract (G) and
32 nasopharyngeal or head (H) are also represented. The alveolar compartment in the model is
33 obviously the most important for long-term retention studies. However, for short-term
3-T consideration, retentions in other lung conipaitniciits may also be significant. The presence of
7/25/00 A-3 DRAFT—DO NOT CITE OR QUOTE
-------
these lung compartments and the two outside compartments in the model therefore provides a
complete description of all clearance processes involved.
3 In Figure A- 1 , r $ r $ and r (l are, respectively, the mass deposition rates of DE material
4 component i (i=l [core], 2 [slowly cleared organics], and 3 [rapidly cleared organics]) in the
5 head, tracheobronchial, and alveolar compartments; and %& represents the transport rate of
6 material component i from any compartment X to any compartment Y. Let the mass fraction of
7 material component i of a diesel particle be /.. Then
=
=firH>
r? =ftrT, (A-2)
8
'A =firA> (A-3)
^L where rH, rT, and rA are, respectively, the total mass deposition rates of DPM in the H, T, and
^^ A compartments, determined from the equations:
11
12
13
rH = c(TV)(RF)(DF)H , (A-4)
14
rT = c(TV)(RF)(DF)T , (A-5)
15
rA = c(TV)(RF)(DF)A . (A-6)
16
17
7/25/00 A-4 DRAFT—DO NOT CITE OR QUOTE
-------
1 In Equations A-4 to A-6, c is the mass concentration of DPM in the air, TV is the tidal
2 volume, RF is the respiratory frequency, and (DF)H, (DF)T, and (DF)A are, respectively, the
3 deposition fractions of DPM in the H, T, and A compartments over a respiratory cycle. The
4 values of (DF)H, (DF)T, and (DF)A, which vary with the particle size, breathing conditions, and
5 lung architecture, were determined from the deposition model of Yu and Xu ( 1 987).
6 The differential equations for m$, the mass of material component i in compartment X as
7 a function of exposure time t, can be written as
8
9 Head (H)
ff _ _(0 l(0_,(0 1 «„,(')
-- rff ~ *HGmH ~ A-HBMff »
1 0 Tracheobronchial (T)
dm®
14-//1 T* /j\ /PJ\ /.-\ fj\ /A /_-\ /-\
fifr
1 1 Alveolar (A)
12
1 3 Lymph nodes (L)
dr
14 Equation A-9 may also be written as
dm®
,, ' A ' nAT"A ~ ^AL"1A ~ "-Atf"A > (A-9)
at
JL_(0
"*" »» »<" . (A-10)
(0 , (J) (i)
-
1 5 where
.
16 is the total clearance rate of material component i from the alveolar compartment. In Equations
1 7 A-7 to A-10. we have assumed vanishing material concentration in the blood compartment tc
1 8 calculate diffusion transport.
7/25/00 A-5 DRAFT— DO NOT CITE OR QUOTE
-------
The total mass of the particle-associated organics in compartment X is the sum of m ^
and m ^;the total mass of DPM in compartment X is equal to
mx = m? + mf + mf (A- 13)
4
5 The lung burdens of diesel soot (core) and organics are defined, respectively, as
6
7
8 and
9
10
Because the clearance of diesel soot from compartment T is much faster than from compartment
A, m (}}< m % a short tune after exposure, Equation A-14 leads to
13
14
1 5 Solution to Equations A-7 to A- 10 can be obtained once all the transport rates /2$ are
1 6 known. When A$ are constant, which is the case in linear kinetics, Equations A-7 to A- 10 will
1 7 have a solution that increases with time at the beginning of exposure but eventually saturates and
1 8 reaches a steady-state value. This is the classical retention model developed by the International
1 9 Commission of Radiological Protection (ICRP, 1979). However, as discussed in Chapter 3, data
20 have shown that when rats are exposed to DPM at high concentration for a prolonged period,
2 1 long-termed clearance is impaired. This is the so-called overload effect, observed also for other
22 insoluble particles. The overload effect cannot be predicted by the classical ICRP model.
23 Soderholm (1 98 1 ) and Strom et al. (1987, 1 988) have proposed a model to simulate this effect by
24 adding a separate sequestering compartment in the alveolar region. In the present approach, a
25 single compartment for the alveolar region of the lung is used and the overload effect is
accounted for by a set of variable transport rates A$> /L%1 and A$ which are functions of mA.
7/25/00 A-6 DRAFT— DO NOT CITE OR QUOTE
-------
23
1 The transport rates A$ and A^ in Equations A-7 to A- 10 can be determined directly from
2 experimental data on lung and lymph node burdens, and Xjk and Jfy from Equation A- 12.
3
4 A.4. SOLUTIONS TO KINETIC EQUATIONS
5 Equation A- 1 1 is a nonlinear differential equation of m % with known function o
6 For diesel soot, this equation becomes
7
8
9 Because clearance of the particle-associated organics is much faster than diesel soot, m^and rd%
1 0 constitute only a very small fraction of the total particle mass (less than 1%) after a long
1 1 exposure, and we may consider Jf^as a function of m^alone. Equation A- 17 is then reduced to a
1 2 differential equation with m^the only dependent variable.
1 3 The general solution to Equation A- 17 for constant rj^at any time, t, can be obtained by
1 4 the separation of variables to give
15
dm
,0)
-7TT = l • (A-18)
16
1 7 If rj^is an arbitrary function oft, Equation A- 17 needs to be solved numerically such as
18 by a Runge-Kutta method. Once m^is found, the other kinetic equations A-7 to A- 10 for both
1 9 diesel soot and the particle-associated organics can be solved readily, as they are linear equations.
20 The solutions to these equations for constant r$, r$ and rf are given below:
21 Head(H)
22
. (!) . if\ . fi\
wnere /.# = A£G + A.^ (A-20)
7/25/00 A-7 DRAFT—DO NOT CITE OR QUOTE
-------
21
Tracheobronchial (T)
? = exp (-J.J? O | ' ( 4° + A«r m» ) exp (k(f t) dt
Jo
, + . (A-22)
4
5 Lymph nodes (L)
6
|
Jo
(A-23)
7
8 In Equations A- 1 9 to A-23 , m $o represents the value of m $> at t = 0.
In the sections to follow, the methods of determining r$, r$ and /^ or (DF)H, (DF)T, and
(DF)A /D$ rfD$ and r^j9 as well as the values of /i^ in the compartmental lung model are
1 1 presented.
12
1 3 A.5. DETERMINATION OF DEPOSITION FRACTIONS
1 4 The mathematical models for determining the deposition fractions of DPM in various
1 5 regions of the respiratory tract have been developed by Yu and Xu (1986, 1987) and are adopted
16 in this report. Yu and Xu consider DPM as a polydisperse aerosol with a specified mass median
1 7 aerodynamic diameter (MMAD) and geometrical standard deviation ag. Each diesel particle is
1 8 represented by a cluster-shaped aggregate within a spherical envelope of diameter de. The
1 9 envelope diameter de is related to the aerodynamic diameter of the particle by the relation
20
7/25/00 A-8 DRAFT—DO NOT CITE OR QUOTE
-------
1 where £ is the bulk density of the particle in g/cm3, £0= 1 g/cm3; is the packing density, which
2 is the ratio of the space actually occupied by primary particles in the envelope to the overall
3 envelope volume; and Cx is the slip factor given by the expression:
4
Cx = 1 + 2-=. [1.257 + 0.4 exp -(——£ )] (A-25)
"x A
5
6 in which A s 8 x 10'6cm3 is the mean free path of ah- molecules at standard conditions. In the
7 diesel particle model of Yu and Xu (1986), £ has a value of 1.5 g/cm3 and a (f) value of 0.3 is
8 chosen based upon the best experimental estimates. As a result, Equation A-24 gives d,/da =
9 1.35. In determining the deposition fraction of DPM, de is used for diffusion and interception
10 according to the particle model.
11
12 A.5.1. Deposition in the Head
13 Particle deposition in the naso- or oropharyngeal region is referred to as head or
14 extrathoracic deposition. The amount of particles that enters the lung depends upon the
15 breathing mode. Normally, more particles are collected via the nasal route than by the oral route
16 because of the nasal hairs and the more complex air passages of the nose. Since the residence
17 time of diesel particles in the head region during inhalation is very small (about 0.1 s for human
18 adults at normal breathing), diffusional deposition is insignificant and the major deposition
19 mechanism is impaction. The following empirical formulas derived by Yu et al. (1981) for
20 human adults are adopted for deposition prediction of DPM:
21 For mouth breathing:
in = °>f°r d* 300° (A-26)
22
0.324 log( 3000 (A-27)
(DF*H, ex = 0, (A-28)
23
uiivj j.or nose
(DF)H in = -0.014 + 0.023 Iog(rffl20, for d2aQ * 337 (A-29)
7/25/00 A-9 DRAFT— DO NOT CITE OR QUOTE
-------
°'399 l°Z(daQ)>f°r dQ > 215 (A-32)
15
16
= -°-959 + °-397 logCrfeX/or 337 (A-30)
= °-003 + °-033 (or d ± 215
3
4 where (DF)H is the deposition efficiency in the head, the subscripts in and ex denote inspiration
5 and expiration, respectively, da is the particle aerodynamic diameter in Jim, and Q is the air
6 flowrate in cm3/sec.
7 Formulas to calculate deposition of diesel particles in the head region of children are
8 derived from those for adults using the theory of similarity, which assumes that the air passage in
9 the head region is geometrically similar for all ages and that the deposition process is
1 0 characterized by the Stokes number of the particle. Thus, the set of empirical equations from
1 1 A-26 through A-32 are transformed into the following form:
For mouth breathing:
,in = Q>f°r daQ * 3000 (A-33)
1 3 and for nose breathing:
14
(DF)H in = -1.117 + 0.972 logK + 0.324 logCg), for dQ > 3000 (A-34)
(DF>a, ex = 0. (A-35)
(DF)H in = - 0.014 + 0.690 log K + 0.023 log(4*0, for d*Q z 337 (A-36)
17
18
(DF)H in = -0.959 + 1.191 log K + 0.397 log 337 (A-37)
19
7/25/00 A- 1 0 DRAFT— DO NOT CITE OR QUOTE
-------
(DF)H>er = 0.003 + 0.099 log K + 0.033 Iog(rf0, for 0 *215 (A-38)
(DF)H, ex = °-851 + 1-197 log K + 0.399 Iog(215 (A-39)
1
2 where K is the ratio of the linear dimension of the air passages in the head region of adults to that
3 of children, which is assumed to be the same as the ratio of adult/child trachea! diameters.
4 For rats, the following empirical equations are used for deposition prediction of DPM in
5 the nose:
(DF)H, in = H, ex = 0-046 + 0.009 log(Q), for d & 13.33 (A-40)
6
7 A.5.2. Deposition in the Tracheobronchial and Alveolar Regions
8 The deposition model adopted for DPM is the one previously developed for
9 monodisperse (Yu, 1978) and polydisperse spherical aerosols (Diu and Yu, 1983). In the model,
(DF)H, in = (DF>H, e* = -O-522 + °-514 l°g(«*«20, for dlQ > 13.33 (A-41)
10
1 1 the branching airways are viewed as a chamber model shaped like a trumpet (Figure A-2). The
1 2 cross-sectional area of the chamber varies with airway depth, x, measured from the beginning of
1 3 the trachea. At the last portion of the trumpet, additional cross-sectional area is present to
1 4 account for the alveolar volume per unit length of the airways. Inhaled diesel particles that
1 5 escape capture in the head during inspiration will enter the trachea and subsequently the
1 6 bronchial airways (compartment T) and alveolar spaces (compartment A).
1 7 Assuming that the airways expand and contract uniformly during breathing, the equation
1 8 for the conservation of particles takes the form:
19
+ A2) + Q = - QCT] (A-42)
dx dx '
20
21 where c is the mean particle concentration at a given x and time t; A, and A2 are. respectively.
22 the summed cross-sectional area (or volume per unit length) of the airways and alveoli at rest: rj
23 is the particle uptake efficiency per unit length of the airway; P is an expansion factor, given by:
7/25/00 A- i 1 DRAFT— DO NOT CITE OR QUOTE
-------
P = 1 + T7 (A-43)
1
2 and Q is the air flow rate, varying with x and t according to the relation
Q _
r*
3
4 where Q0 is the air flow rate at x = 0. In Equations A-43 and A-44, Vt is the volume of new air in
5 the lungs and Vx and Vp are, respectively, the accumulated airway volume from x = 0 to x, and
6 total airway volume at rest.
7
8
louu airway volume ai rest.
Equation A-42 is solved using the method of characteristics with appropriate initial and
boundary conditions. The amount of particles deposited between location x, and x2 from time t,
9 to t2 can then be found from the expression
10
'f?
DF = I I Qcc\dxdt (A-45)
- J \QW
11
1 2 For diesel particles, T| is the sum of those due to the individual deposition mechanisms
1 3 described above, i.e.,
1 4 where T),, T|s, T)P, and T)D are, respectively, the deposition efficiencies per unit length of the
(A-46)
1 5 airway due to impaction, sedimentation, interception, and diffusion. On the basis of the particle
1 6 model described above, the expressions for T|,, T)s, T)P, and T|D are obtained in the following form:
17
I/ = -j—(M)V- (A-47)
18
2 M_ r* r?^ -mi, _ €2/3 + sin-i 6i/3j (A_4g)
71 LJ
7/25/00 A-12 DRAFT—DO NOT CITE OR QUOTE
-------
(A-49)
= -[l-0.819exp(-14.63A) -
i/
0.0976 exp(-89.22A) -
0.0325 exp(-228A) - 0.0509
(A-50)
- 0.444A1/2)
(A-51)
2
3
4
5
6
7
8
Q
10
11
12
13
14
15
16
17
for Reynolds numbers of the flow smaller than 2000, and for Reynolds numbers greater than or
equal to 2000, where ST=c?au/(lSjilR) is the particle Stokes number, 6 = L/(8R), € =
3jJusL/(32uR), P= d/R, and A = DL/(4R2u). In the above definitions u is the air velocity in the
airway; |l is the air viscosity; L and R are, respectively, the length and radius of the airway; us =
C^t/flSp.) is the particle settling velocity; and D = CJcT(3 nf^dj is the diffusion coefficient
with k denoting the Boltzmann constant and T the absolute temperature. In the deposition
model, it is also assumed that T)i and T|P = 0 for expiration, while T|D and T|s have the same
expressions for both inspiration and expiration.
During the pause, only diffusion and sedimentation are present. The combined deposition
efficiency in the airway, E, is equal to:
£=!-(!- Es) (1 -
(A-52)
where ETJ and Es are. respectively, the deposition efficiencies due to the individual mechanisms
of diffusion and sedimentation over the pause period. The expression for ED and Es are given by
-«*-«,-
/ = 1
tf
(A-53)
' J
7/25/00
A-13 DRAFT—DO NOT CITE OR QUOTE
-------
where TD = DT/R2 in which T is the pause time and a,, 1,
7 where Ts = UsT/2R.
8 The values of (DF)T and (DF)A over a breathing cycle are calculated by superimposing
9 DF for inspiration, deposition efficiency E during pause, and DF for expiration in the
tracheobronchial airways and alveolar space. It is assumed that the breathing cycle consists of a
constant flow inspiration, a pause, and a constant flow expiration, each with a respective duration
12 fraction of 0.435, 0.05, and 0.5l5 of a breathing period.
13
14 A.5.3. Lung Models
15 Lung architecture affects particle deposition in several ways: the linear dimension of the
16 airway is related to the distance the particle travels before it contacts the airway surface; the air
17 flow velocity by which the particles are transported is determined by the cross-section of the
18 airway for a given volumetric flowrate; and flow characteristics in the airways are influenced by
19 the airway diameter and branching patterns. Thus, theoretical prediction of particle deposition
20 depends, to a large extent, on the lung model chosen.
21
22 A.5.3.1. Lung Model for Rats
23 Morphometric data on the lung airways of rats were reported by Schum and Yeh (1979).
24 Table A-l shows the lung model data for Long Evans rats with a total lung capacity of
25 13.784 cm3. Application of this model to Fischer rats is accomplished by assuming that the rat
26 has the same lung structure regardless of its strain and that the total lung capacity is proportional
to the body weight. In addition, it is also assumed that the lung volume at rest is about 40% of
7/25/00 A-14 DRAFT—DO NOT CITE OR QUOTE
-------
1 the total lung capacity and that any linear dimension of the lung is proportional to the cubic root
2 of the lung volume.
3
4 A.5.3.2. Lung Model for Human Adults
5 The lung model of mature human adults used in the deposition calculation of DPM is the
6 symmetric lung model developed by Weibel (1963). In Weibel's model, the airways are assumed
7 to be a dichotomous branching system with 24 generations. Beginning with the 1 8th generation,
8 increasing numbers of alveoli are present on the wall of the airways, and the last three
9 generations are completely aleveolated. Thus, the alveolar region in this model consists of all the
1 0 airways in the last seven generations. Table A-2 presents the morphometric data of the airways
1 1 of Weibel's model adjusted to a total lung volume of 3000 cm3.
12
13 A.5.3.3. Lung Model for Children
1 4 The lung model for children in the diesel study was developed by Yu and Xu (1987) on
1.5 the basis of available morphometric measurements. The model assumes a lung structure with
1 6 dichotomous branching of airways, and it matches Weibel's model for a subject when evaluated
17 at the age of 25 years, the age at which the lung is considered to be mature. The number and size
18 of airways as functions of age t (years) are determined by the following equations.
19
20 A.5.3.3.1. Number of airways and alveoli The number of airways N;(t) at generation i for age t
21 is given by
Nfft = 2', for 0 * / *20 (A-57)
22
23
24
25
(A-58)
= 221,
= Nr(f) -221, for 221 < Ntf z 2* (A-59)
= 0,
7/25/00 A-15 DRAFT—DO NOT CITE OR QUOTE
-------
= 221,
21
N22(f) = 2^, for Nr(f) > 221 + 2» (A-60)
23(
#23(0 = W) -221 ~
1 where Nr(t) is the total number of airways in the last three airway generations. The empirical
2 equation for Nr which best fits the available data is
3
4 Thus, Nr(t) increases from approximately 1.5 million at birth to 15 million at 8 years of age and
- {2.036 x 107(1-0.926*-°150, t ± 8
N'(t) \ 1.468 x io7, t > 8 (A
5 remains nearly constant thereafter. Equations A-58 to A-60 also imply that in the last three
6 generations, the airways in the subsequent generation begin to appear only when those in the
7 preceding generation have completed development.
8 The number of alveoli as a function of age can be represented by the following equation
9 according to the observed data:
10
NA(f) = 2.985 x 108(1 -0.919e-°45f) (A-62)
11
12 The number of alveoli distributed in the unciliated airways at the airway generation level
13 is determined by assuming that alveolization of airways takes place sequentially in a proximal
14 direction. For each generation, alveolization is considered to be complete when the number of
15 alveoli in that generation reaches the number determined by Weibel's model.
16
17 A.5.3.3.2. Airway size. Four sets of data are used to determine airway size during postnatal
18 growth: (a) total lung volume as a function of age; (b) airway size as given by Weibel's model;
19 (c) the growth pattern of the bronchial airways; and (d) variation in alveolar size with age. From
20 these data, it is found that the lung volume, LV(t) at age t, normalized to Weibel's model at 4800
21 cm3 for an adult (25 years old), follows the equation
22
LV(f) = 0.959 x 10S(1 - 0.998e-°-0020 (cm3). (A-63)
23
24 The growth patterns of the bronchial airways are determined by the following equations
7/25/00 A-16 DRAFT—DO NOT CITE OR QUOTE
-------
Off) -
(A-64)
2
3
4
(A-65)
where Dj(t) and L;(t) are, respectively, the airway diameter and length at generation i and age t,
Diw and Liw the corresponding values for Weibel's model, CC; and P; are coefficients given by
a, = 3.26 x l(T2exp[-U83 (z + 1)05]
(A-66)
6
7
8
9
10
P,. = 1.05 x l(T6 exp [10.1] (z + 1)'02]
and H(t) is the body height, which varies with age t in the form
H(f) = 1.82
- 0.725e-°140 (cm).
For the growth patterns of the airways in the alveolar region, it is assumed that
D. L.
D.
iw aw
= At), for 17 z i z 23
(A-67)
(A-68)
(A-69)
11
12
13
14
15
16
where Da is the diameter of an alveolus at age t, Daw = 0.0288 cm is the alveolar diameter for
adults in accordance with Weibel's model, and f(t) is a function determined from
16
(LV(i) - ^ TD< «
i = O4
23
(E ^
i = 17'
/A '7A\
-
7/25/00
A-17 DRAFT—DO NOT CITE OR QUOTE
-------
13
14
1 A.6. TRANSPORT RATES
The values of transport rates Aj^ for rats have been derived from the experimental data of
3 clearance for diesel soot (Chan et al., 1981; Strom et al., 1987, 1988) and for the particle-
4 associated organics (Sun et al., 1984; Bond et al., 1986; Yu et al., 1991). These values are used
5 in the present model of lung burden calculation and are listed below:
G = 1.73 (i = 1,2,3) (A-71)
*2 = ^B = 0.00018 (A-72)
= 0.0129 (A-73)
8
1 0) _ ,(3) _ JO) _ , (3) _ 1 « /
~ ~
X. = 0.693 (i = 1,2,3) (A-75)
10
A^ = 0.00068 [1 - exp( - 0.046m j62)] (A-76)
11
*2 = \*$ (i = 23) (A-77)
12
A.^ = 0.012 exp(-0.11wj76) +
(A-78)
0.00068 exp( - 0.046m J62) (i = 1,2,3)
(A-79)
0.012 exp(-0.11/HJ76) + 0.00086
7/25/00 A- 1 8 DRAFT— DO NOT CITE OR QUOTE
-------
f ^AT + *-AB = °-012 exp(-0.11mj76) +•
(A-80)
0.00068 exp(-0.046wa162) + 0.0161
1
A? = ^AL + XAT + ^AB = °-012 exp(-0.11ro]76) +
(A-81)
0.00068 exp(-0.046wj62) + 15.7
2
3 where /?$. is the unit of day'1, and mA = m (A} is the particle burden (in mg) in the alveolar
4 compartment.
5 Experimental data on the deposition and clearance of DPM in humans are not available.
6 To estimate the lung burden of DPM for human exposure, it is necessary to extrapolate the
7 transport rates /?$• from rats to humans. For organics, it is assumed that the transport rates are the
8 same for rats and humans. This assumption is based upon the observation of Schanker et al.
9 (1986) that the lung clearance of inhaled lipophilic compounds appears to depend only on their
10 lipid/water partition coefficients and is independent of species. In contrast, the transport rates of
11 diesel soot in humans should be different from those of rats, since the alveolar clearance rate, AA,
12 of insoluble particles at low lung burdens for human adults is approximately seven times that of
13 rats (Bailey et al., 1982).
14 No data are available on the change of the alveolar clearance rate of insoluble particles in
15 humans due to excessive lung burdens. It is seen from Equation A-79 that A ^for rats can be
16 written in the form
17
A.^ = a exp(-6/n/) + d (A-82)
18
19 where a, b, c, and d are constants. The right-hand side of Equation A-82 consists of two terms,
20 representing, respectively, macrophage-mediated mechanical clearance and clearance by
21 dissolution. The first term depends upon the lung burden, whereas the second term does not.
22 To extrapolate this relationship to humans, we assume that the dissolution clearance term is
23 independent of species and that the mechanical clearance term for humans varies in the same
24 proportion as in rats under the same unit surface particulate dose. This assumption results in the
25 following expression for/? ^in humans
-------
d (A-83)
1
2 where P is a constant derived from the human/rat ratio of the alveolar clearance rate at low lung
3 burdens and S is the ratio of the pulmonary surface area between humans and rats. Equation
4 A-83 implies that rats and humans have equivalent amounts of biological response in the lung to
5 the same specific surface dose of inhaled DPM.
6 From the data of Bailey et al. ( 1 982), a value of A (J} = 0.00 1 69 day1 is obtained for
7 humans at low lung burdens leading to P = 14.4. A value for S of 148 is reported from the data
8 of the anatomical lung model of Schum and Yeh (1979) for rats and Weibel's model for human
9 adults. For humans less than 25 years old, the model assumes the same value for P, but S is
1 0 computed from the data of the lung model for young humans (Yu and Xu 1 987). The value of S
1 1 for different ages is shown in Table A-3.
1 2 The equations for other transport rates that have a lung-burden-dependent component are
1 3 extrapolated from rats to humans in a similar manner. The following lists the values of 2 $
1 4 (in day"1) for humans used in the present model calculation:
15
A^ = 1-73 (i = 1,2,3) (A-84)
16
*-m = *8 = *2 = *2 = °-00018 (A-85)
17
18
*§ = AS = A2 = J.2 = 12.55 (A-87)
19
*.% = 0.693 (i = 1,2,3) (A-88)
20
A.^ = 0.00068 {1 - 0.0694 exp[-0.046(m75)162]} (A-89)
7/25/00 A-20 DRAFT—DO NOT CITE OR QUOTE
-------
(i = 2, 3) (A-90)
4
A.Jj?r = 0.0694 {0.012 exp[-0.11(/n/S)176] +
(A-91)
0.00068 exp[-0.046(/«y5)1-76]} (i = 1, 2, 3)
. .
"•A ~ "-AL A£
(A-92)
0.0694 (0.012 expf-O.HOn^S)1-76]} + 0.00086
3
4 A.7. RESULTS
5 A. 7.1. Simulation of Rat Experiments
6 To test the accuracy of the model, simulation results are obtained on the retention of
7 DPM in the rat lung and compared with the data of lung burden and lymph node burden obtained
8 by Strom etal. (1988). A particle size of 0.19 [imMMAD and a standard geometric deviation,
9 (Jg, of 2.3 (as used in Strom's experiment) are used in the calculation.
1 0 The respiratory parameters for rats are based on their weight and calculated using the
1 1 following correlations of minute volume, respiratory frequency, and growth curve data.
-12
1 3 Minute volume = 0.9W (cm3/min) (A-95)
14
15 Respiratory frequency = 475 W03(l/min) (A-96)
16
1 7 where W is the body weight (in grams) as determined from the equation
18
IS W = 5+537T/(100+T), for T^56 days (A-97)
^ = ^ + ^ ^ ^ = (A-93)
0.0694(0.012 exp[-0.11(»2y,4)176] +
0.00068 exp[-0.046(m/S)176]} + 0.016 (A-94)
20 in which T is the age of the rat measured in days.
7/25/00 A-2 1 DRAFT— DO NOT CITE OR QUOTE
-------
Equation A-95 was obtained from the data of Mauderly (1986) for rats ranging in age
from 3 mo to 2 years old; Equation A-96 was obtained from the data of Strom et al. (1988); and
3 Equation A-97 was determined from the best fit of the experimental deposition data. Figures A-3
4 and A-4 show the calculated lung burden of diesel soot (m ^+ m $ and lymph node burden,
5 respectively, for the experiment by Strom et al. (1988) using animals exposed to DPM at
6 6 mg/m3 for 1, 3, 6, and 12 weeks; exposure in all cases was 7 days/week and 20 h daily.
7 The solid lines represent the calculated accumulation of particles during the continuous exposure
8 phase and the dashed lines indicate calculated post-exposure retention. The agreement between
9 the calculated and the experimental data for both lung and lymph node burdens during and after
10 the exposure periods was very good.
11 Comparison of the model calculation and the retention data of particle-associated BaP in
12 rats obtained by Sun et al. (1984) is shown hi Figure A-5. The calculated retention is shown by
13 the solid line. The experiment of Sun et al. consisted of a 30-min exposure to diesel particles
14 coated with [3H\ benzo[a]pyrene (fH] - BaP) at a concentration of 4 to 6 |ig/m3 of air and
15 followed by a post-exposure period of over 25 days. The fast and slow phase of (fH] - BaP)
16 clearance half-times were found to be 0.03 day and 18 days, respectively. These correspond to
17 AA20 = 0.0385 day1 and A % = 23.1 day1 in our model, where A$>0 is the value of A & at mA ^ 0.
Figure A-5 shows that the calculated retention is in excellent agreement with the experimental
data obtained by Sun et al. (1984).
20
21 A.7.2. Predicted Burdens in Humans
22 Selected results of lung burden predictions in humans are shown in Figures A-6 to A-9.
23 The particle conditions used in the calculation are 0.2 jlm MMAD with Og = 2.3, and the mass
24 fractions of the rapidly and slowly cleared organics are each 10% (f, = f2 = 0.1). Figures A-6
25 and A-7 show, respectively, the lung burdens per unit concentration of diesel soot and the
26 associated organics in human adults for different exposure patterns at two soot concentrations,
27 0.1 and 1 mg/m3. The exposure patterns used in the calculation are (a) 24 h/day and 7 days week;
28 (b) 12 h/day and 7 days/week; and (c) 8 h/day and 5 days/week, simulating environmental and
29 occupational exposure conditions. The results show that the lung burdens of both diesel soot and
30 the associated organics reached a steady-state value during exposure. Because of differences in
31 the amount of particle intake, the steady-state lung burdens per unit concentration were highest
32 for exposure pattern (a) and lowest for exposure pattern (b). Also, increasing soot concentration
33 from 0.1 to 1 mg/m3 increased the lung burden per unit concentration. However, the increase
34 was not noticeable for exposure pattern (c). The dependence of lung burden on the soot
t concentration is caused by the reduction of the alveolar clearance rate at high lung burdens
discussed above.
7/25/00 A-22 DRAFT—DO NOT CITE OR QUOTE
-------
1 Figures A-8 and A-9 show the effect of age on lung burden, where the lung burdens per
2 unit concentration per unit weight are plotted versus age. The data of lung weight at different
3 ages are those reported by Snyder (1975). The exposure pattern used in the calculation is
4 24 h/day and 7 days/week for a period of 1 year at the two soot concentrations, 0. 1 and 1 mg/m3.
5 The results show that, on a unit lung weight basis, the lung burdens of both soot and organics are
6 functions of age, and the maximum lung burdens occur at approximately 5 years of age. Again,
7 for any given age, the lung burden per unit concentration is slightly higher at 1 mg/m3 than at
8 0.1 mg/m3.
9
1 0 A.8. PARAMETRIC STUDY OF THE MODEL
1 1 The deposition and clearance model of DPM in humans, presented above, consists of a
1 2 large number of parameters that characterize the size and composition of diesel particles, the
1 3 structure and dimension of the respiratory tract, the ventilation conditions of the subject, and the
1 4 . clearance half-times of the diesel soot and the particle-associated organics. Any single or
1 5 combined changes of these parameters from their normal values in the model would result in a
1 6 change in the predicted lung burden. A parametric study has been conducted to investigate the
1 7 effects of each individual parameter on calculated lung burden in human adults. The exposure
1 8 pattern chosen for this study is 24 h/day and 7 days/week for a period of 10 years at a constant
1 9 soot concentration of 0.1 mg/m3. The following presents two important results from the
2O parametric study.
21
22 A. 8.1. Effect of Ventilation Conditions
23 The changes in lung burden due to variations in tidal volume and respiratory frequency
24 are depicted in Figures A- 10 and A-l 1. Increasing any one of these ventilation parameters
25 increased the lung burden, but the increase was much smaller with respect to respiratory
26 frequency than to tidal volume. This small increase in lung burden was a result of the decrease in
27 deposition efficiency as respiratory frequency increased, despite a higher total amount of DPM
23 inhaled. The mode of breathing has only a minor effect on lung burden because switching
29 from nose breathin0 dees not produce any appreciable cnan(re in the EinDunt of particle intake
30 into the lung (Yu and Xu, 1987). All lung burden results presented in this report are for nose
3 1 . breathing.
32
33 A.8.2. Effect of Transport Rates
•-)/> -r ------- 1 — j. — i ----- _ ^u-.: ----- .cc.,-*. — .it, --- j. — ^: — _.CT~vr>» * ;- -"-i-,- ' ----- -c* —
OT 1 1 cuispvji i laici navt ail uuvivmo ciicx-i uu illc iCLt>iii.iim ui u\. ivi In ulc ituig, aHci
T-« ------- ____ ___ ____ :_i_. -------- I — ;.ti- .ii. _ i ___ .. ___ _i -------- -.r j: ___ i ___ ^ ___ i ^i. _
ucuauac we cue inaiiu^ vunv&iiicu wiui uic luiig-ttim wcai aiii/c 01 uitoci auui aiiu me
7/25/00 A-23 DRAFT— DO NOT CITE OR QUOTE
-------
associated organics, only the effects of two transport rates, A ^and X%, are studied.
Experimental data of A % from various diesel studies in rats have shown that A % can vary by a
3 factor of two or higher. We use a multiple of 0.5 to 2 for the uncertainty in A ^ and A % to
4 examine the effect on lung burden. Figures A-12 and A-13 show respectively, the lung burden
5 results for diesel soot and the associated organics versus the multiples of/I ^ and A, % used in the
6 calculation. As expected, increasing the multiple of /£ % reduced the lung burden of diesel soot
7 with practically no change in the organics burden (Figure A-12), while just the opposite occurred
8 when the multiple of A % was increased (Figure A-13).
9
10 A.9. OPERATIONAL DERIVATION OF HUMAN EQUIVALENT
11 CONCENTRATIONS (HECs)
12 The model of Yu et al. (1991) is ordered into two parts; one part parameterized on the
13 physiology and anatomy of a 300 g rat and the other part parameterized on the physiology and
14 anatomy of a 25 year old human male. The sequence of steps taken to calculate the human
15 equivalent continuous concentrations (the HECs), outlined in Table A-4, were as follows:
16
17 • The exposure scenario of the rats was entered into the rat portion of the model and the
model ran to obtain the output of lung burden in mg DPM/ rat lung at the time of the
sacrifice of the rats.
20 • The output of mg DPM/ rat lung was normalized to mg DPM/ cm2 of rat lung tissue
21 based on a total pulmonary surface area of 4090 cm2.
22 • The normalized rat lung burdens were used to calculate the corresponding lung
23 burden based on the pulmonary surface area of 627,000 cm2. This operation yielded
24 mg DPM / lung of a 25 year old human male.
25 • Various air concentrations were run in an iterative fashion with the human portion of
26 the model under a continuous exposure scenario of 24 hrs/day, 7d/wk for 70 years
27 with ventilatory parameters set at 0.926 L for tidal volume and 15 breaths per minute
28 as the respiratory frequency to yield a total daily pulmonary volume of 20m3. This
29 was continued until the output (mg DPM/lung) was matched to the mg DPM /human
30 lung obtained from the normalized rat lung burden; the concentration from the model
31 that matched this lung burden was termed the human equivalent continuous
32 concentration, the HEC. The human modeling runs did not consider the preadult
33 status of airway and alveoli number discussed above but rather were ran for 1 to
34 70 years with adult (25 years of age) parameters mentioned above.
7/25/00 A-24 DRAFT—DO NOT CITE OR QUOTE
-------
1 These HEC values address kinetic issues of DPM deposition and retention in the lung by
2 humans. As noted above, these values do not reflect the kinetic variability that may exist in the
3 human population exposed to DPM which includes men and women, young and old. However,
4 the limited parametric analysis of the model clearly shows variability of those parameters most
5 determinative in humans (e.g., tidal volume, respiration rate, and rates of clearance of particles
6 from the airways) were mirrored in the corresponding output of the model (lung burden of DPM).
7 One interpretation of this parallel in parameter-output is that the variability in the physiological
8 characteristics of humans reflects the variability in the model such that, for example, a small tidal
9 volume would be reflected with a decreased lung burden of DPM. Variability among humans of
10 these key parameters such as tidal volume do vary but within an order of magnitude. This would
11 mean that the DPM dose received by different individuals in the population from the same
12 concentration would indeed vary within the extremes of these determinative parameters.
7/25/00 A-25 DRAFT—DO NOT CITE OR QUOTE
-------
Table A-l. Lung model for rats at total lung capacity
Generation
number
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16"
17
18
19
21
22
25
24
Number of
airways
1
2
3
5
8
14
23
38
65
109
184
309
521
877.
1,477
2,487
4,974
9,948
19,896
39,792
79,584
318,336
636,672
Length (cm)
2.680
0.715
0.400
0.176
0.208
0.117
0.114
0.130
0.099
0.091
0.096
0.073
0.075
0.060
0.055
0.035
0.029
0.025
0.022
0.020
0.019
0.017
0.017
Diameter (cm)
0.340
0.290
0.263
0.203
0.163
0.134
0.123
0.112
0.095
0.087
0.078
0.070
0.058
0.049
0.036
0.020
0.017
0.016
0.015
0.014
0.014
0.014
0.014
Accumulative
volume" (cm)
0.243
0.338
0.403
0.431
0.466
0.486
0.520
0.569
0.615
0.674
0.758
0.845
0.948
1.047
1.414
1.185
1.254
1.375
1.595
2.003
2.607
7.554
13.784
"Including the attached alveoli volume (number of alveoli = 3 * 107, alveolar diameter = 0.0086 cm).
bTerminal bronchioles.
7/25/00 A-26 DRAFT—DO NOT CITE OR QUOTE
-------
Table A-2. Lung model by Weibel (1963) adjusted to 3000 cm3 lung volume
Generation
number
0
2
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16"
17
18
19
20
21
22
23
Number of
airways
1
2
4
8
16
32
64
128
256
512
1,024
2,048
4,096
8,192
16,384
32,768
65,536
131,072
262,144
524,283
1,048,579
2,097,152
4,194,304
8,388,608
Length (cm)
10.260
4.070
1.624
0.650
1.086
0.915
0.769
0.650
0.547
0.462
0.393
0.333
0.282
0.231
0.197
0.171
0.141
0.121
0.100
0.085
0.071
0.060
0.050
0.043
Diameter (cm)
1.539
1.043
0.710
0.479
0.385
0.299
0.239
0.197
0.159
0.132
0.111
0.093
0.081
0.070
0.063
0.056
0.051
0.046
0.043
0.040
0.038
0.037
0.035
0.035
Accumulative
volume* (cm)
19.06
25.63
28.63
29.50
31.69
33.75
35.94
38.38
41.13
44.38
48.25
53.00
59.13
66.25
77.13
90.69
109.25
139.31
190.60
288.16
512.94
925.04
i, 694. 16
3,000.00
7/25/00 A-27 DRAFT—DO NOT CITE OR QUOTE
-------
Table A-3. Ratio of pulmonary surface areas between humansand rats as a function
of human age
Age (year)
0
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
27
28
19
20
21
22
23
24
25
Surface area
4.99
17.3
27.6
36.7
44.7
51.9
58.5
64.6
70.4
76.0
81.4
86.6
91.6
96.4
101
106
110
115
119
123
128
132
136
140
144
148
7/25/00
A-28
DRAFT—DO NOT CITE OR QUOTE
-------
-J
O
0
1
K>
p
Tl
1
0
O
H
O
HH
m
0
o
0
3
Table A- = (months of exposure) x '>..)3.
-------
B
HB
LB
Figure A-l. Compartmental model of DPM retention.
7/25/00
A-30 DRAFT—DO NOT CITE OR QUOTE
-------
A2(x)
Trachea
Summed Alveolar Cross Sectional Area
Airway Length x
Summed Airway
Cross Sectional Area A,(x)
Figure A-2. Trumpet model of lung airways.
/ / L-Jl \J\S
A-31 DRAFT—DO NOT CITE OR QUOTE
-------
25
20
en
515
CD
0>
10
0
12 wk
--a H
•O 1-.-Q I
6wk
3wk
^ ^ ^ <
_1_wk
in
0
13
26 39
Time, week
52
65
Figure A-3. The experimental and predicted lung burdens of rats to DPM at a solid and
dashed concentration of 0.6 mg/m3 for different exposure spans. Lines are,
respectively, the predicted burdens during exposure and post-exposure.
Particle characteristics and exposure pattern are explained in the text. The
symbols represent the experimental data from Strom et al. (1988).
7/25/00
A-32
DRAFT—DO NOT CITE OR QUOTE
-------
6wk
-o-
>-o- .-err" . o
1wk
26 39
Time, week
52
65
Figure A-4. Experimental and predicted lymph node burdens of rats exposed to CEPs at a
concentration of 6.0 mg/m3 for different exposure spans. The solid and
dashed lines are, respectively, the predicted burdens during exposure and
post-exposure. Particle characteristics and exposure pattern are explained in
the text. The symbols represent the experimental data from Strom et al.
(1988).
7/25/00
A-33
1—LHJ N(J I Ul i Jb OK
-------
0.8
10
15
Time, day
20
25
30
Figure A-5. Comparison between the calculated lung retention (solid line) and the
experimental data obtained by Sun et al. (1984) for the particle-associated
BaP in rats.
7/25/00
A-34 DRAFT—DO NOT CITE OR QUOTE
-------
700
600
en
"3)
J§. 500
.1 400
§
O
300
o>
1 200
CO
o>
c
_l
100
1 mg/ m
4 6
Time, year
10
Figure A-6. Calculated lung burdens of diesel soot per unit exposure concentration in
human adults exposed continuously to DPM at two different concentrations of
0.1 and 1.0 mg/m3. Exposure patterns are (a) 24 h/day and 7 days/week,
(b) 12 h/day and 7 days/week, and (c) 8 h/day and 5 days/week.
7/25/00
DRAFT—DO NOT CITE OR QUOTE
-------
^ 5
n
2
c 3
03
O
O
O
ffl
O)
3
6
10
Time, year
Figure A-7. Calculated lung burdens of the particle-associated organics per unit exposure
concentration in human adults exposed continuously to DPM at two different
concentrations of 0.1 and 1.0 mg/m3. Exposure patterns are (a) 24 h/day and
7 days/week, (b) 12 h/day and 7 days/week, and (c) 8 h/day and 5 days/week.
7/25/00
A-36
DRAFT—DO NOT CITE OR QUOTE
-------
0.4
O)
E
0-3
•*=
<§ 0.2
o>
O>
0.1
c
o
m
O)
1 mg/m
10
15
Age, year
20
25
Figure A-8. Calculated lung burdens of diesel soot per gram of lung per unit exposure
concentration in humans of different ages exposed continuously for 1 year to
DPM of two different concentrations of 0.1 and 1.0 mg/m3 for 7 days/week and
24 h daily.
V25/00
A-37
DRAFT—DO NOT CITE OR QUOTE
-------
0.01
e
0>
1 mg/m
.008 -
o>
.o
03
~ .006
8
§
O
•s? .004
o>
.002
m
O)
10 15
Age, year
20
25
Figure A-9. Calculated burdens of the particle-associated organics per gram of lung per
unit exposure concentration in humans of different ages exposed continuously
for 1 year to DPM of two different concentrations of 0.1 and 1.0 mg/m3 for
7 days/week and 24 h daily.
7/25/00
A-38
DRAFT—DO NOT CITE OR QUOTE
-------
100
80
60
o>
•*-»"
o
o
40
20
Soot
1.5
o>
if
to
p>
0.5
0.3
0.4 0.5
Tidal Volume, Liter
0.6
0.7
Figure A-10. Calculated lung burdens in human adults versus tidal volume in liters for
exposure to DPM at 0.1 mg/m3 for 10 years at 7 days/week and 24 h daily.
Parameters used in the calculation are: (a) MMAD=0.2 [1m, Og=2.3, /2=0.1,
/3=0.1; (b) respiratory frequency = 14 min"1; and (c) lung volume = 3000 cm3.
7/25/00
A-39
DRAFT—DO NOT CITE OR QUOTE
-------
60
50
40
^" 30
o
to
20
10
Soot
Organics
10
12 14 16
Respiratory Frequency, 1/min.
1.4
1.2
en
0.8 g
'c
CO
0.6 °
0.4
0.2
18
Figure A-ll. Calculated lung burdens in human adults versus respiratory frequency in
bpm for exposure to DPM at 0.1 mg/m3 for 10 years at 7 days/week and 24 h
daily. Parameters used in the calculation are: (a) MMAD=0.2 p.m, Og=2.3,
/2=0.1, /3=0.1; (b) tidal volume = 500 cm3, and (c) lung volume = 3200 cm3.
7/25/00
A-40
DRAFT—DO NOT CITE OR QUOTE
-------
120
100 -
0.6
0.8
1 1.2 1.4
Multiple of X ("
1.6
Figure A-12. Calculated lung burdens in human adults versus multiple of 2 (% for
exposure to DPM at 0.1 mg/m3 for 10 years at 7 days/week and 24 h daily.
Parameters used in the calculation are: (a) MMAD=0.2 |im, Og=2.3, /2=0.1,
/3=0.1; (b) tidal volume = 500 cm3, respiratory frequency = 14 mm'1; and
(c) lung volume = 3200 cm3.
7/25/00
DRAFT—DO NOT CITE OR QU(
-------
60
50
40
I5
B30
o
V)
20
10
Soot
Organics
0.6
0.8
1.2
Multiple of
1.4
1.6
1.8
(2)
1.4
1.2
O)
0.8
0.6
CO
s>
o
0.4
0.2
Figure A-13. Calculated lung burdens in human adults versus multiple of /I (% for
exposure to DPM at 0.1 mg/m3 for 10 years at 7 days/week and 24 h daily.
Parameters used in the calculation are: (a) MMAD=0.2 |lm (7g=2.3, /2=0.1,
/3=0.1; (b) tidal volume = 500 cm3, respiratory frequency = 14 min'1; and (c)
lung volume = 3200 cm3.
7/25/00
A-42
DRAFT—DO NOT CITE OR QUOTE
-------
A.10. REFERENCES
Amann, CA; Siegla, DC. (1982) Diesel particles - what are they and why. Aerosol Sci Technol 1:73-101.
3 Bailey, MR; Fry, FA; James, AC. (1982) The long-term clearance kinetics of insoluble particles from the human
4 lung. Ann Occup Hyg 26:273-289.
5
6 Bond, JA; Sun, JD; Medinsky, MA; et al. (1986) Deposition, metabolism and excretion of l-[MC]nitropyrene and
7 l-[14C]nitropyrene coated on diesel exhaust particles as influenced by exposure concentration. Toxicol Appl
18 Pharmacol 85:102-117.
9
10 Chan, TL; Lee, PS; Hering, WE. (1981) Deposition and clearance of inhaled diesel exhaust particles in the
11 respiratory tract of Fisher rats. J Appl Toxicol 1:77-82.
12
13 Diu, CK; Yu, CP. (1983) Respiratory tract deposition of polydisperse aerosols in humans. Am Ind Hyg Assoc J
14 44:62-65.
15
16 ICRP. (1979) Limits for intakes of radionuclides by workers. Ann ICRP 2. Publication 30, part 1.
17
18 Mauderly, JL. 1986. Respiration of F344 rats in nose-only inhalation exposure tubes. J Appl Toxicol 6:25-30.
19
20 Schanker, LS; Mitchell, EW; Brown, RA. (1986) Species comparison of drug absorption from the lung after aerosol
21 inhalation or intratracheal injection. Drug Metab Dispos 14(l):79-88.
22
23 Scheutzle, D. (1983) Sampling of vehicle emissions for chemical analysis and biological testing. Environ Health
24 Perspect 47:65-80.
25
»Schum, M; Yeh, HC. (1979) Theoretical evaluation of aerosol deposition in anatomical models of mammalian lung
airways. Bull Math Biol 42:1-15.
29 Snyder, WS. (1975) Report of task group on reference man. Oxford, London: Pergamon Press, pp. 151-173.
30
31 Solderholm, SC. (1981) Compartmental analysis of diesel particle kinetics in the respiratory system of exposed
32 animals. Oral presentation at EPA Diesel Emissions Symposium, Raleigh, NC, October 5-7. In: Toxicological effects
33 of emissions from diesel engines (Lewtas J, ed.). New York: Elsevier, pp. 143-159.
34
35 Strom, KA; Chan, TL; Johnson, JT. (1987) Pulmonary retention of inhaled submicron particles in rats: diesel exhaust
36 exposures and lung retention model. Research Publication GMR-5718. Warren, MI: General Motors Research
37 Laboratories.
38
39 Strom, KA; Chan, TL; Johnson, JT. (1988) Inhaled particles VI. Dodgson, J; McCallum, RI; Bailey, MR; et al., eds.
40 London: Pergamon Press, pp. 645-658.
41
42 Sun, JD; Woff, RK; Kanapilly, GM; et al. (1984) Lung retention and metabolic fate of inhaled benzo(a)pyrene
43 associated with diesel exhaust particles. Toxicol Appl Pharmacol 73:48-59.
44
45 Weibel, ER. (1963) Morphometry of the human lung. Berlin: Springer-Verlag.
46 .
47 Xu, GB; Yu, CP. (1987) Desposition of diesel exhaust particles in mammalian lungs: a comparison between rodents
48 and man. Aerosol Sci Tech 7:117-123.
49
50 Yu, CP. (1978) Exact analysis of aerosol deposition during steady breathing. Powder Technol 21:55-62.
51
Yu, CP; Diu, CK; Soong, TT. (1981) Statistical analysis of aerosol deposition in nose and mouth. Am Ind Hyg
Assoc J 42:726-733.
7/25/00 A-43 DRAFT—DO NOT CITE OR QUOTE
-------
1 Yu, CP; Xu, GB. (1986) Predictive models for deposition of diesel exhaust participates in human and rat lungs.
2 Aerosol Sci Technol 5:337-347.
3
4 Yu, CP; Xu, GB. (1987) Predicted deposition of diesel particles in young humans. J Aerosol Sci 18:419-429.
5
6 Yu, CP and Yoon, KJ. (1990) Retention modeling of diesel exhaust particles in rats and humans. Res Rep Health Eff
7 Inst 40:1-33.
8
9 Yu, CP, Yoon, KJ, and Chen,YK. (1991) Retention modeling of diesel exhaust particles in rats and humans. J.
10 Aerosol Med. 4(2): 79-115.
7/25/00 A-44 DRAFT—DO NOT CITE OR QUOTE
-------
Appendix B
Benchmark Concentration Analysis of
Diesel Data
7/25/00 B-1 DRAFT—DO NOT CITE OR QUOTE
-------
B-l. INTRODUCTION TO BENCHMARK
"2 The benchmark dose or benchmark concentration approach, hereafter referred to as the
3 BMC approach, is an alternate to the N/LOAEL option for deriving effect levels. The BMC is
4 currently undergoing extensive consideration by the Agency with promulgation of software and
5 guidelines for application of this methodology (U.S. EPA, 2000). The BMC approach involves
6 fitting a dose-response function to dose and effect information from a single study to derive the
7 best fit of those data. This "best fif'is statistically termed the maximum likelihood estimate but
8 is referred to in the benchmark terminology as the BMC curve. The curve defining the
9 corresponding lower 95% confidence limit of this "best fifestimate is termed the BMCL curve.
10 This BMCL curve is used to predict the dose that will result in a level of response that is defined
11 a priori as the benchmark response "x", BMCLX. In the analyses below,, for example, the
12 benchmark response for a 10% increase in incidence1 of chronic inflammation is defined as a
13 BMCLi0; the corresponding 10% increase as determined from the BMC curve would be termed
14 the BMC10. This BMCL10 would be derived by first using the data and the programs to determine
15 the BMC and BMCL curves. The concentration corresponding to a 10% increase in incidence
16 would then be determined directly from the BMCL. The BMCL10 then would be used as the
17 representative value for the effect level or point of departure in the dose-response assessment.
^B The latest version of the Agency Benchmark Dose Software (BMDS Version 1.2; U.S.
19 EPA, 2000) was used to analyze data on chronic inflammation and pulmonary histopathology
. 20 present in the chronic studies that were amenable to benchmark analysis. At this time, the
21 Agency BMDS offers sixteen different models total that are appropriate for the analysis of
22 dichotomous data (gamma, logistic, probit, Weibull, log-logistic, multistage, log-probit,
23 quantal-linear, quantal-quadratic), continuous data (linear, polynomial, power, Hill) and nested
24 developmental toxicology data (NLogistic, NCTR, Rai & Van Ryzin). Results from all models
25 include a reiteration of the model formula and model run options chosen by the user,
26 goodness-of-fit information, a graphical presentation for visual inspection and the concentration
27 estimate for the response at the designated BMCLX, as well as the corresponding BMCX. More
28 details on the modeling results are described and presented in the analysis on dichotomous data
29 following.
30 The U.S. EPA benchmark dose (BMD/C) methods guidance has not been finalized at this
31 time to provide definitive procedures and criteria (U.S. EPA 1995). Therefore, in this document
32 provisional criteria for minimum data to perform a benchmark analysis are designated such that
For increases in incidence "extra risk" is used which is response incidence (inc) normalized to the
background (BG) incidence; response - BG/l-BG.
7/25/00 B-2 DRAFT—DO NOT CITE OR QUOTE
-------
1 (1) complete quantitative information on the response of interest should be available (e.g.,
2 incidence as number affected / total, means with variability) and that (2) at least two exposure
3 levels with responses that differ from those of the controls are provided, and (3) a benchmark
4 response of 10% is employed such that outcomes are BMCL10s. A response of 10% is at or near
5 the limit of sensitivity in most long-term bioassays as determined from both the typical number
6 of animals used in bioassays and a low spontaneous background rate (e.g., 0.1%) for a given
7 effect (Haseman, 1984; Haseman et al., 1989).
8
9 B-2. DIESEL DATA FOR BENCHMARK ANALYSIS
10 Using the criteria set forth in Section B-1 and the information about the critical effects that
11 have been identified (pulmonary inflammation, pulmonary histopathology including indicators of
12 fibrotic changes such as increases in alveolar-capillary wall thickness) the following rat chronic
13 studies identified in Chapter 6 were analyzed for information suitable for BMC analysis:
14 Ishinishi et al. (1986,1988), Mauderly et al. (1987a,b; 1988); Heinrich et al. (1986,1995), and
15 Nikulaetal. (1995).
16 Results from this analysis yielded only a few data sets from a single study, that of Nikula
17 et al. (1995), that could be used for BMC analysis. The basis for not including data from the
18 other studies varied. Information on pulmonary histopathology hi the studies of Ishinishi et al.
19 (1986, 1988), for example, was supplied only in narrative form with no quantitative information
20 given. A similar situation was found for those reports of the ITRI study; Wolff et al. (1987)
21 reports on clearance alterations due to DPM exposure; Henderson et al. (1988) does give
22 information on hydroxyproline but only in graphical form; the 1988 study of Mauderly et al.
23 deals with pulmonary function as a function of DPM lung loading; the 1987a reference of
24 Mauderly et al. discusses tumor prevalence only and the Mauderly 1987b reference reports on
25 diesel exhaust in developing lung to a single exposure concentration of DPM with no dose-
26 response information available. Those reports on the General Motor study contain extensive
27 information relating not to the critical effects, but mostly to precursors of inflammation such as
28 levels of polymorphonuclear neutrophils and lymphocytes in bronchoalveolar lavage from DPM
29 exposed rats (Strom, 1984) and guinea pigs (Barnhart et al., 1981) as well as information on
30 collagen biosynthesis (Misiorowski et al., 1980) all of which is presented in graphical rather than
31 tabular form amenable for benchmark analysis. The information on noncancer histopathology
32 reported by Heinrich et al. (1995) is in text form only and this author's 1986 study deals
33 primarily with clearance and mortality. Nikula et al. (1995). however, do present extensive
34. rmanHtativp rln^e-rftsnnnse information (incidence / dichotomous data) on several measures of
35 the critical effect including chronic inflamation (presence of focal aggregates of neutrophils),
7/25/00 B-3 DRAFT—DO NOT CITE OR QUOTE
-------
focal fibrosis with epithelial hyperplasia (nodular fibrosis rimmed by hyperplasia), and septal
fibrosis (interstitial fibrosis within alveolar septa) although the study had but 2 exposure
3 concentrations both of which are different from the controls, a minimal number on which
4 benchmark analysis should be performed.
5
6 B-3. BENCHMARK ANALYSIS OF DIESEL DATA
7 These data from Nikula et al. (1995) were extracted, HEC concentrations calculated using
8 the model of Yu et al. (1991; Appendix A), and analyzed using all 9 applicable models for
9 dichotomous data. Because the benchmark models were ran with the HEC, general from the
10 model of Yu et al. (1991), the BMCL10s are also HECs. The results and data are presented in
11 Table B-l. Results were evaluated based on the nature of the data set, visual inspection of the
12 graphical output, and on the goodness-of-fit parameters, including p values and the AIC. When p
13 values were generated for model fits, values for p that were less than 0.1 were considered to
14 reflect a minimal fit to the data and were disqualified from further consideration. However, the
15 small set of only 3 data points was often matched by the number of parameters fitted in several of
16 the models such that the outcome of the model exactly fit the data and thus no p value is
17 generated; these model fits are often referred to as being overparameterized, and are indicated as
1^B "NA" in Table B-l. Values for p that were less than 0.1 were considered to reflect a minimal fit
19 to the data. The AIC (Akaike Information Coefficient; Akaike, 1973; Stone, 1998) is a parameter
20 generated for the models in U.S. EPA (2000) that allows for a general comparison among models
21 run on the same data set. The AIC is defined as -2 log L + 2 p where log L is the log likelihood
22 of the fitted model, and p is the number of parameters estimated; smaller values indicate better
23 fits.
24 The overall results of this mathematical analysis is reasonable in a biologically mechanistic
25 sense in that chronic inflammation is more prevalent and apparently occurs at lower
26 concentrations (i.e., has lower BMCLIO values) than does focal fibrosis. The information on
27 septal fibrosis were not interpretable as the data were not amenable (no or zero background and
28 then total incidence) to any meaningful benchmark or other dose-response analysis. The most
29 sensitive endpoint, chronic inflammation, is therefore the most sensitive benchmark
30 concentration followed by focal fibrosis.
31 The choice for the most appropriate BMCL10 from among the various modeled values for
32 chronic inflammation requires analysis of both the statistical and graphical outputs of the data.
33 The shape of the dose-response curve from information given in Chapter 6 (Table 6-2) gives
evidence of considerable "S" character, e.g., several low HECs without any reported effects up to
about 0.2 mg/m3. The shape of the dose-response curves generated by several of the models,
7/25/00 B-4 DRAFT—DO NOT CITE OR QUOTE
-------
1 including gamma-hit, Weibull, multistage, and quantal linear were all a uniformly upward
2 sloping arc from the origin (graphs not shown) with minimal evidence of any "S" character, a
3 shape not concordant with the data array in Table 6.2. Models that did generate curves with "S"
4 character included log-logistic, logistic, probit, quantal-quadratic, and log-probit. Because of
5 their concordance with this independent data array on dose-response, the latter outputs are further
6 analyzes.
7 The results for both chronic inflammation and focal fibrosis for those models with outputs
8 having appreciable "S" character suggest that females may be more sensitive than males for these
9 endpoints as the incidences are higher and the BMCL10 values are generally lower for females
10 than for males. However, the model fits of the BMCL,0s to the chronic inflammation data
11 segregated by sex were generally inadequate as judged from the p values (most being far less
12 than 0.1) or from visual inspection of the fits to the data, several of which (e.g., log-logistic and
13 log-probit) were lacking any appreciable "S" character. However, combining female and male
14 data improved data fitting as judged by the increased p values to where nearly all were >0.1 and
15 to where the visual fits were concordant with the independent information on dose-response.
16 Too, most of the combined BMCL10s were either intermediate between the female and male
17 values or somewhat closer to the female values such that the combined BMCL10 values were not
18 much different from the females BMCL,0s.
19 From among the combined male and female model outputs in Table B-l, the logistic,
20 probit, and quantal quadratic results were all excluded based on the high AIC value relative to the
21 log-logistic and log-probit results. The log-logistic results were excluded based on the shape of
22 the lower portion of the dose-response curve which was upward sloping near the origin (graph
23 not shown) and not as concordant with the independent dose-response information in Table 6-2
24 as was the fit of the log-probit model (Figure B-l). This leaves the fit of the log-probit model as
25 being most reflective of the information hi Table 6-2. The BMCL10 of the log-probit curve at
26 0.37 mg/m3 remains and, by elimination, appears to be the most defensible choice from among
27 the BMCL,0s arrayed in Table B-l. Figure B-l shows the graphical representation of the log-
28 probit model fit to the data and the origin of the BMCL10. This graph also shows the relationship
29 of the BMCL,0 of 0.37 mg/m3 to the variability that exists around the control value and that the
30 value of 0.37 mg/m3 is not far removed from the outer range of this variability. The log-probit
31 BMCL10 for focal fibrosis (combined) of 1.3 mg/m3 noted as being representative of this lesion
32 from the BMC analysis hi Table B-l.
33 Characterization of this benchmark value indicates that it may not be a suitable candidate
34 for use as a point of departure for development of a dose-response assessment such as the RfC.
35 An attribute of the benchmark method is that the response (such as the 10% as used here) is near
36 the range of the actual experimental values, such that extrapolation is not far below the observed
7/25/00 B-5 DRAFT—DO NOT CITE OR QUOTE
-------
experimental range. However, due to the paucity of data points overall and lack of any values
below an HEC of nearly 2 mg/m3 in the Nikula et al. (1995) study, the extrapolation of this BMC
3 to the 10% response level is considerable, the BMLC10 of 0.37 mg/m3 being > 5-fold below the
4 nearest observed value of 1.95 mg/m3. Also, the high experimental exposures used in this study
5 are in the range of those resulting in pulmonary overload conditions hi rats and therefore in the
6 range of the model assumptions of Yu et al. (1991) about this phenomenon in humans for
7 calculation of the HECs (Chapter 3). The BMCL10 of 0.37 mg/m3 is considerably greater than
8 other NOAELs in the DPM data base of 0.144 mg/m3 and 0.128 mg/m3 (Table 6-2 in Chapter 6),
9 possibly indicating that these NOAELs represent actual incidence levels that are considerably
10 less thanlO%; from the same log-probit model the corresponding BMCLOS was 0.21 mg/m3
11 (near the range of these NOAELs) and the corresponding BMCLOI was 0.07 mg/m3 (below the
12 range of these NOAELs). These limitations on this BMCL10 make it a less than optimal
13 candidate for consideration as a point of departure in the development of dose-response
14 assessments.
15
16 B-4. SUMMARY
17 The recently developed EPA Benchmark dose software (U.S. EPA, 2000) and preliminary
4fe guidance was utilized to analyze diesel data by the benchmark approach. Data from only one of
19 the array of principal studies identified elsewhere (Chapter 6) was found to contain data
20 amenable to benchmark analysis. The data from this study, that of Nikula etal. (1995) on
21 pulmonary inflammation and histopathology, was extracted and analyzed as dichotomous data
22 using all available models and designating a 10% response level such that BMCLI0s were
23 calculated; as the models were ran with HECs, the BMCL10s were also HECs.
24 The analysis resulted in an array of BMCL10s from 3 different effects in two sexes (both
25 separate and combined) with 9 different models. These BMCL10s were each considered from a
26 perspective of biological relevance, known dose-response character, and from the individual fit
27 to the data by the models from statistical parameters and visual judgments. The BMCL10 that
28 emerged after the above considerations was 0.37 mg/m3 for the combined male plus female
29 incidence of chronic active pulmonary inflammation. A BMCL10 of 1.3 mg/m3 for pulmonary
30 focal fibrosis was also noted in this analysis. Characterization of these benchmark values
31 indicates that neither may be a suitable candidate for use as a point of departure in development
32 of a dose-response assessment such as the RfC but that they are concordant with other
33 quantitative dose-response aspects of the DPM database.
7/25/00 B-6 DRAFT—DO NOT CITE OR QUOTE
-------
^J
to
o
o
Table B-l. BMC analysis of pathology incidence data in male and female F344 rats from the study of Nikula et al. (1995)
using the different model!! available from U. S. EPA benchmark dose project (U.S. EPA, 2000) for dichotomous data based on
10% uxtra risk (i.e., a 10% increase relative to a total that has been adjusted for background) and no threshold term. The
concentrations used in the analysis are human continuous equivalent concentrations (HECs) obtained from the interspecies
extrapolation model of Yu et al. (1991). The table listings include the BMCL,0 (the benchmark response level of 10%
obtained from the lower 95% limit of the benchmark curve in mg/m3), the BMC10 (the corresponding estimate at 10%
response from the best fit benchmark curve, also in mg/m3), P = goodness-of-fit values. NA indicates a G-O-F value was not
available, usually due to Ifce Sack of degrees of freedom. AIC = Akaike Information Coefficient (see U.S. EPA, 2000 and
below) which may be used for model comparison on the same data set.
30
i
-J
fl
>
Tl
1
"1
tJ
Z
3
3
Tj
rl
^/
~)
^
r)
Effect (Irom Table 5
and 6, p 86, Nikula Inc @
etal., IS 95) a mg/m3
Chronic active 5/177
inflammation >18mos,
grades 1 -3, male +
female combined
Chronic active 1/86
inflammation >18 mos,
grades 1 -3 in nales
Chronic active 4/91
inflammation >18 mos,
grades 1 -3 in females
Focal filirosiswith 0/177
epithelial hyp:rplasia,
grades 1 -4 in nales and
females combined
Focal filirosiswith 0/86
epitheliil hyp:rplasia,
grades 1 -4 in males
Focal filirosiswith 0/91
epilheliil hypsrplasia,
grades 1 -4 in females
Septal fibrosis, • 1/86
>18moi!, grades 1-4 in
males
Septal fibrosis, 2/91
>I8 mo:, grades 1-4 in
females
Inc@
1 .95 mg/m3
HEC
59/162
19/81
40/81
18/162
5/81
13/81
79/81
75/81
Inc@
5.1 mg/m1
HEC
118/174
54/85
64/89
63/174
19/85
44/89
83/85
87/89
BMCLIO
(BMC10)
log-logistic
0.32(0.64)
P=NA
AIC= 483
0.67(1.16)
P=NA
AIC=217
0.18(0.26)
P=NA
AIC=257
1.25(1.8)
P= 1.000
AIC= 345
1.72(2.7)
P=1.00
AIC= 132
0.80(1.4)
P=1.00
A1C= 199
.003(.008)
P= 0.35
AIC=53
0.009 (.05)
P=NA
AIC=87
BMCLIO
(BMC10)
log-probit
0.37(.70)
P=NA
AIC = 483
0.74(1.22)
P = NA
AIC = 217
.016(.30)
P = NA
AIC = 257
1.3(1.8)
P= 1.000
AIC = 345
1.6(2.7)
P= 1.000
AIC =132
0.87(1.47)
P= 1.000
AIC =199
(failed)
(failed)
BMCLIO
(BMC10)
multi-stage
0.43(.49)
P= 0.982
AIC=481
0.56(.95)
undefined
AIC=217
0.33(.40)
P=0.173
AIC= 257
1.21(1.8)
P= 1.000
AIC=345
1.79(2.8)
undefined
AIC= 134
0.77
P= 0.99
AIC= 199
0.07(.08)
P= 0.000
AIC=65
0.08(.10)
P= 0.003
AIC=91
BMCLIO
(BMC10)-
Weibull
0.43(.49)
P= 0.982
AIC=48I
.56(1.04)
P=NA
AIC=216
0.33(.40)
P= 0.173
AIC=257
1.21(1.8)
P= 1.000
AIC=345
1.79(2.8)
P=1.00
AIC= 132
0.77(1.4)
P=1.0
AIC=199
0.07(.08)
P= 0.000
AIC=65
0.08(.10)
P= 0.000
AIC=91
BMCL,,
(BMCIO)-
gamma
0.43(.49)
P=0.98
AIC=480
.56(1.09)
P=NA
AIC=217
0.33(.40)
P=O.I7
AIC= 257
1.21(1.8)
P=1.0
AIC=345
1.79(2.75
P=1.0
AIC= 132
0.71(1.4)
P=1.00
AIC= 199
0.07(.08)
P= 0.000
A1C=65
O.OS(.IO)
P= 0.003
AIC=91
BMCL10
(BMC,,,) -
quanta!
linear
0.43(.49)
P=.982
AIC=48I
0.50(.61)
P=0.15
AIC=216
0.33(.40)
P=0.173
AIC= 257
1.1(1.3)
P= 0.363
AIC= 345
1.7(2.4)
P=0.70
AIC=131
0.71(.88)
P= 0.445
AIC= 198
0.07(.08)
P= 0.000
AIC=65
O.OS(.IO)
P= 0.003
AIC=91
BMCL10
(BMC10)-
probit
1.06(1.19)
P= 0.000
AIC=499
1.31(1.55)
P=0.05
AIC=219
0.83(.96)
P= 0.0001
A1C= 272
2.32(2.61)
P= 0.013
AIC=353
2.98(3.5)
P=0.199
A1C= 134
1.76
P= 0.037
AIC= 205
0.29(.37)
P= 0.000
AIC= 1 14
0.32(.40)
P= 0.000
AIC=131
BMCL10
(BMCJ-
logisHc
1.12(1.26)
P=0.000
AIC= 502
0.67(1.16)
P=NA
AIC=217
0.85(1.0)
P= 0.000
A1C= 273
2.50(2.8)
P= 0.006
AIC=356
3.17(3.69)
P=0.153
AIC= 135
1.89(2.2)
P=0.02
AIC= 207
0.32(.44)
P= 0.000
AIC=86
0.34(.45)
P= 0.000
AIC= 109
BMCL,,
(BMC,,)
quanta!
quadratic
1.34(1.45)
P= 0.000
AIC = 505
1.42(1.57)
P= 0.055
AIC = 218
1.21(1.35)
P= 0.000
AIC = 279
2.14(2.34)
P= 0.091
AIC = 347
2.68(3.1)
P=0.552
AIC=131
1.7(1.9)
P=0.2i
AIC = 200
0.42(0.47)
P= 0.000
AIC = 100
0.46(.51)
P= 0.000
AIC =119
-------
1
0.8-
0.7-
0.6-
0.4
c
•£ 0.3-
O
2
Ll_ 0.2
0.1-
0-
Best Estimate (log-probit fit)
Lower Bound of Best Estimate
(10% extra risk)
BMCL1(U i BMC
10
012345
Concentration
Figure B-l. Benchmark concentration analysis (log-probit) of chronic pulmonary
inflammation in rats exposed to DPM from Nikula et al. (1995). BMCL10,
the lower confidence estimate of the concentration of DPM associated with
a 10% incidence (extra risk); BMC10, the corresponding estimate from the
best (log-probit) fit. (0) data with 95% error bounds.
7/25/00
B-8
DRAFT—DO NOT CITE OR QUOTE
-------
1 B-5. REFERENCES
2
3 Akaike, H. (1973) Information theory and an extension of the maximum likelihood principle. In: Proceedings of the
4 Second International Symposium on Information Theory, B.N. Petrov and F. Csaki, eds. Akademiai Kiado,
5 Budapest, pp. 267-281
6
7 Barnhart, MI; Chen, S-T; Salley, SO; et al. (1981) Ultrastructure and morphometry of the alveolar lung of guinea
8 pigs chronically exposed to diesel engine exhaust: six months' experience. J Appl Toxicol 1:88-103.
9
10 Baseman, JK. (1984) Statistical issues in the design, analysis, analysis and interpretation of animal carcinogenicity
11 studies. Environ Health Persp 58: 385-392.
12
13 Haseman, JK; Huff, JE; Rao, GN and Eustis, SL. (1989) Sources of variability in rodent carcinogenicity studies.
14 Fund Appl Toxicol 12: 793-804.
15
16 Heinrich, U; Muhle, H; Takenaka, S; et al. (1986) Chronic effects on the respiratory tract of hamsters, mice, and rats
17 after long-term inhalation of high concentrations of filtered and unfiltered diesel engine emissions. J Appl Toxicol
18 6:383-395.
19
20 Heinrich, U; Fuhst, R; Rittinghausen, S; et al. (1995) Chronic inhalation exposure of Wistar rats and two strains of
21 mice to diesel engine exhaust, carbon black, and titanium dioxide. Inhal Toxicol 7:553-556.
22
23 Henderson, RF; Pickrell, JA; Jones, RK; et al. (1988) Response of rodents to inhaled diluted diesel exhaust:
24 biochemical and cytological changes in bronchoalveolar lavage fluid and in lung tissue. Fundam Appl Toxicol
25 11:546-567.
26
27 Ishinishi, N; Kuwabara, N; Nagase, S; et al. (1986) Long-term inhalation studies on effects of exhaust from heavy
28 and light duty diesel engines on F344 rats. In: Ishinishi, N; Koizumi, A; McClellan, RO; et al., eds. Carcinogenic and
29 mutagenic effects of diesel engine exhaust: proceedings of the international satellite symposium on toxicological
30 effects of emissions from diesel engines; July; Tsukuba Science City, Japan. (Developments in toxicology and
31 environmental science: v. 13.) Amsterdam: Elsevier Science Publishers BV; pp. 329-348
32
33 Ishinishi, N; Kuwabara, N; Takaki, Y; et al. (1988) Long-term inhalation experiments on diesel exhaust. In: Diesel
34 exhaust and health risks: results of the HERP studies. Tsukuba, Ibaraki, Japan: Japan Automobile Research Institute,
35 Inc., Research Committee for HERP Studies; pp. 11-84.
36
37 Mauderly, JL; Jones, RK; Griffith, WC; et al. (1987a) Diesel exhaust is a pulmonary carcinogen in rats exposed
38 chronically by inhalation. Fundam Appl Toxicol 9:208-221.
39
40 Misiorowski, RL; Strom, KA; Vostal, JJ; et al. (1980) Lung biochemistry of rats chronically exposed to diesel
41 particulates. In: Pepelko, WE; Danner, RM; Clarke, NA, eds. Health effects of diesel engine emissions: proceedings
42 of an international symposium; December 1979. Cincinnati, OH: U.S. Environmental Protection Agency, Health
43 Effects Research Laboratory; pp. 465-480; EPA report no. EPA-600/9-80-057a. Available from: NTIS, Springfield,
A."1 VA; PBS! -! 73809.
45
46 Nikula, KJ; Snipes, MB; Barr, EB; et al. (1995) Comparative pulmonary toxicities and carcinogenicities of
47 chronically inhaled diesel exhaust and carbon black in F344 rats. Fundam Appl Toxicol 25:80-94.
48
49 Stone, M. (1998) Akaike's Criteria. In: Encyclopedia of Biostatistics, Armitage, P. and Colton, T., eds. Wiley,
50 New York.
51
52 Strom, KA. (1984) Response of pulmonary cellular defenses to the inhalation of high concentrations of diesel
SJ
54
7/25/00 B-9 DRAFT—DO NOT CITE OR QUOTE
-------
U.S. Environmental Protection Agency (U.S. EPA). (1995) The use of the benchmark dose approach in health risk
assessment. Washington, DC: Office of Research and Development, Risk Assessment Forum United States
Environmental Protection Agency.
4
5 U.S. EPA. (2000) Benchmark dose software version 1.2. Washington, DC: National Center for Environmental
6 Assessment, United States Environmental Protection Agency. Available: http://www.epa.gov/ncea/bmds.htm
7 [2000, June 7].
8
Q Wolff, RK; Henderson, RF; Snipes, MB; et al. (1987) Alterations in particle accumulation and clearance in lungs of
10 rats chronically exposed to diesel exhaust. Fundam Appl Toxicol 9:154-166.
11
12 Yu, CP; Yoon, K.J.; Chen, Y.K. (1991) Retention modeling of diesel exhaust particles in rats and humans.
13 J. Aerosol Res. 4 (2): 79-115.
7/25/00 B-10 DRAFT—DO NOT CITE OR QUOTE
-------
Appendix C
Key Particulate Matter (PM) Epidemiologic Findings
Related to PM NAAQS Decisions
C.I Overview of Key Findings Supporting
1997 PM NAAQS Decisions
C.2 Prospective Cohort Studies of Long-Term
Ambient PM Exposure Effects
7/25/00 C-l DRAFT—DO NOT CITE OR QUOTE
-------
C.I. Overview of Key Findings Supporting 1997 PM NAAQS Decisions
In promulgating the 1997 PM NAAQS (Federal Register, 1997), EPA relied mainly on
"3 the relative risk (RR) levels for increased risks of mortality or morbidity associated with acute
4 (short-term) and chronic long-term measures of PM exposure reported in U.S. and Canadian PM
5 epidemiology studies, which provide the most directly pertinent quantitative risk estimates as
6 inputs to U.S. PM NAAQS decisions. These included (a) relative risk (RR) estimates for
7 mortality or morbidity associated with 50 ug/m3 increases in 24-h PM10 concentrations (Table C-
8 1) or with variable increases in fine particle indicators, e.g., 25 ug/m3 increment in 24-h PM2 5
9 concentrations (Table C-2); and (b) analogous relative risk estimates for health effects related to
10 specified increments in long-term (e.g., annual mean or median) levels of fine particle indicators
11 (Table C-3). The study results summarized in these tables reproduced from Chapter 13 of the
12 PM CD (U.S. EPA, 1996a)' were found to provide sufficient evidence for concluding that
13 significant associations of increased mortality and morbidity risks were likely attributable to fine
14 particles, as indexed by various fine particle indicators, e.g., PM2 5, sulfates (SO4), etc.; but
15 possible toxic effects of the coarse fraction of PM10 (i.e., PMI0.25) could not be ruled out. Some
16 inhalable coarse fraction particles subsumed under PM10 do reach the lower respiratory tract, and
17 some health effects of concern are suggested by some epidemiology results.
18 Both the PM CD (U.S. EPA, 1996a) and Staff Paper (U.S. EPA, 1996b) noted the very
9 limited extent of available toxicologic findings by which (a) to identify key PM constituents of
20 urban ambient air mixes that may be causally related to mortality/morbidity effects observed in
21 the community epidemiologic studies; or (b) to delineate plausible biological mechanisms by
22 which such effects could be induced at the relatively low ambient PM concentrations evaluated
23 in the epidemiologic studies. As discussed in the PM CD, several types of mechanisms have
24 been shown to underlie toxic effects observed with acute or chronic exposures to various PM
25 species or mixtures (e.g., acute lung inflammation; impaired respiratory function; impaired
26 pulmonary defense mechanisms, etc.), but generally at much higher PM levels than now
27 typically encountered in U.S. ambient air. As also discussed in the 1996 PM CD, several fine
28 particle constituents were hypothesized as being likely important contributors to ambient PM
29 effects, e.g., acid aerosols (indexed by sulfates; H+ ions, etc.); transition metals (e.g., Fe, Mn,
30 etc.); and ultrafine particles. Nevertheless, despite the lack of more definitive characterization of
31 pertinent underlying biological mechanisms, several aspects of the epidemiologic evidence (e.g.,
32 the consistency and coherence of the epidemiologic findings), as discussed in the PM CD,
33 support the conclusion that exposure to ambient PM, acting alone or in combination with other
'Full reference citations for each study identified in Tables C-l, C-2, and C-3 can be obtained in the
bibliographic listing for Chapter 13 in U.S. EPA (1996a).
7/25/00 C-2 DRAFT—DO NOT CITE OR QUOTE
-------
1 air pollutants, is probably a key causal agent contributing to the 'increased mortality and
2 morbidity risks observed in the epidemiology studies. Figure C-l, from the PM Staff Paper
3 (1996b), illustrates the consistency and coherence of the relative risk findings for PM10.
4 Relative risk estimates shown in Table C-2 for mortality and morbidity effects associated
5 with short-term ambient PM exposures provided the key bases for derivation of the new
6 65 |ig/m3 PM2 5 (24-hr) NAAQS set by EPA in 1997 to protect sensitive human population
7 groups from adverse effects of short-term exposures to fine particles. Of particular importance in
8 substantiating the need for fine particle standards were analyses of Harvard Six City Study data
9 reported by Schwartz et al. (1996a) showing stronger, more consistently statistically significant,
10 associations between acute (24-h) PM2 5 concentrations and increased mortality risks than for
11 24-h concentrations of inhalable coarse fraction particles (PM1S.2 5) in the same cities (see
12 Figure C-2).
13 However, as indicated in Chapter 5 of this document, there is little evidence
14 substantiating the occurrence of health effects due to acute (< 24-hr) exposures to diesel
15 emissions containing DPM at ambient or near-ambient concentrations. Note that 300 ug/m3 is
16 the lowest DPM concentration at which mild irritation and inflammation of respiratory tract
17 tissues (but not pulmonary function decrements) were observed with 1-hr controlled human
18 exposures of healthy adult volunteers to diesel exhaust (see Chapter 5). In contrast, various
19 noncancer (respiratory system) effects have been shown to occur in numerous mammalian
20 species as the result of controlled long-term (subchronic, chronic) exposures to DPM. Thus, key
21 elements forming the basis for derivation of the 15 ug/m3 PM25 annual-average NAAQS set in
22 1997 to protect against health effects associated with long-term fine particle exposures are far
23 more germane here in attempting to relate ambient fine particle health risk estimates to potential
24 ambient DPM exposure risks.
25 As noted in Chapter 6 of this document, the derivation of the 15 ng/m3 PM25 annual-
26 average standard was based, in part, on the assumption that increased mortality and morbidity
27 effects associated with acute (24-h) PM25 exposures were most likely due to PM25 concentrations
28 above the annual mean values for the cities evaluated. Also, it was noted in Chapter 6 that
29 annual mean PM2 5 values typically exceeded 15 ug/m3 for cities where 24-h PM2S levels were
30 found to be statistically significantly related to increased mortality and/or morbidity risks, as
31 shown by several key studies (Schwartz et al., 1996; Thurston et al., 1994; Neas et al., 1995).
32 Other key elements contributing to the derivation of the annual average PM25 NAAQS
33 were several new prospective cohort studies (published in the 1990's) that evaluated associations
34 between long-term exposures to ambient PM and increased risks of mortality or morbidity. The
35 most salient points of the PM CD (U.S. EPA, 1996a) assessment of such prospective cohort
7/25/00 C-3 DRAFT—DO NOT CITE OR QUOTE
-------
1 studies are summarized in Section C.2 below. These are augmented by discussion of pertinent
findings from recent new follow-up analyses for one of the subject prospective cohort studies.
4 C.2. Prospective Cohort Studies of Long-Term Ambient PM Exposure
5 Effects
6 Newer prospective cohort studies (Abbey et al., 1991; Dockery et al., 1993; and Pope
7 et al., 1995) were considered in the PM CD (1996a) as providing more credible evidence on
8 PM-health effects relationships than numerous previous cross-sectional studies. Salient features
9 of those three key prospective studies are summarized in Table C-4 (reproduced from Chapter 12
10 of the 1996PM CD).
11
12 C.2.1. Harvard Six U.S. Cities Study
13 Dockery et al. (1993) analyzed survival probabilities among 8,111 adults first recruited in
14 the mid-1970s in mid-western and eastern U.S. cities, including: Topeka, KS; Portage, WI (a
15 small town north of Madison); St. Louis, MO; Steubenville, OH; (an industrial community on
16 W. VA-PA border); Kingston-Harriman, TN (small towns southwest of Knoxville) and
17 Watertown, MA (western suburb of Boston). These locations comprise a transect across the
18 Northcentral and Northeastern United States, from the upper Midwest through Appalachia, to
^P suburban Boston. In each community, about 2,500 adults (white, aged 25 to 74, at enrollment)
20 were selected randomly, but the final cohorts numbered 1,400 to 1,800 persons in each city.
21 Follow-up periods ranged from 14 to 16 years, during which 13 to 22% of the enrollees died.
22 Of the 1,430 death certificates, 98% of the decedents were located, including persons who had
23 moved away and died elsewhere, but no information was provided on actual locations of death.
24 The analyses reported were mainly based on all-cause mortality; no mention was made of
25 subtracting external causes.
26 Air monitoring data obtained from routine sampling stations and special instruments set
27 up by the research team were used. Individual characteristics of the cohort subjects (and thus of
28 the decedents) considered in statistical analyses included: smoking habits, an index of
29 occupational exposure, body mass index, and completion of high school education. The Cox
30 proportional hazards model was used to estimate coefficients for individual risk factors after
31 stratifying by gender and age (5-year groups). The effects of air pollution were evaluated (a) by
32 estimating the relative risks of residence in each city relative to Portage (the city with the lowest
33 pollution levels for most indices) and (b) by including the community-average air quality levels
34 directly in the models. Since only six different long-term average values were available for each
pollutant, the effective degrees of freedom are small. Most of the air quality measures were
7/25/00 C-4 DRAFT—DO NOT CITE OR QUOTE
-------
1 averaged over the period of study, in an effort to study long-term (chronic) exposure effects; the
2 specific averaging periods varied by pollutant. Steubenville, Kingston-Harriman, and St. Louis
3 were the most polluted cities and also had the oldest and least educated cohorts and the heaviest
4 rates of smoking among the six cities.
5 No consideration was given to possible independent effects of occupation classification,
6 other personal lifestyle variables such as diet or physical activity, migration, or income.
7 Presumably, each subject was characterized by his status at entry to the study; follow-up data on
8 possible changes hi risk factors over time were not mentioned. Since the air quality data used in
9 this study were largely obtained from "private" monitoring rather than from public archives,
10 comparisons of the average levels with routine monitoring data were of some interest; and no
11 serious disagreements were found, except that it might have been preferable to consider peak
12 rather than average levels of ozone, as is more typical in most studies of acute O3 effects on
13 mortality. Also, it is notable that collection of size-classified PM data began in 1980, whereas
14 TSP data began in 1974 and from 1974 to 1980 there were large reductions in TSP (and likely
15 the size-classified particles as well), so that the size-classified data may be less representative
16 than TSP of cumulative exposures. Sulfate appeared to be intermediate hi this regard.
17 A more complete breakdown of relative risk estimates by city, sex, smoking status,
18 education, and body mass index is given hi Table C-5. The mean PM25 values are provided for
19 reference, but the adjusted relative risks used only age, smoking, education, and body mass as
20 covariates. The RR values for men and women combined are plotted hi Figure C-3 for each
21 pollutant. Note that the apparently linear relationship between fine particles and risk is less
22 linear if plotted separately for men and for women, and the confidence intervals also become
23 wider due to smaller sample sizes.
24 Substantial differences in survival rates (expected based on statewide mortality data) were
25 observed across the study's transect of the Northcentral and Northeastern U.S. The long-term
26 average mortality rate in Topeka was 9.7 deaths per 1,000 person-years and in Steubenville was
27 16.2, yielding a range in average (crude) relative risk of 67% among the six cities. After
28 individual adjustment for age, smoking status, education, and body-mass index, the range in
29 average relative risk was reduced to 26%. The relative importance of adjustments for age,
30 smoking, education, and body mass in determining the final ranks of the cities may be seen from
31 the Table C-5. Also, there is more scatter for men and women separately than when combined,
32 presumably because of the reduction in sample size.
33 Dockery et al. (1993) report that "mortality was more strongly associated with the levels
34 of fine, inhalable, and sulfate particles" than with the other pollutants, which they attributed
7/25/00 C-5 DRAFT DO NOT CITE OR QUOTE
-------
primarily to factors of particle size. They provided relative risk estimates and confidence limits
based on the differences between air quality in Steubenville and in Portage for these three PM
indicators. However, it is relatively simple to independently estimate coefficients from the
4 adjusted risks and pollutants levels in each of the six communities. These estimates obtained
5 (see Table C-6) correspond well to those of Dockery et al. (1993), based on output from the Cox
6 proportional hazards model. However, because there are only 6 different values for the air
7 quality data, the resulting confidence limits are considerably wider than those for the risk factors
8 having individual data. The estimates given in Table C-6, allow comparisons of results for
9 various pollutants and combination of pollutants. As in the original paper, the relative risks are
10 based on the difference in air pollution between Steubenville and Portage. The data for 1970
11 TSP (corresponding to a lag of about 12 years) were obtained from Lipfert (1978), assuming that
12 Madison could represent Portage, WI, as was done in the analysis of Schwartz et al. (1996b).
13 Table C-6 shows only small differences among many pollutants, including SO2 and NO2,
14 owing in part to the strong collinearity present. Note that relative risk elevations for the PM15
15 and fine particle indicators (PM2 5, SO4) were statistically significant. The non-sulfate portion of
16 PM2 5 had the tightest confidence limits. In contrast, TSP and the coarse particle variables
17 created by subtracting PM15 from TSP and PM2 5 from PM,5 were hot significant, suggesting that
18 particles > 15 ^m in aerodynamic diameter may be less important; this outcome may reflect in
^fe part greater spatial variability within the communities for coarse versus fine particles. Note also
20 that the estimated 1970 TSP variable performed slightly better than the TSP data (ca. 1982) used
21 by Dockery et al., thus suggesting a role for previous pollution exposure. Dockery et al. noted
22 that mean ozone levels varied little among cities; but this may have been less so if a measure of
23 peak (e.g., 1- or 8-hr) O3 levels had been used instead of daily (24-h) averages. Also, no
24 relationship was found for aerosol acidity (H+), but only limited data were available. Both sulfate
25 and non-sulfate fine particles effects seem rather similar, as shown in Figure C-2, making it
26 plausible that there may be PM effects related to particle size independent of sulfate content or
27 particle acidity.
28 In comparing the most and least polluted cities, Dockery et al. also reported elevated risks
29 for cardiopulmonary causes (RR 1.37; 95% CL 1.11 to 1.68) and lung cancer (RR 1.37; 95% CL
30 0.81 to 2.31, not significant). The relative risk for all other causes of death was 1.01 (0.79 to
31 1.30). When the six cities were considered individually, only Steubenville showed a statistically
32 significant (p < 0.05) elevated risk with respect to the least polluted city (Portage).
33 Comparison of pollution risks among the various cohort subsets considered is one of the
34 most useful outcomes of a study on individuals. Such comparisons must account for the higher
35 variability among subgroups, however, and the study was not capable of distinguishing excess
risks between subgroups less than about 18% (i.e., an excess risk of 1.18 cannot be distinguished
7/25/00 C-6 DRAFT—DO NOT CITE OR QUOTE
-------
1 from one of 1.36, for example). Although none of these subgroup differences were statistically
2 significant, the mortality risks associated with area of residence (and thus air pollution) were
3 higher for females and for smokers, as were risks for those occupationally exposed compared to
4 the nonexposed. Because of reduced uncertainties about exposures of non-smokers and
5 non-occupationally exposed persons to air pollution not reflected in the outdoor monitoring data
6 used in this study, the relative risk estimates for those subgroups might be the most reliable
7 estimates (1.19 and 1.17, respectively).
8 Issues concerning possible residual confounding, age adjustment, and smoking controls
9 were raised, and Dockery and Pope (1994) agreed that confounding is a potential concern but did
10 not address the possibility that variables other than the ones they considered might be important.
11 They dealt with the age adjustment issue quantitatively and pointed out that the air pollution risk
12 estimates were reasonably stable over different subgroups by smoking status. Age is a
13 potentially important covariate because it measures both susceptibility to health effects and
14 cumulative exposure to pollutants. There is also a possible interaction involving age, air
15 pollution, and tune of death, since air pollution concentrations in some communities such as
16 Steubenville and St. Louis decreased substantially during the years preceding and during the
17 period of the study.
18 The authors of the Harvard Six City Study were cautious in their conclusions, stating only
19 that the results suggest that fine-particulate air pollution "contributes to excess mortality in
20 certain U.S. cities." One further caveat is warranted before placing quantitative reliance on the
21 specific relative risk values generated by the study. If the responses to air pollution truly are
22 chronic in nature, it is logical to expect that cumulative exposure would be the preferred metric.
23 Pollution levels 10 years before the Six City study began were much higher in Steubenville and
24 St. Louis, as indexed by TSP from routine monitoring networks; and atmospheric visibility data
25 suggest that previous fine particle levels may have been higher in winter, but not necessarily in
26 summer. These uncertainties argue for caution in accepting and using the quantitative regression
27 results based solely on coincident monitoring data. For example, annual average TSP in 1965 in
28 Steubenville was about three times the value used by Dockery et al.; inclusion of older data in the
29 exposure indices would have reduced implied regression coefficients and relative risk estimates.
30
31 C.2.2. American Cancer Society (ACS) Study
32 Pope et al, (1995) analyzed 7-year survival data (1982 to 1989) obtained by the American
33 Cancer Society (ACS) for about 550,000 adult volunteers. The Cox proportional hazards model
34 was used to define individual risk factors for age, sex, race, smoking (including passive smoke
35 exposure), occupational exposure, alcohol consumption, education, and body-mass index. The
36 deaths (about 39,000 in all) were assigned to geographic locations using 3-digit zip codes for
7/25/00 C-7 DRAFT—DO NOT CITE OR QUOTE
-------
residences listed at enrollment into the ACS study in 1982. Relative risks were then computed
for 151 metropolitan areas defined by these zip codes and compared to corresponding air quality
data (ca. 1980). The sources of air quality data used were (a) the EPA AIRS system data for
4 sulfates, obtained from high-volume sampler filters for 1980, and (b) the Inhalable Particulate
5 Network data for fine particles (PM25) obtained from dichotomous samplers during 1979-81.
6 Pope et al. used the values from this data base reported by Lipfert et al., 1988, but only 50 PM2 5
7 locations could be matched with the death data. The correlation between the two pollutants was
8 0.73. Causes of death considered included all causes, cardiopulmonary causes (ICD-9 401-440,
9 460-519), lung cancer (ICD-9 162), and all other causes.
10 This study took great care with potential confounding factors for which data were
11 available. Several different active smoking measures were considered, as was time exposed to
12 passive smoke. The occupational exposure variable was specific to (any of) chemicals/solvents,
13 asbestos, coal or stone dusts, coal tar/pitch/asphalt, diesel exhaust, or formaldehyde. The
14 education variable was an indicator for having less than a high-school education, and alcohol use
15 and body-mass index were considered as linear predictors of survival. Pope et al. (1995) did not
16 report relative risk coefficients they obtained for these cofactors, which does not allow
17 comparison of findings for the non- pollution variables with exogenous estimates from
18 independent studies. Risk factors not considered by Pope etal. (1995) include: income,
^fc employment status, dietary factors, drinking water hardness and physical activity levels (all
20 shown to affect longevity); and they did not discuss possible influences of other air pollutants.
21 The ACS cohort is not a random sample of the U.S. population; it is 94% white and better
22 educated than the general public, with a lower percentage of smokers than in the Six City Study.
23 The (crude) death rate during the 7.25 years of follow-up was just under 1 % per year, which is
24 about 20% lower than expected for the white population of the U.S. in 1985, at the average age
25 reported by Pope et al. In contrast, the corresponding rates for the Six- Cities Study (Dockery
26 et al., 1993) discussed above tended to be higher than the U.S. average. In spite of these
27 differences, the cause specific-ratios for smoking are not significantly different between the ACS
28 and Six-Cities studies.
29 No mention was made of residence histories for the decedents; matching was done on
30 residence location at time of study entry. The 1979 to 1981 pollution values were assumed to be
31 representative of long-term cumulative exposures, in keeping with the goal of analyzing chronic
32 effects. However, the previous decade was one of extensive pollution cleanup in most of the
33 nation's dirtiest cities (TSP dropped by a factor of 2 in New York City, for example); but PM
34 levels remained relatively constant in cities that already met the standards. Thus, it is reasonable
to expect that the contrast between "clean" and "dirty" cities would have been greater in 1970
than in 1980. For example, the ranges of TSP and SO4 across the U.S. in 1970 were from 40 to
7/25/00 C-8 DRAFT—DO NOT CITE OR QUOTE
-------
1 224 and from 3 to 28 ng/m3, respectively (Lipfert, 1978). In 1980, these ranges decreased to
2 41 -142 and 2-17 jjg/m3 (Lipfert, 1984), suggesting that the dirtiest cities became cleaner while
3 the "clean" cities stayed about the same. The change in pollution range is about a factor of 1.8.
4 If the excess mortality found in the ACS study were in fact due to cumulative exposures, the
5 regression coefficients would have been biased upward (in terms of relative risk per jag/m3) by
6 only using the more recent data. The typically long latency period for lung cancer (ca. 20 yr.)
7 suggests that data on prior exposures may be particularly important for this cause of death.
8 The adjusted total mortality risk ratios (computed for the range of the pollution variables)
9 were 1.15 (95% CL = 1.09 to 1.22) for sulfates and 1.17 (95% CL = 1.09 to 1.26) for PM2 5,
10 suggesting that particle chemistry may be relatively unimportant as an independent risk factor.
11 Pope et al. (1995) found that the PM pollution coefficients were reduced by 10 to 15% when
12 variables for climate extremes were added to the model. No significant excess mortality for the
13 "other" causes of death was attributed to air pollution in this study. Note that Pope et al. found
14 very consistent pollution risks for males and females and for ever-smokers and never-smokers for
15 all-cause mortality. However, the relative risks for air pollution were slightly higher for females
16 for cardiopulmonary causes of death and the sulfate-lung cancer association was only statistically
17 significant for males, except for male never-smokers.
18 The results of the ACS prospective study were qualitatively consistent with those of the
19 Six City Study with regard to their findings for sulfates and fine particles; but relative standard
20 errors were smaller, as expected because of the substantially larger ACS database. However, no
21 other copollutants (e.g., O3, CO, NO2, etc.) were investigated in the ACS analysis, so that it was
22 not possible to provide an analogous type of pollutant comparison given earlier in Table C-6 for
23 the Six Cities Study. In addition, the ACS regression coefficients were about 1/4 to 1/2 of the
24 corresponding Six City values and were much closer to the corresponding values obtained in
25 various acute mortality studies.
26
27 C.2.3. California Seventh-Day Adventists Study
28 In the Abbey et al. (1991) prospective study (the Adventist Health Study of Smog or
29 "AHSMOG"), 6,338 long-term California residents (all white, non-Hispanic, and nonsmoking)
30 were followed for 6 to 10 years, beginning in 1976. Ambient air quality data dating back to 1966
31 were used in analyses restricted to those who lived within 5 miles of their current residence for at
32 least 10 years. Subjects lived either within the 3 major California air basins (San Diego, Los
33 Angeles, or San Francisco) or else were part of a random 10% sample of Adventist Health Study
34 participants in the rest of California. Individual exposure profiles (duration of exposure to
35 specific minimum concentration levels) were created for each participant, by interpolating to
36 their zip code centroids based on the 3 nearest monitoring stations. Monitored pollutants were
7/25/00 C-9 DRAFT—DO NOT CITE OR QUOTE
-------
1 mainly limited to TSP and O3 in this paper; but, total oxidant concentrations were used in the
f early part of the monitoring record. Health endpoints evaluated and the numbers of cases
included: (a) newly diagnosed cancers (incidence at any site) for males, 1 15; (b) any cancer site
4 for females, 175; (c) respiratory cancer, 17; (d) definite myocardial infarction, 62; (e) mortality
5 from any external cause, 845; and (f) respiratory symptoms, 272. The Cox proportional hazards
6 model was used, considering age, sex, past smoking, education, and presence of definite
7 symptoms of asthma, chronic bronchitis, or emphysema of airway obstructive disease (AOD) in
8 1 977 as individual risk factors, together with various exposure indices for TSP or O3 (considered
9 separately). Data on occupational exposures and history of high blood pressure were available
1 0 but not used in the mortality model; nor were data available on climate, body mass, income,
1 1 migration, physical activity levels or diet.
12 Of the above endpoints, only respiratory symptoms and female cancers (any site) were
1 3 reported by Abbey et al. (1991) to be statistically associated with TSP exposure. Neither heart
1 4 attacks or nonextemal mortality were associated with either TSP or O3 / oxidants. The authors
1 5 stated that possible errors in their estimated exposures to air pollution may have contributed to
1 6 the lack of significant findings, and a later version of the data base included estimates of
1 7 attenuation resulting from time spent indoors (Abbey et al., 1 993), but mortality was not
1 8 considered in the 1993 paper. Follow-up analyses (Abbey et al., 1995) considered exposures to
^P PM10 (estimated from site-specific regressions on TSP), PM2 5 (estimated from visibility), sulfates
20 (SO4), and visibility per se (extinction coefficient). No significant associations with nonextemal
2 1 mortality were reported, and only high levels of TSP or PM10 were associated with AOD or
22 bronchitis symptoms.
23 This study used an unique air quality data base developed explicitly for studying effects
24 of long-term cumulative exposures to community air pollution. The technique provided spatial
25 interpolations that were somewhat better for O3 than for TSP, in keeping with the regional nature
26 of O3. TSP may have been an inadequate index of exposure to inhalable particles, especially in
27 this relatively arid region where a large fraction of non-inhalable crustal particles could be
28 expected. Also, no attention was given to temporal matching of air quality and health; the
29 analyses using this data base were intended to evaluate the hypothesis that health is affected by
30 cumulative long-term pollution exposure at some undetermined time, as opposed to acute or
31 coincident exposures. Note that the data base began in 1966 and the mortality follow-up began
32 10 years later. Because air quality generally improved during this period, highest pollutant
33 concentrations likely occurred in the earlier part of the record; and one would not expect
34 spatially-based correlations to also reflect the sum of acute effects, as when air quality and health
data are also matched in time.
S^L
7/25/00 C- 1 0 DRAFT— DO NOT CITE OR QUOTE
-------
1 The PM CD (U.S. EPA 1996a) noted that the finding of Abbey et al. (1991, 1995) of no
2 association between long-term cumulative exposure to ambient TSP or O3 (or to SO4 or estimated
3 PM10 or PM2 5) concentrations and all natural-cause mortality could be interpreted as showing the
4 absence of chronic responses after 10 years but not necessarily the absence of (integrated) acute
5 responses, since coincident air pollution exposures or integrated exposures over the preceding
6 few years were not considered. It is also possible that the exposure measurements or estimates
7 used were inadequate or that the latency period for chronic effects may exceed 10 years and that
8 additional follow-up might still reveal chronic effects.
9 Further such follow-up analyses of the same California AHSMOG database have been
10 reported recently by Abbey et al. (1999). These analyses (not considered in the 1996 PM CD or
11 1997 PM NAAQS decisions) do provide some evidence indicative of increased risk of mortality
12 from contributing non-malignant respiratory causes being associated with long-term PM
13 exposures. Other recent AHSMOG analyses reported by Abbey et al. (1999) and Beeson et al.
14 (1998) are also suggestive of increased risk of mortality from lung cancer possibly being
15 associated with long-term PM10 exposures, as summarized below.
16 Abbey et al. (1999) evaluated the mortality status of AHSMOG subjects after ca. 15-years
17 of follow-up (1977-1992), finding 1,628 deaths (989 female, 639 male) in the cohort. There
18 were 1,575 deaths from all natural (non-external) causes, of which 1,029 were cardiopulmonary
19 deaths, 135 were non-malignant respiratory deaths (ICD9 codes 460-529), and 30 were lung
20 cancer deaths (ICD9 code 162). Abbey et al. (1999) also created an additional death category,
21 "contributing respiratory causes" (CRC). CRC included any mention of nonmalignant
22 respiratory death as either an underlying cause or a contributing cause on the death certificate
23 CRC coded by an exposure-blinded nosologist (the other groups listed only underlying causes),
24 with 410 deaths (246 female and 164 male) being found. Numerous analyses were done for the
25 CRC category, due to the large numbers and relative specificity of respiratory causes as a factor
26 in the deaths. Education was used as an index of socio-economic status, rather than income.
27 Physical activity and occupational exposure to dust were also used as covariates. Migration was
28 not a major concern in this residentially stable cohort.
29 A number of exposure indicators were used: mean values of PMIO (imputed from TSP in
30 the earlier years of the study), SO4, SO,, O3, and NO2; and "threshold'' indicators (i.e., days per
31 year with PM10 > 100 |J.g/m3; and hours per year with O3 > 100 ppb). In summary tables that
32 follow below, the "standard" increments used for PM10 and SO4 are (a) the same as used earlier
33 for the short-term mortality studies (50 ng/m3 for PM10 and 15 |jg/m3 for SO4) and (b) 30 days
34 per year for exceedances of PM,0 above 100 ug/m3. The mean values for PMIO and SO4 during
3R the study period were 51 and 7.2 ug/m3 respectively, and 31 days per year for PM10 exceedances
36 over 100 |ig/m3. The means were much larger than the inter-quartile ranges (IQR) of 24 and
7/25/00 C-11 DRAFT—DO NOT CITE OR QUOTE
-------
1 3.0 ng/m3. IQR is the increment used for other variables. RR and confidence limits using IQR
ffrom Abbey et al. (1999) are shown to 2 decimal places; those estimated for standard increments
are shown to 3 decimal places.
4 Cox proportional hazard models adjusted for a variety of covariates, or stratified by sex,
5 were used in the models. The "time" variable used in most of the models was survival time from
6 date of enrollment, except that age on study was used for lung cancer effects due to the expected
7 lack of short-term effects. A large number of covariate adjustments were evaluated, as shown in
8 Table C-7 and described by Abbey et al. (1999).
9 The CRC RR estimates for 30 days per year with PM10 > 100 ug/m3 for males and
10 females combined are shown in Table C-7. Positive and statistically significant effects are found
11 for almost all models that include age, pack-years of smoking, and body-mass index (BMI) as
12 covariates. Subsets of the cohort also often had elevated risks. Former smokers had higher
13 relative risks than never-smokers (RR for PM10 exceedances for never-smokers was marginally
14 significant by itself, in spite of the reduced sample size). Subjects with low intake of anti-
15 oxidant vitamins A, C, E had significantly elevated risk of response to PMIO whereas those with
16 adequate intake did not, suggesting that dietary factors (or possibly other socio-economic or life
17 style factors for which they are a surrogate) may be important covariates. There also appears to
18 be a gradient of PM10 risk with respect to time spent outdoors, with individuals who had spent at
^B least 16 hours per week outside at distinctly elevated risk from PMIO exceedances. The extent to
20 which time spent outdoors is a surrogate for other variables or is a modifying factor reflecting
21 temporal variation in exposure to ambient air pollution is not certain. For example, males spend
22 about twice as much time outdoors as females, so that outdoor exposure time is confounded with
23 gender.
24 A considerably different picture is shown when the analyses are broken down by gender.
25 Table C-8 shows much lower RR for female CRC deaths for all co-pollutants, with all female
26 RR positive, but not statistically significant. The CRC for males remains significant only for
27 PM10 exceedances, but not for other air pollution metrics. The PM10 exceedance effect for CRC
28 for both sexes is roughly the average of that for males and females. Personal monitoring was not
29 conducted on this part of the cohort, and other factors (e.g., occupational exposure) for which the
30 questionnaire was not adequate may also account for male vs. female differences, along with
31 gender differences in the amount of time spent outdoors. Finally, it is not surprising that
32 individuals reporting respiratory symptoms in 1977 may be at greater risk to PM10 or other
33 environmental insults presumably involved in subsequent CRC deaths, and prior health status
34 may also be gender-related.
7/25/00 C-l2 DRAFT—DO NOT CITE OR QUOTE
-------
1 Table C-9 shows much lower RR for female non-external deaths for all co-pollutants,
2 with no female RR positive nor statistically significant. Deaths from non-external causes for
3 males remains statistically significant for PMIO exceedances, but not for other air pollution
4 metrics. However, the RR estimates for males for other air pollutant metrics are relatively large.
5 Table C-10 shows much lower RR for female cardio-pulmonary deaths for all
6 co-pollutants, with only the female RR for mean SO2 positive and none statistically significant.
7 The RR for deaths from cardiopulmonary causes for males is no longer statistically significant
8 for PM10 exceedances, nor for other air pollution metrics (although the RR estimates for males
9 for air pollutant metrics are relatively large).
10 Table C-l 1 shows a confusing welter of results obtained for lung cancer mortality.
11 For example, the RR's for lung cancer deaths are significant for males for PM10 and O3 metrics,
12 but not for females. In contrast lung cancer deaths are significant for mean NO2 for females, but
13 not for males, but lung cancer metrics for mean SO2 are significant for both males and females.
14 This pattern is not readily interpretable, but may be attributable to the very small numbers of
15 cancer-related deaths (18 for females; 12 for males), resulting in wide RR confidence intervals.
16 In general, this study (Abbey et al., 1999) suggests a pattern of mortality from diverse
17 causes (e.g., CRC, lung cancer) in males, but provides little evidence for female mortality from
18 these causes. The male causes primarily appear to be associated with exposures to PM10 and
19 especially to PM10 > 100 ug/m3. Some other air pollutants (SO2, NO2) appear to be associated
20 with lung cancer deaths in females.
21 The analyses reported here attempted to separate PM,0 effects from those of the other
22 pollutants by use of two-pollutant models, but none of the quantitative findings from these
23 models were reported. The Abbey et al. (1999) text mentions that the PM10 coefficient for CRC
24 remained stable or increased when other pollutants were added to the model. Lung cancer
25 mortality models for males were evaluated for co-pollutant effects in detail. NO2 remained
26 nonsignificant in all two-pollutant models, and the other pollutant coefficients were stable in
27 magnitude. The PM10 and O3 effects remained stable when SO2 was added, suggesting that their
28 effects are independent. However, the effects of PM10 and 03 were hard to separate because
29 these pollutants were highly correlated in this study. When both exceedances PMIO > 100 jag/mj
30 and O3 > 100 ppb were used in the model, both RR were reduced in magnitude, but the O3
31 exceedance RR remained more significant than the RR for the PMIO exceedance. The possibility
32 that the finding of a significant PM,0 effect is partially attributable to correlation with other
33 pollutants such as O3 cannot be precluded. The SO, coefficient for lung cancer mortality in
34 females remained stable in two-pollutant models when PM,0 and O3 exceedances were included.
35 This suggests that the significance of the SO2 effect for females may not be an artifact wholely
36 attributable to collineariry with these co-pollutants.
7/25/00 C-13 DRAFT—DO NOT CITE OR QUOTE
-------
11 . Beeson et aL (1998)
fThis study used essentially the same data as did Abbey et al. (1999), but concentrates on
lung cancer incidence (1977-1992) as an endpoint. There were only 20 female cases and 16 male
4 cases of lung cancer among the 6,338 AHSMOG subjects. The exposure metrics were
5 constructed to be specifically relevant to cancer, being the annual average of the monthly
6 exposure indices from January, 1973 through the following months, but ending 3 years before the
7 date of diagnosis of the case. This represents a 3-year lag between exposure and diagnosis of
8 lung cancer, allowing for a latency period. Therefore, statistical indices for exposure have
9 somewhat different statistics than in Abbey et al. (1999), such as the IQR and mean.
10 The covariates in the Cox proportional hazards model were pack-years of smoking and
11 education, and the time variable was attained age. A number of additional covariates were
12 evaluated for inclusion in the model, but only 'current use of alcohol' met the criteria for
13 inclusion in the final model. Individual pollutants evaluated were PM10, SO2, NO2, and O3.
14 No interaction terms with the pollutants proved to be significant, including outdoor exposure
15 times. Gender-specific relative risk estimates were reported for the various risk factors. Results
16 are shown in Table C-12 for males and Table C-13 for females. Standard increments were used
17 for PM10 mean (50 ug/m3) and exceedances of PM10 > 100 ug/m3 (30 d/y). The RR estimates and
18 confidence limits using IQR from Beeson et al. (1998) are shown to 2 decimal places, those
^A estimated for standard increments are shown to 3 decimal places.
20 The RR estimates for the male lung cancer cases are: positive and statistically significant
21 for all PM10 indicators; positive and predominantly significant for O3 indicators, except for mean
22 O3, number of O3 exceedances > 60 ppb, and in former smokers; and are positive and significant
23 for mean SO2, except when restricted to proximate monitors. The RR for mean NO2 is positive
24 but not significant. The very high RR for mean PM10 for males (31.1) may be attributable to the
25 small number of cases (N = 16) and the large standard increment (50 ug/m3) used. When data
26 are restricted to subjects with at least 80 percent A/B quality data (within 32 km of the
27 residence), the RR is reduced to 9.26 over 50 ug/m3. The RR over the IQR of 24 ug/m3 in the
28 full data set is 5.21, so that the use of the IQR may be more appropriate for the exposure in long-
29 term studies.
30 The female lung cancer RR estimates reported by Beeson et al. (Table C-13) are much
31 smaller than those for males, not being statistically significant for any indicator of PM10 or O3
32 and statistically significant only for mean SO2.
33 Extensive multi-pollutant analyses were also carried out. Regression coefficients for
34 PM10 and SO2 were not reduced when O3 or NO2 were added to the single-pollutant models for
males. The regression coefficients for the two-pollutant model with PMIO and SO2 remained
highly positive and significant, which the authors suggest may be associated with independent
7/25/00 C-14 DRAFT—DO NOT CITE OR QUOTE
-------
effects of PM10 and SO2 on lung cancer incidence. PM10 was more strongly correlated with lung
cancer in males than the other pollutants. For females, the SO2 coefficient remained significant
3 when co-pollutants were added one at a time, and was the air pollutant most strongly associated
4 with female lung cancer cases.
5 The results of Abbey et al. (1999) and Beeson et al. (1998) are somewhat different than
6 those of earlier studies using the same cohort. Abbey et al. (1991) reported completely
7 non-significant relationships between total ('all natural causes') mortality and air pollution. The
8 RR for 1000 h/y of TSP > 200 ng/m3 was 0.99 (CI 0.87-1.13), and for 500 h/y of O3 > 100 ppb
9 was 1.00 (CI 0.89-1.12), after 10 years of follow-up. Also, Abbey et al. (1991) reported no
10 statistically significant increases in all malignant neoplasms for males attributable to air
11 pollution. The RR for 1000 h/y of TSP > 200 ng/m3 was 0.96 (CI 0.68-1.36), and for 500 h/y of
12 O3 > 100 ppb was 1.09 (CI 0.80-1.47), after 10 years of follow-up. However, there was a
13 statistically significant increase in all malignant neoplasms for females. The RR for females
14 attributed to 1000 h/y of TSP > 200 ug/m3 was 1.37 (CI 1.05-1.80). Neoplasms in females
15 attributed to 500 h/y O3 > 100 ppb were much less significant, with RR = 1.03 (CI 0.81-1.32).
16
17 C.2.4. Relationship of AHSMOG to Six Cities and ACS Study Findings
The results of the recent AHSMOG mortality studies (Abbey et al., 1999) are compared
below with the earlier Six Cities Study (Dockery et al., 1993) and ACS Study (Pope et al., 1995).
20 Tables C-14, C-15, and C-16 compare the estimated RR for total, cardiopulmonary, and lung
21 cancer mortality, respectively, among the studies. The PM indices used are the mean PM10
22 concentration for the Six Cities and AHSMOG studies (increment 50 ug/m3), and the mean PM2 5
23 and SO4 concentrations (increments 25 and 15 |ig/m3 respectively) for the ACS study. The
24 comparisons for the Six Cities and ACS studies have been translated from published RR for the
25 most polluted vs. least polluted city for PM10, PM2 5, and SO4. Results are shown by sex and
26 smoking status. The AHSMOG subjects are classified as 'non-smokers', although some former
27 smokers are included. The ACS study combines past and current smokers into an 'ever smoker'
28 category, although long-term past smokers are at much lower risk than current smokers. The
29 number of subjects in these studies varies greatly (6,338 AHSMOG subjects, 8,111 Six Cities
30 Study subjects; compared to 295,223 subjects in the 50 fine particle cities and 552,138 subjectsin
31 the 151 sulfate cities of the ACS study), and may partially account for differences among their
32 results.
33 Table C-14 shows relative risks for total mortality at comparable standard increments.
34 RR is generally highest for the Six Cities Study. The AHSMOG Study found a much smaller
RR for women than did the other studies, whereas the effe.rt for males; was similar tn nnn-
smokers in the ACS Study and marginally significant. RR among the three studies varied
7/25/00 C-15 DRAFT—DO NOT CITE OR QUOTE
-------
substantially with sex and smoking categories. Six of the 16 independent analyses showed
significant positive RR (LCL > 1.0), but subsetting the data allowed less power to detect effects
"3 than the whole data sets would have allowed. Neither of the AHSMOG RR were significant
4 using the mean as the PM10 index, but another PMIO index (exceedances over 100 jig/m3) was
5 significant for males.
6 Table C-l 5 shows relative risks for cardiopulmonary mortality at comparable standard
7 increments. RR is highest for the Six Cities Study, which did not report separate effects by sex
8 and smoking status. The AHSMOG Study found a much smaller cardiopulmonary RR for
9 women than did the other studies. However, the RR for male non-smokers was much more
10 similar to the ACS results than for female non-smokers. RR for the AHSMOG endpoint CRC
11 ('contributing respiratory causes') was more similar to the ACS findings for women, but higher
12 in men, although the confidence intervals are very wide. Seven of 13 of the independent analyses
13 showed significant positive RR (LCL > 1.0). The AHSMOG cardiopulmonary RRs using mean
14 PM10 were not significant for either males or females. However, the 100 ng/m3 exceedance
15 index for males was nearly so.
16 Table C-l6 shows relative risks for lung cancer mortality at comparable standard
17 increments for PM-related variables. The lung cancer mortality RR estimates were highest for
18 ' males in the AHSMOG study, and statistically significant. The AHSMOG study also found a
^P larger RR for women than did the other studies. The only other statistically significant finding
20 for lung cancer mortality was for past and current male smokers in the ACS 151-city sulfate
21 study. The overall pattern of results for lung cancer, then, is a somewhat conflicting set of
22 findings across the three prospective cohort studies assessed here, providing only somewhat
23 suggestive evidence at best for possible ambient PM relationship to increased lung cancer risk.
24 There is no obvious statistically significant relationship between PM effect sizes, gender,
25 and smoking status across these studies. The AHSMOG studies show no statistically significant
26 relationships between PM10 and total mortality or cardiovascular mortality for either sex, and
27 only for male lung cancer incidence and lung cancer deaths in a predominantly non-smoking
28 sample. The ACS results, in contrast, show similar and significant associations with total
29 mortality for both "never smokers" and "ever smokers", although the ACS cohort may include a
30 substantial number of long-term former smokers with much lower risk than current smokers.
31 The Six Cities Study cohort shows the strongest evidence of a higher PM effect in current
32 smokers than in non-smokers, with female former smokers having a higher risk than male former
33 smokers. This study suggests that smoking status is "effect modifier" for ambient PM, just as
34 smoking may be a health effect modifier for ambient ozone (Cassino et al., 1999).
It is interesting to note, in relation to the above discussion, that a comparison of the
Six-Cities Study non-smoker RRs with the Six-Cities results in Table C-l4 for smokers indicates
7/25/00 C-16 DRAFT—DO NOT CITE OR QUOTE
-------
1 that larger and more significant effects of ambient PM pollution are found for smokers than
f non-smokers. This suggests that smoking is an effect modifier that increases the adverse effects
of ambient pollution. This trend is consistent with air pollution effect causality, as smokers
4 represent a compromised population, logically more likely to be adversely affected by air
5 pollution. This may also explain why the reported AHSMOG study RRs are generally not
6 significant, in contrast with the overall Six-Cities Study results (but consistent with the Six-Cities
7 nonsmoker results), as there are no identified smokers among the AHSMOG study group to
8 "drive up" the overall significance of the air pollution effect. This again indicates that more
9 years of follow-up may be required to see any statistically significant total mortality effects in
1 0 both the AHSMOG and Six-Cities studies' non-smoking populations.
11
1 2 C.2.5. Studies by Particulate Matter Size-Fraction and Composition
1 3 Particulate matter mass varies widely over time and from place to place in size and
1 4 chemical composition, and this likely affects the toxicity of that mass. The semi-individual
1 5 cohort studies assessed here investigated the relative roles of various PM components in the air
1 6 pollution association with mortality. As shown in Table C-17, the Harvard Six-Cities study
1 7 (Dockery et al., 1993) results indicated that the PM2S and SO4 RR associations (as indicated by
1 8 their respective 95% CF s and t-statistics) were stronger than those for the coarser mass
1^P components. However, the effects of sulfate and non-sulfate PM25 are indicated to be quite
20 similar. Acid aerosol (H*) exposure was also considered by Dockery et al. (1993), but only less
21 than one year of measurements collected near the end of the follow-up period were available in
22 most cities, so the Six-Cities results were much less conclusive for the acidic component of PM
23 than for these other PM metrics (that, in contrast, were measured over many years during the
24 study). The Six-Cities Study also yielded total mortality RR estimates for the reported range
25 across those cities of PM25 and SO4 concentrations that, although not statistically different, were
26 roughly double analogous RRs for the TSP-PM15 and PMI5.2 5 mass components.
27 Table C-18 presents comparative PM25 and SO4 results from the ACS study that indicate
?8 that, although the RR differences were not statistically significant across pollutants, the SO4 RRs
29 were in every case more strongly significant than those for the PM2 5 across the various mortality
30 cause classifications considered, especially for lung cancer (SO4 t=2.92 vs. t=0.38 for PM25).
3 1 The most recent AHSMOG study analysis (Abbey et al., 1999) employed PM10 as its PM
32 mass index, finding some significant associations with total and by-cause mortality, even after
OO /-»/-»»"* 4-^i-^1 1**t«T •£*-»•*• *^/-»+^»i*-» r»11 f*r^-n-F/tiir*s3v*+f* frt <"•+/•*»•<-• / i f* /•» 1 1 •» y-J i •*+ *-r x-\ + V-«/i»« »trt 1 1 * i*i-\w» + *-A TV*^ i~. f\ »•» rt 1 T r «
^> »-f ^WAALAV/AiAAAft A.\J± M*-* VWAJ.HC**.*. J WV/11..1.VS CAAAVAAA At A.U.WVV/AiJ> \AAAW4 C*.V*.AAi& WVAAWi. LJV/ AJ. kAlUJ.Al.Oy . AlAAtJ CUA14A J
o r\
not as strongly associated as PM10 with mortality, and was not found to be statistically significant
for any mortality category. The significant mortality associations found for PMi0 contrasts with
7/25/00 C- 1 7 DRAFT— DO NOT CITE OR QUOTE
-------
1 } previously published AHSMOG study PM analyses that found weaker mortality associations
with TSP (Abbey et al., 1991). Although the longer follow-up time in this new analysis may
"3 have also contributed, the greater strength of association by PM10 vs. TSP is consistent with the
4 Harvard Six-City study results presented in Table C-17, as well as with the Ozkaynak and
5 Thurston (1987) cross-sectional comparisons of mortality associations with the various PM
6 fractions.
7 Single-pollutant results about PM components are informative, however, as shown in
8 Table C-19 for total mortality, and in Table C-20 for cardiopulmonary causes. The t-statistics are
9 compared for studies where appropriate: mean PM10, PM10_2 s, PM2 5, and sulfate for the Six
10 Cities (Dockery et al., 1993); mean PM2S and sulfate for ACS (Pope et al., 1995); mean PM10 and
11 sulfate, and PM10 exceedances of 100 ug/m3 for AHSMOG (Abbey et al., 1999).
12 Estimates for Six Cities parameters were calculated in two ways: (1) mortality RR for
13 most versus least polluted city in (Table 3, Dockery et al., 1993) adjusted to standard increments;
14 (2) ecological regression fits in (Table 12-18, U.S. Environmental Protection Agency, 1996).
15 The eastern and mid-western Six Cities suggest a strong and highly significant relationship for
16 fine particles and sulfates, a slightly weaker but still highly significant relationship to PM,0, and a
17 marginal relationship to PMI0.2 5. The ACS study looked at a broader spatial representation of
18 cities, and found a stronger statistically significant relationship to PM2 5 than to sulfate (no other
^P pollutants were examined).
20 Overall, the prospective cohort studies conducted to-date collectively confirm cross-
21 sectional study indications that, as opposed to the more coarse mass fractions, the fine mass
22 component of PM (and sometimes including its acidic sulfate constituent) are strongly correlated
23 with mortality.
24 The credibility of the above findings of increased risk of mortality being associated with
25 chronic, long-term exposures to fine particles is enhanced by analogous findings of increased risk
26 of respiratory symptoms and lung function decrements being associated with long-term
27 exposures to fine particles, as illustrated in Figure C-4. That figure graphically depicts results
28 from the study reported on by Razienne et al. (1996), which demonstrate strong positive
29 relationships between decrements in children's lung function and long-term exposure to fine
30 particles (indexed by PM2,), but not to inhalable thoracic coarse particles (PM]0.2,).
31
32 C.2.6. Conclusions
33 A review of the prospective cohort studies summarized in the previous PM AQCD (U.S.
34 Environmental Protection Agency, 1996) indicates that past epidemiologic studies of chronic PM
§ exposures collectively indicate increases in mortality to be associated with long-term exposure to
airborne particles of ambient origins. The PM effect size estimates for total mortality from these
7/25/00 C-18 DRAFT—DO NOT CITE OR QUOTE
-------
1 studies also indicate that a substantial portion of these deaths reflected cumulative PM impacts
f above and beyond those exerted by acute exposure events.
The new AHSMOG study (Abbey et ah, 1999) provides all-cause mortality RR estimates
4 for adult males that are quantitatively and qualitatively consistent with prior semi-individual
5 prospective cohort studies, especially the similarly designed 6-Cities study. Extensive new
6 by-gender, by-cause, and multiple pollutant sensitivity analyses, as well as a more comprehensive
7 analyses of numerous potentially uncontrolled factors in this study (such as of the effects of
8 variations in the time spent outdoors) provide important new evidence that is largely supportive
9 of the mortality associations with PM of ambient origins previously reported by the Six-Cities
10 and ACS studies.
11 With regard to the role of various PM constituents in the PM-mortality association, cross-
12 sectional studies have generally found that the fine particle component, as indicated either by
13 PM2 5 or sulfates, was the PM constituent most consistently associated with mortality.
14 In addition, the Six-Cities prospective semi-individual study also indicates that the fine mass
15 components of PM are more strongly associated with the mortality effects of PM than the coarse
16 PM components.
17 The recent analyses of the long-term AHSMOG study provide some evidence indicative of
18 health effects being associated with ambient PM10 exposure for which a substantially greater
^P level of individualized ambient PMIO information is available, but also demonstrates some
20 differences with the earlier Six Cities and ACS studies (Dockery et ah, 1993; Pope et ah, 1995).
21 Statistically significant increases in lung cancer incidence (Beeson et ah, 1998) and statistically
22 significant increases in lung cancer deaths and deaths associated with any contributing respiratory
23 causes (Abbey et ah, 1999) were found in AHSMOG males, but not females. The results were
24 generally robust to different confounder specifications, population subsets, and inclusion of
25 co-pollutants, and were larger for and more significant for PM exceedance indices (number of
26 days per year with PMIO greater than a cut point, typically 100 ng/m3) than with the mean PMIO
27 concentration. However, PM,0 was estimated from TSP rather than measured in the earlier part
28 of the AHSMOG study and, therefore, the AHSMOG results may not be as credible as those
29 from the other two prospective cohort studies where direct PM,0, PM2 5, or SO4 measurements
30 data were used.
31 Using the same mean PM10 increment of 50 ug/m3, total mortality attributable to long-term
32 ambient PM10 RR was similar to that of the ACS study for PM25 for male nonsmokers (1.24) and
33 smaller than that for the Six Cities study (1.57), albeit only significant fci the ACS study
7/25/00 C-19 DRAFT—DO NOT CITE OR QUOTE
-------
I 1} (Table C-13). The AHSMOG RR for females (Table 6-31) is smaller and non-significant (0.88),
A2 whereas the ACS RR for female non-smokers is significant and only somewhat smaller than the
3 male RR (1.22 in the 50-city PM2 5 study, 1.15 in the 151 -city SO4 study) and 1.28 in the
4 Six Cities.
5 The AHSMOG findings for cardiopulmonary mortality attributable to long-term ambient
6 PM10 are positive for males, but not statistically significant, whereas the ACS findings are
7 significant for female nonsmokers in both studies and in male nonsmokers for the 151 -city study
8 (Table C-14). However, the male RR in AHSMOG (1.22 for cardiopulmonary deaths, 1.54 for
9 CRC deaths) is similar to that of ACS male non-smokers (1.24 for the 50-city study, 1.21 for the
10 151-city study) and smaller than that for all Six Cities subjects (1.74, includes smokers and
11 non-smokers). The ACS female non-smokers have RR of 1.58 and 1.32 respectively, both
12 significant, compared to 0.84 in AHSMOG.
13 Lung cancer mortality attributable to long-term ambient PM]0 is not significant for females
14 in any of the studies, nor for male nonsmokers in ACS, but was reported to be statistically
15 significant for male nonmokers in AHSMOG and male smokers in ACS 151-city. Lung cancer
16 mortality attributable to long-term ambient PM2 5 was not significant for either gender in the ACS
17 and Six Cities studies. Thus, the available overall evidence, from the three prospective cohort
18 studies of PM effects assessed here, definitely is not conclusive and can, at best, be viewed as
indicative of possible ambient PM associations with increased risk of lung cancer or associated
20 mortality.
7/25/00 C-20 DRAFT—DO NOT CITE OR QUOTE
-------
Table C-l. Effect estimates per 50 ug/m3 increase in 24-h PM10 concentrations from
U.S. and Canadian studies
Study Location
RR (± CI) RR (± CI)
Only PM Other Pollutants
in Model in Model
Reported
PM10 Levels
Mean
(Min/Max)t
Increased Total Acute Mortality
Six Cities2
Portage, WI
Boston, MA
Topeka, KS
St. Louis, MO
Kingston/Knoxville, TN
Steubenville, OH
St. Louis, MOC
Kingston, TNC
Chicago, ILh
Chicago, ILg
Utah Valley, UTb
Birmingham, ALd
Los Angeles, CAf
—
1.04(0.98,1.09) —
1.06(1.04,1.09) —
0.98 (0.90, 1.05) —
1.03(1.00,1.05) —
1.05(1.00,1.09) —
1.05(1.00,1.08) —
1.08(1.01,1.12) 1.06(0.98,1.15)
1.09 (0.94, 1.25) 1.09 (0.94, 1.26
1.04(1.00,1.08) —
1.03 (1.02, 1.04) 1.02 (1.01, 1.04)
.1.08(1.05,1.11) 1.19(0.96,1.47)
1.05(1.01,1.10) —
1 .03 (1 .00, 1 .055) 1 .02 (0.99, 1 .036)
18 (±11. 7)
24 (±12.8)
27 (±16.1)
31 (±16.2)
32 (±14.5)
46 (±32.3)
28 (1/97)
30 (4/67)
37 (4/365)
38(NR/128)
47 (1 1/297)
48 (21, 80)
58(15/177)
Increased Hospital Admissions (for Elderly > 65 yrs.)
Respiratory Disease
Toronto, CAN'
Tacoma, WAj
New Haven, CTj
/~>l 1, J /-VTTk
Spokane, WA1
COPD
Minneapolis, MN"
Birmingham. AT,m
Spokane, WA1
Detroit, MP
1.23 (1.02, 1.43)* 1.12 (0.88, 1.36)J
1.10(1.03,1.17) 1.11(1.02,1.20)
1.06(1.00,1.13) 1.07(1.01,1.14)
1 06 '1 00 ^ 1 1 ^
1.08(1.04,1.14) —
1.25(1.10,1.44) —
1.13(1.04.1.22) —
1.17(1.08, 1 ?7) —
1.10(1.02,1.17) —
30-39*
37 (14, 67)
41 (19, 67)
As riQ 77^1
-— ' V 3 - S
46 (16, 83)
36(18,58)
45 (19, 77)
46(16.83)
48 (22, 82)
7/25/00
C-21
DRAFT—DO NOT CITE OR QUOTE
-------
Table C-l. Effect estimates per 50 ug/m3 increase in 24-h PM,0 concentrations from
U.S. and Canadian studies (continued)
Study Location
Pneumonia
Minneapolis, MN"
Birmingham, ALm
Spokane, WA1
Detroit, MI°
Ischemic HP
Detroit, MF
Increased Respiratory
Lower Respiratory
Six Citiesq
Utah Valley, UT
Utah Valley, UTS
Cough
Denver, CO"
Six Citiesq
Utah Valley, UP
RR(±CI)
Only PM
in Model
1.08(1.01,1.15)
1.09(1.03,1.15)
1.06(0.98,1.13)
—
1.02(1.01, 1.03)
Symptoms
2.03(1.36,3.04)
1.28(1.06, 1.56)T
1.01 (0.81, 1.27)*
1.27(1.08,1.49)
1.09(0.57,2.10)
1.51 (1.12,2.05)
1.29(1.12, 1.48)
RR (± CI) Reported
Other Pollutants PM,0 Levels
in Model Mean (Mm/Max)1
— 36 (18,58)
— 45 (19, 77)
— 46(16,83)
1.06(1.02,1.10) 48(22,82)
1.02(1.00,1.03) 48(22,82)
Similar RR 30(13,53)
— 46(11/195)
— 76(7/251)
— 22 (0.5/73)
Similar RR 30(13,53)
— 76(7/251)
Decrease in Lung Function
Utah Valley, UTr
Utah Valley, UTS
Utah Valley, UT"
References:
•Schwartz et al. (1996a).
"Pope et al. (1992, 1994)/O,.
55 (24, 86)**
30 (10, 50)'*
29(7,51)***
'Schwartz (1996).
"Schwartz (1994e).
— 46(11/195)
— 76(7/251)
— 55(1,181)
"Ostroetal. (1991)
'Min/Max 24-h PM,,, in parentheses unless noted
'Dockery et al. (1992)/O3.
"Schwartz (1993).
'Kinney et al. (1995)/O3, CO.
8Ito and Thurston (1996)/O,.
hStyeretal. (1995).
Thurston et al. (1994)/O3.
'Schwartz (l995)/SOj.
kSchwartz et al. (1996b).
"Schwartz (1994f).
"Schwartz (1994d).
'Schwartz and Morris (1995)/O3> CO,
•"Schwartz et al. (1994).
Tope etal. (1991).
'Pope and Dockery (1992).
'Schwartz (1994g)
"Pope and Kanner( 1993).
otherwise as standard deviation (± S.D), 10 and
90 percentile (10, 90). NR = not reported.
SO2. "Children.
"Asthmatic children and adults.
"Means of several cities.
"PEFR decrease in ml/sec.
'"FEV, decrease.
•RR refers to total population, not just>65 years.
7/25/00
C-22
DRAFT—DO NOT CITE OR QUOTE
-------
Table C-2. Effect estimates per variable increments in 24-h concentrations of fine
particle indicators (PM, ^, SO^ H*) from U.S. and Canadian studies
Acute Mortality
Six City3
Portage, WI
Topeka, KS
Boston, MA
St. Louis, MO
Kingston/Knoxville,
TN
Steubenville, OH
Indicator
PM2.5
PM2.5
PM25
PM25
PM25
PM7S
RR (± CI) per 25 ug/m3
PM Increase
1.030(0.993,1.071)
1.020(0.951,1.092)
1.056(1.038,1.0711)
1.028(1.010,1.043)
1.035(1.005,1.066)
1.025(0.998,1.053)
Reported PM
Levels Mean
(Min/MaxV
11. 2 (±7.8)
12.2 (±7.4)
15.7 (±9.2)
18.7 (±10.5)
20.8 (±9.6)
29.6 (±2 1.9)
Increased Hospitalization
Ontario, CAN"
Ontario, CAN0
NYC/Buffalo, NYd
Toronto*1
so:
so:
so:
H+ (Nmol/m3)
so:
1.03(1.02,1.04)
1.03(1.02,1.04)
1.03(1.02,1.05)
1.05(1.01, 1.10)
1.16(1.03,1.30)*
1.12(1.00,1.24)
1.15(1.02,1.78)
R = 3. 1-8.2
R = 2.0-7.7
NR
28.8(NR/391)
7.6 (NR, 48.7)
1 8.6 (NR, 66.0)
Increased Respiratory Symptoms
Southern California6
Six Cities'"
(Cough)
Six Citiesf
(Lowei Rcsp. Sjymp.)
so:
PM2.5
PM2 5 Sulfur
H"
_ PMw^
rivi.25 oiiiiUT
1.48(1.14,1.91)
1.19(1.01,1.42)"
1.23(0.95,1.59)"
1.06(0.87,1.29)"
1.44(1.15-1.82)"
1 00 /"I ">O_-> C
-------
Table C-2. Effect estimates per variable increments in 24-h concentrations of fine
particle indicators (PM2 5, SO^, HT) from U.S. and Canadian studies (continued)
Reported PM
RR (± CI) per 25 fig/m3 Levels Mean
Acute Mortality Indicator PM Increase (Min/Max)t
Decreased Lung Function
Uniontown, PAg PM25 PEFR 23.1 (-0.3, 36.9) (per 25 25/88 (MR/88)
References:
"Schwartz et al. (1996a) tMin/Max 24-h PM indicator level shown in parentheses unless
bBumett et al. ( 1 994) otherwise noted as (± S.D.), 1 0 and 90 percentile ( 1 0,90)
TJumett et al. (1995) O3 or R = range of values from min-max, no mean value reported.
"Thurston et al. ( 1 992, 1 994) 'Change per 1 00 nmoles/m3
eOstro et al (1993) "Change per 20 ug/m3 for PM25; per 5 ug/m3 for
fSchwartz et al. (1994) PM2 5 sulfur; per 25 nmoles/m3 for H+.
et al. (1995) '"50th percentile value (10,90 percentile)
7/25/00 C-24 DRAFT—DO NOT CITE OR QUOTE
-------
Table C-3. Effect estimates per increments8 in annual average levels of fine particle
indicators from U.S. and Canadian studies
Type of Health
Effect & Location
Increased total chronic
Six Cityb
ACS Studyc
(151 U.S. SMSA)
Increased bronchitis in
Six City"
Six Citye
24 Cityf
24 Cityf
24 Cityf
24 Cityf
Southern California8
Indicator
mortality in adults
PM,s,,o
PM25
so:
PM2.5
so:
children
PM.5,,0
TSP
IT
so:
PM2.,
PM10
so:
Change in Health Indicator
per Increment in PM"
Relative Risk (95% CI)
1.42(1.16-2.01)
1.31 (1.11-1.68)
1.46(1.16-2.16)
1.17(1.09-1.26)
1.10(1.06-1.16)
Odds Ratio (95% CI)
3.26(1.13, 10.28)
2.80(1.17,7.03)
2.65(1.22,5.74)
3.02(1.28,7.03)
1.97(0.85,4.51)
3.29(0.81,13.62)
1.39(0.99, 1.92)
Range of City
PM Levels
Means Gig/m3)
18-47
11-30
5-13
9-34'
4-24
20-59
39-114
6.2-41.0
18.1-67.3
9.1-17.3
22.0-28.6
—
Decreased lung function in children
Six City4h
Six City'
24 City'J
24 City'
24 City1
24 City1
PM15/,o
TSP
H+ (52 nmoles/m3)
PM21 (15 ug/m3)
SO: (7 ug/m3)
PMIO(17ug/m3)
NS Changes
NS Changes
-3.45% (-4.87, -2.01) FVC
-3.21% (-4.98, -1.41) FVC
-3.06% (-4.50, -1.60) FVC
-2.42% (-4.30, -.0.511FVC
20-59
39-114
—
—
—
—
"Estimates calculated annual-average PM increments assume: a 100 ug/m3 increase for TSP; a 50 ug/m3
increase for PM10 and PM1S; a 25 ug/m3 increase for PM25; and a 15 ug/m3 increase for SO* except where
noted otherwise: a 100 nmole/m3 increase for H+.
"Dockery et al. (1993) "Abbey et al. (1995a,b,c)
Tope et al. (1995) hNS Changes = No significant changes.
"Dockery et al. (1989) 'Raizenne et al. (1996)
"Ware et al. (1986) JPollutant data same as for Dockery et al. (1996)
fDockery et al. (1996)
* Range of annual median values for subset of 50 cities.
7/25/00
C-25
DRAFT—DO NOT CITE OR QUOTE
-------
[able C-4. Prospective cohort mortality studies
o
o
ON
NJ
ON
Source
Abbey
etal.
(1991)
Docker
y ct al.
(1993)
Pope et
al.
(1995)
Health
Outcome Population
Total mortality Calif. 7th
from disease Day
Adventist
Total mortality White adult
volunteers
in 6 U.S.
cities0
Total mortality American
Cancer
Society,
adult
volunteers
in U.S.
Time Period/ PM
No. Units Indicators
1977-82 24 h
Defined by air TSP >200
monitoring sites
1974-91 PMIS
PM2S
SO,
1982-89 PM2S
PM2 s 50 cities
SO, 151 cities
SO,
PM PM Sites
Mean Range/ Per
(ug/m3) (Std. Dev.) City
102 25-175 NA
(annual avg)
29.9 18-47 1
18 11-30
7.6 5-13
18.2 9-34 1
llc 4-24 1
Total Model
Deaths Type
845 Cox
proportional
hazards
1429 Cox
proportional
hazards
20,765 Cox
proportional
hazard
38,963
PM Lag Other Other Factors
Structure Pollutants
lOyrs none age, sex, race,
smoking,
education,
airway disease
none none age, sex,
smoking,
education,
body mass,
occup.
exposure
hypertension11,
diabetes11
none none age, sex, race,
smoking,
education,
body mass,
occup.
exposure,
alcohol
consumption,
passive
smoking,
climate
Relative
Risk* at
SO, = 15,
PMI5 = 50,
PM!5 = 25
0.99 TSP1
1.42PM15
1.31 PM2i
1.46 SO,
1.17PM2S
1.10 SO,
RR.
Confidence
Interval
(0.87-1.13)'
(1.16-2.01)
(1.11-1.68)
(1.16-2.16)
(1.09-1.26)
(1.06-1.16)
Elasticity
NSb
0.25
0.22
0.23
0.117
0.077
•n
I
O
•z
o
H
o
C
O
H
m
o
!«
o
H-H
m
'For l,000h/yr>200ug/mj.
bNS = non significant, confidence limits not shown.
"Portage, Wl; Topeka, KS; Watertown, MA; Harrisman-Kingston, TN; St. Louis, MO; Steubenville, OH.
dUsed in other regression analyses not shown in this table.
"Value may be affected by filter artifacts.
Source: PM CD (U.S. EPA, 1996a).
-------
Table C-5. Relative mortality risks in six U.S. cities
Adjusted Risks
Risk Factor
Residence
Portage
Topeka
Watertown
Harriman
St. Louis
Steubenville
Smoking Status
Current
Previous
No high school
education
Body mass index
of 4.5
PM2 , Data (p.g/m3)
11.0(1980-7)3b
12.5 (1980-8)
14.9 (1980-5)
20.8 (1980-7)
19.0(1980-6)
29.6 (1980-7)
Crude Risk All*
1.0° 1.0
0.90 1.01
1.16 1.07
1.16 1.17
1.48 1.14
1.51. 1.26
1.59
1.20
1.19
1.08
Men1
1.0
1.04
0.94
1.21
1.15
1.29
1.75
1.25
1.22
1.03
Women'
1.0
0.97
1.22
1.07
1.13
1.23
1.54
1.18
1.13
1.11
'Adjusted for age, smoking, education, and body mass.
bPeriod of PM:5air monitoring.
'Baseline annual crude death rate = 10.73 per thousand population.
Source: Dockery et al. (1993)
7/25/00
C-27
DRAFT—DO NOT CITE OR QUOTE
-------
Table C-6. Estimated relative risks of mortality in six U.S. cities associated with a
range of air pollutants
Species
PM15
PM2.5
so42-
TSP
TSP-PM.s
PM1S-PM25
PM2.5-SO4
PMu-SO,
SO2
NO2
1970 TSP
Regr. Coeff.
0.0085
0.0127
0.0297
0.0037
0.0042
0.0178
0.0255
0.0121
0.0093
0.0126
0.0014
Standard
Error
(0.0026)
(0.0034)
(0.0081)
(0.0014)
(0.0032)
(0.0098)
(0.0029)
(0.0034)
(0.0032)
(0.0046)
(0.00044)
Pollutant
Range
28.3
18.6
8.5
55.8
27.5
9.7
8.4
18.1
19.8
15.8
154.0
Rel. Risk
1.27
1.27
1.29
1.22
1.12
1.19
1.24
1.24
1.20
1.22
1.25
95% CIs (n=6)
(1.04-1.56)
(1.06-1.51)
(1.06-1.56)
(0.99-1.53)
(0.88-1.43)
(0.91-1.55)
(1.16-1.32)
(1.05-1.48)
(1.01-1.43)
(1.00-1.49)
(1.03-1.50)
Source: U.S. EPA (1996a) recalculations based on results of Dockery et al. (1993).
7/25/00 C-28 DRAFT—DO NOT CITE OR QUOTE
-------
Table C-7. Relative risk of mortality from contributing nonmalignant respiratory
causes, for 30 days per year with PM10 > 100 ug/m3
PM Covariate Model
BASE (age, sex)
BASE + pack-years
BASE + pack-years + body-mass-index cats.
BASE + pack-years + body-mass-index cats.+ exercise cats.
STANDARD (age, pack-y., y. lived with smoker, occup., educ., BMI)
STANDARD w. PM10 (100) over last 4 years only
STANDARD, subset for former smokers
STANDARD, subset for never smokers
STANDARD, subset for low anti-oxidant vitamin intake
STANDARD, subset for high anti-oxidant vitamin intake
STANDARD, subset for < 4 h/wk outdoors
STANDARD, subset for 4-16 h/wk outdoors
STANDARD, subset for 16+ h/wk outdoors
STANDARD, subset for reported respiratory symptoms
RR
1.069
1.096
1.122
1.122
1.122
1.102
1.155
1.116
1.175
1.055
1.048
1.122
1.207
1.321
LCL
0.978
1.000
1.022
1.017
1.017
1.001
0.937
0.999
1.008
0.917
0:896
0.928
1.015
1.079
UCL
1.168
1.201
1.233
1.239
1.239
1.214
1.424
1.246
1.370
1.214
1.227
1.358
1.436
1.616
LCL = Lower 95% Confidence Limit.
UCL = Upper 95% Confidence Limit.
Source: Abbey et al. (1999).
7/25/00 C-29 DRAFT—DO NOT CITE OR QUOTE
-------
Table C-8. Relative risk of mortality from contributing nonmalignant respiratory
causes, by sex and air pollutant, with alternative covariate model
Pollution Index
PM10>100, d/yr
PM,0 mean
SO4 mean
O3>100ppb, h/yr
Pollution Incr.
30 days/yr
50 ^g/m3
15 ng/m3
551 h/yr(IQR)
RR
1.069
1.219
1.105
1.01
Females
LCL
0.936
0.739
0.396
0.77
UCL
1.220
2.011
3.086
1.33
RR
1.188
1.537
1.219
1.20
Males
LCL
1.030
0.879
0.411
0.88
UCL
1.370
2.688
3.619
1.64
LCL = Lower 95% Confidence Limit UCL = Upper 95% Confidence Limit.
Source: Abbey et al. (1999).
7/25/00 C-30 DRAFT—DO NOT CITE OR QUOTE
-------
Table C-9. Relative risk of mortality from all nonexternal causes, by sex and air
pollutant, for an alternative covariate model
Pollution Index
PMIO>100, d/yr
PM10 mean
SO4 mean
O3>100ppb,h/yr
SO2 mean
Pollution Incr.
30 days/yr
50 ng/m3
15 ng/m3
551h/yr(IQR)
3.72 (IQR)
RR
0.958
0.879
0.732
0.90
1.00
Females
LCL
0.899
0.713
0.484
0.80
0.91
UCL
1.021
1.085
1.105
1.02
1.10
RR
1.082
1.242
1.279
1.140
1.05
Males
LCL
1.008
0.955
0.774
0.98
0.94
UCL
1.162
1.616
2.116
1.32
1.18
LCL = Lower 95% Confidence Limit UCL = Upper 95% Confidence Limit.
Source: Abbey et al. (1999).
7/25/00
C-31
DRAFT—DO NOT CITE OR QUOTE
-------
Table C-10. Relative risk of mortality from cardiopulmonary causes, by sex and air
pollutant, for an alternative covariate model
Pollution Index
PMIO>100, d/yr
PM,0 mean .
SO4 mean
O3>100ppb, h/yr
O3 mean
SO2 mean
Pollution Incr.
30 days/yr
50 ng/m3
15 ng/m3
551h/yr(IQR)
lOppb
3.72 (IQR)
RR
0.929
0.841
0.857
0.88
0.975
1.02
Females
LCL
0.857
0.639
0.498
0.76
0.865
0.90
UCL
1.007
1.107
1.475
1.02
1.099
1.15
RR
1.062
1.219
1.279
1.06
1.066
1.01
Males
LCL
0.971
0.862
0.002
0.87
0.920
0.86
UCL
1.162
1.616
1018
1.29
1.236
1.18
LCL = Lower 95% Confidence Limit UCL = Upper 95% Confidence Limit.
Source: Abbey et al. (1999).
7/25/00
C-32
DRAFT—DO NOT CITE OR QUOTE
-------
Table C-ll. Relative risk of mortality from lung cancer, by sex and air pollutant, for
an alternative covariate model
Females Males
Pollution Index Pollution Incr. Smoking Category RR LCL UCL RR LCL UCL
PM10>100,d/yr 30 days/yr AH" 1.055 0.657 1.695 1.831 1.281 2.617
PM10mean 50 ^m3 All 1.808 0.343 9.519 12.385 2.552 60.107
NO2mean 19.78 (IQR) All 2.81 1.15 6.89 1.82 0.93 3.57
O3>100ppb,h/yr 551h/yr(IQR) All 1.39 0.53 3.67 4.19 1.81 9.69
never smoker 6.94 1.12 43.08
past smoker 4.25 1.50 12.07
O3mean lOppb All 0.805 0.436 1.486 1.853 0.994 3.453
SO2mean 3.72 (IQR) All 3.01 1.88 4.84 1.99 1.24 3.20
never smokers 2.99 1.66 5.40
"All = both never smokers and past smokers.
LCL = Lower 95% Confidence Limit UCL = Upper 95% Confidence Limit.
Source: Abbey etal. (1999).
7/25/00 C-33 DRAFT—DO NOT CITE OR QUOTE
-------
Table C-12. Relative risk of lung cancer incidence in males, by air pollutant, for
Adventist health study
Pollution Index
PM10>40 ug/m3
PM10>50 ug/m3
PM10>60 ug/m3
PM10>80 ug/m3
PM10>100 ug/m3
PM)0 mean
SO2 mean
NO2 mean
O3>60 ppb
O3>80 ppb
O3>100ppb
O3>120 ppb
O3>150ppb
O3 mean
PM10> 100 ug/m3
O3>100ppb
O3>100ppb
PM10> 100 ug/m3
O3>100ppb
SO2 mean
PM10 mean
SO2 mean
Pollution Incr.
139d/y(IQR)
149 d/y (IQR)
132d/y(IQR)
78 d/y (IQR)
30 d/y
50 ug/m3
3. 7 ppb
2.0 ppb
935 h/y
756 h/y
556 h/y
367 h/y
185 h/y
2.1 ppb
30 d/y
556 h/y
556 h/y
30 d/y
556 h/y
3.7 ppb
50 ug/m3
3.7 ppb
Covariate Model or Sub-Group
standard
standard
standard
standard
standard
standard
standard
standard
standard
standard
standard
standard
standard
standard
never smokers
never smokers
past smokers
high population density
high population density
high population density
> 80% data from monitors within
20 miles of residence
> 80% data from monitors within
20 miles of residence
RR
4.50
4.96
4.72
3.43
2.127
31.147
2.66
1.45
2.14
2.96
3.56
3.75
3.61
2.23
2.102
4.48
2.15
2.865
10.18
3.22
9.256
2.18
LCL
1.31
1.54
1.69
1.71
1.454
3.978
1.62
0.67
0.82
1.09
1.35
1.55
1.78
0.79
1.325
1.25
0.42
1.794
2.44 ,
1.87
1.135
0.92
•UCL
15.44
16.00
13.18
6.88
3.112
243.85
4.39
3.14
5.62
8.04
9.42
9.90
7.35
6.34
3.335
16.04
10.89
4.574
. 42.45
5.54
75.516
5.20
LCL = Lower 95% Confidence Limit.
UCL = Upper 95% Confidence Limit.
Source: Beeson et al. (1998).
7/25/00
C-34
DRAFT—DO NOT CITE OR QUOTE
-------
Table C-13. Relative risk of lung cancer incidence in females, by air pollutant, for
Adventist health study
Pollution Index
PM10>50 ug/m3
PM10>60 ng/m3
SO2 mean
O3>100 ppb
PMIO>100 ng/m3
SO2 mean
PM10 mean
SO2 mean
Pollution Incr.
149 d/y (IQR)
132 d/y (IQR)
3.7 ppb
556 h/y
30 d/y
3.7 ppb
50 ng/m3
3.7 ppb
Covariate Model or Sub-Group
standard
standard
standard
standard
high population density
high population density
> 80% data from monitors
within 20 miles
> 80% data from monitors
within 20 miles
RR
1.21
1.25
2.14
0.94
1.089
2.11
2.425
2.52
LCL
0.55
0.57
1.36
0.41
0.726
1.32
0.310
1.19
UCL
2.66
2.71
3.37
2.16
1.633
3.38
19.004
5.33
LCL = Lower 95% Confidence Limit.
UCL = Upper 95% Confidence Limit.
Source: Beeson et al. (1998).
7/25/00
C-35
DRAFT—DO NOT CITE OR QUOTE
-------
Table C-14. Relative risk (RR) of total mortality in three prospective cohort
studies, by sex and smoking status
Sex Smoking Status Study
F NON-SMOKER Six Cities
ACS
AHSMOG
PAST Six Cities
PAST + CURRENT ACS
CURRENT Six Cities
M NON-SMOKER Six Cities
ACS
AHSMOG
PAST Six Cities
PAST + CURRENT ACS
CURRENT Six Cities
PM
Index
PM.o
PM25
SO4
PMIO
PM10
PM2.5
SO4
PMIO
PM,o
PM2.5
SO4
PM,o
PM,o
PM2.5
S04
PM,0
PM
Inc.
50
25
15
50
50
25
15
50
50
25
15
50
50
25
15
50
RR
1.280
1.215
1.147
0.879
1.999
1.102
1.104
1.442
1.568
1.245
1.104
1.242
1.611
1.164
1.104
1.858
LCL
0.704
1.020
1.045
0.713
0.704
0.898
0.977
0.719
0.674
1.000
0.977
0.955
0.930
1.051
1.037
1.090
UCL
2.345
1.440
1.261
1.085
5.632
1.338
1.240
3.166
3.678
1.554
1.247
1.616
2.825
1.297
1.176
3.166
LCL = Lower 95% Confidence Limit. UCL = Upper 95% Confidence Limit.
Sources: Dockery et al. (1993); Pope et al. (1995); Abbey et al. (1999).
7/25/00 C-36 DRAFT—DO NOT CITE OR QUOTE
-------
Table C-15. Relative risk (RR) of cardiopulmonary mortality in three prospective
cohort studies, by sex and smoking status
Sex Smoking Status Study
F NON-SMOKERS ACS
AHSMOG
AHSMOG - CRC
PAST + CURRENT ACS
M NON-SMOKERS ACS
AHSMOG
AHSMOG - CRC
PAST + CURRENT ACS
F+M ALL Six Cities
PM
Index
PM2.5
SO4
PM10
PM.o
PM2.5
SO4
PM2.5
SO«
PM10
PMIO
PM25
SO4
PM,n
PMlnc.
25
15
50
50
25
15
25
15
50
50
25
15
50
RR
1.585
1.316
0.841
1.219
1.276
1.219
1.245
1.205
1.219
1.537
1.235
1.126
1.744
LCL
1.235
1.147
0.639
0.739
0.918
1.008
0.929
1.023
0.862
0.879
1.061
1.037
1.202
UCL
2.039
1.518
1.107
2.011
1.760
1.465
1.668
1.412
1.616
2.688
1.440
1.233
2.501
LCL = Lower 95% Confidence Limit. UCL = Upper 95% Confidence Limit.
Sources: Dockery et al. (1993); Pope et al. (1995); Abbey et al. (1999).
7/25/00
C-37
DRAFT—DO NOT CITE OR QUOTE
-------
Table C-16. Relative risk (RR) of lung cancer mortality in three prospective cohort
studies, by sex and smoking status
Sex Smoking Status Study
F NON-SMOKERS ACS
AHSMOG
PAST + CURRENT ACS
M NON-SMOKERS ACS
AHSMOG
PAST + CURRENT ACS
F+M ALL Six Cities
ACS
PM Index
PM2.5
SO4
PM,0
PM2.5
SO4
PM2.5
SO4
PMIO
PM2.5
SO4
PM,o
PM2.S
S04
PMInc.
25
15
50
25
15
25
15
50
25
15
50
25
15
RR
0.644
1.432
1.808
0.949
1.074
0.483
1.261
12.385
1.123
1.316
1.744
1.031
1.261
LCL
0.203
0.731
0.343
0.563
0.781
0.086
0.501
2.552
0.827
1.104
0.689
0.796
1.082
UCL
2.091
2.800
9.519
1.595
1.479
2.714
3.190
60.107
1.533
1.577
4.390
1.338
1.465
LCL = Lower 95% Confidence Limit UCL = Upper 95% Confidence Limit.
Sources: Dockery et a). (1993); Pope et al. (1995); Abbey et al. (1999).
7/25/00 C-38 DRAFT—DO NOT CITE OR QUOTE
-------
Table C-17. Comparison of estimated relative risks (RR) for all-cause mortality in six
U.S. cities associated with the reported inter-city range of concentrations of various PM
metrics
PM Species
SO4=
PM2 5 - SO4=
PM2.5
PM15.2.5
TSP-PM,,
Concentration
Range
(fig/m3)
8.5
8.4
18.6
9.7
27.5
Relative Risk
Estimate
1.29
1.24
1.27
1.19
1.12
RR
95% CI
(1.06-1.56)
(1.16-1.32)
(1.06-1.51)
(0.91-1.55)
(0.88-1.43)
Relative Risk
t-Statistic
3.67
8.79
3.73
1.81
1.31
Source: Dockery et al. (1993); U.S. Environmental Protection Agency (1996).
7/25/00 C-39 DRAFT—DO NOT CITE OR QUOTE
-------
Table C-18. Comparison of reported SO4= and PM25 relative risks (RR) for
various mortality causes in the ACS study
Mortality Cause
All Cause
Cardiopulmonary
Lung Cancer
scv
(Range = 19.9 jig/m3)
Relative
Risk
1.15
1.26
1.35
RR
95% CI
(1.09-1.22)
(1.15-1.37)
(1.11-1.66)
RR
t-Statistic
4.85
5.18
2.92
PM2.S
(Range = 24.5 ug/m3)
Relative
Risk
1.17
1.31
1.03
RR
95% CI
(1.09-1.26)
(1.17-1.46)
(0.80-1.33)
RR
t-Statistic
4.24
4.79
0.38
Source: Pope etal. (1995).
7/25/00 C-40 DRAFT—DO NOT CITE OR QUOTE
-------
Table C-19. Comparison of total mortality relative risk (RR) estimates and T-statistics
for PM components in three prospective cohort studies
PM Index
PM10 (50 ng/m3)
PM2.S (25 ug/m3)
S04= (15 ng/m3)
Days/y with PM10>100 (30
days)
PM10.2.5 (25 ug/m3
Study
Six Cities
AHSMOG
Six Cities
ACS (50 cities)
Six Cities
ACS (151 cities)
AHSMOG
AHSMOG
Six Cities
Subgroup
All
Male Nonsmoker
Male Nonsmoker
All
Male Nonsmoker
All
Male Nonsmoker
All
Male Nonsmoker
All
Male Nonsmoker
Male Nonsmoker
Male Nonsmoker
All
Male Nonsmoker
Relative Risk
1.504"; 1.530"
1.280'
1.242
1.364"; 1.379b
1.207"
1.174
1.245
1.504"; 1.567b
1.359
1.111
1.104
1.279
1.082
1.814"; 1.560b
1.434"
t Statistic
2.94';
3.27"
0.8 r
1.616
2.94";
3.73"
0.81'
4.35
1.960
2.94";
3.67"
0.81"
5.107
1.586
0.960
2.183
2.94"-c;
1.816b
0.81"
"Method 1 compares Portage vs. Steubenville (Table 3, Dockery et al., 1993).
bMethod 2 is based on ecologic regression models (Table 12-18, U.S. Environmental Protection Agency, 1996).
cMethod 1 not recommended for PM10-2.5 analysis due to high concentration in Topeka.
7/25/00
C-41
DRAFT—DO NOT CITE OR QUOTE
-------
Table C-20. Comparison of cardiopulmonary mortality relative risk (RR)
estimates and T-statistics for PM components in three prospective cohort studies
("Male Non. - CRC" identifies subjects who died of any contributing nonmalignant
respiratory cause in the AHSMOG study)
PM Index
PM,0(50ug/m3)
PM2.5 (25 ug/m3)
S04=(15ug/m3)
Days/y with
PMIO>100(30days)
PM,0.,,(25ug/m3
Study
Six Cities
AHSMOG
Six Cities
ACS (50 cities)
Six Cities
ACS (151 cities)
AHSMOG
AHSMOG
Six Cities
Subgroup
All
Male Nonsmoker
Male Non. - CRC
All
All
Male
Male Nonsmoker
All
All
Male
Male Nonsmoker
Male Nonsmoker
Male Non. - CRC
Male Nonsmoker
Male Non. - CRC
All
Relative Risk
1.744"
1.219
1.537
1.527"
1.317
1.245
1.245
1.743s
1.190
1.147
1.205
1.279
1.219
1.082
1.188
2.251"
t Statistic
2.94a
1.120
2.369
2.94"
4.699
3.061
1.466
2.94"
5.470
3.412
2.233
0.072
0.357
1.310
2.370
2.94a-b
'Method 1 compares Portage vs. Steubenville (Table 3, Dockery et al., 1993).
""Method 1 not recommended for PM 10-2.5 analysis due to high concentration in Topeka.
7/25/00
C-42
DRAFT—DO NOT CITE OR QUOTE
-------
2.5
2.0-
O)
o
U)
0)
Q.
1.5-
0.5
Legend
x Adults
a All Children
• Symptomatic and/or
Asthmatic Children
_£ Range represents 95%
confidence level.
•I.x. *•
JT T 3f * T
* i * * * x *-
III
.
J I
[llL
Adults
II
c
1
r
i
(C.I. = 3.04)
I
li I il I
i -p — .
0 i ,
f 1
Children
12345 12345 12345 678 910 11128 1314 151617 15161718 15161718
Total Respiratory Cardiovascular Respiratory COPDorlHD Cough Lower Upper
Mortality Mortality Mortality Hospital Hospital Respiratory Respiratory
Admissions Admissions Symptoms Symptoms
1 Pope etal. (1992) Utah Valley, Of
2 Schwartz (1993) Birmingham. AL
3 Styeretal. (1995)Chicago, IL
4 Ostro et al. (1996) Santago, Chile
5 Ito and Thurston (1996) Chicago, IL
6 Schwartz (1995) New Haven, CT
7 Schwartz (1995) Tacoma, WA 13
3 Schwartz (1996) Spokane, WA 14
9 Thurston etal. (1994) Toronto, Canada 15
10 Schwartz et al. (1996b) Cleveland, OH 16
11 Schwartz (1994f) Minneapolis, MN 17
12 Schwartz (1994c) Birmingham, AL 18
Schwartz (1994d) Detroit Ml
Schwartz and Morris (1995) Detroit, Ml
Hoek and Brunekreef (1993) The Netherlands
Schwartz et al. (1994) Six Cities
Pope and Dockery (1992) Utah Valley. UT
Pope et al. (1991) Utah Valley. VT
Figure C-l. Relative risk (RR) estimates for increased mortality and morbidity endpoints
associated with 50 fig/m* increments in PMIO concentrations as derived from
studies cited by numbers listed above each given type of health endpoint.
Note the consistency of RR elevations across studies for given endpoint and
coherence of RR estimates across endpoints, e.g., higher RR values for
symptoms versus hospital admissions and cause-specific mortality.
Source: PM Staff Paper (1996b). See U.S. EPA (1996b) for full reference citations for each study identified in
figure.
7/25/00
C-43
DRAFT—DO NOT CITE OR QUOTE
-------
Relative Risk for 50 ug/m3
in Six City Acute Study
o
Topeka
Portage
Stubenville
St. Louis
Harriman
Boston
.
•
i 1*1 i
1 — 0—1
i f^ \
0.9 1 1
Relative Risk
Relative Risk for 25 ug/m3 Fine Particles
(PM2S) in Six City Acute Study
Topeka
Portage
Stubenville
5 St. Louis
Harriman
Boston
L
\
f> I
— 0—1
h-CH
01,. .1
h-OH
1
Relative Risk for 25 pg/m3Coarse Particles
(PM1S-PM2S) in Six City Acute Study
Topeka
Portage
Stubenville
.$•
" St. Louis
Hamman
Boston
i ^
i
1—
1—
_j
i .».
5 — 1
^A. ,_„_.,!
5 — 1
0.9
1
Relative Risk
1.1
0.9
1
Relative Risk
1.1
Figure C-2. Relative risks of acute mortality in Harvard Six Cities Study, for inhalable
thoracic particles (PM1S/PM10), fine particles (PM2 5), and coarse fraction
particles (PM,S-PM2^). Note that the coarse fraction effects are smaller and
statistically non-significant (i.e., lower 95% confidence intervals do not exceed
relative risk of 1.0), except in Steubenville where there is high correlation
between fine and coarse particles (R2 = 0.69).
Source: PM CD (U.S. EPA, 1996a) graphical depiction of results from Schwartz et al. (1996).
7/25/00
C-44
DRAFT—DO NOT CITE OR QUOTE
-------
Total Particles ^
Ta
I'1
1.0
09
S
H
L
W
P *
30 40 SO 60 70 80 80 101
j^ Total Parades, ug/rrf
I
Total Particles , 2
Divided into Inhalable and i
B t 1
Non-lnhalable Particles |
1.0
0.0
S
H
L
W
P T
1.2
S
S 1 1
I
t.o
0.9
S
H
L
W
P T
15 20 25 30 35 40 45 SO
Inhalable Particles, pj/m'
10 20 30 40
Non-lnhalabte Particles, v
Fine Particles
Divided into
Sulfate and
Non-Sulfate
Particles
Inhalable Particles
Divided into Fine . 1-2
and Coarse Particles 1 1.1
EC
1.0
0.9
1
JL
***•
1.2
1.1
1.0
S
H
L
W
TP
1.3
1.2
£
I-
sc
1.0
0.8
S
H
L
W
PT
0 15 20 25 3
Fine Partides. pg/m'
S
H
L
W
P T
1.3
1.2
« 11
e
1.0
0.9
0 (
S
H
L
W
P T
t 10 12 14 16
Coarse Partides. pg/rn
4 6 3 10 12
Sulfate Paftides. pg/m'
57 9 11 13 15 17
Non-Suttate Rne Particles. ugW
Figure C-3. Adjusted relative risks for mortality are plotted against each of seven long-
term average particle indices in the Harvard Six City Study, from largest
range (total suspended particles, upper right) through sulfate and nonsulfate
fine particle concentrations (lower left). Note that a relatively strong linear
relationship is seen for fine particles, and for its sulfate and non-sulfate
components. Topeka, which has a substantial coarse particle component of
inhalable (thoracic) particle mass, stands apart from the linear relationship
between relative risk and inhalable particle concentration.
Source: U.S. EPA (1996a) replotting of results from Dockery et al. (1993).
7/25/00
C-45
DRAFT—DO NOT CITE OR QUOTE
-------
22 City Fine Mass vs. % Children <85% FVC
8-
7-
5-
U>
«
V
8
^ 3-
2-
1-
0
% FVC .85
Linear (% FVC .85)
I I
2 46 8 10 12 14 16 18 20 22 24
PM 2.1 (HO/m3)
22 City Coarse Fraction Mass vs. % Children <85% FVC
8-
7-
10
3-
2-
1-
0
* % FVC .85
— Linear (% FVC .85)
6 8
PM 10-2.1 (M9/m3)
10
12
14
Figure C-4. Percent of children with <85% normal FVC versus annual-average fine
(PM2,) particle concentrations and coarse fraction (PM10_2,) levels for
22 North American cities. Note much stronger relationship of fine particles
to lung function decrements (top panel) versus for coarse fraction particles
(bottom panel).
Source: PM Staff Paper (1996b) graphical depiction of results from Razienne et al. (1996).
7/25/00
C-46 DRAFT—DO NOT CITE OR QUOTE
-------
1 C.3. REFERENCES
2
3 Abbey, DE; Mills, PK; Petersen, FF; et al. (1991) Long-term ambient concentrations of total suspended participates
4 and oxidants as related to incidence of chronic disease in California Seventh-Day Adventists. Environ Health
5 Perspect 94:43-50.
6
7 Abbey, DE; Hwang, BL; Burchette, RJ; et al. (1995) Estimated long-term ambient concentrations of PM]0 and
8 development of respiratory symptoms in a nonsmoking population. Arch Environ Health 50:139-152.
9
10 Abbey, DE; Nishino, N; McDonnell, WF; et al. (1999) Long-term inhalable particles and other air pollutants related
11 to mortality in nonsmokers. Am J Respir Crit Care Med 159:373-382.
12
1 3 Beeson, WL; Abbey, DE; Knutsen, SF. (1998) Long-term concentrations of ambient air pollutants and incident lung
14 cancer in California adults: results from the AHSMOG study. Environ Health Perspect 106:813-823.
15
16 Dockery, DW; Pope, CA, III; Xu, X; et al. (1993) An association between air pollution and mortality in six U.S.
17 cities. N Engl J Med 329:1753-1759.
18
19 Federal Register. (1987) Revisions to the national ambient air quality standards for particulate matter. F R
20 52:24,634-24,669.
21
22 Federal Register. (1997) National ambient air quality standards for particulate matter; final rule. F R
23 62:38,652-38,752.
24
25 Lipfert, FW. (1978) The association of human mortality with air pollution: statistical analyses by region, by age, and
26 by cause of death. Mantua, NJ: Eureka Publications.
27
28 Lipfert, FW. (1984) Air pollution and mortality: specification searches using SMSA-based data. J Environ Econ
29 Manage 11:208-243.
30
31 Neas, LM; Dockery, DW; Koutrakis, .P; et al. (1995) The association of ambient air pollution with twice daily peak
32 expiratory flow rate measurements in children. Am J Epidemiol 141:111-122.
33
34 Ozkaynak, H; Thurston, GD. (1987) Associations between 1980 U.S. mortality rates and alternative measures of
35 airborne particle concentration. Risk Anal 7:449-461.
36
37 Pope, CA, III; Thun, MJ; Namboodiri, MM; et al. (1995) Particulate air pollution as a predictor of mortality in a
38 prospective study of U.S. adults. Am J Respir Crit Care Med 151:669-674.
39
40 Raizenne, M; Neas, LM; Damokosh, Al; et al. (1996) Health effects of acid aerosols on North American children:
41 pulmonary function. Environ Health Perspect 104:506-514.
42
43 Schwartz, J; Dockery, DW; Neas, LM. (1996) Is daily mortality associated specifically with fine particles? J Air
44 Waste Manas« Assoc 46:927-939.
45
46 Schwartz, J. (1996b) Air pollution and hospital admissions for respiratory disease. Epidemiology 7:20-28.
47
48 Thurston, GD; Ito, K; Hayes, CG; et a!. (1994) Respiratory hospital admissions and summertime haze air pollution in
49 Toronto, Ontario: consideration of the role of acid aerosols. Environ Res 65:271-290.
50
51 U.S. Environmental Protection Agency (U.S. EPA). (1996) Review of the national ambient air quality standards fnr
52 paiticuiitic marcer: policy assessment of scientific and technical information. OAQPS staff paper. Research Triangle
53 Park, NC: Office of Air Quality Plannino anH Standards; report r.c. EPA/452/R-9C-G13. Available from: NTIS,
54 Springfield, VA; PB97-115406REB.
7/25/00 C-47 DRAFT—DO NOT CITE OR QUOTE
-------
Appendix D
A SUMMARY REVIEW OF CANCER DOSE-RESPONSE
ANALYSES ON DIESEL EXHAUST
7/25/00 D-l DRAFT—DO NOT CITE OR QUOTE
-------
APPENDIX D
A SUMMARY REVIEW OF CANCER DOSE-RESPONSE
ANALYSES ON DIESEL EXHAUST
1 D.I. INTRODUCTION
2 Several individuals and organizations have previously conducted dose-response
3 assessments to estimate quantitatively the cancer risk from exposures to DE. Estimations were
4 performed on the basis of either epidemiologic and/or experimental data. As concluded in
5 Section 8.5, EPA finds that available epidemiologic data are too uncertain to confidently derive a
6 unit risk estimate for DE-induced lung cancer, and that rat data are not suitable for estimating
7 human risk. Nevertheless, a review of historical dose-response evaluations is provided here as
8 background information. This information is not intended to constitute endorsement or a
9 recommendation for use in quantitative risk assessment.
10 Early analyses to quantitatively assess the carcinogenicity of DE were hindered by a lack
11 of positive epidemiologic studies and long-term animal studies. One means of overcoming these
12 obstacles was the use of comparative potency methods based on combined epidemiologic and
13 experimental data. By the late 1980s, the availability of dose-response data from animal
^P bioassays and epidemiologic studies provided an opportunity for the derivation of both animal
15 and human data-based estimates, although considerable uncertainties were generally
16 acknowledged by the authors of these assessments.
17
18 D.2. COMPARATIVE POTENCY METHODS
19 In this method, the potency of diesel particulate matter (DPM) extract is compared with
20 other combustion or pyrolysis products for which epidemiology-based unit risk estimates have
21 been developed. Comparisons are made using short-term tests such as skin painting, mutations,
22 and mammalian cell transformation. The ratio of the potency of DPM extract to each of these
23 agents is then multiplied by their individual unit risk estimates to obtain the unit risk for DE. If
24 epidemiology-based estimates from more than one pollutant are used, the derived potencies are
25 generally averaged to obtain an overall mean. Major uncertainties of this method include the
26 assumptions that (1) the cancer potency of DE can be determined on the basis of the relative
27 effectiveness of the organic fraction alone; (2) the relative potency in short-term tests is an
28 accurate predictor of lung cancer potency; and (3) DPM extracts are similar in chemical
29 composition and proportion as combustion or pyrolysis products.
In the study by Albert et al. (1983), epidemiology-based unit cancer risk estimates for
coke oven emissions, cigarette smoke condensate, and roofing tar were used. Samples of DPM
32 were collected from three light-duty engines (a Nissan 220 C, an Oldsmobile 350, and a
7/25/00 D-2 DRAFT—DO NOT CITE OR QUOTE
-------
1 cancer potency estimates were then derived by multiplying the epidemiology-based cancer
2 potency estimates for both coke oven emissions and roofing tar by the ratio of their potencies
3 compared with DPM extract in each of the three bioassays. Harris (1983) derived an overall
4 mean relative risk value of 3.5 x 10'5per |Ig/m3 for the three"light-duty engines with a 95% upper
5 confidence limit of 2.5 x 10"4. Individual mean values for each engine were not reported.
6 McClellan (1986), Cuddihy et al. (1981, 1984), and Cuddihy and McClellan (1983)
7 estimated a risk of about 7.0 x 10~5 per |ig/m3 DPM using a comparative potency method similar
8 to those reported in the preceding paragraph. The database was similar to that used by Albert et
9 al. (1983) and Harris (1983).
10
11 D.3. EPIDEMIOLOGY-BASED ESTIMATION OF CANCER RISK
12 The first lung cancer risk estimates based on epidemiologic data were derived by Harris
13 (1983). He assessed the risk of exposure to DE using data from the London Transport Worker
14 Study reported by Waller (1981). Five groups of employees from the London Transport
15 Authority (LTA) were used: bus garage engineers, bus drivers, bus conductors, engineers in
16 central works, and motormen and guards. The first group was considered to have received the
17 highest exposure; the next two, intermediate; and the last two groups, none. When cancer death
'rt
18 rates for the high-exposure group were compared with those of London males, there was no
19 increase in the observed-to-expected (O/E) ratios. The author, in fact, considered the results to
20 be negative. However, because the low rate of lung cancer in all the LTA exposure groups may
21 have been the result of a "healthy worker" effect, Harris (1983) compared the exposed groups
22 with internal controls. He merged the three exposed groups and compared them with the two
23 groups considered to be unexposed. An adjustment was made for the estimated greater exposure
24 levels of garage engineers compared with bus drivers and conductors. Using this method, the
25 relative risk of the exposed groups was greater than 1 but was statistically significant only for
26 garage engineers exposed from 1950 to 1960. In that case, the O/E ratio was 29% greater than
27 the presumed unexposed controls.
28 Harris (1983) identified a variety of uncertainties relative tc potency assessment based on
29 this study. These included:
30 • small unobserved differences in smoking incidences among groups, which could have
31 a significant effect on lung cancer rates;
32 • uncertainty about the magnitude of exposure in the exposed groups;
33 • uncertainty regarding the extent of change in exposure conditions over time;
34 " random effects arising from the stochastic nature of the cancer incidence; and ^
35 • uncertainty in the mathematical specification of the model.
7/25/00 D-4 DRAFT—DO NOT CITE OR QUOTE
-------
D.4. ANIMAL BIO ASSAY-BASED CANCER POTENCY ESTIMATES
'2 With the availability of chronic cancer bioassays, a considerable number of potency
3 estimates were derived using lung tumor induction in rats. A high degree of uncertainty exists in
4 the use of the rat data to predict human risk. Major uncertainties include: (1) differences in -
5 particle deposition patterns between rats and humans, (2) differences in sensitivity between rats
6 and humans to the carcinogenic action of DE, and (3) extrapolation of rat lung tumor responses at
7 high concentrations to ambient concentrations without a clear understanding of the mode of
8 action of DE. It is now widely recognized that the rat lung tumor response associated with any
9 insoluble particles at high concentrations is mediated by a particle-overload mechanism (ILSI,
10 2000), suggesting that rat data for DE are not suitable for estimating human risk at low
11 environmental concentrations.
12 The first risk estimate was reported by Albert and Chen (1986), based on the chronic rat
13 bioassay conducted by Mauderly et al. (1987). Using a multistage model and assuming
14 equivalent deposition efficiency in humans and rats, they derived a 95% upper confidence limit
15 of 1.6 x 10'5 for lifetime risk of exposure to 1 \ig/m\. Pott and Heinrich (1987) also used a linear
16 model and data reported by Brightwell et al. (1989), Heinrich et al. (1986), and Mauderly et al.
17 (1987). They reported risk estimates ranging from 6,x 10'5 to 12 x 10'5 per [lg/m3. Smith and
^ Stayner (1990), using time-to-tumor models based- on the data of Mauderly et al. (1987), derived
19 point (MLE) estimates ranging from 1.0 x 10"4 to 2.1 x 10"4 per |J.g/m3 after converting from
20 occupational to environmental exposure scenario.
21 Pepelko and Chen (1993) developed unit risk estimates based on the data of Brightwell et
22 al. (1989), Ishinishi et al. (1986), and Mauderly et al. (1987) using a detailed dosimetry model to
23 extrapolate dose to humans and a linearized multistage (LMS) model. Taking the geometric
24 mean of individual estimates from the three bioassays, they derived unit risk estimates of 1.4 x
25 10"5 per |lg/m3 when dose was based on carbon particulate matter per unit lung surface area rather
26 than whole DPM, and 1.2 x 10"4 per |ig/m3 when based on lung burden per unit body weight.
27 Hattis and Silver (1994) derived a maximum likelihood estimate for occupational
28 exposure of 5.2 x 10'5 per |lg/m3 based on lung burden and bioassay data reported by Mauderly et
29 al. (1987) and use of a five-stage Armitage-Doll low-dose extrapolation model. California EPA
30 (CAL-EPA, 1998) derived a geometric mean estimate of 6 x 10"5per }lg/m3 from five bioassays
31 using an LMS model.
32 To demonstrate the possible influence of particle effects as well as particle-associated
33 organics, an additional modeling approach was conducted by Chen and Oberdorster (1996).
Employing a biologically based two-stage model and using malignant tumor data from Mauderly
et al. (1987), the upper-bound risk estimate for exposure to 1 |lg/m3 was estimated to be
7/25/00 _ D-6 DRAFT—DO NOT CITE OR QUOTE
-------
1 Garshick, E; Schenker, MB; Munoz, A; et al. (1988) A retrospective cohort study of lung cancer and DE exposure in
2 railroad workers. Am Rev Respir Dis 137:820-825.
3
4 Harris, JE. (1983) Diesel emissions and lung cancer. Risk Anal 3:83-100.
5
6 Hattis, D; Silver, K. (1994) Use of mechanistic data in occupational health risk assessment: the example of diesel
7 particulates. In: Chemical risk assessment and occupational health. Smith, MC; Christiani, DC; Kelsey, KT, eds.
8 Westport, CT: Auburn House, pp. 167-178.
9
10 Heinrich, U; Muhle, H; Takenaka, S; et al. (1986) Chronic effects on the respiratory tract of hamsters, mice and rats
11 after long-term inhalation of high concentration of filtered and unfiltered diesel engine emissions. J Appl Physiol
12 6:383-395.
13
14 Ishinishi, N; Kuwabara, N; Nagase, S; et al. (1986) Long-term inhalation studies on effects of exhaust from heavy
15 and light duty diesel engines on F344 rats, hi: Carcinogenic and mutagenic effects of diesel engine exhaust. Ishinishi,
16 N; Koizumi, A; McClellan, RO; et al., eds. Amsterdam: Elsevier, pp. 329-348.
17
18 Mauderly, JL; Jones, RK; Griffith, WC; et al. (1987) DE is a pulmonary carcinogen in rats exposed chronically by
19 inhalation. Fundam Appl Toxicol 9:208-221.
20
21 McClellan, RO. (1986) Health effects of DE: a case study in risk assessment. Am Ind Hyg Assoc J 47:1-13.
22
23 McClellan, RO; Cuddihy, RG; Griffith, WC; et al. (1989) Integrating diverse data sets to assess the risks of airborne
24 pollutants. In: Assessment of inhalation hazards. Mohr, U, ed. New York: Springer Verlag, pp. 3-22.
25
26 Pepelko, WE; Chen, C. (1993) Quantitative assessment of cancer risk from exposure to diesel engine emissions.
27 Regul Toxicol Pharmacol 17:52-65.
28
29 Pott, F; Heinrich, U. (1987) Dieselmotorabgas und Lungenk auf die GefShrdung des Menschen. In: Umwelthygiene,
30 vol. 19. Med. Institut f. Umwelthygiene, Annual Report 1986/87. Dusseldorf, F.R.G., pp. 130-167.
31
32 Smith, RA; Stayner, L. (1990) An exploratory assessment of the risk of lung cancer associated with exposure to DE
33 based on a study with rats. Final report. Division of Standards Development and Technology Transfer; Cincinnati,
34 OH: NIOSH.
35
36 Steenland, NK; Silverman, DT; Hornung, RW. (1990) Case-control study of lung cancer and truck driving in the
37 Teamsters Union. Am J Publ Health 80:670-674.
38
39 Steenland, K; Deddens, J; Stayner, L. (1998) DE and lung cancer in the trucking industry: exposure-response
40 analysis and risk assessment. Am J Ind Med 34:220-228.
41
42 Valberg, PA; Crough, EAC. ( ) Meta analysis of rat lung tumors from lifetime inhalation of diesel exhaust.
43 Environ Health Perspect (in press).
44
45 Waller, RE. (1981) Trends in lung cancer in London in relation to exposure to diesel fumes. Environ Int 5:479-483.
46
47 Woskie, SR; Smith, TJ; Hammond, SK; et al. (1988) Estimation of the DE exposures of railroad workers: II.
48 National and historical exposures. Am J Ind Med 13:395-404.
49
50 Zaebst, D; Clapp, D; Blade, L; et al. (1991) Quantitative determination of trucking industry workers' exposures to
51 diesel particles. Am Ind Hyg Assoc J 52:529-541.
7/25/00 D-8 DRAFT—DO NOT CITE OR QUOTE
------- |