United States Env i ronmental Protec tion Agenc y


# Health Assessment Document For Diesel Engine ExhaustEPA/600/8-90/057F May 2002



# Health Assessment Document for Diesel Engine Exhaust

National Center for Environm ental Assessm ent Office of Research and Developm ent U.S. Environm ental Protection Agency W a shington, DCii DISCLAIMER

This document has been reviewed in accordance with U.S. Environmental Protection Agency policy and approved for publication. Me ntion of trade names or commercial products
does not constitute endorsement or recommendation for use.

ABSTRACT

This assessment examined information regarding the possible health hazards associated with exposure to diesel engine exhaust (DE), wh ich is a mixture of gase s and particles. The assessment concludes that long-t erm (i.e., chronic) inhalation ex posure is likely to pose a lung cancer hazard to humans, as well as damage th e lung in other ways depending on exposure. Short-term (i.e., acute) exposures can cause irritation and inflammatory symptoms of a transient nature, these being highly variable across th e population. The assessmen t also indicates that evidence for exacerbation of existing allergies and asthma symptoms is emerging. The assessment recognizes that DE emissions, as a mixture of many constituents, also contribute to
ambient concentrations of several criteria air pollutants including nitrogen oxides and fine particles, as well as other air toxics. The assessment’s health hazard conclusions are based on exposure to exhaust from diesel engines built prior to the mid-1990s. The health hazard conclusions, in general, are app licable to engines currently in use, which include many older engines. As new diesel engines with cleaner exhaust emissions replace existing engines, the applicability of the conclusions in this Health Assessment Document will need to be reevaluated.

Preferred citation: U.S. Environmental Protection Agency (EPA). (2002) Health assessment document for diesel engine exhaust. Prepared by the National Center for Environmental Asse ssment, Washington, DC, for the Office of Transportation and Air Quality; EPA/600/8-90/057F. Available from: Natio nal Technical Information Service, Springfield, VA; PB2002-107661, and <http:// www.epa.gov/n cea>.CONTENTS

L I S T O F T A B L E S

........................................................... vi i i L I S T O F F I G U R E S

........................................................... xi F O R E W O R D

............................................................... xi v P R E F A C E

................................................................. xvi AUTHORS, CONTRIBUTORS, AND REVIEW ERS

.............................. xvi i ACKNOW LEDGMENTS

.................................................... xxi i

1. E X E C U T I V E S U M M A R Y

................................................. 1- 1 1.1. INTRODUCTION

..................................................... 1- 1 1.2. COMPOSITION OF DIESEL EXHAUST

.................................. 1- 1 1.3. DIESEL EXHAUST AS A COMPONENT OF AMBIENT PARTICULATE MATTER

............................................................ 1- 2 1.4. ATMOSPHERIC TRANSFORMATION OF DIESEL EXHAUST

............... 1- 2 1.5. EXPOSURE TO DIESEL EXHAUST

..................................... 1- 3 1.6. HEALTH EFFECTS OF DIESEL EXHAUST

............................... 1- 3 1.6.1. Acute (Short-Term Exposure) Effects

................................ 1- 4 1.6.2. Chronic (Long-Term Exposure) Noncancer Respiratory Effects

............ 1- 4 1.6.3. Chronic (Long-Term Exposure) Carcinogenic Effects

.................... 1- 4 1.7. SOURCES OF UNCERTAINTY

......................................... 1- 6

2. DIESEL EXHAUST EMISSIONS CHARACTERIZATION, ATMOSPHERIC TRANSFORMATION, AND EXPOSURES

.................................... 2- 1 2.1. INTRODUCTION

..................................................... 2- 1 2.2. PRIMARY DIESEL EXHAUST EMISSIONS

.............................. 2- 3 2.2.1. History of Dieselization

........................................... 2- 3 2.2.2. Diesel Com bustion and Form ation of Prim ary Em issions

................. 2- 9 2.2.3. Diesel Em ission Standards and Em ission Trends Inventory

.............. 2- 15 2.2.4. Historical Trends in Diesel Fuel Use and Im pact of Fuel P r ope r t i e s on E m i s s i ons

.......................................... 2- 25 2.2.5. Chronological Assessm ent of Em ission Factors

........................ 2- 29 2.2.6. Engine Technology Description and Chronology

...................... 2- 43 2.2.7. Air Toxic Em issions

............................................. 2- 53 2.2.8. Physical and Chem ical Com position of Diesel Exhaust Particles

.......... 2- 59 2.3. ATMOSPHERIC TRANSFORMATION OF DIESEL EXHAUST

.............. 2

- 8 4 2.3.1. Gas-Phase Diesel Exhaust

........................................ 2

- 8 4 2.3.2. Particle-Phase Diesel Exhaust

..................................... 2

- 9 0 2.3.3. Diesel Exhaust Aging

............................................ 2

- 9 3 2.4. AMBIENT DIESEL EXHAUST CONCENTRATIONS AND EXPOSURES

..... 2

- 9 4 2.4.1. Diesel Exhaust Gases in the A m bient At m o sphere

..................... 2

- 9 4 2.4.2. A m bient Concentrations of DPM

.................................. 2

- 9 5 2.4.3. Exposures to Diesel Exhaust

..................................... 2

- 10 6 2.5. SUMMARY AND DISCUSSION

...................................... 2

- 1 18 2.5.1. History of Diesel Engine Use, Standards, and Technology

.............. 2

- 1 19 2.5.2. Physical and Che m ical Co m position of Diesel Exhaust

................ 2

- 1 2 0

iiiCONTENTS (continued)

2.5.3. At m o spheric Transfor m a tion of Diesel Exhaust

...................... 2

- 12 3 2.5.4. A m bient Concentrations and Exposure to Diesel Exhaust

.............. 2

- 12 4 R E F E R E N C E S F O R C H A P T E R 2

............................................ 2

- 12 6

3. DOSIMETRY OF DIESEL PARTICULATE MATTER

........................... 3- 1 3.1. INTRODUCTION

..................................................... 3- 1 3.2. CHARACTERISTICS OF INHALED DIESEL PARTICULATE MATTER

....... 3- 2 3.3. REGIONAL DEPOSITION OF INHALED DIESEL PARTICULATE MATTER

. . . 3-2 3.3.1. Deposition Mechanism s

........................................... 3- 3 3.3.2. Particle Clearance and Translocation Mechanis m s

...................... 3

- 9 3.3.3. Translocations of Particles to Extra-Alveolar Macrophage Co m p art m ent S i t e s

.......................................................... 3

- 2 2 3.4. PARTICLE “OVERLOAD”

............................................ 3

- 2 6 3.4.1. Introduction

.................................................... 3

- 2 6 3.4.2. Relevance to Hum a ns

............................................ 3- 28 3.4.3. Potential Mechanism s for an AM Sequestration Com p artm ent f or P a r t i c l e s D u r i ng P a r t i c l e O v e r l o a d

............................... 3

- 3 0 3.5. BIOAVAILABILITY OF ORGANIC CONSTITUENTS PRESENT ON DIESEL EXHAUST PARTICLES

....................................... 3

- 3 1 3.5.1. In Vivo Studies

................................................. 3

- 3 2 3.5.2. In Vitro Studies

................................................. 3

- 3 4 3.5.3. Modeling Studies

............................................... 3

- 3 6 3.5.4. Su m m a ry and Bioavailability

...................................... 3

- 3 7 3.6. MODELING THE DEPOSITION AND CLEARANCE OF PARTICLES IN T H E R E S P I R A T O R Y T R A C T

.......................................... 3

- 38 3.6.1. Introduction

.................................................... 3

- 38 3.6.2. Dosi m e try Models for DPM

....................................... 3

- 38 3.7. SUMMARY AND DISCUSSION

........................................ 3

- 5 4 R E F E R E N C E S F O R C H A P T E R 3

............................................. 3

- 5 6

4. M U T A G E N I C I T Y

........................................................ 4- 1 4.1. GENE MUTATIONS

.................................................. 4- 2 4.2. CHROMOSOME EFFECTS

............................................. 4- 5 4.3. OTHER GENOTOXIC EFFECTS

........................................ 4- 7 4.4. SUMMARY AND DISCUSSION

......................................... 4- 8 R E F E R E N C E S F O R C H A P T E R 4

.............................................. 4- 9

5. NONCANCER HEALTH EFFECTS OF DIESEL EXHAUST

...................... 5- 1 5.1. HEALTH EFFECTS OF W HOLE DIESEL EXHAUST

....................... 5- 2 5.1.1. Hum a n Studies

.................................................. 5- 2 5.1.2. Traffic Studies

.................................................. 5- 23 5.1.3. Laboratory Anim al Studies

........................................ 5- 24 5.2. MODE OF ACTION OF DIESEL EXHAUST-INDUCED NONCANCER E F F E C T S

........................................................... 5

- 8 2 5.2.1. Co m p arison of Health Effects of Filtered and Unfiltered Diesel Exhaust

.... 5

- 8 2

ivCONTENTS (continued)

5.2.2. Mode of Action for the Noncarcinogenic Effects of DPM

................ 5- 8 9 5.3. INTERACTIVE EFFECTS OF DIESEL EXHAUST

......................... 5- 90 5.4. COMPARATIVE RESPONSIVENESS AMONG SPECIES TO THE HISTOPATHOLOGIC EFFECTS OF DIESEL EXHAUST

................... 5- 92 5.5. DOSE-RATE AND PARTICULATE CAUSATIVE ISSUES

.................. 5- 9 3 5.6. SUMMARY AND DISCUSSION

........................................ 5- 9 7 5.6.1. Effects of Diesel Exhaust on Hu m a ns

............................... 5- 9 7 5.6.2. Effects of Diesel Exhaust on Laboratory Ani m als

...................... 5- 9 9 5.6.3. Co m p arison of Filtered and Unfiltered Diesel Exhaust

.................. 5- 102 5.6.4. Interactive Effects of Diesel Exhaust

............................... 5- 10 3 5.6.5. Conclusions

................................................... 5- 10 3 R E F E R E N C E S F O R C H A P T E R 5

............................................ 5- 10 4

6. ESTIMATING HUMAN NONCANCER HEALTH RISKS OF DIESEL EXHAUST

. . . 6-1 6.1. INTRODUCTION

.................................................... 6- 1 6.2. THE INHALATION REFERENCE CONCENTRATION APPROACH

.......... 6- 3 6.3. CHRONIC REFERENCE CONCENTRATION FOR DIESEL EXHAUST

........ 6- 5 6.3.1. Principal Studies for Dose-Response Analysis: Chronic, Multiple-Dose Level Rat Studies

.................................... 6- 6 6.3.2. Derivation of Hum a n Continuous Equivalent Concentrations, HECs

........ 6- 9 6.3.3. Dose-Response Analysis—Choice of an Effect Level

.................. 6- 11 6.3.4. Uncertainty Factors (UF) for the RFC—A Co m posite Factor of 30

........ 6

- 1 4 6.3.5. Derivation of the RfC for Diesel Exhaust

............................ 6

- 1 6 6.4. EPIDEMIOLOGICAL EVIDENCE AND NAAQS FOR FINE PM

............. 6

- 1 7 6.4.1. Epide m iological Evidence for Fine PM

.............................. 6

- 1 8 6.4.2. NAAQS for Fine PM

............................................ 6

- 2 5 6.4.3. DPM as a Co m ponent of Fine PM

.................................. 6

- 3 0 6.5. CHARACTERIZATION OF THE NONCANCER ASSESSMENT FOR DIESEL EXHAUST

......................................................... 6

- 3 0 6.6. SUMMARY

........................................................ 6

- 3 2 R E F E R E N C E S F O R C H A P T E R 6

............................................. 6

- 3 3

7. CARCINOGENICITY OF DIESEL EXHAUST

................................. 7- 1 7.1. INTRODUCTION

.................................................... 7- 1 7.1.1. Overview

...................................................... 7- 1 7.1.2. Am bient PM-Lung Cancer Relationships

............................. 7- 1 7.2. EPIDEMIOLOGIC STUDIES OF THE CARCINOGENICITY OF EXPOSURE TO DIESEL EXHAUST

................................................ 7- 3 7.2.1. Cohort Studies

.................................................. 7- 6 7.2.2. Case-Control Studies of Lung Cancer

............................... 7- 32 7.2.3. Su m m a ries of Studies and Meta-Analyses of Lung Cancer

.............. 7- 6 1 7.2.4. Su m m a ry and Discussion

......................................... 7- 6 6

vCONTENTS (continued)

7.3. CARCINOGENICITY OF DIESEL EXHAUST IN LABORATORY A

N I M A L S

......................................................... 7

- 8 3 7.3.1. Inhalation Studies ( W hole Diesel Exhaust)

........................... 7

- 8 4 7.3.2. Inhalation Studies (Filtered Diesel Exhaust)

......................... 7

- 10 8 7.3.3. Inhalation Studies (DE Plus Cocarcinogens)

......................... 7

- 10 9 7.3.4. Lung I m plantation or Intratracheal Instillation Studies

................. 7

- 11 1 7.3.5. Subcutaneous and Intraperitoneal Injection Studies

................... 7

- 11 7 7.3.6. Der m al Studies

................................................ 7

- 11 9 7.3.7. Su m m a ry and Conclusions of Laborato r y Ani m al Carcinogenicity Studies

. 7 -1 21 7.4. MODE OF ACTION OF DIESEL EXHAUST-INDUCED CARCINOGENESIS

................................................ 7

- 12 8 7.4.1. Potential Role of Organic Exhaust Co m ponents in Lung Cancer Induction

. 7-12 9 7.4.2. Role of Inflam m a tory Cytokines and Proteolytic Enzym e s in the Induction of Lung Cancer in Rats by Diesel Exhaust

.................. 7

- 13 2 7.4.3. Role of Reactive Oxygen Species in Lung Cancer Induction by Diesel E xh a us t

..................................................... 7

- 13 3 7.4.4. Relationship of Physical Characteristics of Particles to Cancer Induction

. . 7 -13 6 7.4.5. Integrative Hypothesis for Diesel-Induced Lung Cancer

............... 7

- 13 7 7.4.6. Su m m a ry

.................................................... 7

- 1 39 7.5. W E IGHT-OF-EVIDENCE EVALUATION FOR POTENTIAL HUMAN CARCINOGENICITY

....................................... 7

- 1 40 7.5.1. Hu m a n Evidence

.............................................. 7

- 1 41 7.5.2. Ani m al Evidence

.............................................. 7

- 14 2 7.5.3. Other Key Data

............................................... 7

- 14 3 7.5.4. Mode of Action

............................................... 7

- 14 3 7.5.5. Characterization of Overall W e ight of Evidence: EPA’s 1986 Guidelines for Carcinogen Risk Assess m ent

.................................. 7

- 14 4 7.5.6. W e ight-of-Evidence Hazard Narrative: EPA’s Proposed Guidelines for Carcinogen Risk Assess m ent (1996b, 1999)

...................... 7

- 14 4 7.6. EVALUATIONS BY OTHER ORGANIZATIONS

........................ 7

- 14 6 7.7. CONCLUSION

..................................................... 7

- 14 7 R E F E R E N C E S F O R C H A P T E R 7

............................................ 7

- 14 8

8. DOSE-RESPONSE ASSESSMENT: CARCINOGENIC EFFECTS

................. 8- 1 8.1. INTRODUCTION

.................................................... 8- 1 8.2. MODE OF ACTION AND DOSE-RESPONSE APPROACH

.................. 8- 2 8.3. USE OF EPIDEMIOLOGIC STUDIES FOR QUANTITATIVE R I S K A S S E S S M E N T

.................................................. 8- 4 8.3.1. Sources of Uncertainty

............................................ 8- 4 8.3.2. Evaluation of Key Epidem iologic Studies for Potential Use in Quantitative Risk Estim ates

.......................... 8- 5 8.3.3. Conclusion

.................................................... 8- 11 8.4. PERSPECTIVES ON CANCER RISK

................................... 8- 11 8.5. SUMMARY AND DISCUSSION

....................................... 8- 16 R E F E R E N C E S F O R C H A P T E R 8

............................................. 8- 17

viCONTENTS (continued)

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

- 2 9.2.1. Diesel Exhaust Com ponents of Possible Health Concern

................. 9

- 2 9.2.2. “Fresh” Versus “Aged” Diesel Exhaust

............................... 9

- 4 9.2.3. Changes of Diesel Exhaust Em issions and Com position Over Tim e

........ 9

- 5 9.3. AMBIENT CONCENTRATIONS AND EXPOSURE TO DIESEL EXHAUST

.... 9

- 6 9.4. HAZARD CHARACTERIZATION

...................................... 9

- 8 9.4.1. Acute and Short-Term Exposures

................................... 9

- 8 9.4.2. Chronic Exposure

............................................... 9

- 1 0 9.5. DOSE-RESPONSE ASSESSMENT

..................................... 9

- 1 6 9.5.1. Evaluation of Risk for Noncancer Health Effects

...................... 9

- 1 7 9.5.2. Evaluation of Cancer Risks

....................................... 9

- 2 0 9.6. SUMMARY AND CONCLUSIONS

..................................... 9

- 2 4 R E F E R E N C E S F O R C H A P T E R 9

............................................. 9

- 2 7

APPENDIX A: CALCULATION OF HUMAN EQUIVALENT CONTINUOUS E X P O S U R E C O N C E N T R A T I O N S ( H E C s )

....................... A

- 1

APPENDIX B: BENCHMARK CONCENTRATION ANALYSIS OF D I E S E L D A T A ...............................................B

- 1

APPENDIX C: A SUMMARY REVIEW OF CANCER DOSE-RESPONSE ANALYSES ON DIESEL EXHAUST .............................C

- 1

viiLIST OF TABLES

2- 1. V e hi c l e c l a s s i f i c a t i on a nd w e i ght s f or on- r oa d t r uc ks

......................... 2- 5 2-2. Total (gas and diesel) diesel trucks in the fleet in 1992

........................ 2- 5 2- 3. T ypi c a l c he m i c a l c om pos i t i on of f i ne pa r t i c ul a t e m a t t e r

...................... 2- 14 2- 4. U .S

. e m i s s i on s t a nda r ds : H D hi ghw a y di e s e l e ngi ne s

........................ 2- 16 2-5. U.S. em ission standards: locom o tives (g/bhp-hr)

........................... 2- 17 2-6. U.S. em ission standards for nonroad diesel equipm ent (g/bhp-hr)

............... 2- 18 2-7. Com p arison of in-use truck fleet with truck fleet tested on chassis dynam om e ter, pe r c e nt of t ot a l ve hi c l e s

............................................... 2- 31 2-8. Diesel engine e m issions data from engine dyna m o m e ter tests

.................. 2

- 3 4 2-9. HD diesel em issions results from tunnel tests

.............................. 2- 40 2-10. Rem o te sensing results for HD vehicles

................................... 2- 41 2-11. Sum m a ry of CDD/CDF em issi ons from diesel-fueled vehicles

................. 2- 57 2-12. Baltim ore Harbor Tunnel Study: estim ated CDD/CDF em ission factors for HD

ve hi c l e s

............................................................ 2- 60 2-13. Organic and ele m ental carbon fractions of diesel and gasoline engine PM exhaust

. 2-7 0 2-14. E m ission rates of PAH ( m g/ m i ) from LD and HD diesel vehicles

............... 2

- 7 3 2-15. Polycyclic arom atic hydrocarbons identified in extracts of diesel particles from L D d i e s e l e ng i ne e xh a us t

.............................................. 2

- 7 4 2-16. Em ission rates of particle-bound PAH ( 

### : g/ m i ) from diesel and gasoline engines

. . 2 -7 6 2-17. Concentrations of nitro-PAHs identified in LD diesel particulate extracts

........ 2- 7 7 2-18. Average em ission rates for polycyclic arom atic hydrocarbons for different fuel
t ype s

.............................................................. 2- 8 1 2-19. Classes of co m pounds in diesel exhaust

................................... 2- 8 6 2-20. Calculated atm o spheric lifetim es for gas-phase reactions of selected com pounds
pr e s e nt i n a ut om ot i ve e m i s s i ons w i t h i m por t a nt r e a c t i ve s pe c i e s

............... 2- 8 7 2-21. Major com ponents of gas-phase diesel engine em issions, their known atm o spheric transfor m a tion products, and the biological i m pact of the reactants and products

. . . 2-8 8 2-22. Major com ponents of particle-phase diesel engine em issions, their known
atm o spheric transform a tion products, and the biological im pact of the reactants and products

........................................................ 2- 9 2 2-23. A m bient DPM concentrations reported from che m ical m a ss balance m odeling

.... 2- 9 8 2-24. Am bient diesel particulate m a tter concentrations from elem ental carbon

m e a s ur e m e nt s i n ur ba n l oc a t i ons

....................................... 2- 10 3 2-25. A m bient diesel particulate m a tter concentrations from dispersion m odeling

..... 2

- 10 5 2-26. Nationwide a m bient diesel particulate m a tter concentrations for 1996 from the National Air Toxics Assess m ent National-Scale Assess m ent d i s p e r s i on m od e l i ng

................................................. 2

- 10 7 2-27. Occupational exposure to DPM

........................................ 2

- 1 0 9 2-28. Ranges of occupational exposure to DPM by job category with estim ates of equivalent environ m ental exposures

............................................. 2

- 11 1 2-29. Annual average nationwide DPM exposure estim ates ( 

### : g/m 3 ) from on-road sources for rural and urban dem ographic groups in 1990, 1996, and 2007 using H A P E M

- M S 3

...................................................... 2

- 11 4

viiiLIST OF TABLES (continued)

ix 2-30. Draft annual average, 25 th , and 75 th percentile nationwide DPM exposure estimates (μg/m 3 ) from on-road and nonroad sources for rural and urban counties in 1996 using HAPEM4 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-113 2-31. 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 . . . . . . . . . . . . . . . . . . . . 2-114 2-32. Modeled and estimated concentrations of DPM in microenvironments for California for all sources . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2-117 2-33. Estimated indoor air and total air exposures to DPM in California in 1990 . . . . . . . . . 2-118

3-1. Predicted doses of inhaled DPM per minute based on total lung volume (M), total airway surface area (M 1 ), or surface area in alveolar region (M 2 ) . . . . . . . . . . . . . . . . . . . 3-9 3-2. Alveolar clearance in laboratory animals exposed to DPM in whole exhaust . . . . . . . . 3-17 3-3. Model comparison for deposition of DPM under equi valent conditions . . . . . . . . . . . . 3-50 3

- 4. Co m p a r a tive m od e l e s ti m a t e s of D P M d e position in hu m a n lun g s fr o m e x posu r e to 5μg/m3 c ontinuously f o r one y ea r

.................................. .... ... 3

- 5 3

5-1. Human studies of exposure to diesel exhaust . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5-18 5-2. Short-term effects of diesel exhaust on laboratory animals . . . . . . . . . . . . . . . . . . . . . . 5-26 5-3. Effects of chronic exposures to diesel exhaust on survival and growth of laboratory animals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5-31 5-4. Effects of chronic exposures to diesel exhaust on organ weights and organ-to-body-weight ratios . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5-33 5-5. Effects of diesel exhaust on pulmonary function of laboratory animals . . . . . . . . . . . . 5-36 5-6. Histopathological effects of diesel exhaust in the lungs of laboratory animals . . . . . . . 5-41 5-7. Effects of exposure to diesel exhaust on the pulmonary defense mechanisms of laboratory animals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5-51 5-8. Effects of inhalation of diesel exhaust on the immune system of laboratory animals . . 5-60 5-9. Effects of diesel particulate matter on the immune response of laboratory animals . . . 5-64 5-10. Effects of exposure to diesel exhaust on the liver of laboratory animals . . . . . . . . . . . . 5-70 5-11. Effects of exposure to diesel exhaust on the hematological and cardiovascular systems of laboratory animals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5-72 5-12. Effects of chronic exposures to diesel exhaust on serum chemistry of laboratory animals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5-74 5-13. Effects of chronic exposures to diesel exhaust on microsomal enzymes of laboratory animals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5-76 5-14. Effects of chronic exposures to di esel exhaust on behavior and neurophysiology . . . . 5-79 5-15. Effects of chronic exposures to diesel exhaust on reproduction and development in laboratory animals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5-81 5-16. Composition of expos ure atmospheres in studies comparing unfiltered and filtered diesel exhaust . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5-84

6-1. Histopathological effects of diesel exhaust in the lungs of laboratory animals . . . . . . . . 6-7 6-2. Human equivalent continuous concentrations: 70-year HECs calculated with the model of Yu et al. (1991) from long-term studies of rats repeatedly

e x po s e d t o D P M

........................................................ 6

- 2 2LIST OF TABLES (continued)

6-3. Effect estim ates per 50 

### : g/m 3 increase in 24-h PM 10 concentrations from U.S. and C a n a d i a n s t ud i e s

..................................................... 6

- 2 0 6-4. Effect estim ates per variable increm ents in 24-h concentrations of fine particle indicators (PM 2.5 , S O = 4 , H + ) f r om U

. S

. a nd C a n a d i a n s t ud i e s

.................. 6

- 2 3 6-5. Effect estim ates per increm ents in annual average levels of fine particle i nd i c a t o r s f r om U

. S

. a nd C a n a d i a n s t ud i e s

................................. 6

- 2 4 6-6. Decision sum m a ry for the quantitative noncancer RfC assessm ent for continuous exposure to diesel particulate m a tter (DPM)

............................... 6

- 3 3

7-1. Epide m iologic studies of the health effects of exposure to DE: cohort m o rtality

s t udi e s

............................................................. 7- 27 7-2. Epidem iologic studies of the health effects of exposure to DE: case-control studies of lung cancer

................................................. 7- 5 5 7-3. Su m m a ry of ani m al inhalation carcinogenicity studies

....................... 7- 8 5 7-4. Tum o r incidences in rats f o llowing intratracheal instillation of DE particles (DPM), extracted DPM, carbon black (CB), benzo[ a ]pyrene (B[ a ]P), or particles plus B [ a ] P

.......................................................... 7- 9 6 7-5. Tu m o rigenic effects of der m al application of acetone extracts of DPM

.......... 7- 9 6 7-6. Tum o r incidence and survival tim e of rats treated by surgical lung im plantation with fractions from DE condensate (35 rats/group)

......................... 7

- 11 1 7-7. Der m al tu m o rigenic and carcinogenic effects of various e m ission extracts

...... 7

- 12 0 7-8. Cu m u lative (concentration × ti m e) exposure data for rats exposed to whole DE

. . 7-12 2 7-9. Evaluations of DE as to hu m an carcinogenic potential

...................... 7

- 14 4

8-1. DPM exposure m a rgins (ratio of occupational ÷ environm ental exposures)

....... 8- 12

xLIST OF FIGURES

2-1. Diesel truck sales (dom estic) for the years 1939-1997

......................... 2- 6 2-2. Diesel truck sales as a percentage of total truck sales for the years 1939-1997

...... 2- 6 2- 3. P e r c e nt a ge of t r uc k m i l e s a t t r i but a bl e t o di e s e l t r uc ks

......................... 2- 7 2-4. Model year distribution of in-use HD truck fleet in 1997

...................... 2- 8 2-5. Model year distribution of vehicle m iles traveled by the in-use HD truck fleet in 1997

......................................................... 2- 8 2-6. A com p arison of IDI (A) and DI (B) com bustion system s of high-speed HD diesel
t r

uc k e ngi ne s

........................................................ 2- 10 2- 7. S c he m a t i c di a gr a m of di e s e l e ngi ne e xha us t pa r t i c l e s

........................ 2- 12 2-8. Typical chem ical com position for diesel particulate m a tter (PM 2.5 ) from new (post-1990) HD diesel vehicle exhaust

.................................... 2- 14 2-9. Trends in PM 10 em issions from on-road and nonroad engines com b ined and other anthropogenic sources of PM10 from 1970 to 1998 (excludes m i scellaneous and

na t ur a l s our c e s )

...................................................... 2- 19 2-10. Trends in PM 10 em issions from on-road and nonroad diesel engines from 1970 to 1998 and projections of em issions to 2007 and 2030

....................... 2- 21 2-11. Trends in NO

x em issions from on-road and nonroad diesel engines com b ined and other anthropogenic sources of NO

x from 1970 to 1998 (excludes m i scellaneous

a nd na t ur a l s our c e s )

.................................................. 2- 22 2-12. Trends in NO

x em issions from on-road and nonroad diesel engines from 1970 to 1998

........................................................ 2- 23 2-13. Trends in SO 2 em issions from on-road diesel engines from 1970 to 1998 and nonroad diesel engines from 1990 to 1998

................................. 2- 24 2-14. Trends in VOC em issions from on -road and nonroad diesel engines from 1970 to 1998

............................................................ 2- 24 2-15. Trends in CO em issions from on -road and nonroad diesel engines from 1970 to 1998

........................................................ 2- 25 2-16. Percentage of total m o tor fuel use that is on-road diesel fuel since 1949

......... 2- 26 2-17. On-highway diesel fuel consum ption si nce 1943, values in thousands of gallons

. . . 2-26 2-18. Model year trends in PM, NO

x , HC, and CO em issions from HD diesel vehicles ( g/ m i l e )

............................................................ 2- 30 2-19. Diesel engine certification data for NO

x e m i s s i ons a s a f unc t i on of m ode l ye a r

.... 2- 33 2-20. Diesel engine certification data for PM em issions as a function of m odel year

..... 2- 37 2-21. Em ission factors from HD dies el vehicles from tunnel studies

................. 2- 42 2-22. Line-haul and switch em issions data

..................................... 2- 44 2-23. Effect of turbocharging and aftercooling on NO

x a nd P M

..................... 2

- 4 9 2-24. Com p arison of diesel engine dynam om e ter PM em issions for four-stroke,

na t ur a l l y a s pi r a t e d, a nd t ur boc ha r ge d e ngi ne s

.............................. 2- 50 2-25. An exam ple of uniflow scavenging of a two-stroke diesel engine with a positive displacem ent blower

.................................................. 2- 52 2-26. Com p arison of two- and four-stroke ve hicle diesel PM em issions from chassis dynam om e ter studies

................................................. 2- 53 2-27. Com p arison of two- and four-stroke engine diesel PM em issions from engine dynam om e ter studies

................................................. 2- 54 2-28. Diesel engine dynam om e ter SOF em issions from two- and four-stroke engines

. . . . 2-54

xiLIST OF FIGURES (continued)

2-29. Diesel engine aldehyde em issions m easured in chassis dynam om e ter studies

..... 2- 56 2-30. Diesel engine aldehyde em issi ons from engine dynam om e ter studies

............ 2- 56 2-31. Trend in SOF em issions based on chassis dynam om e ter testing of HD diesel

ve hi c l e s

............................................................ 2- 66 2-32. Trend in SOF em issions for transient engine dynam om e ter testing of HD diesel

e ngi ne s

............................................................ 2- 66 2-33. Trend in SOF em issions as a percent of total PM based on chassis dynam om e ter t e s t i ng of H D d i e s e l v e h i c l e s

........................................... 2

- 6 8 2-34. Trend in SOF em issions as a percentage of total PM from engine dynam om e ter t e s t i ng

............................................................. 2

- 6 8 2-35. EC e m ission rates for diesel vehicles

..................................... 2- 7 2 2-36. EC content as percent of fine PM for DPM sam p les obtained in chassis dyna m om e ter studies

................................................. 2- 7 2 2-37. Diesel engine em issions of benzo[ a ]pyrene and 1-nitropyrene m easured in chassis dyna m om e ter studies

................................................. 2- 7 8 2-38. Diesel engine dynam om e ter m easurem ents of benzo[ a ]pyrene and 1-nitropyrene

e m i s s i ons f r om H D di e s e l e ngi ne s

....................................... 2- 7 9 2-39. Particle size distribution in DE

.......................................... 2- 8 2

3-1. Schem a tic representation of m a jor m echanism s , including diffusion, involved in particle deposition

..................................................... 3- 4 3-2. Generalized regional deposition fractions of various sized particles in the hum an

r e s pi r a t or y t r a c t

....................................................... 3- 6 3-3. Modeled deposition distribution patterns of inhaled DE particles in the airways of di f f e r e nt s pe c i e s

...................................................... 3- 8 3-4. Diagram of known and suspected clearance pathways for poorly soluble particles de pos i t i ng i n t he a l ve ol a r r e gi on

......................................... 3- 10 3-5. Modeled clearance of poorly soluble 4- 

### : m particles deposited in tracheobronchial

a nd a l ve ol a r r e gi ons i n hum a n s.......................................... 3- 12 3-6. Short-term thoracic clearance of inhaled particles as determ ined by m odel prediction

a nd e xpe r i m e nt a l m e a s ur e m e nt

.......................................... 3- 13 3-7. Clearance from lungs of rats of 134 Cs-FAP fused alum inosilicate tracer particles inhaled after 24 m onths of DE exposure at concentrations of 0 (control), 0.35 (low), 3.5 (m edium ) , and 7.1 (high) m g DPM/m 3

....................... 3- 18 3-8. Lung burdens (in m g DPM soot/g lung) in rats chronically exposed to DE at 0.35 (low), 3.5 (m edium ) , and 7.1 (high) m g ppm /m 3

........................ 3- 27 3-9. Modeled estim ates of lung burden in hum ans after a sim u lated lifetim e exposure to DPM using the Yu et al. (1991) and ICRP66 m odels

....................... 3- 51

6- 1. F l ow di a gr a m of pr oc e dur e f or c a l c ul a t i ng H E C s

........................... 6- 10 6-2. Relative risk (RR) estim ates for increased m o rtality and m o rbidity endpoints associated with 50 

### : g/m 3 increm ents in PM 10 concentrations as derived from studies cited by num bers listed above each given type of health endpoint

........ 6- 22 6-3. Relative risks of acute m o rtality in Harvard Six Cities Study, for inhalable thoracic particles (PM 15 /PM 10 ), fine particles (PM 2.5 ), and coarse fraction

xiiLIST OF FIGURES (continued)
particles (PM -PM )

................................................. 2.5 15 6

- 2 6 6-4. Adjusted relative risks for m o rtality are plotted against each of seven long-term average particle indices in the Harvard Six Cities Study, from largest range (total suspended particles) through sulfate and nonsul fate fine particle concentrations

. . . 6-27 6-5. Percent of children with <85% norm a l FVC versus annual-average fine (PM 2.1 ) particle concentrations and coarse fraction (PM 10-2.1 ) levels for 22 North Am erican cities

...................................................... 6

- 2 8

7-1. Pooled relative risk estim ates and heterogeneity-adjusted 95% confidence intervals for all studies and subgroups of studies included in the m e ta-analysis

........... 7

- 6 2 7-2. Pooled estim ates of relative risk of lung cancer in epidem iologic studies involving occupational exposure to DE (random -effects m odels)

............... 7

- 6 4 7-3. Pathogenesis of lung disease in rats with chronic, high-level exposures to
p a r t i c l e s

........................................................... 7

- 1 3 7

xiiiFOREWORD

The diesel engine has been a vital workhorse in the United States, powering m a ny of its large trucks, buses, and farm , railroad, m a rine, and construction equipm ent. Expectations are that diesel engine use in these areas will increase due to the superior perf orm a nce characteristics of the engine. Diesel engine exhaust (DE), however, contains harm ful pollutants in a com p lex

m i xture of gases and particulates. Hum a n exposure to this exhaust com e s from both highway uses (on-road) as well as nonroad uses of the diesel engine. EPA started evaluating and regulating the gaseous em issions from the heavy-duty highway use of diesel engines in the 1970s and particle em issions in the 1980s. The reduction of
harm ful exhaust em issions has taken a large step forward because of standards issued in 2000 which will bring about very large reductions in exhaust em issions for m odel year 2007 heavyduty engines used in trucks, buses, and other on-road uses. A draft of this assessm ent, along with the peer review com m e nts of the Clean Air Scientif ic Advisory Com m ittee, was part of the scientific basis for EPA’s regulation of heavy-duty highway engines com p leted in Decem ber
2000. The inform ation provided by this assessm ent was useful in developing EPA’s understanding of the public health im plications of exposure to DE and the public health benefits of taking regulatory action to control exhaust em issions. EPA anticipates developing sim ilarly stringent regulations for other diesel engine us es, including those used in nonroad applications. Until these regulations take ef f ect, EPA is partnering with state and local agencies to retrofit older, dirtier, engines to m a ke them run cleaner and to develop m odel program s to reduce

em issions from idling engines. In addition, EPA and local authorities are working to ensure early introduction of effective technologies for particulate m a tter control and the availability of low- sulfur fuel where possible in advance of the 2007 requirem e nts. Today, at least one engine

m a nufacturer is producing new engines with partic ulate traps that, when coupled with low-sulfur fuel, m eet 2007 particulate em ission levels. Th e Agency expects significant environm ental and public health benefits as the environm ental perform a nce of diesel engines and diesel fuels
im proves. The health assessm ent concludes that long-term (i.e., chronic) exposure to DE is likely to pose a lung cancer hazard as well as dam a ge the lung in other ways depending on exposure. The
health assessm ent’s conclusions are based on exposure to exhaust from diesel engines built prior
to the m i d-1990s. Short-term (i.e., acute) exposures can cause transient irritation and
inflam m a tory sym p tom s , although the nature and extent of these sym p tom s are highly variable across the population. The assessm ent also states that evidence is em erging that diesel exhaust

xivxvPREFACE

This docum ent is the U.S. Environm ental Protection Agency’s science-based Health Assessment Document for Diesel Engine Exhaust

. The assessm ent was prepared by the National
Center for Environm ental Assessm ent which is the health risk assessm ent program in EPA’s Office of Research and Developm ent. The assessm ent broadly supports activities authorized in
the 1990 Clean Air Act. This assessm ent was specifically prepared for EPA’s Office of Transportation and Air Quality which requested inf o rm ation regarding the potential health hazards associated with diesel engine exhaust (D E) exposure. As DE em issions also contribute
to urban air toxics and am bient particulate m a tter, other EPA air program s also have an interest
in this assessm ent. This docum ent was preceded by five earlier drafts: a W o rkshop Review Draft (EPA/600/8-90/057A, July 1990), an External Review Draft (EPA/600/8-90/057B, Decem ber 1994), an SAB Review Draft (EPA/600/8-90/057C , February 1998), an SAB Review Draft (EPA/600/8-90/057D, Novem b er 1999), and an SAB Review Draft (EPA/600/8-90/057E, July 2000). There was an SAB Environm ental Health Com m ittee Review in 1990 of the July 1990
draft. The Science Advisory Board’s Clean Air Scientif ic Advisory Com m ittee (CASAC) reviewed the 1994 draft in public sessions in May 1995, the 1998 draft in May 1998, the 1999
draft in Decem ber 1999, and the July 2000 draft in October 2000. Public com m e nt periods also were conducted concurrently with the CASAC revi ews. In addition m a ny reviewers, both within and outside the Agency, provided assistance at various review stages. This is the final version of
the assessm ent which was prepared in response to CASAC advice and public com m e nts received
on the 2000 draft.
The scientific literature search for this assessm ent is generally current through January 2000, although a few later publications have been included.

xviAUTHORS, CONTRIBUTORS, AND REVIEWERS

The National Center for Environm ental Assessm ent (NCEA) within EPA’s Office of Research and Developm ent (ORD) was responsible for the preparation of this docum ent.

CHAPTER 1. EXECUTIVE SUMMARY

Authors NCEA Diesel Team

CHAPTER 2. DIESEL EMISSIONS CHARACTERIZ ATION, ATMOSPHERIC TRANSFORMATION, AND EXPOSURES

Author Marion Hoyer, Office of Transportation and Air Quality, U.S. Environm ental Protection Agency, Ann Arbor, MI.

Contributors Chad Bailey, Office of Transportation and Air Qu ality, U.S. Environm ental Protection Agency, Ann Arbor, MI.

Tom Baines, Office of Transportation and Air Qu ality, U.S. Environm ental Protection Agency, Ann Arbor, MI.

David Cleverly, National Center for Environm en tal Assessm ent, U.S. Environm ental Protection Agency, W a shington, DC.

W illiam Ewald, National Center f o r Environm ental Assessm ent, U.S. Environm ental Protection Agency, Research Triangle Park, NC.

Robert McCorm ick, Colorado School of Mines, Golden, CO.

Joseph McDonald, Office of Transportation and Air Quality, U.S. Environm ental Protection Agency, Ann Arbor, MI.

Joseph Som e rs, Office of Transportation and Air Quality, U.S. Environm ental Protection Agency, Ann Arbor, MI.

Janet Yanowitz, Colorado School of Mines, Golden, CO.

Barbara Zielinska, Desert Research Institute, Reno, NV.

xviiAUTHORS, CONTRIBUTORS, AND REVIEWERS (continued)

CHAPTER 3. DOSIMETRY OF DIESEL PARTICULATE MATTER

Authors Gary Fourem an, National Center for Environm en tal Assessm ent, U.S. Environm ental Protection Agency, Research Triangle Park, NC.

Jam e s McGrath, National Center for Environm ental Assessm ent, U.S. Environm ental Protection Agency, Research Triangle Park, NC.

W illiam Pepelko, National Center for Environm ental Assessm ent, U.S. Environm ental Protection Agency, W a shington, DC.

CHAPTER 4. MUTAGENICITY

Author Lawrence Valcovic, National Center for Environm ental Assessm ent, U.S. Environm ental Protection Agency, W a shington, DC.

CHAPTER 5. NONCANCER HEALTH EFFECTS OF DIESEL EXHAUST

Author Jam e s McGrath, National Center for Environm ental Assessm ent, U.S. Environm ental Protection Agency, Research Triangle Park, NC.

Contributor Gary Fourem an, National Center for Environm en tal Assessm ent, U.S. Environm ental Protection Agency, Research Triangle Park, NC.

CHAPTER 6. QUANTITATIVE APPROACHES TO ESTIMATING

HUMAN NONCANCER HEALTH RISKS OF DIESEL EXHAUST

Authors Gary Fourem an, National Center for Environm en tal Assessm ent, U.S. Environm ental Protection Agency, Research Triangle Park, NC.

Lester Grant, National Center for Environm en tal Assessm ent, U.S. Environm ental Protection Agency, Research Triangle Park, NC.

Contributors Karen Martin, Of f i ce of Air Quality Planni ng and Standards-OAR, U.S. Environm ental Protection Agency, Research Triangle Park, NC.

xviiiAUTHORS, CONTRIBUTORS, AND REVIEWERS (continued)

xix 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 Envi ronmental 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 Nationa l 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 Agency, Washington, DC.

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 TeamAUTHORS, CONTRIBUTORS, AND REVIEWERS (continued)

xx REVIEWERS The Science Advisory Board’s Clean Ai r Scientific Advisory Committee (CASAC) reviewed the 1994 draft in public sessions in May 1995, the 1998 draft in May 1998, the 1999 draft in December 1999, and the July 2000 draft in October 2000. Public comment periods also were conducted concurrently with the CASAC re views. In addition, many reviewers both within and outside the Agency provided assistance at various review stages. This is the final version of the assessment which was prepared in response to the latest CASAC advice and public

comments. The authors wish to thank all those who sought to improve the quality of this report with their comments and are particularly grateful to the CASAC for its advice. The following members of the SAB’s CASAC participated in the review of the July 2000
draft.

Panel Chair Dr. Joe Mauderly 1 , Vice President, Senior Scientist, and Director of National Environmental Respiratory Center, Lovelace Respirator y Research Institute, Albuquerque, NM.

CASAC Members 2 Mr. John Elston , Administrator, Office of Air Quality Management, State of New Jersey, Department of Environmental Protection and Energy, Trenton, NJ.

Dr. Philip K. Hopke 3 , R.A. Plane Professor of Chemistry, Clarkson University, Department of Chemistry, Potsdam, NY (CASAC Chair).

Dr. Eva J. Pell 4 , Steimer Professor of Agriculture Sciences, The Pennsylvania State University, University Park, PA.

Dr. Arthur C. Upton, M.D. , Director, Independent Peer Review, UMDNJ-CRESP, Environmental and Occupational Health Sciences Institute, New Brunswick, NJ.

Dr. Sverre Vedal
, M.D. , University of British Columbia, Vancouver Hospital, Vancouver, BC, Canada.

Dr. Warren White 5 , Senior Research Associate, Washington University, Chemistry Department, St. Louis, MO.

CASAC Consultants 6 Dr. David Diaz-Sanchez , Department of Medicine, UCLA, Los Angeles, CA.

Dr. Eric Garshick, M.D. , Staff Physician, Pulmonary and Critical Care Section, West Roxbury Virginia Medical Center, West Roxbury, MA.Dr. Roger O. McClellan , Advisor, Toxicology and Hum a n Health Risk Analysis, and President Em eritus, Chem ical Industry Institute of Toxicology (CIIT), Albuquerque, NM.

Dr. Gunter Oberdörster , University of Rochester Medical Center, Departm e nt of Environm ental Medicine, Rochester, NY.

Dr. Leslie Stayner 7 , National Institute f o r Occupational Saf e ty and Health (NIOSH), Risk Evaluation Branch, Taft Laboratories, Cincinnati, OH.

Dr. Ron Wyz ga , Electric Power Research Institute (EPRI), Palo Alto, CA.

Science Advisory Board Staff Mr. Robert Flaak , Designated Federal Official (DFO) and Team Leader, Com m ittee Operations Staff, EPA Science Advisory Board ( 1400A), 1200 Pennsylvania Avenue, NW , U.S. Environm ental Protection Agency, W a shington, DC 20460

Ms. Diana Poz u n , Managem e nt Assistant, Com m ittee Operations Staff, and Program Specialist, Office of the Staff Director, EPA Scien ce Advisory Board (1400A), 1200 Pennsylvania Avenue, NW , U.S. Environm ental Protection Agency, W a shington, DC 20460

_____________________ 1 Appointm e nt as Chair of CASAC e nded on October 30, 2000. Appointed ex officio P a s t C h air until Septem ber 30, 2001.

2 CASAC M e m b ers are experts appointed by the Adm i nistrator to two- year term s to serve on the Clean Air S c ientific A dvis o ry C o m m ittee. 3 Appointed as Chair of CASAC on October 30, 2000. 4 Resigned from CASAC on Septem ber 28, 2000. 5 Appointm e nt as M e m b er of CASAC ended on October 30, 2000. 6 CASAC Consultants are experts appointed by the Scie nce Advisory Board Staff Director to a one-year
term to serve on ad hoc Panels form ed to address a particular issu e; in this case, the CASAC Review of EPA’s Health Assessm ent Docum e nt for Diesel Exhaust.
7 Federal Expert.

xxiACKNOWLEDGMENTS

The authors would like to acknowledge the contributions of several people who have

m a de this report and the previous drafts possible.

Document Review Vanessa Vu National Center f o r Environm ental Assessm ent U.S. Environm ental Protection Agency W a shington, DC

Document Production Terri Konoza Judy Theisen National Center for Environm ental Assessm ent U.S. Environm ental Protection Agency W a shington, DC

Kay Marshall Joanna Taylor Eric Sorensen Clara Calderon-Laucho The CDM Group, Inc. Chevy Chase, MD Sam Napolitano O f f i ce of Transportation and Air Quality U.S. Environm ental Protection Agency W a shington, DC

xxii1. EXECUTIVE SUMMARY

1.1. INTRODUCTION
This Health Assessm ent Docum e nt for Diesel Engine Exhaust (DE) represents EPA’s
first com p rehensive review of the potential health effects from am bient exposure to exhaust from diesel engines. The assessm ent was developed to provide inform ation about the potential for DE to pose environm ental health hazards, inform ation that would be useful in evaluating regulatory needs under provisions of the Clean Air Act. The assessm ent identifies and characterizes the potential hum an health hazards of DE (i.e, hazard assessm ent) and seeks to estim ate the relationship between exposure and disease response for the key health effects (i.e., dose-response

assessm ent). A full exposure assessm ent and risk characterization, the other two com ponents of

a com p lete risk assessm ent, are beyond the scope of this docum ent.

The report has nine chapters and three appendices. Chapter 2 provides a characterization
of diesel em issions, atm o spheric transform a tion, and hum an exposures to provide a context for
the hazard evaluation of DE. Chapters 3, 4, 5, and 7 provide a review of relevant inform ation for the evaluation of potential health hazards of DE, including dosim etry (Chapter 3), m u tagenicity (Chapter 4), noncancer effects (Chapter 5), and car cinogenic effects (Chapter 7). Chapters 6 and 8 contain dose-response analyses to provide in sight about the significance of the key noncancer
and cancer hazards. Chapter 9 sum m a rizes and characterizes the overall nature of the health hazard potential in the environm ent and the overall confidence and/or uncertainties associated
with the conclusions.

1.2. COMPOSITION OF DIESEL EXHAUST
DE is a com p lex m i xture of hundreds of constituents in either a gas or particle form

. Gaseous com ponents of DE include carbon dioxide, oxygen, nitrogen, water vapor, carbon

m onoxide, nitrogen com pounds, sulfur com pounds, and num erous low-m o lecular-weight hydrocarbons. Am ong the gaseous hydrocarbon com ponents of DE that are individually known to be of toxicologic relevance are the al dehydes (e.g., form aldehyde, acetaldehyde, acrolein), benzene, 1,3-butadiene, and polycyclic arom atic hydrocarbons (PAHs) and nitro-PAHs. The particles present in DE (i.e., diesel particulate m a tter [DPM]) are com posed of a
center core of elem ental carbon and adsorbed organic com pounds, as well as sm all am ounts of sulfate, nitrate, m e tals, and other trace elem ents. DPM consists of fine particles (fine particles
have a diam eter <2.5 

### : m ) , including a subgroup with a large num ber of ultrafine particles (ultrafine particles have a diam eter <0.1 

### : m )

. Collectively, these particles have a large surf ace area which m a kes them an excellent m e dium for adsorbing organics. Also, their sm all size

m a kes them highly respirable and able to reach the deep lung. A num ber of potentially

1-1toxicologically relevant organic com pounds are on the particles. The organics, in general, range
from about 20% to 40 % of the particle wei ght, though higher and lower percentages are also reported. Many of the organic com pounds present on the particle and in the gases are individually known to have m u tagenic and carcinoge nic properties. For example, PAHs, nitroPAHs, and oxidized PAH derivatives are present on the diesel particles, with the PAHs and their
derivatives com p rising about 1% or less of the DPM m a ss. DE em issions vary significantly in chem ical com position and particle sizes between different engine types (heavy-duty, light-duty) , engine operating conditions (idle, accelerate, decelerate), and fuel form ulations (high/low sulf ur fuel). Also, there are em ission differences between on-road and nonroad engines sim p ly because the nonroad engines to date are generally of older technology. The m a ss of particles em itted and the organic com ponents on the particles
from on-road diesel engines have been reduced over the years. Available data for on-road engines indicate that toxicologically relevant organic components of DE (e.g., PAHs, nitroPAHs) em itted f r om older vehicle engines are still present in em issions f r om newer engines, though relative am ounts have decreased. There is currently insufficient inform ation to characterize the changes in the com position of DE from nonroad diesel engines over tim e.

1.3. DIESEL EXHAUST AS A COMPONENT OF AMBIENT PARTICULATE MATTER
DE is em itted from “on-road” diesel engines (vehicle engines) or “nonroad” diesel engines (e.g., locom o tives, m a rine vessels, heavy-duty equipm ent, etc.). Nationwide, data in
1998 indicated that DE as m easured by DPM m a de up about 6% of the total am bient PM 2.5 inventory (i.e., particles with aerodynam i c diam et er of 2.5 m i crom eters or less) and about 23% of the inventory, if natural and m i scellaneous sources of PM 2.5 are excluded. Estim ates of the DPM percentage of the total inventory in urban centers are higher. For exam ple, estim ates range from 10% to 36% in som e urban areas in Califor nia, Colorado, and Arizona. Available data also indicate that over the years there have been significant reductions in DPM em issions from the exhaust of on-road diesel engines, whereas lim ited data suggest that exhaust em issions from nonroad engines have increased.

1.4. ATMOSPHERIC TRANSFORMATION OF DIESEL EXHAUST
After em ission from the tailpipe, DE undergoes dilution and chem ical and physical
transform a tions in the atm o sphere, as well as dispersion and transport in the atm o sphere. The

atm o spheric lifetim e for som e com pounds present in DE ranges from hours to days. DPM is directly em itted f r om diesel-powered engines (prim a ry particulate m a tter) and can be f o rm ed
from the gaseous com pounds em itted by diesel engines (secondary particulate m a tter). Lim ited
inform ation is available about the physical and chem ical transform a tion of DE in the

1-2atm o sphere. It is not clear what the overall toxi cological consequences of DE’s transform a tions
are because som e com pounds in the DE m i xture are altered to m o re toxic form s while others are

m a de less toxic.

1.5. EXPOSURE TO DIESEL EXHAUST
DPM m a ss (expressed as 

### : g DPM/m 3 ) has historically been used as a surrogate m easure of exposure for whole DE. Although uncertainty exists as to whether DPM is the m o st appropriate param e ter to correlate with hum an health effects, it is considered a reasonable choice
until m o re definitive inform ation about the m echanism s of toxicity or m ode(s) of action of DE
becom e s available. In the am bient environm ent, hum an exposure to DE com e s fromboth onroad and nonroad engine exhaust. A large percentage of the U.S. population also is exposed to

am bient PM 2.5 , of which DPM is typically a significant constituent. Although this docum ent does not provide an exposure assessm ent, DE exposure inform ation is included to provide a
context for the health effects inform ation. Exposure estim ates for the early to m i d-1990s suggest that national annual average DE exposure from on-road engines alone was in the range of about
0.5 to 0.8 

### : g DPM/m 3 of inhaled air in m a ny rural and urban areas, respectively. Exposures could be higher if there is a nonroad DE source that adds to the exposure from on-road vehicles.
For exam ple, prelim inary estim ates show that, on a national average basis, accounting for
nonroad DE em issions adds another twofold to the on-road exposure. For localized urban areas where people spend a large portion of their tim e outdoors, the exposures are higher and, for

exam ple, m a y range up to 4.0 

### : g DPM/m 3 of inhaled air.

1.6. HEALTH EFFECTS OF DIESEL EXHAUST
Available evidence indicates that there are hum an health hazards associated with exposure to DE. The hazards include acute exposure-related sym p tom s , chronic exposurerelated noncancer respiratory effects, and lung cancer. The health hazard conclusions are based
on exhaust em issions from diesel engines built prior to the m i d-1990s. W ith current engine use including som e new and m a ny m o re older engines (engines typically stay in service for a long
tim e), the health hazard conclusions, in general, are applicable to engines currently in use. As new and cleaner diesel engines, together with di fferent diesel fuels, replace a substantial num ber of existing engines, the general applicability of th e health hazard conclusions will need to be reevaluated. W ith new engine and fuel tec hnology expected to produce significantly cleaner engine exhaust by 2007 (e.g., in response to new federal heavy duty engine regulations), significant reductions in public health hazards are expected for those engine uses affected by the regulations.

1-31.6.1. Acute (Short-Term Exposure) Effects
Inf o rm ation is lim ited f o r characterizing the potential health ef f ects associated with
acute or short-term exposure. However, on the basis of available hum an and anim al evidence, it
is concluded that acute or short-term (e.g., episodic) exposure to DE can cause acute irritation (e.g., eye, throat, bronchial), neurophysiological sym p tom s (e.g., lightheadedness, nausea), and respiratory sym p tom s (cough, phlegm )

. There also is evidence for an im m unologic effect–the exacerbation of allergenic responses to known allergens and asthm a -like sym p tom s

. The lack of adequate exposure-response inform ation in the acute health effect studies precludes the developm ent of recom m e ndations about levels of exposure that would be presum ed safe for
these effects.

1.6.2. Chronic (Long-Term Exposure) Noncancer Respiratory Effects
Inf o rm ation f r om the available hum an studies is inadequate f o r a def i nitive evaluation of possible noncancer health effects from chronic exposure to DE. However, on the basis of
extensive anim al evidence, DE is judged to pose a chronic respiratory hazard to hum ans. Chronic-exposure, anim al inhalation studies s how a spectrum of dose-dependent inflam m a tion and histopathological changes in the lung in several anim al species including rats, m i ce, ham s ters, and m onkeys.
This assessm ent provides an estim ate of inhalation exposure of DE (as m easured by DPM) to which hum ans m a y be exposed throughout their lifetim e without being likely to experience adverse noncancer respiratory effects. This exposure level, known as the reference concentration (RfC) for DE of 5 

### : g/m 3 of DPM was derived on the basis of dose-response data
on inflam m a tory and histopathological changes in the lung from rat inhalation studies. In recognition of the presence of DPM in am bient PM 2.5 , it also is appropriate to consider the
wealth of PM 2.5 hum an health effects data. In this regard, the 1997 National Am bient Air Quality Standard f o r PM 2.5 of 15 

### : g/m 3 (annual average concentration) also would be expected to provide a m easure of protection from DPM, reflecting DPM’s current approxim a te proportion
to PM 2.5 .

1.6.3. Chronic (Long-Term Exposure) Carcinogenic Effects
This assessm ent concludes that DE is “likely to be carcinogenic to hum ans by inhalation” and that this hazard applies to environm ental exposures. This conclusion is based on the totality
of evidence from hum an, anim al, and other supporting studies. There is considerable evidence
dem onstrating an association between DE exposure and increased lung cancer risk am ong workers in varied occupations where diesel e ngines historically have been used. The hum an evidence from occupational studies is considered strongly supportive of a finding that DE

1-4exposure is causally associated with lung cancer, though the evidence is less than that needed to definitively conclude that DE is carcinogenic to hum ans. There is som e uncertainty about the degree to which confounders are having an influence on the observed cancer risk in the occupational studies, and there is uncertainty evolving from the lack of actual DE exposure data
for the workers. In addition to the hum an evidence, there is supporting evidence of DPM’s carcinogenicity and associated DPM organic com pound extracts in rats and m i ce by noninhalation routes of exposure. Other supporting evidence includes the dem onstrated

m

u tagenic and chrom o som a l effects of DE and its organic constituents, and the suggestive evidence for bioavailability of the DPM organics in hum ans and anim als. Although highexposure chronic rat inhalation studies show a significant lung cancer response, this is not thought predictive of a hum an hazard at lower environm ental exposures. The rat response is
considered to result from an overload of particles in the lung resulting from the high exposure, and such an overload is not expected to occur in hum ans at environm ental exposures. Although the available hum an evidence shows a lung cancer hazard to be present at occupational exposures that are generally higher than environm ental levels, it is reasonable to

presum e that the hazard extends to environm ental exposure levels. W h ile there is an incom p lete understanding of the m ode of action for DE-induced lung cancer that m a y occur in hum ans, there is the potential for a nonthreshold m u tagenic m ode of action stem m i ng from the organics in the
DE m i xture. A case for an environm ental hazard also is shown by the sim p le observation that
the estim ated higher environm ental exposure levels are close to, if not overlapping, the lower range of occupational exposures for which lung cancer increases are reported. These considerations taken together support the prudent public health choice of presum ing a cancer
hazard for DE at environm ental levels of exposure. Overall, the evidence for a potential cancer
hazard to hum ans resulting from chronic inhalation exposure to DE is persuasive, even though

assum p tions and uncertainties are involved. W h ile the hazard evidence is persuasive, this does
not lead to sim ilar confidence in understanding the exposure/dose-response relationship. Given a carcinogenicity hazard, EPA typically perform s a dose-response assessm ent of
the hum an or anim al data to develop a cancer unit risk estim ate that can be used with exposure
inform ation to characterize the potential cancer disease im pact on an exposed population. The
DE hum an exposure-response data are considered too uncertain to derive a confident quantitative
estim ate of cancer unit risk, and with the chronic rat inhalation studies not being predictive for
environm ental levels of exposure, EPA has not developed a quantitative estim ate of cancer unit
risk. In the absence of a cancer unit risk, sim p le exploratory analyses were used to provide a perspective of the range of possible lung cancer risk from environm ental exposure to DE. The analyses m a ke use of reported lung cancer risk increases in occupational epidem iologic studies,

1-5and the differences between occupational and environm ental exposure. The purpose of having a risk perspective is to illustrate and have a sense of the possible significance of the lung cancer
hazard from environm ental exposure. The risk perspective cannot be viewed as a definitive quantitative characterization of cancer risk nor is it suitable for estim ation of exposure-specific population risks.

1.7. SOURCES OF UNCERTAINTY Even though the overall evidence for potential hum an health effects of DE is persuasive,

m a ny uncertainties exist because of the use of assum p tions to bridge data and knowledge gaps
about hum an exposures to DE and the general lack of understanding about underlying

m echanism s by which DE causes observed toxicities in hum ans and anim als. A notable uncertainty of this assessm ent is whether the health hazards identified from studies using

em issions from older engines can be applied to present-day environm ental em issions and related

exposures, as som e physical and chem ical charact eristics of the em issions from certain sources have changed over tim e. Available data are not suf f i cient to provide def i nitive answers to this question because changes in DE com position over tim e cannot be confidently quantified, and the relationship between the DE com ponents and the m ode(s) of action for DE toxicity is/are
unclear. W h ile recognizing the uncertainty, for this assessm ent a judgm ent is m a de that prioryear toxicologic and epidem iologic findings can be applied to m o re current exposures, both of
which use DPM m a ss in air as the m easure of DE exposure.
Other uncertainties include the assum p tions that health effects observed at high doses

m a y be applicable to low doses, and that toxico logic findings in laboratory anim als generally are predictive of hum an responses. In the absence of a m o re com p lete understanding of how DE

m a y cause adverse health effects in hum ans and laboratory anim als, related assum p tions (i.e., the presence of a biological threshold for chronic respiratory effects based on cum u lative dosage and absence of a threshold for lung cancer stem m i ng from subtle and irreversible effects) are considered reasonable and prudent. Although parts of this assessm ent, particularly the noncancer RfC estim ate, have been derived with a generic consideration of sensitive subgroups within the population, the actual spectrum of the population that m a y have a greater susceptibility to DE is unknown and cannot
be better characterized until m o re inf o rm ation is available regarding the adverse ef f ects of DPM
in hum ans. Increased susceptibility, for exam ple, could result from above-average increases in DE deposition and retention in the respiratory system or intrinsic differences in respiratory system tissue sensitivity. There is no DE-specif i c inf o rm ation that provides direct insight to the question of differential hum an susceptibility. Given the nature of DE’s noncancer effects on the respiratory system it would be reasonable, for exam ple, to consider possible vulnerable

1-6subgroups to include infants/children, the elderly, or individuals with preexisting health conditions, particularly respiratory conditions. In developing a perspective on the possible significance of the environm ental cancer hazard of DE, this assessm ent uses inform ation about the differences in the m a gnitude of DE exposures between the occupational and environm ental settings. Although an appreciation for differences in exposure is needed only at an order-of-m a gnitude level for this assessm ent, one should recognize that individual exposure is a functi on of both the variable concentrations in the
environm ent and the related breathing and particle retention patterns of the individual. Because of variations in these factors across the popul ation, different subgroups could receive lower or higher exposure to DE than those groups m e ntioned in this assessm ent. Lastly, this assessm ent considers only potential heath effects from exposures to DE alone. Effects of DE exposure could be additive to or synergistic with concurrent exposures to

m a ny other air pollutants. However, in the absence of m o re def i nitive data dem onstrating interactive effects (e.g., potentiation of allergen icity effects, potentiation of DPM toxicity by

am bient ozone and oxides of nitrogen) from com b ined exposures to DE and other pollutants, it is not possible to address this issue. Further research is needed to im prove the knowledge and data on DE exposures and potential hum an health effects, and thereby reduce uncertainties of future

assessm ents of the DE health effects data.

1-72. DIESEL EXHAUST EMISSIONS CHARACTERIZ ATION, ATMOSPHERIC TRANSFORMATION, AND EXPOSURES

2.1. INTRODUCTION This chapter provides background inform ation relating to the diesel engine, the pollutants
it em its, the history of its use in highway vehicles and railroad locom o tives, diesel exhaust

com position and em issions trends, and air pollution regulatory standards for diesel engines in the
United States. The chapter also provides specific inform ation about the physical and chem ical

com position of diesel exhaust, descriptions of its atm o spheric transform a tions, observations of

m easured and m odeled am bient concentrations (considered alone and as a com ponent of

atm o spheric particles in general), som e estim ates of population exposures as well as a

com p arison of DPM with am bient fine particulate m a tter (PM 2.5 ). In addition, this chapter gives background inform ation that is used in conjunction with toxicology and epidem iology data to
form ulate conclusions about hum an health hazards that are discussed in later chapters of this
docum ent. The exposure inform ation does not represent a form al or rigorous exposure

assessm ent; it is intended only to provide a context for the health effects data and health hazard findings. For the purposes of this docum ent, carbonaceous m a tter, diesel exhaust, diesel particulate

m a tter, elem ental carbon, organic carbon, soluble or ganic fraction, and soot are defined below.

Carbonaceous matter : Carbon-containing com pounds that are associated with particulate m a tter in diesel exhaust. In this docum ent, the term carbonaceous m a tter includes all organic and elem ental carbon-containing com pounds that are found in the particle phase. In other docum ents, this term is som e tim es used interchangeably to refer to the insoluble fraction of diesel particulate m a tter or the soot fraction.

Diesel engine exhaust (DE) : Gaseous and particle-phase em issions resulting from the

com bustion of diesel fuel in an internal-com bustion, com p ression-ignition engine. DE
includes em issions from a diesel engine or diesel vehicle (inclusive of aftertreatm e nt devices), but does not include em issions from brake and tire wear.

Diesel particulate matter (DPM) : The particle-phase com pounds em itted in DE. DPM can refer to both prim ary em issions and secondary particles that are form ed by

atm o spheric processes. In this docum ent, DPM refers to prim ary particles. Prim ary diesel particles are considered f r esh af ter being em itted and aged af ter

2-1undergoing oxidation, nitration, or other chem ical and physical changes in the

atm o sphere. As used in this docum ent, DPM refers to both fresh and aged DPM unless a
distinction is m a de.

Elemental carbon (EC) : Carbon that has undergone pyrolysis (i.e., has been stripped of hydrogen). In pure form , EC contains only carbon atom s, although EC as it exists in

com bustion particulate m a tter is likely to contain som e hydrogen atom s.

Organic carbon (OC) : Carbon- and hydrogen-containing m o lecules em itted in DE largely as the result of unburned diesel fuel and, to a lesser extent, from engine
lubrication oil. OC com pounds also can contain oxygen, nitrogen, and sulfur, as well as
other elem ents in sm all quantities.

Soluble organic fraction (SOF) : The organic portion of DPM that can be extracted
from the particle m a trix into solution. Extraction solutions and procedures vary and are
described in Section 2.2.8.1.

Soot : Agglom erations of EC and OC particles. Soot also is often characterized as the insoluble portion of DPM, and is therefore considered to be m a inly EC by som e investigators.

This chapter begins with a history of dieselization for on-road vehicles and locom o tives, followed by an introductory discussion of the form a tion of prim ary diesel em issions to assist the reader in understanding the com p lex factors that influence the form ation of particulate m a tter (PM) and other DE em issions. The next secti on is a sum m a ry of EPA em ission standards for
on-road and locom o tive diesel engines and a description of the national trends in em issions from on-road and nonroad diesel engine sources base d on inventory m odeling. The chapter continues
with a discussion of diesel fuel use and the im pact of fuel properties on em issions. The chronological assessm ent of em issions factors is presented in sum m a ries of chassis and engine dynam o m e ter testing and tunnel tests. This is followed by a description of engine technologies
and their effect on em issions, and a description of the chem ical and physical nature of em issions. The data describing the im portant atm o spheric transform a tions of DE are sum m a rized. The chapter concludes with a sum m a ry of the available literature regarding atm o spheric concentrations of DPM and exposures to DE. EPA has assessed national and urban-area annual average exposure to DPM using the Hazardous Air Pollutant Exposure Model, and this

assessm ent is presented in Section 2.4.3. A full exposure assessm ent would include the

2-2distribution of am bient DE exposures in di fferent geographic regions and am ong different
dem ographic groups, the m o st highly exposed (90 th percentile), exposures in m i croenvironm ents for short and long durations, the m a xim u m exposure range (98 th percentile), and the num ber of

m a xim u m

- exposed individuals. However, such an assessm ent is not currently available. EPA is developing tools to provide a m o re com p lete exposure assessm ent.

2.2. PRIMARY DIESEL EXHAUST EMISSIONS
2.2.1. History of Dieseliz ation The diesel engine was patented in1892 by R udolf Diesel, who conceived it as a prim e

m over that would provide m u ch im proved fuel efficiency com p ared with spark-ignition (SI) engines. To the present day, the diesel engine' s excellent fuel econom y rem a ins one of its strongest selling points. In the United States, the diesel engine is used m a inly in trucks, buses, agricultural and other nonroad equipm ent, locom o tives, and ships. The chief advantages of the diesel engine over the gasoline engine are its fuel econom y and durability. Diesel engines, however, em it m o re PM per m ile driven com p ared with gasoline engines of a sim ilar weight. Over the past decade, m odifications of engine com ponents have substantially reduced particle em issions from both diesel and gasoline engines (Ham m e rle et al., 1994; Sawyer and Johnson, 1995). The diesel engine com p resses air to high pressure and tem p erature. Fuel, when injected
into this com p ressed air, autoignites, releasing its chem ical energy. The expanding com bustion gases do work on the piston before being exhausted to the atm o sphere. Power output is controlled by the am ount of injected fuel rather than by throttling the air intake. Com p ared to its SI counterpart, the diesel engine' s superior efficiency derives from a higher com p ression ratio and no part-load throttling. To ensure structural integrity for prolonged reliable operation at the higher peak pressures brought about by a higher com p ression ratio and autoignition, the structure of a diesel engine generally is m o re m a ssive than its SI counterpart. Diesel engines (also called com p ression-ignition) m a y be broadly identified as being either two- or four-stroke cycle, injected di rectly or indirectly, and naturally aspirated or supercharged. They also are classified accordi ng to service requirem e nts such as light-duty (LD) or heavy-duty (HD) autom o tive/truck, sm all or large industrial, and rail or m a rine. All diesel engines use hydraulic fuel injection in one form or another. The fuel system

m

u st m eet four objectives if a diesel engine is to function properly over its entire operating

range. It m u st: (1) m e ter the correct quantity of f

u el, (2) distribute the f u el to the correct cylinder, (3) inject the fuel at the correct tim e, and (4) inject the fuel so that it is atom ized and

m i xes well with the in-cylinder air. The first two objectives are functions of a well-designed injection pum p, and the last two are m o stly functions of the injection nozzle. Fuel injection

2-3system s are m oving toward the use of electronic com ponents for m o re flexible control than is available with purely m echanical system s to obtain lower exhaust em issions without dim i nishing fuel efficiency.
Both the f u el and the lubricants that service diesel engines are highly finished petroleum

- based products com b ined with chem ical additives. Diesel fuel is a m i xture of m a ny different hydrocarbon m o lecules from about C 7 to about C 35 , with a boiling range from roughly
350 °F to 650 °F. Many of the fuel and oil proper ties, such as specific energy content (which is higher than gasoline), ignition quality, and specific gravity, are related to hydrocarbon

com position. Therefore, fuel and lubrican t com position affect m a ny aspects of engine perform a nce, including fuel econom y and exhaust em issions.
Com p lete and incom p lete com bustion of fuel in the diesel engine results in the form ation
of a com p lex m i xture of gaseous (gas-phase hydrocarbons, CO, CO 2 , NO, NO 2 , SO 2 ) and particulate exhaust (carbonaceous m a tter, sulfate, and trace elem ents). Because of concerns over health effects associated with DE, EPA began regulating em issions from diesel engines in 1970 (for sm oke) and then added regulations for gase ous em issions. EPA first regulated particulate

em issions from HD diesels in 1988.

2.2.1.1. Dieselization of the On-Road Fleet Because of their durability and fuel econom y, the use of diesel engines, particularly in long-distance applications, has increased over the years. The Census of Transportation, Truck Inventory and Use Survey (TIUS) indicates that am ong Class 3-8 trucks, diesel engine use has
increased m o re rapidly than gasoline engine use in the past 20 years. Truck classes are defined by gross vehicle weight as described in Table 2-1. Dieselization first occurred am ong Class 7
and 8 trucks. The TIUS indicates that 81.5% of diesel trucks on the road in 1963 were Class 7 or 8 trucks (Table 2-2). Class 7 sales becam e predom inantly (>50%) diesel in the 1970s and Class
8 sales becam e predom inantly diesel in the 1960s. Diesels did not m a ke up a m a jority of class 5 and 6 sales until the 1990s (Figures 2-1 and 2-2). HD trucks have historically constituted the

m a jority of diesel sales and m ileage. However, an increasing num ber of LD diesel trucks have
been sold dom estically in recent years. In the 1990s, approxim a tely one in three diesel trucks
sold was a Class 1 or Class 2 vehicle. Diesel truc ks have historically been driven m o re m iles per truck than gasoline trucks. For exam ple, the TIUS indicates that 59% of diesel trucks were
driven m o re than 50,000 m iles in 1963, com p ared with 3% of gasoline trucks.

2-41

2

3

4

5

6

7 Table 2-1. Vehicle classification and w e ights for on-road trucks

Class G ross vehicle w e ight (lb)

< 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

8A a 33,001–60,000

8B a >60,000
Medium duty (MD) 10,001–19,500 (sam e as Classes 3–5) Light-heavy duty (LHD) 19,501–26,000 (sam e as Class 6) Heavy-heavy duty (HHD) >26,001 (sam e as Class 7–8)

a Class 8A and Class 8B are often considered together.

Table 2-2. Total (gas and diesel) diesel trucks in the fleet in 1992 Truck class 1992 gas and diesel trucks 1992 diesel trucks % Diesels

Class 1 and 2 (Light duty) 55,193,300 1,387,600 3

Class 3, 4, and 5 (Medium duty) 1,258,500 326,300 26

Class 6 (Light heavy-duty) 732,300 273,800 37

Class 7 and 8 (Heavy heavy-duty) 2,016,600 1,725,300 86

Source: Census of Transportation, 1995.

2-50 20

, 0 0 0 40

, 0 0 0 60

, 0 0 0 80

, 0 0 0 10 0, 0 0 0 12 0, 0 0 0 14 0, 0 0 0 16 0, 0 0 0 18 0, 0 0 0 20 0, 0 0 0

19 30 19 40 19 50 19 60 19 70 19 80 19 90 20 00Number Sold C l a ss 5
C l a ss 6
C l a ss 7
C l a ss 8

Figure 2-1. Diesel truck sales (domestic) for the years 1939-1997.
Source: AAMA, 1927-1974 and 1975-1998.

0. 0 % 20

. 0 % 40

. 0 % 60

. 0 % 80

. 0 % 10 0. 0%

19 30 19 40 19 50 19 60 19 70 19 80 19 90 20 00% Diesel Cl a s s 5
Cl a s s 6
Cl a s s 7
Cl a s s 8

Figure 2-2. Diesel truck sales as a percentage of total truck sales for the years 1939-1997.
Source: AAMA, 1927-1974 and 1975-1998.

2-6Am ong com b ination trucks, consisting of tractor-trailers and single-unit trucks with trailers,
diesel vehicles have driven a m a jority of the m iles since at least 1963, the first year in which TIUS was conducted (Figure 2-3).

Figure 2-3. Percentage of truck miles attributable to diesel trucks. VMT= vehicle miles traveled. Source: U.S. Bureau of the Census, 1999b.

The longevity of diesel trucks is an im portant factor to understand past, current, and projected exposures to DE because older vehicles are subject to less stringent regulations and

m a y rem a in in use for several decades after their m a nufacture. Am erican Autom obile Manufacturers Association publications (AAMA, 1927-1997) indicate that 53% of trucks from

m odel years 1947-1956 were still on the road after 14 years. The proportion of trucks in use after 14 years was 63% for m odel years 1974-1983, suggesting that the lifespan of trucks built in later years is longer. According to the 1997 TIUS, vehicles older than 10 years m a de up 40% of
Class 7 and 8 trucks and 16% of Class 7-8 vehicl e m iles traveled (VMT) (Figures 2-4 and 2-5).
Alm o st all Class 7 and 8 trucks were diesel vehicles in the period 1982-1997 (93% in 1982 and 99% in 1997).

2.2.1.2. Dieselization of Railroad Locomotive Engines Early in the 20th century the political and econom ic pressure on the railroads to replace

steam locom o tives was substantial. Railroads were losing business to other form s of transport.
The diesel-electric locom o tive provided 90% in-service tim e, com p ared with only 50% for steam
locom o tives, and had three tim es the therm a l efficiency (Klein, 1991; Kirkland, 1983).

2-7Figure 2-4. Model year distribution of in-use HD truck fleet in 1997. Source: U.S. Bureau of the Census, 1999b.

Figure 2-5. Model year distribution of vehi cle miles traveled by the in-use HD truck fleet in 1997. Source: U.S. Bureau of Census, 1999b.

2-8Additionally, several cities had passed laws barring steam locom o tives within the city lim its because the large quantities of sm oke obscured visibility, creating a safety hazard. The first prototype diesel locom o tive was com p leted in 1917. By 1924 General Electric (GE) was producing a standard line of switching locom o tives on a production basis. Electro-Motive Corporation was founded the sam e year to produce diesel locom o tives in com p etition with GE.
This com p any was purchased in 1929 by General Motors (GM) and becam e the Electro-Motive
Division. After this acquisition, GM began to deve lop the two-stroke engine for this application. Up to this tim e, all locom o tive diesel engines were four-stroke. Two-strokes offered a m u ch higher power-to-weight ratio, and GM’s strategy was to get a large increase in power by m oving to the two-stroke cycle. The first true high- speed, two-stroke, diesel-electric locom o tives were produced by GM in 1935. However, because of the econom ic clim ate of the Great Depression,
few of these were sold until after the Second W o rld W a r. At the end of the war, m o st locom o tives were still steam -driven but were m o re than 15 years old, and the railroads were ready to replace the entire locom o tive fleet. Few, if any, steam locom o tives were sold after 1945 because the entire fleet was converted to diesel (Coifm an, 1994).
The locom o tive fleet has included significant percentages of both two- and four-stroke engines. The four-stroke diesel engines were naturally aspirated in the 1940s and 1950s. It is unlikely that any of the two-stroke engines used in locom o tive applications were strictly naturally aspirated. Nearly all two-stroke diesel locom o tive engines are uniflow scavenged, with a positive-displacem ent blower for scavenging assistance. In 1975, it was estim ated that 75% of
the locom o tives in service were two-stroke, of which about one-half used one or m o re turbochargers in addition to the existing positive-displacem ent blower for additional intake boost

pressure. Alm o st all of the four-stroke locom o tive engines were naturally aspirated in 1975. Electronic fuel injection for locom o tive engines was first offered in the 1994 m odel year (U.S. EPA, 1998b). All locom o tive engines m a nufactured in recent years are turbocharged, aftercooled or intercooled four-stroke engines. In part, this is because of the som e what greater durability of four-strokes, although im pending em issions regulations m a y have also been a factor
in this shif t. The typical lif espan of a locom o tive has been estim ated to be m o re than 40 years (U.S. EPA, 1998b). Many of the sm aller railroads are still using engines built in the 1940s, although the engines m a y have been rebuilt several tim es since their original m a nufacture.

2.2.2. Diesel Combustion and Fo rmation of Primary Emissions A basic understanding of diesel com bustion processes can assist in understanding the

com p lex factors that influence the form ation of DPM and other DE em issions. Unlike SI

com bustion, diesel com bustion is a fairly nonhom ogenous process. Fuel is sprayed at high

2-9Figure 2-6. HD diesel truck engines. these applications by the early 1980s. A comparison of IDI (A) and DI (B) combustion systems of high-speed DI engines almost completely replaced IDI engines for pressure into the com p ressed cylinder contents (prim a rily air with som e residual com bustion products) as the piston nears the top of the com p ression stroke. The turbulent m i xing of fuel and air that takes place is enhanced by injection pressure, the orientation of the intake ports (inducem ent of intake-swirl tangential to the cylinder wall), piston m o tion, and piston bowl shape. In som e cases, fuel and air m i xing is induced via injection of the fuel into a turbulencegenerating pre-cham ber or swirl cham ber located adjacent to the m a in cham ber (prim a rily in older, higher speed engines and som e LD diesels)

. Exam ples of typical direct injection and indirect injection com bustion system s are com p ared in Figure 2-6. Diesel com bustion can be considered to consist of the following phases (Heywood, 1988; W a tson and Janota, 1982):

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. (IDI = indirect injection, DI=direct injection)

2-10• A n ignition delay period, which starts after the initial injection of fuel and continues
until the initiation of com bustion. The delay period is governed by the rate of fuel
and air m i xing, diffusion, turbulence, heat transfer, chem ical kinetics, fuel vaporization, and fuel com position. Fuel cetane rating is an indication of ignition delay.

• R apid, prem ixed burning of the fuel and air m i xture from the ignition delay period.

• D iffusion-controlled burning, where the fuel burns as it is injected and diffuses into the cylinder.

• A very sm all am ount of rate-controlled burning during the expansion stroke, after the end of injection.

Engine speed and load are controlled by the quantity of f

u el injected. Thus, the overall fuel-to-air ratio varies greatly as engine speed and load vary. On a m acro scale, the cylinder contents are always fuel-lean. Depending on the tim e available for com bustion and the proxim ity of oxygen, the fuel droplets are either com p letely or partially oxidized. At

tem p eratures above 1,300 K, m u ch of the unburned fuel that is not oxidized is pyrolized (stripped of hydrogen) to form EC (Dec and Espey, 1995). In addition to EC, other
carbonaceous m a tter is present, largely from unburned fuel. The agglom eration of elem ental and
OC form s particles that are frequently referred to as “soot” particles. In this docum ent, the term s “EC” and “OC” are used to refer to the carbon-containing com ponents of DPM, and collectively, they are referred to as the carbonaceous fraction of a diesel particle. Carbonaceous particle form ation occurs prim arily during the diffusion-burn phase of

com bustion, and is highest during high load and ot her conditions consistent with high fuel-air
ratios. Most of the carbonaceous m a tter form ed (80% to 98%) is oxidized during com bustion,

m o st likely by hydroxyl radicals (Kittelson et al., 1986; Foster and Tree, 1994). DPM is defined by the m easurem ent procedures sum m a rized in the Code of Federal Regulations, Title 40 CFR, Part 86, Subpart N (C FR 40:86.N). These procedures define DPM

em issions as the m a ss of m a terial collected on a f ilter at a tem p erature of 52 °C or less af ter dilution of the exhaust with air. DPM is form ed by a num ber of physical processes acting in
concert as the exhaust is cooled and diluted. These are nucleation, coagulation, condensation, and adsorption. The core DE particles are form ed by nucleation and coagulation from prim ary spherical particles consisting of solid carbonaceous (EC) m a terial and ash (trace m e tals and other
elem ents). To these, through coagulation, adsorption, and condensation, are added organic and
sulfur com pounds (sulfate) com b ined with other condensed m a terial (Figure 2-7). Because of

2-11

### S o lid C a rb on a ceou s /A s h Pa rt icle wi t h a d sor b e d hy dr oc ar bo n/ sul f ate l a y e r

S u l fu ric A c id Part icles

H y dr o c a r bon/ Sul fa te P a r t i c l e s

0.2 μ m Figure 2-7. Schematic diagram of diesel engine exhaust particles.

Source: Modified from Kittelson, 1998. their size, <0.5 m m , these particles have a very large surface area per gram of m a ss, which

m a kes them able to adsorb large quantities of ash, organic com pounds, and sulfate. The specific
surface area of the EC core has been m easured to be approxim a tely 30–50 m 2 /g (Frey and Corn, 1967). Pierson and Brachaczek (1976) report that after the extraction of adsorbed organic

m a terial, the surface area of the diesel particle core is approxim a tely 90 m 2 /g. The organic m a terial associated with diesel particles originates from unburned fuel, engine lubrication oil, and sm all quantities of partial com bustion and pyrolysis products. This is frequently quantified as the SOF, which is discussed in m u ch m o re detail in Section 2.2.7. The
form ation of sulfate in DE depends prim arily on fuel sulfur content. During com bustion, sulfur

com pounds present in the fuel are oxidized to sulfur dioxide (SO 2 ). Approxim a tely 1% to 4% of
fuel sulfur is oxidized to form sulfuric acid (H 2 SO 4 ) (W all et al., 1987; Khatri et al., 1978; Baranescu, 1988; Barry et al., 1985). Upon cooling, sulfuric acid and water condense into an
aerosol that is nonvolatile under am bient conditions. The m a ss of sulfuric acid DPM is m o re than doubled by the m a ss of water associated with the sulfuric acid under typical DPM

m easurem ent conditions (50% relative hum idity, 20–25 °C) (W all et al., 1987).

2-12Em issions from com bustion engines produce oxide of nitrogen (NO

x ) prim arily (at least initially) as of NO. High com bustion tem p eratures cause reactions between oxygen and nitrogen
to form NO and som e NO 2 . Most NO 2 form ed during com bustion is rapidly decom posed. NO
can also decom pose to N 2 and O 2, but the rate of decom position is very slow (Heywood, 1988;
W a tson and Janota, 1982). Thus, alm o st all of the NO

x em itted is NO.
Som e organic com pounds from unburned fuel a nd from lubricating oil consum ed by the engine can be trapped in crevices or cool spot s within the cylinder and thus are not sufficiently available to conditions that would lead to their oxidation or pyrolysis. These com pounds are

em itted f r om the engine and either contribute to gas-phase organic em issions or to DPM

em issions, depending on their volatility. W ithin the exhaust system , tem p eratures are suf f i ciently high that these com pounds are entirely present within the gas phase (Johnson and Kittelson, 1996). Upon cooling and m i xing with am bient air in the exhaust plum e, som e of the less volatile organic com pounds can adsorb to the surfaces of the EC agglom erate particles. Lacking sufficient EC adsorption sites, the organic com pounds m a y condense on sulfuric acid nuclei to
form a heterogeneously nucleated organic aerosol (Abdul-Khalek et al., 1999). Although not unique to DE, the high content of EC associated with typical DPM

em issions has long been used by som e investigator s to distinguish diesel engine sources of this particle from other com bustion aerosols. Dies el particles from newer HD engines are typically

com posed of ~75% EC (EC can range from 33% to 90%), ~20% OC (OC can range from 7% to 49%), and sm all am ounts of sulfate, nitrate, trace elem ents, water, and unidentified com ponents (Figure 2-8). Metallic com pounds from engine com ponent wear, and from com pounds in the fuel and lubricant, contribute to DPM m a ss. Ash from oil com bustion also contributes trace

am ounts.
Am bient PM 2.5 m easured in the eastern United States is dom inated by sulfate (34%),
whereas am bient PM 2.5 in the western United States is dom inated by OC (39%) (Table 2-3) (U.S. EPA, 1999a). Many sources contribute to am bient PM 2.5 , and these sources and their relative
contribution to am bient PM 2.5 can be identified on the basis of the chem ical species present. The OC fraction of DPM is increasingly being used to assist investigators in identif ying the contribution of diesel engine em issions to am bient PM 2.5 . In particular, hopane and sterane

com pounds (arom a tic com pounds, >C 30 ) have been used in addition to other polycyclic arom atic hydrocarbons (PAHs) and long-chain alkanes to di stinguish DPM from other m obile source PM
and from am bient PM (Schauer et al., 1996; Fujita et al., 1998). Although PAH com pounds

m a ke up 1% or less of DPM m a ss, diesel em issions have been observed to have elevated
concentrations of m e thylated naphthalenes a nd m e thylated phenanthrene isom ers com p ared to
other com bustion aerosols (Benner et al., 1989; Lowenthal et al., 1994; Rogge et al., 1993).
Enrichm e nt of benzo[ a ]anthracene and benzo[ a ]pyrene (B[ a ]P) in DPM has also been

2-13Figure 2-8. Typical chemical composition for diesel particulate matter (PM 2.5 ) from new (post-1990) HD diesel vehicle exhaust.

Table 2-3. Typical chemical composition of fine particulate matter

Eastern U.S. Western U.S. Diesel PM 2.5
Elemental carbon 4% 15% 75%

OC 21% 39% 19%
Sulfate, nitrate, ammonium 48% 35% 1%

Minerals 4% 15% 2%

Unknow n 23% – 3 %

Source: U.S. EPA, 1999a.

2-14observed under som e conditions and has been used to assess the relative contribution of DE to

am bient PM. Although specific OC species are being used to help distinguish DPM aerosols from
other com bustion aerosols, up to 90% of the organic fraction associated with DPM is currently
classified as unresolvable com p lex m a terial. Ultrafine DPM (5–50 nm ) accounts for the m a jority (50% to 90%) of the num ber of particles but only 1% to 20% of the m a ss of DPM. A study conducted by Gertler (1999) in the Tuscarora Mountain tunnel dem onstrated an increase in 20

nm diam eter particles as the fraction of diesel vehicles in the tunnel increased from 13% to 78%.
The contribution of nuclei-m ode particles from a freeway on an am bient aerosol size distribution was reported by W h itby and Sverdrup (1980).
In sum m a ry, four m a in characteristics of DP M are (1) the high proportion of EC, (2) the large surface area associated with the carbonaceous particles in the 0.2 

### : m size range, (3)
enrichm e nt of certain polycyclic organic com pounds, and (4) 50%–90% of the num ber of DPM particles in diesel engine exhaust are in the nuclei-m ode size range, with a m ode of 20 nm

.

2.2.3. Diesel Emission Standards and Emission Trends Inventory
EPA set a sm oke standard for on-road HD diesel engines beginning with the 1970 m odel year and added a carbon m onoxide (CO) standard and a com b ined hydrocarbon (HC) and NO

x standard for the 1974 m odel year (Table 2-4). Beginning in the 1979 m odel year, EPA added a HC standard while retaining the com b ined HC and NO

x standard. All of the testing for HC, CO,
and NO

x was com p leted using a steady-state test procedure. Beginning in the 1985 m odel year,EPA added a NO

x standard (10.7 g/bhp-hr), dropped the com b ined HC and NO

x standard,
and converted from steady-state to transient testing for HC, CO, and NO

x em issions. EPA introduced a particulate standard for 1988 m odel year diesel engines using the transient test (0.6 g/bhp-hr). Transient testing involves running an engine on a dynam o m e ter over a range of load and speed set points.
Since the 1985 m odel year, only the NO

x and particulate standards have been tightened for on-road diesel engines. For truck and bus engines, the particulate standard was reduced to 0.25 g/bhp-hr in 1991, and it was reduced again in 1994 for truck engines to 0.1 g/bhp-hr. For urban bus engines, the particulate standard was reduced in 1994 to 0.07 g/bhp-hr and again in 1996 to 0.05 g/bhp-hr. The NO

x standard was reduced to 4.0 g/bhp-hr in 1998 for all on-road diesel engines (bus and truck engines). The standards for nonm ethane hydrocarbon (NMHC)
and NO

x com b ined were further lowered in a 1997 rulem a king, to take effect in 2004. EPA has recently finalized a regulation that will further reduce NO x, NMHC, and PM em issions from diesel engines starting in 2007.

2-15Table 2-4. U.S. emission standard s: HD highw ay diesel engines
Model

year Pollutant (g/bhp-hr) Smok e a

HC CO NO x HC + NO x Particulate (PM) t=truck , b=bus, ub=urban bus 1970 — — - — - — - — A:40%; L:20%

1974 — 4 0

— - 1 6 b
— A :20%; L:15%; P:50%

1979 1.5 2 5

— - 1 0 b
— A :20%; L:15%; P:50%

1985 c 1.3 15.5 10.7 — — A :20%; L:15%; P:50%

1988 1.3 15.5 10.7 — 0.60 A:20%; L:15%; P:50%

1990 1.3 15.5 6 .0 — 0 .60 A :20%; L:15%; P:50%

1991 1.3 15.5 5 .0 — 0 .25 A :20%; L:15%; P:50%

1993 1.3 15.5 5 .0 — 0 .25 t, 0.10 b A :20%; L:15%; P:50%

1994 1.3 15.5 5 .0 — 0 .10 t, 0.07 ub A:20%; L:15%; P:50%

1996 1.3 15.5 5 .0 — 0 .10 t, 0.05 ub A:20%; L:15%; P:50%

1998 1.3 15.5 4 .0 — 0 .10 t, 0.05 ub A:20%; L:15%; P:50%

2004 1.3 15.5 — 2.4 NMHC d 0.10 t, 0.05 ub A:20%; L:15%; P:50%

2007 15.5 0 .2 0.14 NMHC 0.01 A:20%; L:15%; P:50%

a Em issions m easured in percent opacity during differe nt operating m odes: A=acceleration; L=lug; P=peaks during either m ode.

b Total HC.

c In 1985, test cycle changed from steady-state to transient operation for HC, CO, and NO x m easurem ent and in 1988 for PM.

d Or 2.5 plus a lim it of 0.5 nonm ethane hydrocarbon (NMHC).

In Decem ber 1997, EPA adopted em ission standards for NO

x , HC, CO, PM, and sm oke for newly m a nufactured and rem a nufactured railroad locom o tives and locom o tive engines. The
rulem a king, which took effect in the year 2000, applies to locom o tives originally m a nufactured in 1973 or after, and any tim e they are m a nufactured or rem a nufactured (locom otives originally

m a nufactured before 1973 are not regulated). Three sets of em ission standards have been adopted (Tier 0, 1, and 2); they apply to locom o tives and locom o tive engines originally

m a nufactured from 1973 through 2001 (Tier 0), from 2002 through 2004 (Tier 1), and in 2005 and later (Tier 2) (Table 2-5; see EPA web page at http://www.epa.gov/om swww/ or http://www.dieselnet.com / standards/ for current inform ation on m obile source em ission standards). The em issions are m easured over two steady-state test cycles that represent two

2-16Table 2-5. U.S. emission standards: locomotives (g/bhp-hr)

Year a CO HC NO x PM

Line-haul 1973-2001 (Tier 0) 5.0 1 .0 9.5 0 .6

Switch 1973-2001 (Tier 0) 8.0 2 .1 14.0 0 .72

Line-haul 2002-2004 (Tier 1) 2.2 0 .55 7 .4 0.45

Switch 2002-2004 (Tier 1) 2.5 1 .2 11.0 0 .54

Line-haul 2005 + (Tier 2) 1.5 0 .3 5.5 0 .20

Switch 2005 + (Tier 2) 2.4 0 .6 8.1 0 .24

a Date of engine m a nufacture.

different types of service, including line-haul (long-distance transport) and switch (involved in all transfer and switching operations in switchyards) locom o tives.
Em ission standards for nonroad equipm ent are not as stringent as current standards for on-road equipm ent and are being phased in within the next decade. Currently, Federal PM standards exist for nonroad equipm ent of several horsepower ratings. For equipm ent between 175 and 750 horsepower, the PM standard was set at 0.4 g/bhp-hr in 1996 and will decrease to 0.15 g/bhp-hr between 2001 and 2003 depending on the power rating (Table 2-6). This equipm ent includes construction, agricultural, and industrial such as bulldozers, graders, cranes,
and tractors. The current PM standard for this equipm ent is only slightly lower than the 0.6 g/bhp-hr PM standard in place for on-road HD diesel engines in the late 1980s.
The EPA em ission trends report (U.S. EPA, 2000a) provides em ission inventories for criteria pollutants (PM 10 , PM 2.5 , SO 2 , NO
x , volatile organic com pounds [VOC], CO, Pb, and
NH 3 ) from point, area, and m obile sources, whic h indicate how em issions have changed from
1970 to 1998. The em ission trends are based on the EPA m obile source inventory m odels MOBILE, PART5, and the draft NONROAD m odel. PART5 derives particulate em ission rates for HD diesel vehicles using data generated for new engine certification purposes. PART5 is currently being m odified to account for deterioration, in-use em issions, poor m a intenance, and

tam p ering effects, all of which would increase em ission factors. PM, SO 2 , NO
x , and VOC

em issions trends from the report are discusse d below. Am bient urban/suburban PM sam p les rarely reflect the large fraction of natural and m i scellaneous sources suggested by the national inventory, owing to rem oval of a large portion of these em issions close to their sources as well as dispersion from these sources to urban/suburban sites. The rem oval of natural and m i scellaneous
PM 10 (largely fugitive dust) near their source is a result of the lack of inherent therm a l buoyancy, low release height, and interaction with their surroundings (im p action and filtration by vegetation).

2-17Table 2-6. U.S. emission standards fo r nonroad diesel equipment (g/bhp-hr)

Power rating Model year Pollutant (g/bhp-hr) S mok e % a

HC CO NO x NMHC + NO x PM

11 < hp 2000 — 6 .0 — 7 .8 (ABT) 0 .74 (ABT)

2005+ — 6 .0 — 5 .6 (ABT) 0 .60 (ABT)

11 # hp < 25 2000 — 4 .9 — 7 .0 (ABT) 0 .60 (ABT)

2005+ — 4 .9 — 5 .6 (ABT) 0 .60 (ABT)

25 # hp < 50 2000 — 4 .1 — 7 .0 (ABT) 0 .60 (ABT)

2005+ — 4 .1 — 5 .6 (ABT) 0 .44 (ABT)

50 # hp <100 1998+ — — 6.9 (ABT) — — 20/15/50

2004 — 3 .7 5.6 (ABT) 0 .30 (ABT)

2008+ — 3 .7 3.5 (ABT) —

100 # hp <175 1997+ — — 6.9 (ABT) — — 20/15/50

2003 — 3 .7 — 4 .9 (ABT) 0 .22 (ABT)

2007+ — 3 .7 — 3 .0 (ABT) —

175 # hp < 750 1996+ 1.0 8 .5 6.9 (ABT) — 0.4 20/15/50

175 # hp < 300 2003 — 2 .6 — 4 .9 (ABT) 0 .15 (ABT)

2006+ — 2 .6 — 3 .0 (ABT) —

300 # hp < 600 2001 — 2 .6 — 4 .8 (ABT) 0 .15 (ABT)

2006+ — 2 .6 — 3 .0 (ABT) —

600 # hp < 750 2002 — 2 .6 — 4 .8 (ABT) 0 .15 (ABT)

2006+ — 2 .6 — 3 .0 (ABT) —

$ 750 hp 2000+ 1.0 8 .5 6.9 (ABT) — 0.4 20/15/50

2006+ — 2 .6 — 4 .8 (ABT) 0 .15 (ABT)

a Em issions m easured in percent opacity during dif f e rent operating m odes: accel eration/lug/peaks during either m ode.
ABT = aver age banking and tr ading

. Note: The standards for engines less than 50 hp also apply to diesel m a r i ne engines

.

2-18�������������������������������

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������������������������������� ������������������������������� ������� For the sum m a ries presented here, natural and m i scellaneous sources are excluded from the
national PM and NO

x inventories.
From 1970 to 1998, PM 10 em issions decreased from slightly over 12,200,000 tons to just over 2,800,000 tons (Figure 2-9). PM 10 em issions from on-road and nonroad diesel engines
increased from 320,000 tons to m o re than 521,000 tons during this sam e period, so that in 1970 diesel engine em issions were 3% of the PM 10 inventory whereas in 1998, diesel engine em issions
were 18% of the PM 10 inventory. Diesel engines also contribute to secondary PM form ation
from NO
x and SO 2 em issions that are converted to nitrate and sulfate. VOCs from diesel engines also contribute to secondary organic particle fo rm ation. The contribution of secondary PM is not
included in the national trends inventories cited here. Mobile sources of PM include both gasoline- and diesel-powered on-road vehicles and a variety of nonroad equipm ent. Nonroad diesel engine sources include construction equipm ent, agricultural equipm ent, m a rine vessels, locom o tives, and other sources. The EPA em ission trends report (U.S. EPA, 2000a) indicates that, excluding natural and m i scellaneous sources,

m obile sources were responsible for 25% of PM 10 em issions in 1998. Diesel engines (on-road
and nonroad com b ined) were estim ated to contribute 72% of m obile-source PM 10 em issions.

� �������������������������� ���� � �������������������������� ���� � �������������������������� ���� � �������������������������� ���� � �������������������������� ���� � �������������������������� ���� � �������������������������� ���� � �������������������������� ���� � �������������������������� ���� � �������������������������� ���� � �������������������������� ���� � �������������������������� ���� � �������������������������� ���� � �������������������������� ���� � �������������������������� ���� � �������������������������� ���� � �������������������������� ���� � �������������������������� ���� � �������������������������� ���� � �������������������������� ���� ������������������������������� � ������������������������� ����� � ������������������������� ����� � ������������������������� ����� � ������������������������� ����� � ������������������������� ����� � ������������������������� ����� � ������������������������� ����� � ������������������������� ����� � ������������������������� ����� ������������������������������� � ������������������������� ����� � ������������������������� ����� � ������������������������� ����� ������������������������������� � ������������������������� ����� � ������������������������� ����� � ������������������������� ����� ������������������������������� 0 200 0 400 0 600 0 800 0 100 00 120 00 140 00

197 0 198 0 199 0 199 8

C a l e nda r Y e a

rThousands of short tons Di e s e l P M 1 0

������� O t

he r So ur c e s o f P M 1 0

Figure 2-9. Trends in PM 10 emissions from on-road and nonroad engines combined and other anthropogenic sources of PM 10 from 1970 to 1998 (excludes miscellaneous and natural sources).

Source: U.S. EPA, 2000a, National Air Pollutant Em ission Trends, 1900-1998.

2-19Because of the high concentration of fine partic les in engine em issions, diesel engines (on-road
and nonroad com b ined) were estim ated to contribute 77% of m obile-source PM 2.5 em issions and
23% of total PM 2.5 in 1998 (excluding natural and m i scellaneous em issions). If natural and

m i scellaneous PM 2.5 sources are included in the inventory, diesel PM 2.5 contributes 6% to the national inventory.
Gram per m ile particulate em issions from diesel vehicles are m u ch greater than those
from gasoline-fueled vehicles, accounting for the la rge contribution of diesel engine em issions to the national inventory in spite of the sm aller num ber of diesel engines in use. Particulate

em issions (PM 10 ) from gasoline-fueled engines decreased dram atically in 1975 with the widespread introduction of unleaded gasoline. Particulate em issions from diesel highway vehicles have decreased recently because of EPA em ission standards for new m odel year HD
diesel trucks that were first im plem ented in 1988 and becam e increasingly stringent in 1991, 1994, and 2000, as presented in Table 2-4. A decrease in on-road HD DPM em issions since the

m i d-1980s is confirm e d by in-use vehicle testing, as described in Section 2.2.5. Because of the
im plem entation of existing regulations, DPM em issions from on-road sources are expected to
decrease 37% from 1998 to 2007; however, nonroad DPM em issions are expected to increase
15% in the sam e period (Figure 2-10).
The EPA em ission trends report (U.S. EPA, 2000a) indicates that annual on-road vehicle
PM 10 em issions decreased from 397,200 tons to 257,080 tons from 1980 to 1998. 1 Passenger car particulate em issions decreased 53% (from 119,000 to 56,000 tons) in this tim efram e , while onroad diesel vehicle PM 10 em issions decreased 27% (from 208,000 to 152,000 tons) (Figure 2-10). Nonroad diesel engine PM 10 em issions increased 17% (from 314,000 tons in 1980 to 69,000 tons in 1998). Em issions data for PM 2.5 are available only for the period from 1990 to 1998. Between 1990 and 1998, PM 2.5 em issions from m obile sources decreased by 14%, largely as the
result of decreased on-road em issions.
From 1970 to 1998, NO
x em issions increased from 20,598,000 tons to 24,126,000 tons (Figure 2-11). NO
x em issions from on-road and nonroad diesel engines increased from 1,748,000 tons to 4,753,000 tons during this sam e peri od, so that in 1970 diesel engine em issions
were 8% of the NO

x inventory while in 1998, diesel engine em issions were 20% of the NO

x

1 Exhaust em issions constitute the m a jority of PM em issions from m obile sources, with tire and brake wear contributing the rem a inder. To com p are trends estim ates fro m past years with future projections (which are provided
for exhaust em issions only), the fraction of brake and tir e wear would need to be om itted from these estim ates as reported in the em ission trends report (U.S. EPA, 2000a). On average in the late 1990s 39% and 64% of gasoline vehicle particulate em issions originated from exhaust and 95% and 98% of on-road diesel em issions originated from
exhaust for PM 10 and PM 2.5 , respectively.

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0 19 70 19 80 19 90 19 98 20 07 20 30 � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� ����������������� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� ���������������� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� ���������������� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� ���������������� � ����������� ���� � ����������� ���� � ����������� ���� � ����������� ���� � ����������� ���� � ����������� ���� ���������������� � ���������� ���� � ���������� ���� � ���������� ���� � ���������� ���� � ���������� ���� � ���������� ���� � ���������� ���� � ���������� ���� � ���������� ���� � ���������� ���� � ���������� ���� � ���������� ���� � ���������� ���� ��������������� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� ���������������� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� ���������������� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� � ����������� ����� ����������������� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� ���������������� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� � ���������� ����� ����������������

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Figure 2-10. Trends in PM 10 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 Em ission Trends, 1900-1998. *Projection to 2030 includes im plem entation of the recently finalized regulation “Control of Air Pollution from New Motor Vehicles: Heavy-Duty Engine and Vehicle Standards and Highway Diesel Fuel Sulfur Control Requirem e nts” U.S. EPA, 2000b.

inventory. As m e ntioned above, som e of this NO

x will be converted to particulate nitrate in the

atm o sphere, and this contribution to am bient PM is not quantified in national inventories. In 1998, 53% of total em itted NO

x cam e from m obile sources, with diesels responsible
for 57% of the m obile-source contribution. Overall, NO

x em issions from m obile sources have

rem a ined relatively constant over tim e, increasing an estim ated 7% from 1980 to 1998. W h ereasThousands of short tons

2-21���������������������������������� ����������������������������������� ���������������������������������� ���������������������������������� �������

� ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� ���������������������������������� � ����������������������������� ����� � ����������������������������� ����� � ����������������������������� ����� � ����������������������������� ����� � ����������������������������� ����� � ����������������������������� ����� � ����������������������������� ����� � ����������������������������� ����� � ����������������������������� ����� � ����������������������������� ����� � ����������������������������� ����� � ����������������������������� ����� � ����������������������������� ����� � ����������������������������� ����� � ����������������������������� ����� � ����������������������������� ����� � ����������������������������� ����� � ����������������������������� ����� � ����������������������������� ����� � ����������������������������� ����� � ����������������������������� ����� � ����������������������������� ����� ����������������������������������� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� ���������������������������������� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� � ���������������������������� ����� ���������������������������������� 

### 0 5000 10000 15000 20000 25000 30000

1970 1980 1990 1998

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Figure 2-11. Trends in NO
x emissions from on-road and nonroad diesel engines combined and other anthropogenic sources of NO
x natural sources). from 1970 to 1998 (excludes miscellaneous and

Source: U.S. EPA, 2000a, National Air Pollutant Em ission Trends, 1900-1998.

NO
x from LD gasoline vehicles decreased from 1980 to 1998, resulting in an overall decrease in
on-road NO

x em issions of 9%, NO

x from diesel trucks and buses increased 7% (from 2,463,390 tons in 1980 to 2,630,120 tons in 1998), owing to the illegal use of electronic control devices that bypassed the trucks’ em ission control system s, as discussed in Section 2.2.5. NO
x em issions
from nonroad diesel engines (including com m e rcial m a rine and locom o tives) have increased 46% (from 3,251,600 tons in 1980 to 4,752,800 tons in 1998) (Figure 2-12).
About 7% of SO 2 cam e from m obile sources in 1998, with diesels responsible for 74% of
that total. EPA regulations for on-road diesel fuel sulfur content (which started in 1993) have significantly reduced SO 2 em issions from highway diesels. SO 2 em issions from highway diesel

2-22Figure 2-12. Trends in NO
x emissions from on-road and nonroad diesel engines from 1970 to 1998.

Source: U.S. EPA, 2000a, National Air Pollutant Em ission Trends, 1900-1998.

engines have decreased 72% (from 303,000 tons in 1980 to 85,000 tons in 1998) (Figure 2-13).
Sim ilar trends are not apparent for nonroad diesels, although in 1998 nonroad diesel engines, excluding com m e rcial m a rine vessels, em itted 785,000 tons of SO 2 , accounting for 56% of

m obile-source SO 2 em issions in 1998. Diesel engines are not a large source of VOC em issions com p ared with gasoline engines.
VOC em issions from diesel engines in 1998 were es tim ated at 2% of the total em issions from all

sources. VOC em issions from diesel m obile sources decreased 9% (from 779,000 tons in 1980 to 721,000 tons in 1998) (Figure 2-14). Diesel engines are also not a large source of CO em issions com p ared with gasoline engines. In 1998, m obile sources em itted 79% of all CO, and diesel engines accounted for 4% of
the m obile-source CO. CO em issions from on-road diesel vehicles increased 34% between 1980 and 1998, during which tim e nonroad diesel em issi ons of CO increased 45% (Figure 2-15).

2-23Figure 2-13. Trends in SO 2 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 em ission trends, 1900-1998.

Figure 2-14. Trends in VOC emissions from on-road and nonroad diesel engines from 1970 to 1998.

Source: U.S. EPA, 2000a, National air pollutant em ission trends, 1900-1998.

2-24Figure 2-15. Trends in CO emissions from on-road and nonroad diesel engines from 1970 to 1998.

Source: U.S. EPA, 2000a, National Air Pollutant Em ission Trends, 1900-1998.

2.2.4. Historical Trends in Diesel Fuel Use a nd Impact of Fuel Properties on Emissions Use of diesel fuel increased steadily in the second half of the 20 th century. According to
statistics from the Federal Highway Adm i ni stration (1995, 1997), in 1949 diesel fuel was approxim a tely 1% of the total m o tor fuel used, and in 1995 it was about 18%. Over the sam e tim e, diesel fuel consum ption in the United States increased from about 400 m illion gallons to 26 billion gallons per year, an increase by a f actor of m o re than 60 (Figures 2-16 and 2-17).
The chem istry and properties of diesel fuel ha ve a direct effect on em issions of regulated pollutants from diesel engines. Researchers have studied the NOx and DPM effect of sulfur content, total arom atic content, polyarom a tic content, fuel density, oxygenate content, cetane

num ber, and T90 on em issions of regulated pollutants. T90 is the 90% distillation point

tem p erature. An increase in T90 has been observed to cause an increase in DPM em issions (Cunningham et al., 1990; Sienicki et al., 1990). Cetane num ber is a m easure of the ignition quality, or ignition delay tim e, of a diesel fuel. The percent of cetane (less com m only referred to as hexadecane, C 16 H 34 ) by volum e in a blend with alpha-m ethylnaphthalene (C 10 H 7 CH 3 ) defines
the cetane num ber that provides the sam e ignition delay tim e as the fuel in use.

2-25������� �������� ��������� ����� �������� ���� ����� ���� �������� ����� �������� ���� ���� ����� ����� ���� �������� ����� �������� ��������� ���� �������� ��������� ����� �������� ���� ��������� ���� ��������� ����� �������� 20Thousands ofGallonsPercentage Diesel Fuel of Total Motor Fuel Use

��� ������� ���� ���� ���� ���� ���� ���� �������� ���� ���� �������� ���� ���� ���� ����� ����� ����� ����� ����� ����� ����� ����� ����� ��������� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ��������� ����� ����� ����� ����� ����� ����� ��������� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� �������� ���� ���� ���� ���� ���� ���� ���� ���� ��������� ����� ����� ����� ����� ����� ����� ����� ����� ��������� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ����

���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ��������� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ��������� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� �������� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ��������� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ��������� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� �������� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� �������� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ��������� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ��������� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� �������� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ��������� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ����� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� ���� 18

16

14

12

10

8

6

4

2

0 1949 1952 1955 1958 1961 1964 1967 1970 1973 1976 1979 1982 1985 1988 1991 1994

Figure 2-16. Percentage of total motor fuel use that is on-road diesel fuel since 1949.

Source: Federal Highway Adm i nistration, 1995.

30, 000, 000

25, 000, 000

20, 000, 000

15, 000, 000

10, 000, 000

5, 000, 000

0



### Y ear Figure 2-17. On-highw ay diesel fuel consumption since 1943, values in thousands of gallons.

Source: Federal Highway Adm i nistration, 1995.195119431955195319591957196319611967196519711969197519731979197719831981198719851991198919951993

2-26Before 1993, diesel fuel sulfur levels were not federally regulated in the United States, although the State of California had such regulations

. Industry practices that were in place (e.g., the ASTMD 975 specification for No. 2 oils) lim ited sulfur to 0.5%. During the years 1960 to 1986, fuel sulfur content showed no chronological increasing or decreasing trends and ranged
from 0.23 to 0.28 wt% (NIPER, 1986). A m a xim u m allo wable on-road diesel fuel sulfur content
in the United States was established at 0.05 m a ss % in 1993, in advance of the 1994 0.10 g/bhphr PM standard for HD on-highway trucks. Na tionally, on-road fuels averaged 0.032% sulfur in 1994 while nonroad fuels averaged 10-fold the sulfur level of on-road fuel, or 0.32% (Dickson
and Sturm
, 1994). The reduction in diesel fuel sulfur reduced total DPM m a ss em issions through reduction of sulfate PM (prim a rily present as sulfuric acid). Considerably higher sulfuric acid PM em issions are possible with DE aftertreatm e nt system s containing precious m e tals (oxidation catalysts, lean NO

x catalysts, catalyzed DPM traps). At tem p eratures over 350 °C to 500 °C (depending on device), SO 2 in the exhaust can be oxidized to sulfuric acid (McClure et al., 1992; McDonald et al., 1995; W a ll, 1998). Sulfur

content rem a ins at unregulated levels for off-hi ghway diesel fuels and fuels used in railroad
locom o tives.
The chem ical m a keup of diesel fuel has changed over tim e, in part because of new regulations and in part because of technological developm ents in refinery processes. EPA currently regulates on-road diesel fuel and requires the cetane index (a surrogate for actual

m easurem ents of cetane num ber) to be greater than or equal to 40, or the m a xim u m arom atic content to be 35% or less (CFR 40:80.29). EPA recently finalized a regulation that will lim it the
sulfur content of on-road diesel fuel to 15 ppm starting in 2006 (U.S. EPA, 2000b). California has placed additional restrictions on the arom atic content of diesel fuel (California Code of Regulations, Title 13, Sections 2281-2282) and requires a m i nim u m cetane num ber of 50 and an

arom atics cap of 10% by volum e, with som e exceptions for sm all refiners and alternative
form ulations as long as equivalent em issions are dem onstrated. Diesel fuel from larger refiners
is lim ited to 10% arom atic content, and for three sm all refiners (a sm all fraction of diesel sales)
to 20% arom atic content. The refiners can also certify a fuel with higher arom atic content as being em issions-equivalent to the 10% (or 20%) arom atic content fuels by perform i ng a 7-day engine dynam o m e ter em issions test. This m e thod is chosen by m o st, if not all, California refiners, and so a typical California diesel fuel has an arom atic content above 20%. Em issions equivalence has been obtained through use of cetane enhancers, oxygenates, and other proprietary additives. Nonroad diesel f u el is not regulated, and consequently, cetane index,

arom atic content, and sulfur content vary widely with nom inal values for cetane num ber around 43, 31% arom atics, and sulfur approxim a tely 3,000 ppm

.

2-27The average cetane num ber of U.S. diesel fuel declined steadily from 50.0 to 45.1, or about 0.2% per year, from 1960 to 1986 (NIPER, 1986). The decline in cetane num ber was likely accom p anied by an increase in arom atic content and density (Lee et al., 1998). A num ber of EPA-sponsored studies refer to fuels with nom inally 22% arom atics content as “national average fuel” during the 1970s (Hare, 1977; Springer, 1979), whereas by the 1980s a so-called national average fuel contained 30% arom a tics (Martin, 1981a,b). Shelton (1979, 1977) has reported a trend of increasing T90 from 1960 through the late 1970s, which is consistent with increasing density, arom atic content, and polyarom a tic content. Unfortunately, arom atic content

was not com m only m easured before the 1980s.
Studies m easuring the em issions im pact of changes in cetane num ber and arom atic content for roughly 1990 m odel year engine technol ogy find that increasing the arom atic content
from 20% to 30%, with an accom p anying decrease in the cetane num ber from 50 to 44, results in
a 2% to 5% increase in NO

x and a 5% to 10% increase in total DPM (McCarthy et al., 1992;
Ullm an et al., 1990; Sienicki et al., 1990; Gr aboski and McCorm ick, 1996). These ranges m a y be reasonable upper bounds for the effect of changes in fuel quality on NO

x and DPM em issions during the years 1960–1990. In the northern United States during wintertim e, on-road No. 2 diesel m a y contain som e percentage of No. 1 diesel to im prove cold-flow properties. Discussions with refiners indicate that a typical wintertim e No. 1 diesel blending level is 15 volum e %; however, this num ber m u st be taken as a rough estim ate. Blending of No. 1 m a y lower the arom atic content, resulting in
im proved em issions perform a nce. Nationally, on-highway No. 1 fuels averaged 17% arom atic content in 1994 (Dickson and Sturm
, 1994). Thus, there m a y also be som e sm all but perceptible seasonal changes in em issions from diesel engines. Railroad-grade diesel fuel is currently unre gulated. Typically, railroad-grade diesel fuel is a blend of approxim a tely 10% on-road fuel and 90% nonroad diesel fuel. There are no recent
data on the com position of railroad-grade diesel fuel. Som e what dated diesel fuel oil surveys (Shelton, 1979) reported that railroad-grade diesels had lower cetane num ber, higher density, and higher T90. Also, the cetane index for these fuels can be as m u ch as 9 cetane units higher
than the cetane num ber, an indication of a high arom atic content in railroad-grade diesels.
Fuel chem istry is also im portant for em ission of particle-associated PAHs. In studies perform e d over m o re than a decade, W illiam s a nd Andrews of the University of Leeds have
shown that the solvent-extractable PAHs from dies el particulate originate alm o st entirely in the fuel (W illiam s et al., 1987; Andrews et al ., 1998; Hsiao-Hsuan et al., 2000). The PAH

m o lecules are relatively refractory, so a significant fraction survive the com bustion process and
condense onto the DPM. These studies have been confirm e d by other research groups (Crebelli et al., 1995; Tancell et al., 1995). There is a consensus am ong these researchers that

2-28pyrosynthesis of PAHs occurs only at the highe st tem p erature operating conditions in a diesel engine. Under these conditions, m o st of the DPM and other pyrolysis products are ultim ately burned before exiting the cylinder. These results indicate that em issions of PAHs are m o re a function of the PAH content of the fuel than of engine technology. For a given refinery and crude oil, diesel fuel PAH correlates with total arom atic content and T90. Representative data

on arom atic content for diesel fuels in the United States do not appear to be available before the

m i d-1980s. However, the decreasing trend in cetane num ber, increasing trend in T90, and the increasing use of light cycle oil from catalytic cracking beginning in the late 1950s suggest that diesel PAH content has increased over the past 40 years. Because PAHs have been im plicated as one potential contributing com ponent to the observed toxicity of DE, changes in PAH content of
diesel fuel over tim e, as well as differences betw een diesel fuels used in different applications (on-road, nonroad, locom o tive), m a y influence the hazard observed in exposed populations from different occupations. However, such a relati onship would be difficult to differentiate in an epidem iologic study because there are several other properties of DE that m a y be contributing to the observed toxicity. Historical trends in P AH-m easured em issions are discussed in Section
2.2.8.2.

2.2.5. Chronological Assessment of Emission Factors
2.2.5.1. On-Road Vehicles
Num e rous studies have been conducted on em issions from in-use on-road HD diesel
vehicles. HD vehicles are defined as having a rated gross vehicle weight (GVW R) of greater than 8,500 lb, and m o st over-the-road trucks ha ve a GVW R of 80,000 lb. Em issions of regulated pollutants from these studies have been reviewed (Yanowitz et al., 2000); the review findings,
which encom p ass vehicles from m odel years 1976 to 1998, are sum m a rized below. In addition, a large am ount of engine dynam o m e ter data on HD diesel engines have been published since the

m i d-1970s. These data are used below to confirm and expand upon the findings from in-use vehicle testing. Figure 2-18 shows chassis dynam o m e ter data for m o re than 200 different vehicles (approxim a tely one-half of which are transit buses), reported in 20 different published studies, as well as a large am ount of additional data collected by W e st Virginia University (Yanowitz et al., 1999; W a rner-Selph and Dietzm ann, 1984; Dietzm ann et al., 1980; Graboski et al., 1998a,b;
McCorm ick et al., 1999; Clark et al., 1995, 1997; Bata et al., 1992; Brown and Rideout, 1996, Brown et al., 1997; Dunlap et al., 1993; Ferguson et al., 1992; Gautam et al., 1992; Katragadda

2-29ModelYear197519801985199019952000

PM Em i s s i o n s ( g / m i l e )0.1110ModelYear197519801985199019952000

N O x E m i s si o n s ( g / m i l e )10100slopeloModelYear197519801985199019952000

H C E m i s si o n s ( g / m i l e )0.1110ModelYear197519801985199019952000

C O E m i s si ons ( g / m i l e)110100Figure 2-18.x, HC, and CO emissions from HD diesel vehicles (g/mile).Source: Yanowitz et al., 2000.

2-30Model year trends in PM, NOet al., 1993; Rideout et al., 1994; W a ng et al ., 1993, 1994; W illiam s et al., 1989; W h itfield and Harris, 1998; W e st Virginia University data available on the W o rld W i de W e b at www.afdc.nrel.gov). The results from vehicles tested m o re than once using the sam e test cycle, and without any additional m ileage accum u lated between tests, are averaged and reported as one data point. Buses were tested using the Cent ral Business District (CBD) cycle, while m o st trucks were tested using the Urban Dynam o m e ter Driving Schedule (UDDS), also known as the Schedule 1d cycle. Som e of the trucks were tested using the W e st Virginia 5-peak cycle, which generates considerably lower g/m i em issions than the CBD or UDDS (Yanowitz et al., 1999).
Em issions results from vehicles tested under differe nt test cycles or at different points in the engine' s life cycle have been reported as separate data points. Note that all NO

x m a ss em issions data are reported as equivalent NO 2 . Table 2-7 com p ares the m a ke-up of the fleet of trucks that was tested with the in-use truck fleet according to the 1997 Vehicle Inventory and Use Survey (U.S. Bureau of the Census, 1999a). The tested fleet is m o stly vehicles in the 33,000-60,000 lb

range. Analysis of the tested fleet also shows that the m odel year distribution is skewed toward
newer vehicles. The 1997 Vehicle Inventory and Use Survey indicates a flat distribution with roughly the sam e num ber of in-use vehicles for each of the m odel years in the decade preceding
1997. The 1992 Truck Inventory and Use Survey (U.S. Bureau of the Census, 1995) shows the

sam e trend, as shown in Figure 2-1. Analysis of odom eter m ileage for the tested fleet shows that 45% of the vehicles had less than 50,000 m iles at the tim e of testing. Only 10% of the vehicles
had m o re than 250,000 m iles. Although the m ileage distribution of the in-use fleet is unknown,
it seem s unlikely to be as heavily weighted to low-m ileage vehicles. Because of the relatively
low m ileage of m o st of the vehicles tested, deteri oration of em issions m a y not be reflected in the

Table 2-7. Comparison of in-use truck fleet

w i th truck fleet tested on chassis dynamometer, percent of total vehicles

Class I n-use trucks, 1995 census Tested trucks

3 17.7 1

4 & 5 13.3 0

6 & 7 25.0 17

8A 20.9 52

8B 23.1 30

2-31results. Yanowitz and co-workers (2000) report that average em issions of regulated pollutants for vehicles of the different classes listed in Table 2-7 are approxim a tely the sam e

. This is clearly a reflection of the sm all num ber of vehicles in the lighter weight classes for this dataset,
but it also indicates no real difference in em issi ons for vehicles in Classes 6–8. The data are

m a inly for vehicles of 19,500 lb and greater GVW R (Classes 6 and 7 and heavier), and predom inantly for vehicles of 33,000 lb and greater GVW R (Class 8 trucks and buses). Figure 2-18 shows em issions trends in g/m i

. Least-squares linear regressions and 95% confidence intervals are plotted on each graph and yield the following equations for predicting

em issions trends (applicable to the years 1976–98):

Log NO
x (g/m ile) = (Model year *



### ! 0.008) + 16.519 R 2 = 0.024 (2-1) Log PM (g/m ile) = (Model year *



### ! 0.044) + 88.183 R 2 = 0.28 (2-2) Log HC (g/m ile) = (Model year *



### ! 0.055) + 109.39 R 2 = 0.27 (2-3) Log CO (g/m ile) = (Model Year *



### ! 0.041) + 82.876 R 2 = 0.22 (2-4)

As shown in Figure 2-18, changes in NO

x em issions have been relatively sm all, with an

em ission rate averaging about 26 g/m i

. The data reported in Figure 2-18 are real-world, in-use

em issions m easurem ents and therefore m o re accurately reflect em ission factors than engine test data during this period. There are two potential causes for the relative constancy of NO

x
em issions as described by Figure 2-18. The first is em issions deterioration due to engine wear.
W eaver and Klausm eier (1988) have shown that di esel engine deterioration results in lower NO

x
em issions and higher DPM em issions, and this finding has recently been confirm e d by
McCorm ick and co-workers (2000). W ear of m echanical devices that lim it sm oke, fuel pum ps, and f u el injectors alters the ef f ective injection tim ing to decrease NO

x . Because deterioration is

m o re a function of m a intenance than vehicle age or m ileage, deterioration introduces a wide range in NO

x em ission factors m easured in the chassis dynam o m e ter studies. The lack of a decreasing trend in NO

x em issions can also be attributed to the use of illegal em issions control devices that bypassed the trucks’ em ission control system s under som e driving conditions such as steady-state cruise. EPA has reached a settle m e nt with the diesel engine m a nufacturers to
discontinue use of these devices. The illegal devices produced low NO

x em issions on the transient test (HD FTP) but operated in a high-NO

x /high-fuel-econom y m ode in use under highway cruise conditions. Figure 2-19 shows engine certification data for NO

x em issions reported in the m a ny studies that have em ployed the transient test over the past 25 years. The engine testing data are
also listed in Table 2-8. The data com p iled in Figure 2-19 show a significant decline in NO

x

2-3214

12

10

8

6

4

2 19 75 19 80 19 85 19 90 19 95 20 00

E n gi n e M o d e l Y e ar

Figure 2-19. Diesel engine certification data for NO
x emissions as a function of model year.

Source: Data are from the transient test results provided in Table 2-8.

em issions, and all engines would appear to m eet the regulatory standards for their year of

m a nufacture because of the illegal em issions devices. From 1980 to 1997, the EPA em issions trends report (U.S. EPA, 1998a) predicted a decline in NO

x em issions from HD diesel vehicles because these data are based on engine test data. The em issions trend includes the growth in
vehicle m iles traveled over tim e as well as changes in em ission factors. The m o re recent trends inventory (U.S. EPA, 2000a, discussed earlier) includes em ission from the illegal em issions devices and accordingly dem onstrates a slight increase in NO

x em issions from on-road HD diesel
vehicles from 1990 to 1998. DPM, CO, and THC em issions, although widely variable within any m odel year, have shown a pronounced declining trend (Figure 2-18). DPM em issions from chassis dynam o m e ter tests decreased from an average of 3-4 g/m i in 1977 to an average of about 0.5 g/m i in 1997, suggesting a decrease in DPM em issions of a factor of about 6. Note that these data are for vehicles or engines tested on in-use or industry-average fuel at the tim e they were tested. Indications are that the observed decline in DPM is caused prim arily by changes in engineNOxEmission, g/bhp-h

2-33------Table 2-8.data from engine dynamometer testsReferenceEngineaYearTestbNOxg/bhphrPMg/bhphrCOg/bhphrTHCg/bhphrSOFg/bhphrSOFMethcTotalaldehyde,mg/bhp-hrB[a]Pug/bhp-hrd1-NP (NPAH)ug/bhp-hreHare, 1977Cat 3208 (NA)1976SS7.980.8714.041.110.103c-hexane0.76DDC 6V71 (blower)1976SS10.241.926.550.710.937c-hexane0.24Springer, 1979Mack ETAY(B)673A (DI,TC,AC)1977SS6.6130.611.5880.4760.098Benz/cyc652.23Cat 3208 (EGR, NA)1977SS3.7472.216.2001.163Benz/cyc1611.72Cat 3406 (DI, TC, AC)1977SS9.790.352.340.350.063Benz/cyc730.15Cat 3406 (DI, TC, AC, EGR)1977SS5.490.934.810.170.181800.08Cat 3406 (IDI, TC, AC)1977SS5.140.281.260.120.031Benz/cyc800.11DB OM-352A (DI, TC, AC)1977SS8.930.560.190Benz/cyc2800.87DB OM-352A (DI, NA)1977SS7.460.990.287Benz/cyc2801.07Perez, 1980Cat (DI, NA)1978SS8.120.775.920.770.19DCM1.08Cat (DI, EGR)1978SS5.161.215.370.570.079DCM4.34Cat (DI, TC, AC)1978SS7.660.332.200.270.037DCM0.34Martin, 1981aCat 32081978T7.831.06Cummins NTC3501976T11.410.81DDC 6V92T (2S)1978T9.550.72Cummins NTCC3501979T6.580.52DDC 8V71N (2S)1978T7.150.92DDC 6V92TA (2S)1979T7.800.65IH DTI466B1979T7.460.48Mack ETAY(B)673A1979T9.010.77Mack ETSX676-011980T6.900.85Cummins VTB-9031979T8.100.53Cat 34061979T11.280.69Cat 3406PCTA1979T7.240.49Cummins BigCamNTC3501979T9.970.54IH DT4661979T7.910.71DDC 6V92TA (2S)1979T11.660.73DDC 8V71TA (2S)1979T9.810.51Cummins NTC2901979T11.100.78Cummins NH-2501979T10.870.97Martin, 1981bCummins VTB-9031980T5.590.672.02.230.228DCMDDC 8V71TA (2S)1980T7.910.442.280.730.176DCMIH DTI466B1980T4.410.622.350.870.186DCMUllman et al., 1984DDAD 6V-71 (2S)1980T6.090.563.861.420.298DCM23Martin, 1984Cummins NTC3001981T8.130.452.701.36Barry et al., 1985Cat 3406B1985T6.580.482.10.50.061DCM701DDC 8V-92 TA (2S)1980T8.150.452.610.53Diesel engine emissions(PAH)

2-34 2-34----------------------------------------------|||Table 2-8.Diesel engine emissions data from engine dynamometer tests(continued)ReferenceEngineaYearTestbNOxg/bhphrPMg/bhphrCOg/bhphrTHCg/bhphrSOFg/bhphrSOFMethcTotalaldehyde,mg/bhp-hrB[a]Pug/bhp-hrd1-NP (NPAH)ug/bhp-hreSS6.640.361.830.380.0255Enga et al., 1985DDC 8V-71 TAC (2S)1984T5.851.262.991.48Baines, 1986Cummins NTCC-4001985T4.620.553.210.53Wachter, 1990Iveco 84601991T0.220.0957?McCarthy et al., 1992Navistar DTA466 ES2101993T4.930.0821.30.280.0237SFEPerez and Williams,Engine 11982T0.930.179DCM260.831989Engine 21982T0.860.145DCM5.8Engine 31982T0.590.185DCM4.90.89Engine 41982T0.960.325DCM261.2Engine 51982T1.060.076DCM5.3Engine 61982T0.880.344DCMNeedhamet al., 1989Average of 16 engines1988T0.370.12DCMAverage of 3 engines1991T0.240.10DCMKreso et al., 1998Cummins L10-3001988SS5.150.1030.260.030DCMCummins L10-3101991SS4.700.0350.0670.022DCMCummins M11-330E1995SS3.820.0370.160.016DCMBagley et al., 1998Cat 3304 (IDI, NA) non-road1983SS0.560.319Benz/cyc1.5(133)2.2Graboski, 1998bDDC 6V-71N-77 (MUI, 2S)1977T9.960.833.592.010.729DCM(and references therein)DDC 6V-92TA-91 (DDECII)1991T4.230.1971.510.720.0788?DDC-6V-92TA-87 (2S)1987T10.770.590.71DDC-6V92TA-83 (MUI, 2S)1983T5.620.2651.190.4350.133DCMDDC 6V-92TA -88 (DDECII, 2S)1988T8.520.21.60.60.116Tol/EtOHDDC 6V-92TA-91 (DDECII, 2S)1991T4.40.2761.650.420.07Tol/EtOHDDC 6V-71N-77 (MUI, 2S)1977T11.720.2823.180.860.212DCMDDC 6V-92TA-81/89 (MUI, 2S)1981T10.060.2682.160.420.144DCMDDC 6V-92TA-91 (DDECII, 2S)1991T4.840.2271.510.44DDC 6V-92TA-89 (DDECII, 2S)1989T4.8550.3382.4990.526DDC Series 60-91 DDECII1991T4.6350.3004.4580.164Cummins L-10-87 (MUI)1987T5.640.3092.330.89DDC Series 60-91 (DDECII)1991T4.680.2202.260.080.066DCMCummins N-14-87 (MUI)1987T6.320.3692.200.580.100?DDC Series 60-89 (DDECII)1989T5.1280.2524.0080.154DDC Series 60-91 (DDECII)1991T4.3030.1822.0040.3920.061Tol/EtOHCummins B5.91995T4.370.1061.470.300.05DCM0.24((18.5)Spreen et al., 1995Navistar DTA4661994T4.7790.0900.9890.1810.035DCM26Norbeck et al., 1998bCummins L101991T4.770.2242.260.538020(1725)1.95(4.92)DDC Series 601994T4.890.1121.4020.0650.043DCM17Sienicki et al., 1990Navistar DTA4661991T5.250.220.230.05DCM(PAH)

2-35--Table 2-8.Diesel engine emissions data from engine dynamometer tests(continued)EngineaYearTestbNOxg/bhphrPMg/bhphrCOg/bhphrTHCg/bhphrSOFg/bhphrSOFMethcTotalaldehyde,mg/bhp-hrB[a]P(PAH)ug/bhp-hrd1-NP (NPAH)ug/bhp-hreDDC Series 601991T4.5520.1882.1020.508Cat 3406E1997TDCM0.07(30)0.34Cummins NTCC4001988T4.470.422.220.53DDC Series 601994T4.430.1112.170.220.021DCM34(141)0.04(0.12)Navistar DTA4661994T4.860.0991.100.340.046DCM560.11(242)0.3(0.6)Unknown1994SS4.9340.1430.8070.3520.036DCM.076Scania1990SS9.300.1570.031DCMValmet1990SS8.670.157Volvo1990SS9.870.262Volvo1995SS4.560.135ReferenceUllman et al., 1990Kado et al., 1998Ullman, 1988Mitchell et al., 1994Tanaka et al., 1998Rantanen et al., 1993aNA=naturally aspirated.TC=turbocharged (engines not designated asNA or TC are turbocharged).AC=aftercooled. DI=direct injection.IDI=indirectinjection.EGR=exhaust gas recirculation.2S=two-stroke (engines not designated as2S are four-stroke).MUI=mechanical unitinjector (not electronicallycontrolled).DDEC=Detroit Diesel Corporation’s engine control module (electronic control).bSS=various single or multimode steady-state tests.T=heavy-duty FTP (transient test).cSOF extraction method.SFE=Supercritical fluid extraction.Allothers by Soxhlet extraction using the indicated solvents (?for unreported).DCM=dichloromethane.Tol/EtOH=toluene/ethanol mixture.Benz/cyc=benzene/cyclohexane mixture.C-hexane=cyclohexane.dNumber in parentheses is the total PAH emissionobtained by summing emissions of all PAHs reported.eNumber in parentheses is the total NPAH emissionobtained by summing emissions of all NPAHs reported.technology that often result from em ission standards, as well as by the lowering of on-road diesel
fuel sulfur content in 1993. As the discussion above indicates, there is a reasonable am ount of data upon which to
base em ission factor estim ates for late 1970s and later HD vehicles. However, very little transient test data are available on engines earlier than the m i d-1970s. The lim ited data available
from six pre-1976 vehicles tested using the tr ansient cycle suggests that PM em ission rates ranged from 1.6 g/m i to 9.0 g/m i , which is a substantially greater range than in post-1976 engines (Fritz et al., 2001). Although a substantial decreasing trend in DPM em issions from in-use chassis dynom om eter testing and engine testing (Figure 2-20) is evident, these data reflect a wide range
in em ission factors within any given m odel year. For exam ple, em ission factors for m odel year 1996 range from less than 0.1 g/m i to m o re than 1 g/m i (Yanowitz et al., 2000; Graboski et al., 1998b). The high variability in DPM em issions m easured in the chassis dynam o m e ter tests is observed because of several factors, including differences in m e asurem ent m e thods and test conditions at the various testing facilities, de terioration, and engine-to-engine variation. Although there can be excellent agreem ent between chassis dynam o m e ter testing facilities (Graboski et al., 1998a), there is no standard HD chassis dynam o m e ter Federal test procedure, and no detailed procedures for such testing are described in any authoritative source such as the Code of Federal Regulations, which does contain such procedures for engine dynam o m e ter

1.5 0

1.2 5

1.0 0

0.7 5

0.5 0

0.2 5

0.0 0 19 75 19 80 19 85 19 90 19 95 20 00

E n gin e M odel Y e a r Figure 2-20. 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.PMEmission,g/bhp-h

2-37testing used for EPA em ission regulations. Therefore, each facility has developed its own approach to HD testing. Clark et al. (1999) report that the test cycle can have a substantial
effect on DPM em issions, with higher DPM em issions reported from test cycles that incorporate full-power accelerations. Test cycles incorpor ating full-power accelerations reflect urban HD vehicle driving for several types of vehicles (garbage trucks, buses) operating in urban areas. Clark et al. (1999) also report that aggressi ve acceleration produces higher DPM em ission rates than does conservative acceleration, and Clark and co -workers suggest that real in-use driving is

m o re likely to m i m i c aggressive acceleration. Although figures are currently unquantified, it is generally believed that the m a jority of DPM is generated under transient conditions such as heavy acceleration.
W eaver and Klausm eier (1988) have exam ined potential causes and frequency of DPM

em issions deterioration for in-use HD diesel vehicles. Potential causes include m a nufacturing
defects and m a lfunctions such as retarded tim ing, fuel injector m a lfunction, sm oke-lim iting

m echanism problem s, clogged air filter, wrong or worn turbocharger, clogged intercooler, engine

m echanical failure, excess oil consum ption, and electronics that have been tam p ered with or
have failed. The recent report by McCorm ick and co-workers (2000) indicates that m a ny of
these m a lfunctions can have very large effects on DPM em issions, resulting in DPM increases of typically 50% to 100%. Although Yanowitz and co-workers (1999) found that DPM em issions were positively correlated with odom eter m ileage for a fleet of 21 vehicles, it is m o re likely that
the vehicle state of m a intenance will be m o re im portant than m ileage f o r determ ining the degree
of em issions deterioration. In fact, in a sim ilar analysis perform e d on the chassis dynam o m e ter results included in the review of Yanowitz et al. (2000), DPM em issions could not be correlated
with odom eter m ileage. Differences in testing m e thods between various facilities as well as varying states of m a intenance for vehicles of the sam e m ileage and m odel year probably account
f o r this lack of correlation. It is difficult, given current inform ation, to quantitatively assess the contribution of highem itting or sm oking diesel vehicles to am bient DPM. Em ission m odels used to prepare diesel particulate em ission inventories do not account fo r deterioration. The relative contribution of high-em itting diesel vehicles to the total m a ss and overall chem ical com position of diesel particulates is being quantified. Som e studies report num erous sm oking diesel trucks. A study
of the sm oke opacity-based inspection and m a intenance program in California found failure rates of 20% and higher, suggesting that high-em itting vehicles are not uncom m on (CARB/EEAI, 1997). In the Northeast, sm oke opacity testing conducted on 781 HD trucks found that 15% of
the vehicles failed the sm oke standard (40% opacity for 1991 and newer HD diesel vehicles and 50% opacity for pre-1991 HD diesel vehicles) (Cooper, 1999). Although the correlation
between sm oke and particulate em issions tends to be qualitative or sem i quantitative (discussed

2-38below), there is a good correlation between opacity and EC concentrations, and it is expected that high-em itting diesel vehicles m a y be an im portant part of the DPM em ission inventory.
Others have attem p ted to determ ine if the effects of deterioration could be detected for
in-use vehicles. In a study of 21 vehicles (Yanowitz et al., 1999), a linear m u ltivariate regression analysis found that DPM em issions were positively correlated with odom eter m ileage (several other correlation factors were also identified, including m odel year). A sim ilar analysis perform e d on the chassis dynam o m e ter results included in the review of Yanowitz et al. (2000)
found that DPM em issions could not be correlated with odom eter m ileage, probably because of differences in testing m e thods between the various facilities. Other approaches for m easuring em issions from in-use on-road diesel vehicles include
tunnel tests and rem o te sensing, the latter of which m easures gaseous, but not DPM, em issions. The literature reports of those studies are sum m a rized in Tables 2-9 and 2-10. Several tunnel test studies have reported DPM em ission fact ors (Pierson and Brachaczek, 1976; Japar et al., 1984; Pierson et al., 1983; Kirchstetter et al., 1999; Gertler et al., 1996, 1999).
The m e thod for determ ining em ission rates for vehicles traveling through a tunnel is explained in detail by Pierson et al. (1996). Brie fly, the em issions of a species are determ ined by

m easuring the concentration of a pollutant entering and leaving a tunnel along with knowledge
of the cross-section of the tunnel and m easurem en ts of the wind flux at the inlet and outlet of the
tunnel. The em ission rate is calculated by dividing the m a ss of the pollutant by the num ber of vehicles that passed through the tunnel and the length of the tunnel. The diesel and gasoline
vehicle contributions to the total em ission of the pollutant are separated by a sim p le regression analysis where the intercepts (100% HD and 100% LD) are the diesel and gasoline em ission rates, respectively.
Em ission factors from tunnel studies provide a snapshot of real-world em issions under driving conditions experienced in the tunnel and reflect em ission factors representative of the

m i x of in-use vehicles and the atm o spheric dilution and short-term transform a tion processes of
DE. Em ission factors derived from tunnel studies are often used as one source of inform ation to study the im pact of im proved technology and fleet turnover on em issions because they allow
random sam p ling of large num bers of vehicles, including a range of ages and m a intenance
conditions. However, tunnel studies are lim ited in that they represent driving conditions on a single roadway passing through a tunnel and represent m o stly steady-state driving conditions,
whereas m o st DPM is generated during transien t m odes of operation; also, tunnel studies do not include cold-start operations. Both of these factors need to be assessed to understand em ission rates for DPM to which people are exposed (U.S. EPA, 1992, 1995). DPM em ission factors
from in-use fleets derived from tunnel studies in the 1970s and 1980s com p ared with the 1990s

2-39Table 2-9.HD diesel emissions results from tunnel tests (adapted from Yanowitzet al., 1999)aNOxreported as NO2.eSlope of tunnel unknown, so used average fuel efficiency for the UnitedbCalculated fromobserved CO2emissions assuming fuel density 7.1 lb/galStates.and C is 87% of diesel fuel by weight.fPM3.cSince CO2emissions not available, fuel efficiency assumed to be the samegPM2.5.as in slightly uphill tunnel (Fort McHenry).hUncertainty reported as ±1.0 standard deviation, except where literaturedReported as black carbon, assumed that50% of total PM emissions arereport did not specify standard deviation; in those cases uncertainty listed asBC.reported.DPM(g/gal)4.9-9.89.49±1.038.1 ± 0.547.6 ± 0.47.0 ± 1.07.5 ± 1.407.0 ± 0.436.5 ± 0.167.6 ± 0.199.0± 1.3d3.5± 0.1f7.7± 0.6gCO(g/gal)55± 9578± 1778± 3039± 10NMHC(g/gal)-1 ± 711± 214± 64± 1NOxa(g/gal)157± 34129± 23111± 4122± 5125± 5129± 16CO2(g/mi)1,280± 40897± 481,897± 1681,596± 78DPM(g/mi).90-1.801.75 ± 0.191.5 ± 0.101.4 ± 0.071.3 ± 0.191.39 ± 261.3 ± 0.081.2 ± 0.031.4 ± 0.041.67± 0.24d0.62± 0.02f1.43± 0.12g0.29CO(g/mi)6.79± 11.786.8± 1.514.3± 5.56.03± 1.61NMHC(g/mi)-0.16± 0.880.92± 0.212.55± 1.050.68± 0.20NOxa(g/mi)19.50± 4.2223.82 ±4.179.66± 0.3222.50± 1.0019.46± 0.8523.82 ±2.98Fuelefficiency(mi/gal)5.42b8.03b5.42c5.60e11.46b5.42b6.44b5.42cTunnel location,year of studyAllegheny, 1974Allegheny, 1975Allegheny, 1976Allegheny, 1976Tuscarora, 1976Tuscarora, 1976Allegheny, 1977Allegheny, 1979Allegheny, 1979Cassiar Tunnel,1995, VancouverCaldecott Tunnel,1996, San FranciscoGubrist Tunnel,1993, ZurichFort McHenryTunnel, downhill,1992, BaltimoreFort McHenryTunnel, uphill,1992, BaltimoreTuscarora Tunnel1992, PennsylvaniaCaldecott Tunnel,1997, San FranciscoTuscarora Tunnel,1999, PennsylvaniaTestPierson andBrachaczeck, 1983Rogak et al., 1998Miguel et al., 1998Weingartner et al.,1997bPierson et al., 1996Pierson et al., 1996Pierson et al., 1996Kirchstetter et al.,1999Gertler,1999

2-40Table 2-10. Remote sensing results for HD vehicles

Reference Year study conducted Emissions (g/gal)

NO
x Jim e nez et al., 1998 1997 150 a,b,c Cohen et al., 1997 1997 108 a,b,c Countess et al., 1999 1998 187 a,b,c

CO Bishop et al., 1996 1992 59 b Cohen et al., 1997 1997 54 b Countess et al., 1999 1998 85 b THC B ishop et al., 1996 1992 0.002 HC/CO 2 m o le ratio d Cohen et al., 1997 1997 0.00073 HC/CO 2 m o le ratio d

a Rem o te sensing m easures NO. The reported value was corrected to a NO

x (as NO 2 ) value by

assum i ng 90% (m ole fraction) of NO

x is NO.

b Em issions in g/gal calculated by assum i ng that fuel density is 7.1 lb/gal and C is 87% by weight of fuel.

c No hum idity correction factor is included.

d In order to calculate em issions in g/gal, an average m o lecular weight is needed.

Source: Yanowitz et al., 1999.
suggest approxim a tely a fivefold decrease in DPM m a ss em ission factors over that tim e, with the

m o st recent data from 1999 reporting an em issi on factor of 0.29 g/m i for the on-highway HD diesel fleet (Figure 2-21).
Em ission factors vary substantially for the various tunnels, with NO

x em issions ranging
from 9.7 to 23.8 g/m i in the 1990s, CO em issions ranging from 6 to 14 g/m i , and THC em issions ranging from 0.16 to 2.55 g/m i

. Rem o te sensing reports em ission factors in term s of pollutant em issions per unit of fuel, not on a per-m ile basis. Agreem ent between rem o te sensing and tunnel studies for NO

x
em issions is reasonably good for the fleet as a whole, suggesting an average level for the fleet of about 130 g/gal, com p arable to the average em issions factor m easured in chassis dynam o m e ter studies (rem o te sensing can m easure em issions from an individual vehicle, whereas tunnel
studies m easure em issions from the fleet as a whole). Generally, chassis dynam o m e ter tests and engine dynam o m e ter test results are corrected for am bient hum idity, in accordance with the Federal Test Procedure (CFR 40, Subpart N). Tunnel tests and rem o te sensing tests have typically not included corrections for hum idity. Appropriate hum idity corrections for NO

x and DPM can be greater than 20% and 10%, respectively (or a total difference of m o re than 45% and 20%, respectively, between low- and high-hum idity areas), under norm a lly occurring clim atic
conditions. Additionally, the rem o te sensing literature has not addressed how to determ ine the

2-410. 10 0. 30 0. 50 0. 70 0. 90 1. 10 1. 30 1. 50 1. 70 1. 90

1 970 1 975 1 980 1 985 1 990 1 995 2 000

Y ear o f Mea s u r em e n tg/mi A l l e g h e n y / Tu s c a r or a - P A

F o r t M c H e n r y - MD

C a l d e c o tt - C A

Figure 2-21. Emission factors from HD diesel vehicles from tunnel studies.

Source: Data from Pierson and Brachaczek, 1976; Japar et al., 1984; Pierson et al., 1996; Kirchstetter et al., 1999; Gertler et al., 1995, 1996; Gertler, 1999.

correct value for the NO/NO

x ratio, and there is reason to believe that this value m a y differ system atically from site to site, although alm o st all of the NO

x is NO as it leaves the vehicle.
In addition to the hum idity correction discussed above, several factors m u st be taken into
account when com p aring DPM m easurem ents from tunnel tests to chassis dynam o m e ter

m easurem ents (Yanowitz et al., 2000): (1) Chassis testing m easures only tailpipe em issions;
tunnel tests can include em issions from other sources (tire wear, etc.), and (2) tunnel tests typically m easure em issions under steady-speed freeway conditions, whereas m o st chassis dynam o m e ter tests are m easured on cycles that are m o re representative of stop-and-go urban driving conditions. This latter lim itation also applies to rem o te sensing readings, which m easure instantaneous em issions versus em issions over a representative driving cycle.
Because THC em issions for diesel vehicles are very low in total m a ss in com p arison with gasoline vehicles, tunnel test results for THC ha ve a high degree of uncertainty. A regression analysis to determ ine the contribution of the lim ited num ber of HD vehicles to THC em issions is unstable; sm all errors in the total m easurem ents can change estim ates substantially. Sim ilarly,

2-42CO em issions are com p arable to autom obile em issions on a per-vehicle-m ile basis, but because there are generally m a ny m o re autom obiles than HD diesels in tunnel tests, CO m easurem ents from diesels m a y also have a high degree of uncertainty.

2.2.5.2. Locomotives
Locom o tive engines generally range from 1,000 horsepower up to 6,000 horsepower.
Sim ilar to the m u ch sm aller truck diesel engines, the prim ary pollutants of concern are NO

x , DPM, CO, and HC. Unlike truck engines, m o st locom o tive engines are not m e chanically coupled to the drive wheels. Because of this decoupling, locom o tive engines operate in specific steady-state m odes rather than the continuous transient operation norm a l for trucks. Because the
locom o tive engines operate only at certain speeds and torques, the m easurem ent of em issions is considerably m o re straightforward for locom o tive engines than for truck engines. Em issions

m easurem ents m a de during the relatively brief transition periods f r om one throttle position to another indicate that transient effects are very short and thus could be neglected for the purposes
of overall em issions estim ates.
Em issions m easurem ents are m a de at the various possible operating m odes with the engine in the locom o tive, and then weighting f actors f o r typical tim e of operation at each throttle position are applied to estim ate total em issions under one or m o re reasonable operating
scenarios. In the studies included in this analysis, two scenarios were considered: line-haul (m ovem e nt between cities or other widely separated points) and switching (the process of

assem b ling and disassem b ling trains in a switchyard).
The Southwest Research Institute m a de em issions m easurem ents f o r three dif f e rent engines in locom o tives in 1972 (Hare and Springer, 1972) and five m o re engines in locom o tives using both low- and high-sulfur fuel in 1995 (Fritz, 1995). Two engine m a nufacturers (the Electro-Motive Division of GM, and GE Transpor tation System s) tested eight different engine

m odels and reported the results to EPA (U.S. EPA, 1998b). All available data on locom o tives

are sum m a rized in the regulatory im pact assessm ent and shown in Figure 2-22.

2.2.6. Engine Technology Description and Chronology
NO
x em issions, DPM em issions, and brake-specific fuel consum ption (BSFC) are am ong the param e ters that are typically considered during the developm ent of a diesel engine. Many engine variables that decrease NO

x can also increase DPM and BSFC. One m a nifestation of the interplay am ong NO
x , DPM, and BSFC is that an increase in com bustion tem p eratures will tend
to increase NO form ation. Higher tem p eratures will also often im prove therm a l efficiency, can
im prove BSFC, and can increase the rate of DPM oxidation, thus lowering DPM em issions. One

exam ple of this is the tradeoff of DPM em issions and BSFC versus NO

x em issions with fuel

2-432-44 0 0 . 2 0

. 4 0 . 6 0

. 8 1 0 10 20 30 40

P M 

### S w i t c h C y c l e E m i s s i o n s D a t a N O x an d P M ( g / b hp

- h r )

Figure 2-22. itch emissions data.

Source: Line-haul and sw

U.S. EPA, 1998a.injection tim ing. Many recent advances in reducing the em issions of diesel engines without
aftertreatm e nt are com b inations of technologies that provide increm ental im provem e nts in the
tradeoffs am ong these em issions and fuel consum ption. The sum total, however, can be considerable reductions in regulated em issions within acceptable levels of fuel consum ption.
The m a jority of current HD diesel truck engines certified for use in the United States
utilize:

• A four-stroke cycle

• D irect-injection, high-pressure (1,200 bar to >2,000 bar) fuel injection system s with electronic control of injection tim ing and, in som e cases, injection rate

• C entrally located m u ltihole injection nozzles

• T hree or four valves per cylinder

• T urbochargers

• I n m a ny cases, air-to-air aftercooling

• I n som e cases, the use of an oxidation catalyst.

These features have phased into use with HD truck engines because they offer a relatively good com b ination of fuel consum ption, torque-rise, em issions, durability, and the ability to better “tune” the engines for specific types of applications. Fuel consumption, torquerise, and drivability have been m a intained or im proved while em issions regulations have becom e

m o re stringent. Many Class 8a and 8b dies el truck engines are now capable of 700,000 to 1,000,000 m iles of driving before their first rebuild and can be rebuilt several tim es because of their heavy construction and the use of rem ovable cylinder liners. These engines are expected to last longer and therefore have a useful life longe r than the regulatory estim ate of full useful life for HD engines (~1,000,000 m iles) previously used by EPA (for 1980 engines that were driven less than 300,000 m iles between rebuilds and were rebuilt up to three tim es). Current fourstroke locom o tive engines use engine technology sim ilar to on-highway diesel engines, except that electronic controls have only recently been introduced. It is difficult to separate the com ponents of current high-speed diesel engines for
discussion of their individual effects on em issions

. Most of the com ponents interact in num erous ways that affect em issions, perform a nce, and fuel consum ption.

2.2.6.1. Indirect and Direct Injection High-Speed Diesel Engines Prior to the 1930s, diesel engine design was lim ited to relatively low-speed applications because sufficiently high-pressure fuel injection equipm ent was not available. W ith the advent of high-speed and higher pressure pum p-line-nozzle system s, introduced by Robert Bosch in the

2-451930s, it becam e possible to inject the fuel directly into the cylinder for the first tim e, although indirect injection (IDI) diesel engines conti nued in use for m a ny years. As diesels were
introduced into the heavy truck fleet in the 1930s through the 1950s, both IDI and direct injection (DI) naturally aspirated variants were evident. A very low-cost rotary injection pum p technology was introduced by Roosa-Master in the 1950s, reducing the cost of DI system s and allowing their introduction on sm aller displacem ent, higher speed truck engines. After this tim e, only a sm all fraction of truck engines used an IDI system

. DI diesel engines have now all but re placed IDI diesel engines for HD on-highway applications. 2 IDI engines typically required m u ch m o re com p licated cylinder head designs but generally were capable of using less sophisticated, lower pressure injection system s with less expensive single-hole injection nozzles. IDI com bustion system s are also m o re tolerant of lower grades of diesel fuel. Fuel injection system s are likely the single m o st expensive com ponent of

m a ny diesel engines. Caterpillar continued producing both turbocharged and naturally aspirated IDI diesel engines for som e on-highway applications into the 1980s. Caterpillar and Deutz still produce engines of this type, prim arily for use in underground m i ning applications. IDI

com bustion system s are still used in m a ny sm all-displacem ent (<0.5 L/cylinder), very highspeed (>3,000 rpm rated speed) diesel engines for sm all nonroad equipm ent (sm a ll im ported tractors, skid-steer loaders), auxiliary engines, and sm all generator sets, and they were prevalent
in diesel autom o tive engines in the 1980s; IDI designs continue to be used in autom o tive diesel engines. IDI engines have practically no prem ixed burn com bustion and thus are often quieter and
have som e what lower NO

x em issions than DI engines. Electronic controls, high-pressure injection (e.g., GM 6.5), and four-valve/cylinder designs (e.g., the six-cylinder Daim ler LD engine) can be equally applied to IDI diesel engines as in DI, but they negate advantages in cost over DI engines. DI diesel engines of the sam e power output consum e 15% to 20% less fuel than IDI engines (Heywood, 1988). Considering the sensitivity of the HD truck m a rket to fuel costs, this factor alone accounts for the dem i se of IDI diesel engines in these types of applications. Throttling and convective heat transfer through the cham ber-connecting orifice, and heat rejection from the increased surface area of IDI com bustion system s, decrease their
ef

f i ciency and can cause cold-start dif f i culties when com p ared to DI designs. Most IDI diesel engine designs require considerably higher than optim um com p ression ratios (from an efficiency standpoint) to aid in cold-starting (19:1 to 21:1 fo r IDI engines vs. ~15:1 to 17:1 for DI engines).

2 The GM Powertrain/AM General 6.5L el ect roni cal l y cont rol l e d, t u rbocharged IDI-swi rl cham ber engi ne,
cert i f i e d as a l i ght HD di esel t r uck engi ne, i s t h e l a st rem a i n i ng HD on-hi ghway IDI engi ne sol d i n t h e Uni t e d
States.

2-46Because of the early introduction of DI technology into truck fleets, it is likely that by the end of the 1960s, only a sm all fraction of the HD diesel engines sold for on-highway use were IDI engines. It is unlikely that the shift from IDI to DI engine designs through the 1950s and 1960s occurred rapidly and likely that this shift had little significant im pact on em issions. Springer (1979) reports a com p arison of nearly identical Caterpillar 3406 engines (turbocharged and aftercooled) in DI and IDI configurations tested on an engine dynam o m e ter under steadystate conditions, which lim its the usefulness of these data. There was no significant difference in

em issions of DPM, SOF, aldehydes, or DPM-associated B[ a ]P (Table 2-8). Note that IDI designs continue to be used in autom o tive diesel engines.

2.2.6.2. Injection Rate Decreasing the duration of diffusion com bus tion and prom oting EC oxidation during the expansion stroke can reduce form ation of EC agglom erates (Stone, 1995) and reduce the particulate carbon fraction at high load (Needha m et al., 1989). Both of these effects are enhanced by increasing the fuel injection rate. The prim ary m eans of accom p lishing this is by increasing fuel injection pressure. In 1977 Robe rt Bosch introduced a new type of high-pressure

pum p capable of producing injection pressures of 1,700 bar at the nozzle (Voss and Vanderpoel, 1977). This increased fuel injection pressure by roughly a factor of 10. Unit injection, which

com b ines each fuel injection nozzle with individual cam -driven fuel pum ps, can achieve very high injection pressures (>2,000 bar). The first com b ination of unit injectors with electronically
controlled solenoids for tim ing control was offered in the United States by Detroit Diesel Corporation in the 1988 m odel year (Ham es et al., 1985). Replacem ent of the injection cam with hydraulic pressure, allowing a degree of injection rate control, was m a de possible with the hydraulic-electronic unit injection jointly developed by Caterpillar and Navistar, introduced on the Navistar T444E engine (and variants) in 1993. It is widely known that high fuel injection pressures have been used to obtain com p liance with the PM standards that went into effect in 1988 (Zelenka et al., 1990). Thus, it is likely that a transition to this technology began in the 1980s, with the vast m a jority of new engine sales

em ploying this technology by 1991, when the 0.25 g/bhp-hr Federal PM standard went into
effect. The use of electronic control of injection ra te is rapidly increasing on m e dium HD diesel engines. Engines are currently under developm ent, perhaps for 2002–2004 introduction, that use

com m on-rail fuel injection system s with even m o re flexible control over injection pressure and
tim ing than previous system s. Increased injection rate and pressure can significantly reduce EC em issions, but it can
also increase com bustion tem p eratures and cause an increase in NO

x em issions (Springer, 1979;

2-47W a tson and Janota, 1982; Stone, 1995). Low NO
x brake m ean engine pressure (BMEP) are possible when com b ined with turbocharging, aftercooling, and injection tim ing retard. , low DPM, and relatively good BSFC and

2.2.6.3. Turbocharging, Charge-Air Cooling, and Electronic Controls Use of exhaust-driven turbochargers to increase intake m a nifold pressure has been applied to both IDI and DI diesel engines for m o re than 40 years. Turbocharging can decrease
fuel consum ption com p ared with a naturally aspirated engine of the sam e power output. Turbocharging utilizes otherwise wasted exhaust heat and pressure to generate intake boost. The boosted intake pressure effectively increases air displacem ent and increases the am ount of fuel that can be injected to achieve a given fuel-air ratio. Turbocharging increases the power density of an engine. Boosting intake pressure via tu rbocharging and reducing fuel-to-air ratio at a constant power can significantly increase both intake tem p eratures and NO

x em issions. Increased boost pressure can significantly reduce ignition delay, which reduces VOC and DPM
SOF em issions (Stone, 1995) and increases the flexibility in selection of injection tim ing. Injection tim ing on turbocharged engines can be retarded further for NO

x em ission control with
less of an effect on DPM em issions and fuel consum ption. This allows a rough parity in NO

x
em issions between turbocharged (non-aftercooled) and naturally aspirated diesel engines (W atson and Janota, 1982). Turbocharging perm its the use of higher initial injection rates (higher injection pressure),

Figure 2-23. Effect of turbocharging and aftercooling on NO
x and PM.

Source: Mori, 1997.

2-48which can reduce particulate em issions. Although this m a y offer advantages for steady-state operation, hard accelerations can tem porarily cause overly fuel-rich conditions because the turbocharger speed lags behind a rapid change in engine speed (turbo-lag). This can cause significant increases in DPM em issions during accel erations. Before the advent of electronic controls, the effect of acceleration on DPM em issions could be lim ited by m echanically delaying
dem a nd for m a xim u m fuel rate with a “sm oke-puff elim inator.” Because this device also lim ited engine response, there was considerable incentive for the end-users to rem ove or otherwise
render the device inactive. Charge-air cooling, for example, using an air-to-air aftercooler (aircooled heat exchanger) between the turbocharger com p ressor and the intake m a nifold, can greatly reduce intake air and peak com bustion tem p eratures. W h en com b ined with injection
tim ing retard, charge-air cooling allows a significant reduction in NO

x em issions with acceptable
BSFC and DPM em issions when com p ared to either non-aftercooled or naturally aspirated diesel engines (Hardenberg and Fraenkle, 1978; Pisc hinger and Cartellieri, 1972; Stone, 1995). The use of charge-air cooling effectively shifts the NO

x -DPM tradeoff curve, as shown in Figure 2-23. Electronic control of fuel injection tim i ng allowed engine m a nufacturers to carefully tailor the start and length of the fuel injection events m u ch m o re precisely than through

m echanical m eans. Because of this, newer on-highway turbocharged truck engines have virtually no visible sm oke on acceleration (although em issions of DPM are substantial during this driving m ode). Electronic controls also allowed fuel injection retard under desirable
conditions for NO

x reduction, while still allowing tim ing optim ization for reduced VOC

em issions on start-up, acceptable cold-weather perform a nce, and acceptable perform a nce and durability at high altitudes. Previous m echanical unit injected engines (e.g., the 1980s Cum m i ns L10, the Non-Electronic Control Detroit Diesel 6V 92) were capable of reasonably high injection pressures, but they had fixed injection tim ing that only varied based on the hydraulic param e ters of the fuel system

. Many other engines with m echanical in-line or rotary injection pum ps had only coarse injection tim ing control or fixed injection tim ing. Precise electronic control of injection tim ing over differing operating conditions also allowed HD engine m a nufacturers to retard injection tim ing to obtain low NO

x em issions during highly transient urban operation, sim ilar to that found during em issions certification. HD engine

m a nufacturers also advanced injection tim ing dur ing less transient operation (such as freeway driving) for fuel consum ption im provem e nts (~3% to 5%) at the expense of greatly increased
NO
x em issions (approxim a tely three to four tim es regulated levels). This particular situation
resulted in the recent consent decree settlem e nts between the Federal Governm e nt and m o st HD engine m a nufacturers to ensure effective NO

x control in all driving conditions, including on2-49highway high-speed steady-state driving. Turbocharged engines entered the m a rket very slowly beginning in the 1960s. Data for
DPM em issions from naturally aspirated engines of m odel years 1976 to 1983 are com p ared with
DPM em issions from turbocharged engines in Fi gure 2-24. There is no consistent difference in
DPM em issions between turbocharged and naturally aspirated engines. Although not plotted, the
data also show no difference in em issions of NO

x , DPM SOF, or DPM-associated B[ a ]P and 1- nitropyrene (1-NP). Charge-air cooling was introduced during the 1960s and was initially perform e d in a heat exchanger using engine coolant. Cooling of the charge air using am bient air as the coolant was introduced into heavy trucks by Mack in 1977 with production of the ETAY(B)673A engine (Heywood, 1988). Use of am bient air allowed cooling of the charge air to m u ch lower

tem p eratures. Most HD diesel engines sold toda y em ploy som e form of charge air cooling, with air-to-air aftercooling being the m o st com m on. Johnson and co-workers (1994) have presented a

com p arison of sim ilar engines that differ in that the charge air is cooled by engine coolant (1988 engine) and by am bient air, with a higher boost pressure for the second (1991 engine). The 1991 engine also used higher pressure fuel inject ors. The 1991 engine exhibited both lower DPM

em issions (50% lower than the 1988 engine) and lower NO

x em issions. Higher injection pressure is likely to have enabled the reduced DPM em issions, whereas the lower charge-air

3

2

1

0 19 74 19 76 19 78 19 80 19 82 19 84 T u r boc ha r g e d Na t u ral l y A s p i r a t e d

Eng i ne M o d e l Ye a r
Figure 2-24. Comparison of diesel engine dynamometer PM emissions for four-stroke, naturally aspirated, and turbocharged engines.

Source: Data are from Table 2-8.PMEmissions,g/bhp-h

2-50tem p erature and the ability to electronically retard the injection tim ing under som e conditions likely enabled the lower NO

x em issions. It is apparent on the basis of both the literature and certif ication data that turbochargers with aftercoolers can be used in HD engines in conjunction with other changes to produce a
decrease in em issions. On the advent of a NO

x standard in 1985, NO

x was probably reduced on the order of 10% to 30% in turbocharged aftercooled engines with retarded injection tim ing. This decrease is not evident in the in-use chassis testing data because of deterioration and the use of illegal em issions devices as described above. Overall, it is expected that engines in the 1950s

to m i d-1970s tim efram e would have sim ilar DPM em ission rates, whereas post-1970 engines
would have som e what lower DPM em ission rates.

2.2.6.4. Two-Stroke and Four-Stroke High-Speed Diesel Engines A detailed discussion of the two- and four-stroke engine cycles can be found in the literature (Heywood, 1988; Taylor, 1990; Stone, 1995). Nearly all high-speed two-stroke diesel engines utilize unif l ow scavenging assisted by a positive-displacem ent blower (Figure 2-25). Uniflow-scavenged two-stroke diesels use poppet exhaust valves sim ilar to those found in fourstroke engines. The intake air enters the cylinder through a pressurized port in the cylinder wall. A crankshaft-driven, positive-displacem ent blower (usually a roots-type) pressurizes the intake port to ensure proper scavenging. A turbocharger m a y be added to the system to provide additional boost upstream of the blower at higher speeds and to reduce the size and parasitic losses associated with the positive-displacem ent blower. Two-stroke diesel engines can achieve effici ency com p arable to four-stroke counterparts and have higher BMEP (torque per unit displacem ent) (Heywood, 1988). It is useful to note that two-stroke cycle fires each cylinder once every revolution, whereas the four-stroke cycle fires every other revolution. Thus, for a given engi ne size and weight, two-strokes can produce m o re
power. However, two-stroke diesel engines are less durable than their four-stroke counterparts. Lubricating oil is transferred from the piston rings to the intake port, which causes relatively high oil consum ption relative to f our-stroke designs. Durability and low oil consum ption are desirable for on-highway truck applications. Th is m a y be why four-stroke engines have been favored for these applications since the beginni ng of dieselization in the trucking industry, with the notable exception of urban bus applications. Although it is no longer in production, the
Detroit Diesel 6V92 series of two-stroke dies el engines is still the m o st popular for urban bus applications, where the high power density allows the engine to be m o re easily packaged within
lim ited spaces. The prim ary reason that two-stroke engines like the 6V92 are no longer offered for urban bus applications is excessive DPM em issions. The lubricating oil control with twostrokes tends to be lower than for four-stroke engines, and therefore, em issions have higher VOC

2-51Figure 2-25. An example of uniflow scavenging of a tw ostroke diesel engine w i th a positive displacement blow er. Scavenging is the process of simultaneously emptying the cylinder of exhaust and refilling w i th fresh air.

Source: Adapted from Taylor, 1990.
and organic DPM em issions relative to four-stroke designs. This was particularly problem atic for urban bus applications because urban bus e ngines m u st m eet tighter Federal and California
PM em issions standards. The current urban bus PM standard (0.05 g/bhp-hr) is one-half of the current on-highway HD diesel engine PM standa rd, although EPA is in the process of proposing

m o re strict standards for HD diesel truck engi nes along with further reductions in diesel fuel
sulfur levels. No two-stroke diesel engine de signs have been certified to m eet the m o st recent urban bus PM em issions standards, and Detroit Di esel Corporation has not certified a two-stroke diesel engine for on-highway truck use since 1995.
A com p rehensive review of em issions from hundreds of vehicles (1976–98 m odel years) that had been tested on chassis dynam o m e ters found that DPM em issions vary substantially within a given m odel year and that within that variation there are no discernible differences in
DPM em issions between two- and four-stroke ve hicles (Figure 2-26) (Yanowitz et al., 2000).
DPM em ission factors reported for engine tests al so indicate that two- and four-stroke engines
have com p arable em ission factors, as these engines all had to m eet the sam e regulatory standard (Figure 2-27). In contrast to DPM em issi ons, evidence suggests that m i d-1970s two-stroke engines exhibited very high SOF levels com p ared with four-stroke engines, with later m odel years showing sim ilar SOF em issions for two- and four-stroke engines (Figure 2-28). For aldehydes, benzo[a]pyrene, and 1-nitropyrene, data are available for only one two-stroke engine,

2-525

4

3

2

1

0 197 8 198 0 198 2 198 4 198 6 198 8 199 0 199 2 199 4 199 6 199 8 

### 4- S t rok e E ngines 2- S t rok e E ngines

Engine Emis s i ons M o de l Y e a r

Figure 2-26. Comparison of tw o- and four-strok e vehicle diesel PM emissions from chassis dynamometer studies.

Source: Yanowitz et al., 2000.

but they indicate no significant difference in em i ssions from com p arable m odel year four-stroke engines. Overall, regulated em issions changes attributable to changing proportions of two- and four-stroke engines in the in-use fleet do not a ppear to have influenced DPM em ission levels, but the transition to four-stroke engines in the 1970s would have decreased the fraction of SOF
associated with the DPM. It appears that the proportion of two-stroke engines in the in-use fleet was relatively constant until the late 1980s, when it began to decline.

2.2.7. Air Toxic Emissions HD diesel vehicle exhaust contains several substances that are known, likely, or possible
hum an or anim al carcinogens, or that have serious noncancer health effects. These substances include, but are not lim ited to, benzene, form aldehyde, acetaldehyde, 1,3-butadiene, acrolein, dioxin, PAH, and nitro-PAH (the com p lete list of chem ically characterized com pounds present in DE is provided in Section 2.3.1). Very few hist orical data are available to exam ine changes in

em ission rates over tim e. In this section, trends in aldehyde em issions over tim e and a sum m a ryPMEmission, g/mi

2-533

2

1

0 19 75 19 80 19 85 19 90 19 95 20 00
Eng i n e M o d e l Yea r
Figure 2-27. Comparison of tw o- and four-str oke engine diesel PM emissions from engine dynamometer studies.

1.00

0.75

0.50

0.25

0.00 19 75 19 80 19 85 19 90 19 95 20 00 4-S t roke 2-S t roke

Eng i ne M o d e l Ye ar

Figure 2-28. Diesel engine dynamometer SOF emissions from tw o- and four-stroke engines. SOF obtained by dichloromethane extraction in most studies.

Source: Data are from Table 2-8.SOF Emission, g/bhp-hPMEmission,g/bhp-h 4-S t rok e 2-S t rok e

2-54of dioxin em ission factors are presented. PAH and nitro-PAH em ission factors are discussed in
Section 2.2.8.2.

2.2.7.1. Aldehyde Emissions
Am ong the gaseous com ponents em itted by diesel engines, the aldehydes are particularly
im portant because they constitute an im portant fraction of the gaseous em issions and they are probable carcinogens that also produce noncancer health effects. Form aldehyde m a kes up the

m a jority of the aldehyde em issions (65% to 80%), with acetaldehyde being the second m o st abundant aldehyde in HD diesel em issions. To tal aldehyde em issions reported from chassis dynam o m e ter testing suggest that aldehyde em issions have declined since 1980; however, only two tests reported aldehydes from engines m a de after 1985 (Figure 2-29). Engine dynam o m e ter studies also suggest a downward trend in the em issions of aldehydes in the tim e period from 1976 to 1994 (Figure 2-30). Engine dynam o m e ter studies report aldehyde em ission levels of
150–300 m g /bhp-hr for late 1970s engines with no si gnificant effect of turbocharging, or IDI
versus DI. High-pressure fuel injection m a y have resulted in a m a rginal increase in aldehyde

em issions (Springer, 1979). By com p arison, 1991 m odel year engines (DI, turbocharged) exhibited aldehyde em issions in the 30–50 m g /bhp-hr range (Mitchell et al., 1994).

2.2.7.2. Dioxin and Furans
Ballschm iter et al. (1986) reported detecti ng polychlorinated dibenzo-p-dioxins (CDDs) and polychlorinated dibenzofurans (CDFs) in used m o tor oil and thus provided som e of the first
evidence that CDDs and CDFs m i ght be em itted by the com bustion process in diesel-f ueled engines. Incom p lete com bustion and the presence of a chlorine source in the f o rm of additives in the oil or the fuel were speculated to lead to the form ation of CDDs and CDFs. Since 1986,
several studies have been conducted to m easure or estim ate CDD/CDF concentrations in

em issions from diesel-fueled vehicles. Th ese studies can be characterized as direct

m easurem ents from the engine exhaust and indirect m easurem ents from the sam p ling of air within transportation tunnels.
Table 2-11 is a sum m a ry of various CDD/CD F em ission characterization studies reported in the United States and Europe for diesel-fue led cars and trucks. Hagenm aier et al. (1990) reported an em ission f actor f o r LD diesel vehicles of 24 pg TEQ per liter of diesel f u el

consum ed. TEQ, or the toxic equivalency factor, rates each dioxin and furan relative to that of 2,3,7,8-TCDD, which is arbitrarily assigned a TEQ of 1.0 based on anim al assays. Schwind et al. (1991) and Hutzinger et al. (1992) studied em issions of CDDs/CDFs from Germ an internal

com bustion engines running on com m e rcial diesel fuels and reported a range of CDD/CDF

em ission rates across the test conditions (in units of pg TEQ per liter of diesel f u el consum ed) of 10–130 pg TEQ/L for diesel car exhaust and 70–81 pg TEQ/L for diesel truck exhaust.

2-5510 00

80 0

60 0

40 0

20 0

0

19 75 19 80 19 85 19 90 19 95 20 00
Eng i ne M o d e l Ye a r 4- st r o ke 2- st r o ke

Figure 2-29. Diesel engine aldehyde emission s measured in chassis dynamometer studies.

Source: Data are from W a rner-Selph a nd Dietzm ann, 1984; Schauer et al., 1999; Unnasch et al., 1993.

30 0

25 0

20 0

15 0

10 0

50

0 19 76 19 78 19 80 19 82 19 84 19 86 19 88 19 90 19 92 19 94 19 96 4- st r o ke 2- st r o ke

Eng i ne M o d e l Ye a r Figure 2-30. Diesel engine aldehyde emi ssions from engine dynamometer studies.

Source: Data from Table 2-8.AldehydeEmission, mg/bhp-hAldehyde Emissions,mg/mile

2-56Table 2-11.Summary of CDD/CDF emissions from diesel-fueled vehiclesStudyCountryVehicletestedNumberof testvehiclesEmission factor(pg TEQ/km driven)Drivingcycle; sampling locationCARB, 1987; Lew, 1996United StatesDiesel truck1663–1,3006-hr dynamometer test at 50 km/hrMarklund et al., 1990SwedenDiesel truck1not detected (<18)aU.S. Federal mode 13 cycle; before mufflerHagenmaier et al., 1990GermanyDiesel car12.4aComparable to FTP-73 test cycle; in tailpipeHagenmaier, 1994GermanyDiesel bus1not detected (< 1 pg/L)On-the-road testingOehme et al., 1991 (tunnelstudy)Norway—(b)520c38cavg = 2809,500c720cavg = 5,100Cars moving uphill (3.5% incline) at 60 km/hrCars moving downhill (3.5% decline) at 70 km/hrTrucks moving uphill (3.5% incline) at 60 km/hrTrucks moving downhill (3.5% decline) at 70 km/hrSchwind et al., 1991Hutzinger et al., 1992GermanyDiesel carDiesel truck115.0–13a13–15aVarious test conditions (i.e., loads and speeds)Various test conditions (i.e., loads and speeds)Gertler et al., 1996 (tunnelstudy)United StatesDiesel trucks(d)mean = 172Mean of seven 12-hour samplesGullett and Ryan, 1997United StatesDiesel truck1mean - 29.0Mean of five sample routes

2-57aResults reported were in units of pg TEQ/literof fuel.For purposes of this table, the fuel economyfactor used byMarklund et al.(1990),10 km/Lor24 miles/gal,was used to convertthe emission ratesinto units of pg TEQ/kmdriven forthe cars.Forthe diesel-fueled truck,the fueleconomyfactorreported in CARB (1987a)fora 1984 heavy-dutydiesel truck,5.5 km/L(or13.2 miles/gal),was used.bTests wereconducted overportions of 4 days,with traffic rates of 8,000-14,000 vehicles/day.Heavy-dutyvehicles (defined asvehicles over7 metersin length)ranged from4% to 15% of total.cEmission factors are reported in units of pg Nordic TEQ/kmdriven;the values in units of I-TEQ/kmare expected to be about 3%to 6% higher.dTests wereconducted over5 dayswith heavy-dutyvehicle rates of 1,800-8,700 vehiclesper12-hoursampling event.Heavy-dutyvehicles accounted for21% to 28% of all vehicles.In 1994, Hagenm aier reported CDD/CDF em issi ons from a diesel-fueled bus and found no detectable levels in the exhaust (at a detection lim it of 1 pg/L of fuel consum ed) for individual congeners. In 1987, the California Ai r Resources Board (CARB) produced a draft report of a HD engine tested under steady-state conditions indicating a TEQ em ission factor of 7,290 pg/L of fuel burned (or 1,300 pg/km driven) if nondetected values are treated as one-half
the detection lim it. Treating nondetected values as zeros yields a TEQ concentration equivalent to 3,720 pg/L of fuel burned (or 663 pg/km driven) (Lew, 1996). Norbeck et al. (1998c) reported

em ission factors for dioxin and furans from a Cum m i ns L10 HD diesel engine running on pre1993 fuel of 0.61 pg/L and 0.41 pg/L for the sam e e ngine running on reform ulated fuel. The low

em ission factors reported by Norbeck et al. (1998c) were attributed to losses of dioxin and furan

com pounds to the dilution tunnel walls. EPA has directly sam p led the exhaust from a HD diesel truck for the presence and occurrence of CDDs/CDFs (Gullett and Ryan, 1997). The average of five tests (on highway and city street driving conditions) was 29.0 pg TEQ/km with a standard deviation of 38.3 pg TEQ/km ; this standard deviation reflects the 30-fold variation in the two city driving route tests.
Tunnel studies are an indirect m eans of m easuring contam inants that m a y be associated
with em issions from cars and trucks. In these studies, scrapings of carbonaceous m a tter from the interior walls of the transportation tunnel or the tunnel air are sam p led and analyzed for the target contam inants. Several European studies and one recent U.S. study evaluated CDD/CDF

em issions from vehicles by m easuring the presence of CDDs/CDFs in tunnel air. This approach has the advantage of allowing random sam p ling of large num bers of vehicles passing through the tunnel, including a range of ages and m a intenance levels. The disadvantage of this approach is
that it relies on indirect m easurem ents (rather than tailpipe m easurem ents), which m a y introduce unknown uncertainties into the interpretation of results.
Oehm e et al. (1991) reported the em ission rates associated with HD diesel trucks as
follows: uphill = 9,500 pg TEQ/km ; downhill = 720 pg TEQ/km ; m ean = 5,100 pg TEQ/km

. The m ean values are the averages of the em ission rates corresponding to the two operating

m odes: vehicles m oving uphill on a 3.5% incline at an average speed of 37 m i/hr and vehicles

m oving downhill on a 3.5% decline at an average speed of 42 m i/hr.
W e vers et al. (1992) m easured the CDD/CDF content of air sam p les taken during the winter of 1991 inside a tunnel in Antwerp, Belgium

. The results obtained indicated that the tunnel air had a dioxin TEQ concentration about twice as high as the outside air (80.3 fg TEQ/m 3 for tunnel air vs. 35 fg TEQ/m 3 for outside air for one set of m easurem ents and 100 fg TEQ/m 3 for tunnel air vs. 58 fg TEQ/m 3 for outside air for a second set of m easurem ents).

2-58During October/Novem b er 1995, Gertler et al . (1996, 1998) m easured CDDs/CDFs in the Fort McHenry Tunnel in Baltim ore, Maryland. The em ission factors calculated, assum i ng that
all CDDs/CDFs em itted in the tunnel were from HD vehicles, are presented in Table 2-12. The average TEQ em ission factor was reported to be 172 pg TEQ/km

. The m a jor uncertainties in the study were tunnel air volum e m easurem ent, sam p ler flow volum e control, and analytical

m easurem ent 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; (2) virtually no vehicle using the tunnel used leaded gasoline, which is associated with past

em issions of CDDs and CDFs from gasoline-power ed vehicles; (3) the tunnel walls and streets
were cleaned 1 week before the start of sam p ling, and in addition, the study analyzed road dust
and determ ined that resuspended road dust contributed only about 4% of the estim ated em ission factors; and (4) HD vehicles m a de up, on av erage 25.7% of vehicles using the tunnel. Using the em issions factor from the Gertler et al. studies, the EPA Office of Research and Developm ent’s dioxin source em ission inventor y estim ates that 33.5 g of dioxin TEQ (total 2,3,7,8-TCDD equivalents) were em itted from HD U.S. trucks in 1995. This is a very sm all contribution (1.2%) com p ared with the national annual em ission of 2,800 g CDDs/CDFs.

2.2.8. Physical and Chemical Composition of Diesel Exhaust Particles
DPM is def i ned by the m easurem ent procedures sum m a rized in Title 40 CFR, Part 86, subpart N. This definition and the basic characteristics of DPM have been sum m a rized in
Section 2.2.2. As described there, DE particles ar e aggregates of prim ary spherical particles that
consist of solid carbonaceous m a terial and ash and contain adsorbed organic and sulfur

com pounds (sulfate) com b ined with other condensed m a terial. The organic m a terial includes unburned fuel, engine lubrication oil, and lo w levels of partial com bustion and pyrolysis products. The organic m a terial is absorbed to the EC core and is also found in heterogeneously
nucleated aerosol. This fraction of the DPM is frequently quantified as the SOF (i.e., the fraction that can be extracted by an organic solvent). Because of the toxicological significance of the organic com ponents associated with DPM, it is im portant to understand, to the extent possible, the historical changes in the com position of SOF and potential changes in the fraction of SOF
associated with DPM.
Various researchers have attem p ted to apportion the SOF to unburned oil and fuel sources by ther m ogravi m etric analysis and have found that the results vary with test cycle
and engine (Abbass et al., 1991; W achter, 1990). Kittelson (1998) esti m ates that a typical

co m position of SOF is about one-fourth unburned fuel and three-fourths unburned

2-59Table 2-12.Baltimore Harbor Tunnel Study:estimated CDD/CDF emission factors for HD vehiclesCongener/congener groupRun-specific emissionfactorsMeanemissionfactors(pg/km)Run no.2(pg/km)Run no.3(pg/km)Run no.5(pg/km)Run no.6(pg/km)Run no.8(pg/km)Run no.9(pg/km)Run no.10(pg/km)2,3,7,8-TCDD1,2,3,7,8-PeCDD1,2,3,4,7,8-HxCDD1,2,3,6,7,8-HxCDD1,2,3,7,8,9-HxCDD1,2,3,4,6,7,8-HpCDDOCDD24.540.218.237.553.60061.620.625.228.256.54013,3610.015.446.564.391.67293,38221.25.68.319.648.41111,12037.838.464.51532802,4389,73040.10.00.071.11269635,82954.983.01231863702,0807,62034.329.040.880.01479604,4352,3,7,8-TCDF1,2,3,7,8-PeCDF2,3,4,7,8-PeCDF1,2,3,4,7,8-HxCDF1,2,3,6,7,8-HxCDF1,2,3,7,8,9-HxCDF2,3,4,6,7,8-HxCDF1,2,3,4,6,7,8-HpCDF1,2,3,4,7,8,9-HpCDFOCDF0024.515.40.327.715.212.60094.348.975.713975.114.882.528058.523967.672.613120473.775.615244560.8401152.823.646.693.851.0055.715431.1175155.853.385.012461.320.693.031325.041673.40.063.916454.437.286.83542.353461.743.310816695.563.511130834.937086.534.576.412958.834.285.226730.4305Total 2,3,7,8-CDDTotal 2,3,7,8-CDFTotal TEQ17495.773.83,9541,1081754,3281,6841701,3357849612,7431,3472357,0281,37115310,5151,3623035,7251,107172Total TCDDTotal PeCDDTotal HxCDDTotal HpCDDTotal OCDDTotal TCDFTotal PeCDFTotal HxCDFTotal HpCDFTotal OCDFTotal CDD/CDF245110677000124136001,291021.9080233619011193192232395,98714083.37531,4983,3821,3141,15285281440110,39016535.654.51421,12065678.467.61441752,6383111742,0095,6969,7302,4161,05544451341622,7661090.01,6661,9335,8291,00728271935453412,43497.31652,9714,3777,62068762661963737018,16815284.21,1622,0644,43599749145138430510,525HD vehicles as % of totalvehicles21.222.022.634.028.824.227.425.7

2-60Notes:(1) Listed values are based on the difference between the calculated chemical massentering the tunnel and the mass exiting thetunnel.(2) All calculated negative emission factors were set equal to zero.(3)All CDD/CDF emissions wereassumed to result fromheavy-dutydiesel-fueled vehicles.The table presents in the last row the percent oftotal traffic that was heavy-duty vehicles.Source: Gertleret al.,1996.engine oil. Partial combustion and pyrolysis products represented a very small fraction
of the SOF on a mass co m bustion and pyrolysis products represented a very s m all fraction
of the SOF on a mass basis (Kittelson, 1998), which is confir m e d in nu m erous other studies.
A num ber of investigators have tried to separate the organic fraction into various classes
of com pounds. Schuetzle (1983) analyzed the dichlorom e thane extract of DPM from a LD diesel engine and found that approxim a tely 57% of the extracted organic m a ss is contained in the nonpolar fraction. About 90% of this fraction cons ists of aliphatic HCs from approxim a tely C 14 to about C 40 (Black and High, 1979; Pierson and Brachaczek, 1983). PAHs and alkyl-substituted
PAHs account for the rem a inder of the nonpolar m a ss. The m oderately polar fraction (~9% w/w of extract) consists m a inly of oxygenated PAH species, substituted benzaldehydes, and nitrated
PAH. The polar fraction (~32% w/w of extract ) is co m posed m a inly of n-alkanoic acids, carboxylic and dicarboxylic acids of PAH, hydr oxy-PAH, hydroxynitro-PAH, and nitrated Ncontaining heterocyclic com pounds (Schuetzle, 1983; Schuetzle et al., 1985). Rogge et al. (1993) reported the com position of the extractable portion of fine DPM

em itted from two HD diesel trucks (1987 m odel year). The DPM filters were extracted twice with hexane, then three tim es with a benzene/2-propanol m i xture. The extract was analyzed by capillary gas chrom a tography/m a ss spectrom e try (GC/MS) before and after derivatization to convert organic acids and other com pounds having an active H atom to their m e thoxylated analogues. Unidentified organic com pounds m a de up 90% of the eluted organic m a ss and were
shown to be m a inly branched and cyclic HCs. From the m a ss fraction that was resolved as discrete peaks by GC/MS, ~42% were identified as specific organic com pounds. Most of the identified resolved organic m a ss (~60%) consisted of n-alkanes, followed by n-alkanoic acids (~20%). PAH accounted for ~3.5% and oxy-PAH (ketones and quinones) for another ~3.3%.
The distribution of the em issions between the gaseous and particulate phases is
determ ined by the vapor pressure of the individual species, by the am ount and type of the DPM present (adsorption surface available), and by the tem p erature (Ligocki and Pankow, 1989). Two-ring and sm aller com pounds (e.g., naphthalene) exist prim arily in the gas phase, whereas five-ring and larger com pounds (e.g., benzo[a]pyrene) are alm o st com p letely adsorbed on the particles. Three- and four-ring com pounds are distributed between the two phases. The vapor pressures of these interm ediate PAHs can be significantly reduced by their adsorption on various
surfaces. Because of this phenom enon, the am ount and type of DPM present play an im portant role, together with tem p erature, in the vapor-particle partitioning of sem i volatile organic

com pounds (SOCs).
The m easurem ents of gas/particulate phase distribution are often accom p lished by using a high-volum e filter followed by an adsorbent such as polyurethane foam (PUF), Tenax, or XAD-2 (Cautreels and Van Cauwenberghe, 1978; Thrane and Mikalsen, 1981; Yam a saki et al., 1982).

2-61The pressure drop behind a high-volum e f ilter or cascade im pactor can contribute to volatilization of the three- to five-ring PAHs from the PM proportional to their vapor pressures.
The m a gnitude of this blow-off artifact depends on a num ber of factors, including sam p ling

tem p erature and the volum e of air sam p led (Van Vaeck et al., 1984; Coutant et al., 1988). Despite these problem s from volatilization, m easurem ents with the high-volum e filters followed by a solid adsorbent have provided m o st estim ates of vapor-particle partitioning of SOCs in

am bient air, as well as insights into the factors influencing SOC adsorption onto aerosols. Significant fractions of phenanthrene, anthracen e, and their alkylated derivatives, along with fluoranthene and pyrene, exist in the gas phase. PAHs with m o lecular weight greater than that of pyrene are typically not observed on PUF sam p les. During the collection of particulate organic

com pounds, adsorption of sem i volatile PAHs can also occur, as well as chem ical transform a tion
of the sem i volatile com pounds (Schauer et al ., 1999; Cantrell et al., 1988; Feilberg et al., 1999; Cautreels and Van Cauwenberghe, 1978).
Most of the sulfur in the fuel is oxidized to SO 2 , but a sm all am ount (1% to 4%) is
oxidized to sulfuric acid in the exhaust. Sulfate em issions are roughly proportional to sulfur in
the fuel. Since the reduction of the allowable sulfur content in diesel fuel in 1993, sulfate

em issions have declined from roughly 10% of the DPM m a ss to around 1%. Particulate

em issions from num erous vehicles tested using low-sulfur fuel were found to have a sulfate content of only about 1% (Yanowitz et al., 1999). W a ter content is on the order of 1.3 tim es the

am ount of sulfate (W all et al., 1987).
Metal com pounds and other elem ents in the fuel and engine lubrication oil are exhausted
as ash. Hare (1977) exam ined 1976 Caterpillar 3208 and Detroit Diesel Corporation 6V-71 engines and found the m o st abundant elem ents em itted from the 6V-71 engine were silicon, copper, calcium , zinc, and phosphorus. From the Caterpillar engine the m o st abundant elem ents were lead, chlorine, m a nganese, chrom i um , zinc, and calcium

. Calcium
, phosphorus, and zinc were present in the engine lubrication oil. The two-stroke 6V-71 engine had higher engine
lubrication oil em issions and therefore em itte d higher levels of zinc, calcium , and phosphorus than the Caterpillar 3208 engine. Other elem ents m a y have been products of engine wear or

contam inants from the exhaust system

. Springer (1979), in his study of 1977 Mack ETAY(B)673A and Caterpillar 3208 (EGR) engines, found that calcium was the m o st abundant

m e tallic elem ent in DPM sam p les, with levels ranging from 0.01 to 0.29 wt% of the DPM. Phosphorus and silica were the next m o st abunda nt elem ents reported, and sodium , iron, nickel,
barium , chrom i um , and copper were either present at very low levels or were below detection
lim its. Roughly 1 wt% of the total DPM was represented by the analyzed m e tals. There was no
consistent difference in m e tal em issions between the engines tested by Springer or between

m odes. Springer tested both engines on a 13-m ode steady-state test. Dietzm ann and co-workers

2-62(1980) exam ined m e tal em ission rates from four HD vehicles tested using the UDDS chassis cycle. For the single two-stroke engine te sted (1977 Detroit Diesel Corporation 8V-71),
calcium , phosphorus, and zinc em ission rates were m o re than 10 tim es higher than m e tal levels
observed for three 1979 m odel year four-stroke engi nes because of higher engine lubrication oil

em issions. Metals accounted for 0.5% to 5% of total DPM, depending on engine m odel. In addition to these studies, other source profiles for HD diesel engine em issions report levels of
chrom i um , m a nganese, m e rcury com pounds, and nickel at levels above the detection lim it (Cooper et al., 1987).
In m o re recent studies, Hildem ann and co-workers (1991) exam ined m e tals in DPM from
the sam e two 1987 trucks (four-stroke engines) studied by Rogge and co-workers (1993).
Alum inum , silicon, potassium , and titanium were the only m e tals observed at statistically significant levels. Taken together these m a de up less than 0.75 wt% of total DPM m a ss. Lowenthal and co-workers (1994) also report m e tals em ission rates for a com posite sam p le of
several diesel vehicles. The m o st abundant m e tals were zinc, iron, calcium , phosphorus, barium
, and lanthanum

. Together these represented less than 0.3% of total DPM m a ss, with an em issions
rate of 3.3 m g /m i. Norbeck and co-workers (1998b) report engine transient test em issions of

m e tals for a 1991 Cum m i ns L10 engine. Silicon, iron, zinc, calcium , and phosphorus were observed and together m a de up about 0.5% of total DPM, with an em issions rate of 1.2 m g /bhphr.

2.2.8.1. Organic and EC Content of Particles
2.2.8.1.1

. Measurement of the organic and EC fraction. Various m e thods have been used to quantify the organic fraction of DPM. The m o st com m on m e thod has been Soxhlet extraction with an organic solvent. Following extraction, the solvent can be evaporated and the m a ss of
extracted m a terial (the SOF) determ ined, or alternatively the PM f ilter is weighed bef o re and
after extraction and the extracted m a terial can be further analyzed to determ ine concentrations of individual organic com pounds. Vacuum oven sublim ation is used to m easure a com p arable quantity, the volatile organic f r action (VOF), which can be f u rther speciated by GC with a f l am e ionization detector. Other m e thods have al so been em ployed, including therm a l m e thods,

m i crowave extraction, sonication with an organic solvent, supercritical fluid extraction,
therm ogravim etric analysis, and therm a l desorption GC. Abbass et al. (1991) com p ared various

m e thods, including vacuum oven sublim ation and 8 hours of Soxhlet extraction, with 4:1
benzene/m e thanol solvent for determ ination of SOF and found reasonably good agreem ent between the two m e thods. The VOF value was typically 10% higher; however, this variation
was less than the coefficient of variation between m easurem ents using the sam e m e thod.

2-63Levson (1988) reviewed literature regarding the extraction efficiency of various solvents and found contradictory results in m a ny cases. He concluded that there is strong evidence that
the m o st com m only used solvent, dichlorom e thane, leads to poor recoveries of higher m o lecular weight PAH. More recently, Lucas et al . (1999) reported the effect of varying
dichlorom e thane/benzene ratios in the solven t (from 25% to 100% dichlorom e thane) and changing extraction tim es and found that the m o st effective extraction (i.e., the largest extracted

m a ss) utilized a 70% dichlorom e thane/30% benzene m i xture and extraction tim es several tim es longer than the com m only used 8-hour extraction period. Extractions of 70 hours using pure
dichlorom e thane were found to result in about twice as m u ch SOF as extractions of only 12
hours. Between 6 and 24 hours of extraction tim e (the typical range of extraction tim es used), the SOF recovered increased by about one-third. Using the m o st effective extraction conditions (Soxhlet, 70 hours, 70:30 dichlorom e thane:benzene ratio), Lucas et al. (1999) were able to

extract m o re than 90% of the total particulate m a ss. Other researchers have investigated the relative quantities of m a ss rem oved by sequential extraction by polar, m oderately polar, and nonpolar solvents. The extracted nonpolar fraction (cyclohexane) ranged from 56% to 90% of the SOF, the m oderately polar (dichlorom ethane)
from 6% to 22%, and the polar fraction (aceton itrile) from 4% to 29% (Dietzm a nn et al., 1980).
W a ter and sulfate are not soluble in cyclohexane or dichlorom e thane but are soluble in
acetonitrile. Although the reports on the extraction efficienci es for PAHs are in part contradictory, it appears that Soxhlet extraction and the binary solvent system com posed of arom atic solvent and alcohol yield the best recovery of PAHs, as determ ined by C-B[ a ]P 14 (benzo[a]pyrene) spiking experim e nts (Schuetzle and Perez, 1983). Lim ited recovery studies have shown that there is little degradation or loss of diesel POM on the HPLC colum n. More than 90% of the m a ss and
70% to 100% of the Am es S. typhimurium -active m a terial injected onto the colum n has been recovered (Schuetzle et al., 1985).
Two therm a l m e thods of organic and EC analysis include therm a l optical reflectance (TOR) and therm a l optical transm ittance (TOT). The extractable portion of total carbon, although com m only used as a m easure of organic com pound content, is not equivalent to the OC
fraction as m easured by TOR or TOT. In addition, m e thodological differences between TOR and TOT also give rise to significant differe nces in the fraction of total carbon reported as organic and EC (Birch, 1998; Norris et al., 2000; Chow et al., 2000). Although total carbon reported using TOR or TOT provides results that are com p arable (within 10%) (Norris et al., 2000) the EC content of sam p les analyzed by TOR is higher than that m easured by TOT. This difference is prim arily attributed to the tem p erature used to evolve carbon from the quartz filter
onto which it is collected. In an analysis of urban PM 2.5 sam p les, Norris et al. (2000) found that

2-64the EC content of sam p les analyzed by TOR was a factor of two higher than the EC content of
the sam e sam p les analyzed by TOT. Experim e nts are ongoing to test specific source m a terials (including DPM) because som e of the difference between m e thods appears to depend on the type of OC present on the sam p le. The analytical technique used to m easure OC and EC can have a significant effect on the quantity of reported. In the discussion that f o llows, every ef f o rt has been m a de to com p are only studies using com p arable m e thods and to state the analysis m e thod em ployed.

2.2.8.1.2. Trends in SOF emissions. SOF em ission values are highly dependent on the test cycle used. Various studies have shown that SOF generally increases at light engine loads and high engine speeds because these conditions lead to low exhaust tem p eratures, where fuel and oil are not as effectively oxidized (Scholl et al., 1982; Kittelson, 1998; Springer, 1979; Schuetzle and Perez, 1983; Martin, 1981b; Shi et al., 2000). These conditions are m o re typically observed in LD diesel vehicle applications, and thus DP M from these vehicles typically has a higher SOF

com ponent than HD diesel vehicles (Norbeck et al., 1998c). Acceleration m odes norm a lly cause
increased em ission of EC and an increase in total DPM em issions, whereas organic com ponents

are m o re dom inant when m o toring (W achter, 1990). Additionally, cold-start test em issions of SOF have been shown to be approxim a tely 25% higher than hot-start em issions (W achter, 1990). The quantity of sulfur in diesel fuel has been suggested to have a role in the quantity of
SOF em itted (Sienicki et al., 1990; Tanaka et al., 1998). Sienicki et al. (1990) reported an approxim a te 25% increase in SOF when sulfur concentrations are increased from 0.08% to
0.33%. The cause is unclear but several explan ations have been put forth, including increased absorption of organic com pounds from the vapor phase onto the DPM by sulfates or sorbed
sulfuric acid. Alternatively, it has been proposed that the m easured SOF m a y include som e sulfate, so that the apparent increase in organic m a terial is due instead to sulfate. Other fuel
effects include an increase in SOF em issions with a higher T90 (or T95) and with an increase in

arom atic content (Barry et al., 1985; Sienicki et al., 1990; Tanaka et al., 1998; Rantanen et al., 1993). Figures 2-31 and 2-32 show SOF em issions as a function of year for transient em issions tests on chassis and engines, respectively. Both figures suggest a significant decline in SOF

em issions of approxim a tely a factor of 5 sin ce about 1980. The highest SOF em issions are for two-stroke engines built in the 1970s (up to approxim a tely 1.2 g/m i ). These data indicate that
SOF em ission factors for newer m odel year vehi cles are lower than SOF em ission factors for pre-1990 m odel year vehicles and that this decrease is sim ilar to that observed for em issions of total DPM by m odel year. In a recent test of six pre-1976 HDDVs, Fritz et al. (2001) reported

2-651. 4

1. 2

1.0

0. 8

0. 6

0. 4

0. 2

0.0

1975 1980 1985 1990 1995 2000 W a rn er- S el ph et al

. , 1 984 D i etz m a nn et al ., 198 0 Gr a b o s k i et al.

, 19 98 R o g ge et al ., 199 3

E n gin e M odel Y ear

Figure 2-31. Trend in SOF emissions based on chassis dynamometer testing of HD diesel vehicles. Warner-Selph and co-w orkers: dichloromethane for 8 hours. Dietz m an and cow o rkers: hexane follow e d by dichloromethane, extraction times not reported. Graboski and co-w orkers: VOF by vacuum sublimation at 225 B C for 2.5 to 3 hours. Rogge and cow o rkers: cyclohexane follow e d by a benz ene/2-propanol mixture that may extract significantly more organic matter.

0. 8

0. 6

0. 4

0. 2

0. 0

19 76 19 78 19 80 19 82 19 84 19 86 19 88 19 90 19 92 19 94 19 96 Ull m an et al

. , 1 984 M c Car t hy et al.

, 1992 Pe r e z a nd W i ll iam s , 1 989 Need ham et al

. , 19 89 G r abosk i

, 1998 Sp r e en et al.

, 199 5 Si eni c k i et al.

, 1990 M a r t i n , 1 981 M i tch e l l e t a l ., 1 9 9 4 Ba r r y et al.

, 1985

En g i ne M o d e l YearSOF(g/bhp-h)SOF orVOF(g/mi)

Figure 2-32. Trend in SOF emissions for transient engine dynamometer testing of HD diesel engines. Various extraction methods used; see Table 2-8.

2-66the volatile organic fraction (VOF) ranged from 0.4 g/ m i to 4.5 g/ m i

. These data highlight the wide range in em ission rates for of OC, as have been observed for total PM. Steady-state testing conducted on late-1970s e ngines reported SOF at levels between 0.1 and 0.9 g/bhp-hr, whereas engines from the late 1980s and 1990s all em itted 0.03 g/bhp-hr or less (Table 2-8). Hori and Narusawa (1998) m easured em issions from engines produced two decades apart, using identical analytical proce dures, and found that SOF em ission factors and the percentage contribution of SOF to DPM were lower in the new engine com p ared with the old engine, under all tested engine load and speed c onditions and with different fuels. The authors reported that the decrease in SOF was due to lower em issions of both lubricating oil and
unburned fuel. To m eet the 1991 and 1994 U.S. em ission standards, SOF em ission rates would
need to be reduced from the levels of the previous decade, although one m a y expect differences
in SOF f r actions of DPM with transient cycles used to determ ine com p liance with em ission standards verus steady-state conditions used in earlier test program s (Kawatani et al., 1993;
W achter, 1990). Finally, in the past three decades, for econom ic reasons engine m a nufacturers
have m a de efforts to reduce oil consum ption and increase the fuel efficiency of diesel engines, both of which would be expected to reduce SOF em issions. Problem s in achieving SOF
reductions from two-stroke engines were one factor leading to the phaseout of these engines for on-road use during the 1990s. No data are available prior to 1976 on SOF em issions from HD
diesel vehicles. The engine technology changes that occurred between the m i d-1950s and m i d1970s (high-pressure direct injection and turbocha rging, prim arily) m i ght be expected to increase the efficiency of com bustion and thereby reduce fuel-related SOF. SOF em issions levels in the

m i d- to late 1970s m a y be used as a conservative (low) estim ate of SOF em issions during the preceding two decades.
The fraction of DPM attributed to SOF from chassis dynam om e ter studies also shows a decreasing trend over tim e, from SOFs that ranged up to approxim a tely 50% in the 1980s to 20% SOF or less in the 1990s (Figure 2-33). The recent study by Fritz et al. (2001) reported the
fraction of DPM attributed to VOF from 10% to 60% for HDDVs of m odel years 1951-1974. The wide range in SOF as a percent of DPM displayed in Figure 2-33 is suspected to result from factors such as engine deterioration and test cycl e. The vehicle em issions data reported in Figure 2-33 do not overrepresent buses that are likely to em it DPM with a greater fraction of SOF than
other vehicles. Figure 2-34 presents SOF as a fraction of DPM from the sam e engine dynam om e ter studies reported in Figure 2-32. These data do not reflect a downward trend in
SOF as a fraction of DPM. Because sim ilar extraction m e thods were used in reports of the SOF in both the chassis and engine dynam om e ter studies, this does not appear to be a source of the wide variability observed in the f r action of SOF reported. In som e of the engine studies,
im proved air:fuel ratio control was tested in an attem p t to lower carbonaceous DPM form ation.

2-6710 0

80

60

40

20

0 19 75 19 80 19 85 19 90 19 95 20 00 W a rner-Sel ph

, 19 84 Di etz m a nn, et al

. , 198 0 G r ab os ki

, et a l

. , 1 998 Ro gge , e t al

. , 19 93

E n gin e M o del Y e ar

Figure 2-33. Trend in SOF emissions as a percent of total PM based on chassis dynamometer testing of HD diesel vehicles. Warner-Selph and co-w orkers: dichloromethane for 8 hours. Dietz m an and co-w orkers: hexane follow e d by dichloromethane, extraction times not reported. Graboski and co-w orkers: VOF by vacuum sublimation at 225 B C for 2.5 to 3 hours. Rogge and co-w orkers: cyclohexane follow e d by a benz ene/2-propanol mixture that may extract significantly more organic matter.

100

80

60

40

20

0 197 6 197 8 198 0 198 2 198 4 198 6 198 8 199 0 199 2 199 4 199 6

E n gin e M o del Y earSOF,%ofTotalPMSOF,%of TotalPM

Figure 2-34. Trend in SOF emissions as a percentage of total PM from engine dynamometer testing. Data are from Table 2-8. (See Figure 2-32 for figure key.)

2-68Therefore, substantial differences in SOF as a percent of total DPM could be the result of different engine technology or test conditions. The engine dynam o m e ter results presented in Figures 2-32 and 2-34 are from new, or relativ ely new, engines, that is, engines with no deterioration, whereas the older engines tested on a chassis dynam o m e ter m a y have experienced significant deterioration that would increase SOF em issions as a percent of DPM. One of the

m a in differences suspected for the lack of a decreasing trend in the percent of SOF in the engine dynam o m e ter studies is the test cycle used. The engine dynam o m e ter tests typically include test

m odes, such as high speed and low load, or low-speed lugging m odes, that produce m u ch higher SOF relative to DPM than the driving cycles used on the chassis tests. It appears that as a fraction of total DPM, SOF from new m odel year HD diesel vehicles
is lower than that from older (pre-1990) HD diesel vehicles. However, as with total DPM

em issions, a wide range in the fraction of SOF can be observed under different driving
conditions and from vehicles with extensive engine wear. In general, DPM em issions have a lower fraction of organic m a tter com p ared to gasoline PM (Table 2-13). Recent testing of HD engines at the Desert Research Institute suggests that the OC fraction of DPM is approxim a tely 19%, whereas earlier studies reported in the U.S. EPA SPECIATE database suggest a slightly higher organic fraction of DPM from HD diesel vehicles, ranging from 21% to 36%. The SPECIATE database represents older vehicles th at, as discussed above, tend to have higher SOF

em issions. The OC em issions from LD diesel vehicles recently reported by Norbeck et al. (1998c) and those reported by the U.S. EPA SPECI ATE suggest that LD diesel vehicles em it DPM with a slightly higher organic content than that from HD diesel vehicles, ranging from 22%
to 43%. Gasoline engine PM em issions have recently been analyzed at the Desert Research Institute by Fujita et al. (1998) and W a tson et al. (1998) for hot stabilized, visibly sm oking vehicles, and cold-starts. These data all indicate that LD gas vehicles em it PM with a higher fraction of organic m a tter than diesel vehicles, with the highest organic content m easured from

sm oking and high-em itting gasoline vehicles (averaging 76% OC). One new f i nding f r om the data reported by Fujita et al. (1998) is the roughl y equivalent em ission of organic and EC from
cold-start em issions of gasoline vehicles. Additional inform ation is needed to characterize a range of OC f o r DPM f r om sm oking and high-em itting diesel vehicles as well as cold-start HD
diesel vehicles.

2.2.8.1.3. Trends in EC content. Because EC is a m a jor com ponent of the chem ical source profile of DE, it is com m only used to determ ine the contribution of diesel vehicles to am bient
PM sam p les (i.e., in source apportionm ent via chem ical m a ss balance m odeling). EC is not, strictly speaking, a regulated pollutant, and so EC em issions are not routinely m easured in tests of diesel vehicles and engines. The scant da ta available on m easured EC em issions from HD

2-69Table 2-13. Organic and elemental carbon fractions of diesel and gasoline engine PM exhaust

Engine type % OC % Elemental

carbon
HD diesel engines a 19 ± 8 75 ± 10 HD diesel engines (SPECIATE) b 21-36 52-54 LD diesel engines c 30 ± 9 61 ± 16 LD diesel engines (SPECIATE) b 22-43 51-64 Gasoline engines (hot stabilized) a 56 ± 11 25 ± 15 Gasoline engines (sm oker and high em itter) a,c 76 ± 10 7 ± 6 Gasoline engines (cold start) a 46 ± 14 42 ± 14

a Fuji t a et al ., 1998, and W a t s on et al ., 1998.

b U.S. EPA SPECIATE database.

c Norbeck et al ., 1998c.

diesel vehicles are plotted in Figure 2-35. Diffe rent analytical m e thods were em ployed for these studies, m a king the com p arison of em ission rates di fficult. Results from the three studies, all perform e d on HD trucks, suggest a decline in EC em ission rates by m odel year since the early
1980s. In a study conducted in 1992, four HD vehi cles of unknown vintage were tested and a

com b ined EC em ission rate of 0.81 g/m i was reported, which is consistent with the 1990
tim efram e in Figure 2-35 (Lowenthal et al., 1994). EC as a percentage of total DPM in these studies ranged from 30% to 90%, m o st likely as a result of different testing cycles and different engines and different analytical m e thods. Figure 2-36 presents these data as EC fraction of total fine PM. The EC content of DPM varied widely in the 1980s from approxim a tely 20% to 90%, whereas in m o re recent years, the data suggest a sm aller range in the EC fraction, from approxim a tely 50% to 90% (with one data point at 30%). Recent em ission profiles for HD di esel vehicles suggest that 75% ± 10% of the DPM is attributable to EC, whereas approxim a tely 25% of gasoline PM is com posed of EC, except for PM em issions during gasoline vehicle cold-starts, which were found to have an EC content of approxim a tely 42% (Table 2-13). These data also provide evidence that newer m odel year HD engines generally em it DPM that is m o re rich in EC than older HD engines.

2.2.8.2. PAHs and Nitro-PAH Emissions PAHs, nitro-PAHs, and oxidized derivatives of these com pounds have attracted
considerable attention because of their known m u tagenic and, in som e cases, carcinogenic

2-7014

12

10

8

6

4

2

0 Benz o( a)pyr ene 1- Ni tropyrene

1975 1980 1985 1990 1995 2000

E n gin e M o d e l Y e ar

Figure 2-35. EC emission rates for diesel vehicles.

10 0

90

80

70

60

50

40

30

20

10

0 82 84 86 88 90 92 94 96 Z i el i n s k a , e t al

. , 19 98 Sc h aue r, e t a l

., 19 99 Ro gge , e t al

. , 19 93

E n gin e M o del Y e ar

Figure 2-36. EC content as percent of fine PM for DPM samples obtained in chassis dynamometer studies.EC, %ofTotal PM1-NPorB(a)P Emission, ug/mi

2-71character (National Research Council, 1982). In this section, PAH and nitro-PAH
concentrations and em ission rates and trends in em issions over tim e are presented.

2.2.8.2.1. PAHs identified in DE. At least 32 PAHs have been identified in the exhaust of LD diesel vehicles and HD diesel vehicles (Table 2-14) (W atson et al., 1998; Zielinska et al., 1998). Table 2-15 lists the PAHs and thioarenes identified in three LD diesel vehicles’ DPM extracts, reported as ng/g of DPM (Tong et al., 1984). SOF fractions accounted for 11% to 15% of the
total DPM m a ss for the LD diesel vehicles reported by Tong et al. (1984), which is lower than the LD diesel vehicles organic fraction re ported by Norbeck et al. (1998c) in Table 2-13.
Am ong the PAHs reported by W a tson et al. (1998) and Zielinska et al. (1998), the higher

m o lecular weight com pounds (pyrene through coronene) that are expected to partition to the particle phase have em ission rates from HD dies el vehicles ranging from below detection lim its up to 0.071 m g /m i. HD diesel vehicle em ission rates for the lower m o lecular weight PAHs ranged up to 2.96 m g /m i for dim e thylnaphthalenes. In general, am ong the vehicles tested, PAH

em ission rates were higher for LD diesel vehicles com p ared with HD diesel vehicles. Table 2-16 presents em ission rates of four representative pa rticle-phase PAHs from HD diesel vehicles, LD diesel vehicles, and gasoline (with and without cat alytic converter) engines. Em ission rates for benzo[a]pyrene were higher in diesel em issions com p ared with gasoline em issions, except for the report by Rogge et al. (1993), who used extr action m e thods different from those in other studies (discussed above).

2.2.8.2.2. Nitro-PAHs identified in DE. Positive isom er identif ication f o r 16 nitro-PAHs has
been m a de utilizing the GC retention tim es of authentic standards and low- and high-resolution

m a ss spectra as identification criteria. These include 1-nitropyrene; 2-m e thyl-1-nitronaphthalene; 4-nitrobiphenyl; 2-nitrofluorene; 9-n itroanthracene; 9-m e thyl-10- nitroanthracene; 2- nitroanthracene; 2-nitrophenanthrene; 1-m e thyl -9-nitroanthracene; 1-m e thyl-3-nitropyrene; 1-

m e thyl-6-nitropyrene; 1-m e thyl-8-nitropyren e; 1,3-, 1,6-, and 1,8-dinitropyrene; and 6- nitrobenzo[a]pyrene. In addition, two nitrated heterocyclic com pounds were identified, 5- and 8- nitroquinoline. Forty-five additional nitro-PAHs were tentatively identified in this diesel particulate extract (Paputa-Peck et al., 1983). The concentration of nitro-PAHs adsorbed on diesel particles varies substantially from sam p le to sam p le. Usually 1-nitropyrene is the predom inant com ponent, and concentrations ranging from 7 to 165 μg/g of particles are reported (Levson, 1988). Table 2-17 gives the approxim a te concentrations of several of the abundant nitro-PAHs quantified in the early 1980s LD diesel particulate extracts (with the exception of

2-72Table 2-14. Emission rates of PAH (mg/m i) from LD and HD diesel vehicles
PAH L ight-duty diesel Heavy-duty diesel Napht hal e ne 5.554 ± 0.282 2.451 ± 0.154
2-M e napht hal e ne 3.068 ± 0.185 2.234 ± 0.152
1-M e napht hal e ne 2.313 ± 0.134 1.582 ± 0.103
Di m e t h y l napht hal e nes 5 .065 ± 0.333 2.962 ± 0.488
B i pheny l 0 .743 ± 0.041 0.505 ± 0.037
2-M e t h y l bi pheny l 0 .203 ± 0.015 0.049 ± 0.024
3-M e t h y l bi pheny l 1 .048 ± 0.063 0.401 ± 0.036
4-M e t h y l bi pheny l 0 .447 ± 0.028 0.144 ± 0.021
Tri m et hy l n apht hal e nes 6 .622 ± 0.563 1.940 ± 0.221 Acenapht hy l e ne 0.422 ± 0.024 0.059 ± 0.087 Acenapht hene 0.096 ± 0.008 0.030 ± 0.040 Phenant h rene 1.411 ± 0.072 0.084 ± 0.011
Fl uorene 0 .442 ± 0.038 0.066 ± 0.022
M e t h y l fl uorenes 1.021 ± 0.091 0.071 ± 0.055
M e t h y l phenant hrenes 1.115 ± 0.064 0.124 ± 0.069
Di m e t h y l phenant hrenes 0.637 ± 0.047 0.090 ± 0.096 Anthracene 0 .246 ± 0.025 0.052 ± 0.016 9-Methylanthracene 0 .013 ± 0.002 0.434 ± 0.082
Fl uorant h ene 0 .213 ± 0.014 0.044 ± 0.026 Py rene 0.245 ± 0.020 0.071 ± 0.017
M e t h y l (py r enes/ f l uorant h enes) 0 .548 ± 0.045 0.022 ± 0.082
B e nzonapht hot hi ophene 0.002 ± 0.002 0.001 ± 0.027 Benz[a]anthracene 0 .020 ± 0.005 0.066 ± 0.046
C h ry sene 0.029 ± 0.005 0.009 ± 0.021
B e nz[b+j+k] f l uorant h ene 0 .056 ± 0.005 0.009 ± 0.022
B e nzo[e] py rene 0.019 ± 0.003 0.010 ± 0.014
B e nzo[a] py rene 0.013 ± 0.004 0.013 ± 0.044 Indeno[1,2,3-cd] p y r ene 0 .010 ± 0.003 0.001 ± 0.037 Dibenzo[a]anthracene 0 .002 ± 0.003 0.000 ± 0.053
B e nzo[b] chry sene 0.001 ± 0.002 0.001 ± 0.027
B e nzo[ghi ] p erl y ne 0.018 ± 0.004 0.013 ± 0.048
C o ronene 0.006 ± 0.006 0.001 ± 0.095

2-73Table 2-15. Polycyclic aromatic hydrocarbons identified in extracts of diesel particles from LD diesel engine exhaust

Compound Molec. wt

. Concentration ng/ m g extract Acenapht hy l e ne 152 30 Tri m et hy l n apht hal e ne 170 140–200 Fl uorene 166 100–168 Di m e t h y l bi pheny l 182 30–91 C 4 -Napht hal e ne 184 285–351 Tri m et hy l b i pheny l 196 50 Di benzot hi ophene 184 129–246 Phenant h rene 178 2,186–4,883 Anthracene 178 155–356 M e t h y l di benzot hi ophene 198 520–772 M e t h y l phenant hrene 192 2,028–2,768 Methylanthracene 192 517–1,522 Et hy l phenant hrene 206 388–464 4H-C y c l opent a[ def ] phenant hrene 190 517–1,033 Et hy l d i b enzot h i ophene 212 151–179 2-Pheny l napht hal e ne 204 650–1,336 Dim e thyl(phenanthrene/anthracene) 206 1,298–2,354 Fl uorant h ene 202 3,399–7,321 B e nzo[ def ] d i b enzot h i ophene 208 254–333 Benzacenaphthylene 202 791–1,643 Py rene 202 3,532–8,002 Et hy l m et hy l (phenanthrene/anthracene) 220 590–717

M e t h y l (fl uorant h ene/ py rene) 216 1,548–2,412 B e nzo[ a ] f l uorene/ benzo[ b ] f l uorene 216 541–990 B e nzo[ b ] n apht ho[2,1- d ] t hi ophene 234 30–53 C y cl opent apy r ene 226 869–1,671 B e nzo[ ghi ] f l uorant h ene 226 217–418 B e nzonapht hot hi ophene 234 30–126 Ben z [ a ]anthracene 228 463–1,076 C h ry sene or t r i pheny l e ne 228 657–1,529 1,2-B i napt hy l 254 30–50 M e t h y l benz[ a ]anthracene 242 30–50 3-M e t h y l chry sene 242 50–192 Phenyl(phenanthrene/anthracene) 254 210–559 B e nzo[ j ] f l uorant h ene 252 492–1,367 B e nzo[ b ] f l uorant h ene 252 421–1,090 B e nzo[ k ] f l uorant h ene 252 91–289 B e nzo[ e ] p y r ene 252 487–946 B e nzo[ a ] p y r ene 252 208–558 B e nzo[ ah ]anthracene 278 50–96 Indeno[1,2,3-[ cd ] p y r ene 276 30–93 B e nzo[ ghi ] p ery l ene 276 443–1,050 Di benzopy rene 302 136–254 Source: Tong et al ., 1984.

2-74Table 2-16. Emission rates of particle-bound PAH (μg/mi) from diesel and gasoline engines

PAH D iesel engines G asoline engines
HDD L DD Noncatalyst C atalyst (a) ( b) (c) ( a) (d) ( c) (e) ( a) (c) Py rene 71 17.6 36.2 245 66 49.6 4 5 248 4.0 Fluoranthene 4 4 27.2 20.8 213 50 77.3 3 2 196 3.6 Benzo[ a] py rene 13 <0.1 2.1 1 3 N A 69.6 3 .2 1.0 3 .0 Benzo[ e] py rene 10 0.24 4.2 1 9 N A 73.3 4 .8 1.0 3 .6 (a) W a t s on et al ., 1998 i n cl uded gas-phase PAH . (b) W e st erhol m et

al ., 1991. (c) R ogge et al ., 1993. (d) Sm i t h , 1989; 1986 M e rcedes B e nz. (e) Al sberg et al ., 1985.

3-nitrobenzanthrone, reported recently) in μg/g of particles. Concentrations for som e of the nitro-PAHs identified range from 0.3 μg/g for 1,3-dinitropyrene to 8.6 μg/g for 2,7-dinitro-9- fluorenone and 75 μg/g for 1-nitropyrene. More recent nitro-PAH and PAH data for HD diesel engines are reported in units of g/bhp-hr or m a ss/volum e of exhaust, m a king it im possible to directly com p are them to the older data (Norbeck et al., 1998b; Bagley et al., 1996, 1998;
Baum gard and Johnson, 1992; Opris et al., 1993; Hansen et al., 1994; Harvey et al., 1994; Kantola et al., 1992; Kreso et al., 1998; McClure et al., 1992; Pataky et al., 1994).

2.2.8.2.3. PAH and nitro-PAH emission changes over time. It is difficult to com p are PAH

em issions from different studies because not all investigators analyze for total PAH or the sam e suite of PAH com pounds. Most studies have reported em issions of B[ a ]P or 1-nitropyrene (1- NP) because of their toxicological activity. The results of chassis dynam o m e ter studies in which B[ a ]P or 1-NP were m easured are displayed in Figure 2-37. Dietzm ann and co-workers (1980)

exam ined four vehicles equipped with late 1970s turbocharged DI engines. Em issions of B[ a ]P
from particle extracts ranged from 1.5 to 9 μg/m i. No relationship between engine technology (one of the engines was two-stroke) and B[ a ]P em issions was observed. Rogge and co-workers (1993) reported total particle-associated PAH and B[ a ]P em issions from two 1987 m odel year trucks (averaged together, four-stroke and turbocharged engines). The total particle-phase PAH

em ission rate was 0.43 m g /m i and the B[ a ]P em ission rate was 2.7 μg/m i. Particle-phase PAH in the Rogge et al. (1993) study accounted for approxim a tely 0.5% of total DPM m a ss. Schauer and co-workers (1999) recently reported a particle-phase PAH em ission rate of 1.9 m g /m i (accounting for about 0.7% of total DPM m a ss) for a 1995 MD turbocharged and aftercooled
truck. B[ a ]P em issions were not reported, but em i ssions of individual species of sim ilar

2-75Table 2-17. Concentrations of nitro-PAHs identified in LD diesel particulate extracts

Nitro-PAH a Concentration (μ g/g of particles)
4-nitrobipheny l 2 .2

2-nitrofluorene

~ 1.8

2-nitroanthracene 4 .4

9-nitroanthracene 1 .2 9-nitrophenanthrene 1 .0 3-nitrophenanthrene 4 .1

2-m e thy l -l-nitroanthracene 8 .3

1-nitrofluoranthene 1 .8

7-nitrofluoranthene 0 .7

3-nitrofluoranthene 4 .4

8-nitrofluoranthene 0 .8 1-nitropy rene 18.9; 75 b
6-nitrobenzo[ a] py rene 2.5 1,3-dinitropy rene b 0.30 1,6-dinitropy rene b 0.40 1,8-dinitropy rene b 0.53
2,7-dinitrofluorene c 4.2; 6.0
2,7-dinitro-9-fluorenone c 8.6; 3.0

3-nitrobenzanthrone d 0.6 to 6.6

a From Cam pbell and Lee (1984) unless noted otherwis e. Concentrations recalculated from μg/g of

ex tract to μg /g o f p a rticles u s in g a v a lu e o f 4 4 % fo r ex tractab le m a terial (w/w).

b From Paput a-Peck et al , 1983.

c From Schuet z l e , 1983.

d From Eny a et al ., 1997 (Isuzu M odel 6HEL 7127cc).

2-7614

12

10

8

6

4

2

0 Benz o( a)pyr ene 1- Ni tropyrene

1975 1980 1985 1990 1995 2000

E n gin e M o d e l Y e ar

Figure 2-37. Diesel engine emissions of benz o[ a ]pyrene and 1-nitropyrene measured in chassis dynamometer studies.

Source: Schuet z l e and Perez, 1983; Zi el i n ska et al ., 1988; Kado et al ., 1996; Di et zm ann et al ., 1980; W a rner-Sel ph and Di et zm ann, 1984; R ogge et al ., 1993; Schauer et al ., 1999.

m o lecular weight were approxim a tely 10 μg/m i. Schauer et al. (1999) also reported a gas-phase
PAH em ission rate of 6.9 m g /m i for the sam e truck. Measurem ents of particle- and gas-phase PAHs conducted for the Northern Front Range Air Quality Study in Colorado (Zielinska et al., 1998) showed an average B[ a ]P em ission rate of 13 μg/m i for 15 vehicles ranging from 1983 to
1993 m odel years. The com b ined gas- and particle-phase PAH em ission rate reported for the NFRAQS study was 13.5 m g /m i. B[ a ]P em issions from chassis studies are sum m a rized in Figure 2-37. Zielinska (1999) reports a decreasing trend in particle-associated DE PAH from 11

m easurem ents m a de on vehicles from m odel year 1984 to 1993 with a low correlation coefficient
of 0.29. B[ a ]P em issions reported from diesel engine dynam o m e ter studies are sum m a rized in Figure 2-38. Springer (1979) com p ared B[ a ]P em issions from naturally aspirated and turbocharged engines and found that naturally aspirated engines em itted about 1 μg B[ a ]P/bhphr, and DI and IDI engines em itted about 0.15 μg B[ a ]P/bhp-hr (Table 2-8). The difference between 1 and 0.15 μg/bhp-hr could not be attri buted to specific technology changes. The

m a jority of engine test data indicate that B[ a ]P em issions have generally ranged from approxim a tely 1 to 4 μg/bhp-hr over the past 25 years.1-NPorB(a)P Emission, ug/mi

2-77Em issions reported for 1-NP from diesel engines tested by chassis dynam o m e ter range from 0.1 to 12 μg/m i (Figure 2-37), and diesel engine dynam o m e ter studies report 1-NP

em ission factors ranging from 1 to 4 μg/bhp-hr (Figure 2-38). Too few m easurem ents are available to discern trends in the em ission rates of these com pounds.

30Benzo(a)pyreneor1-Nitropyrene Emission, ug/bhp-h

25

20

15

10

5 T r an si en t T e st B( a ) P S t e a d y -S tate Tes t B ( a)P T r an si en t T e st 1

- NP S t e a d y -S tate Tes t 1

- N P

0

19 70 19 75 19 80 19 85 19 90 19 95 20 00
Eng i ne Mo d e l Ye a r Figure 2-38. Diesel engine dynamometer measurements of benz o[ a ]pyrene and 1- nitropyrene emissions from HD diesel engines.

Source: Data are from Table 2-8.

As discussed in Section 2.2.4, W illiam s et al. (1987) and Andrews et al. (1998) of the University of Leeds have dem onstrated that the solvent-extractable PAH from diesel particulate originates prim arily in the fuel. PAH m o lecules are relatively refractory, so a significant fraction
survives the com bustion process and is exhausted as DPM. These studies have been confirm e d by other research groups (Crebelli et al., 1995; Tancell et al., 1995) that included the use of isotopic labeling of fuel PAH. Additionally, engi ne oil was found to be a reservoir for PAH that originates in the fuel. Pyrosynthesis of PAH o ccurs during very high tem p erature conditions in a diesel engine, and under these conditions m a ny of the DPM and other pyrolysis products are
ultim ately oxidized before exiting the cylinder. Thus, pyrogenic form ation of PAH is thought to
contribute a sm all fraction of the total PAH in dies el engine exhaust. As discussed above, fuel

2-78PAH content is expected to have slowly increased over a 30-year period until 1993, after which PAH content of diesel fuel is expected to have rem a ined constant. Increasing use of catalytic cracking over tim e m a y lead to increasing proportions of PAH in distillates; however, fuel
standards lim it the arom aticity of fuel to 35% (Section 2.2.4). Recently, Norbeck et al. (1998a) reported on the effect of fuel arom atic content on PAH

em issions. Three diesel fuels were used in a Cum m i ns L10 engine: pre-1993 fuel containing
33% arom atic HC and 8% PAH; low arom atic fuel containing a m a xim u m content of 10%

arom atic HC and m a xim u m of 1.4% PAH; and a reform ulated fuel containing 20% to 25%

arom atic HC and 2% to 5% PAH. The investigators found that em ission rates for the lowm o lecular-weight PAHs (PAHs with three or fewer rings) were significantly lower when the engine was tested using the low arom atic fuel com p ared to when the engine was run on the pre1993 or reform ulated fuel (Table 2-18). A lthough em ission rates reported for several higher

m o lecular weight (particle-associated) PAHs were lower (ranging from 4% to 28% lower) for
the low arom atic f u el com p ared with the other two f u els, the dif f e rences were not statistically significant except for coronene.
On the basis of these lim ited data it is dif f i cult to draw a precise, quantitative conclusion regarding how PAH, B[ a ]P, or 1-NP em issions have changed over tim e and in response to fuel and engine changes. A decrease in the em i ssions of PAH from post-1990 m odel year vehicles and engines com p ared with pre-1990 vehicles and engines is suggested by the data; however, the data also suggest that differences in a vehicle’ s engine type and m a ke, general engine condition,
fuel com position, and test conditions can influence the em ission levels of PAH.

2.2.8.3. Particle Size Figure 2-39 shows a generic size distribution for diesel particulate based on m a ss and particle num ber. Approxim a tely 50% to 90% of the num ber of particles in DE are in the ultrafine size range (nuclei-m ode), with the m a jority of diesel particles ranging in size from
0.005-0.05 μ m and the m ode at about 0.02 μ m

. These aerosol particles are form ed from exhaust constituents and consist of sulfuric acid droplets, ash particles, condensed organic m a terial, and prim ary carbon spherules (Abdul-Khalek et al ., 1998; Baum gard and Johnson, 1996). Although
it accounts for the m a jority of particles, ultrafine DPM accounts for only 1% to 20% of the m a ss of DPM. Approxim a tely 80% to 95% of diesel particle m a ss is in the size range from 0.05 to 1.0

μ m , with a m ean particle diam eter of about 0.2 μ m

. The EC core has a high specific surface area of approxim a tely 30 to 50 m 2 /g (Frey and Corn, 1967), and Pierson and Brachaczek (1976) report that after the extraction of adsorbed organic m a te rial, the surface area of the diesel particle core

2-79Table 2-18. Average emission rates for polycyclic aromatic hydrocarbons for different fuel types (units are μg/bhp-hr)

PAH P re-1993 di esel fuel Cetane No. >40 Arom at i c 33% v. PAH 8% wt. L o w aromati c diesel fuel Cetane No. >48 Arom at i c 10% v.

PAH 1.4% wt. Reformulated diesel blend C e t a ne No. 50-55 Arom at i c 20%-25% v. PAH 2%-5% wt. 2,3,5-t r i m et hy l napht hal e ne 283.68 ± 5.27 14.77 ± 2.42 56.21 ± 2.82 Phenant h rene 336.71 ± 9.08 160.92 ± 15.54 220.73 ± 52.68 Anthracene 38.89 ± 1.43 18.54 ± 2.13 26.16 ± 6.86 Methylphenanthrenes/anthracenes 331. 32 ± 16.07 25.17 ± 1.41 111.98 ± 28.74
Fl uorant h ene 128.45 ± 7.60 132.36 ± 18.30 123.07 ± 26.21 Py rene 193.03 ± 16.51 211.19 ± 37.35 206.82 ± 39.04
B e nzo[c] phenant hrene 3 .03 ± 0.24 1.74 ± 0.14 1.54 ± 0.26
B e nzo[ghi ] f l uorant h ene 24.84 ± 2.68 18.93 ± 2.14 16.94 ± 2.31
C y cl opent a[cd] p y r ene 21.44 ± 4.11 26.15 ± 3.12 21.25 ± 3.46 Benz[a]anthracene 16.42 ± 1.67 10.57 ± 1.15 10.96 ± 2.42
C h ry sene + t r i pheny l e ne 17.36 ± 1.66 10.38 ± 0.54 12.20 ± 2.72
B e nzo[b+j+k] fl uorant h ene 31.05 ± 4.17 23.17 ± 1.98 29.18 ± 7.93
B e nzo[e] py rene 16.71 ± 2.72 14.55 ± 1.34 18.99 ± 5.58
B e nzo[a] py rene 20.46 ± 3.27 16.48 ± 1.56 20.59 ± 5.75 Pery l e ne 4.32 ± 0.88 3.71 ± 0.74 4.18 ± 1.16 Indeno[1,2,3-cd] f l uorant h ene 0 .34 ± 0.07 0.21 ± 0.02 0.17 ± 0.00
B e nzo[c] chry sene 0.29 ± 0.05 0.18 ± 0.05 0.14 ± 0.04 Dibenz[a,h]anthracene 0 .93 ± 0.05 0.55 ± 0.10 0.67 ± 0.09 Indeno[1,2,3-cd] p y r ene 19.45 ± 2.71 14.04 ± 1.99 22.16 ± 9.11 Dibenz[a,h+a,c]anthracene 1 .54 ± 0.15 0.87 ± 0.12 1.48 ± 0.67
B e nzo[b] chry sene 0.40 ± 0.01 0.15 ± 0.05 0.27 ± 0.05
B e nzo[ghi ] p ery l ene 49.17 ± 9.63 39.81 ± 7.22 60.74 ± 26.60
C o ronene 9.49 ± 3.13 4.93 ± 0.47 7.48 ± 1.59
Di benzo[a,l ] py rene 2.84 ± 0.45 1.25 ± 0.15 2.31 ± 0.48
Di benzo[a,e] py rene 1.10 ± 0.29 0.61 ± 0.06 1.13 ± 0.15
Di benzo[a,i ] py rene 0.91 ± 0.21 0.27 ± 0.09 0.71 ± 0.15
Di benzo[a,h] py rene 1.33 ± 0.25 0.75 ± 0.07 0.84 ± 0.20

Source: Norbeck et al., 1998a.

2-80Figure 2-39. Particle siz e distribution in DE.

Source: Kittelson, 1998.

is approxim a tely 90 m 2 /g. Because these particles have a very large surface area per gram of

m a ss, it m a kes them excellent carriers for adsorbed inorganic and organic com pounds; potentially enhancing penetration of such com pounds to lower portions of the respiratory tract upon inhalation. In addition, ultrafine aerosols can also reach the sam e areas of the lung. Considerable caution is required when reporting particle size m easurem ents from diesel engine exhaust because dilution conditions during the m easurem ent process significantly affect size distributions (i.e., the size distribution is largely a function of how it was m easured), and DPM size distributions obtained in dilution tunnel system s m a y not be relevant to size distributions resulting from the physical transfor m a tion of engine exhaust in the atm o sphere.
Measurem ents m a de on diluted DE typically show higher num bers of nuclei-m ode particles than
do m easurem ents m a de on raw exhaust because of condensation to form nuclei-m ode aerosol upon cooling of the exhaust. To understand particle size distributions em itted from diesel engines, investigators em ploy various dilution techniques, none of which have been
standardized. Dilution ratio, sam p ling tem p erature, hum idity, relative concentrations of carbon
and volatile m a tter, and other sam p ling f actors can theref ore have a large im pact on the num ber
and m a keup of nuclei-m ode particles (Abdul-Kha lek et al., 1999; Shi and Harrison, 1999; Lüders

2-81et al., 1998; Brown et al., 2000). Dilution air te m p erature and hum idity can have a large effect on particle num ber and size distribution, especially in the size range below 0.05 μ m ( a l s o referred to as nanoparticles). Shi and Harrison (1999) report that a high dilution ratio and high
relative hum idity favor the production of ultrafine particles in diesel engine exhaust. Abdul-Khalek et al. (1998) report that an increase in the residence tim e of the exhaust during
dilution resulted in an increase in the num ber of particles in exhaust. Khatri et al. (1978) report increased gas-phase HC condensation to DPM with a decrease in dilution air tem p erature. Som e studies report no peak in diesel particles in the ultrafine size range (Kleem an et al., 2000). Kittelson (2000) reports that nanoparticle form ation can be prevented by an oxidizing catalyst, which burns organic com ponents of the exhaust, m a king them unavailable for nucleation or
condensation to form an aerosol. Experim e nts conducted in a dilution tunnel represent the atm o spheric behavior of DE only under the conditions specific to the dilution tunnel and do not represent the full range of

atm o spheric conditions. Gertler (1999) dem onstrated an increase in 0.02 μ m particles as the
fraction of diesel vehicles in the Tuscarora M ountain tunnel increased from 13% to 78%. These data suggest that the m ode at 0.02 μm for ultrafine DPM from DE is evidenced under som e real-world conditions. Several groups have shown that decreasing sulfur content decreases the num ber of
nuclei-m ode particles m easured in the exhaust, assum i ng tem p erature is low enough and
residence tim e is long enough for nucleation and condensation of sulfate aerosol and water in the dilution tunnel (Baum g ard and Johnson, 1992, 1996; Opris et al., 1993; Abdul-Khalek et al., 1999). The application of this finding to real-world conditions is difficult to predict, as the

num ber of nuclei-m ode particles f o rm ed f r om sulf ate and water in the atm o sphere will be
determ ined by atm o spheric conditions, not by dilution tunnel conditions. W ith all other factors held constant, it appears that reducing fuel sulfur content reduces the number of sulfate nucleim ode particles. Thus, the reduction in on-road fuel sulfur content that occurred in 1993 reduced
the am ount of sulfur dioxide and sulfate available for particle form ation. As discussed above,
the contribution of sulfate to total DPM m a ss ranges from 1% to 5% and is therefore not a substantial portion of DPM m a ss. More controversial is the suggestion that the DPM em ission size distribution from newer technology engines (1991 and later) m a y be shifted to a m u ch higher num ber concentration of
nuclei-m ode particles, independent of fuel su lfur content (Kreso et al., 1998; Abdul-Khalek et al., 1998; Baum gard and Johnson, 1996; Bagley et al., 1996). For example, Kreso and coworkers (1998) com p ared em issions from a 1995 m odel year engine with m easurem ents m a de on
1991 and 1988 m odel year engines in earlier studies (Bagley et al., 1993, 1996). Nuclei-m ode particles m a de up 40% to 60% of the num ber fraction of DPM em issions for the 1988 engine and

2-8297%+ of the DPM from the 1991 and 1995 engines. Num b er concentrations were roughly two
orders of m a gnitude higher for the newer engines. SOF m a de up 25% to 30% of DPM m a ss in the 1988 engine and 40% to 80% of DPM m a ss for the newer engines. Total DPM m a ss was significantly reduced for the newer engines. It was suggested that increased fuel injection pressure leads to im proved fuel atom ization and evaporation, in turn leading to sm aller prim ary carbonaceous particles. Dilution conditions (relatively low tem p erature, low prim ary dilution ratio, long residence tim e of m o re than 3 seconds) strongly favor the form ation of nucleation products. The 1991 and 1988 engines were tested with 100 ppm sulfur fuel whereas the 1995 engine was tested with 310 ppm sulfur fuel, which m a y confound the results to som e extent. The results of Kreso and co-workers (1998) and of Bagley and co-workers (1993, 1996) have been called into question because the high level of SOF em itted by the 1991 engine, particularly at high-load test m odes, was inconsistent with SOF values m easured for other engines using sim ilar types of technology (Last et al., 1995; Ullm an et al., 1995). Kittelson (1998) notes that there is far less carbonaceous DPM form ed in newer engines com p ared with older engines. Accum u lation-m ode particles m a y have provided a high surface area for adsorption of sulfate and unburned organic com pounds. In the absence of this surface area for adsorption, higher num ber concentrations of sm all particles are form ed from nucleation of HCs
and sulfuric acid. A study perform e d at EPA by Pagan (1999) sugge sted that increased injection pressure
can lead to the form ation of m o re nuclei-m ode particles in the exhaust. Particle size distributions

were m easured for diluted exhaust from an engine in which injection pressure could be varied
from roughly 35 to 110 MPa (about 5,000–16,000 psi), com p arable to pressures obtained with injection technology introduced in the 1980s. The dilution system and particle size m e asurem ent setup were identical in all experim e nts, rem oving som e of the uncertainty in earlier studies that

com p ared engine tests perform e d years apart. The results showed a clear increase in the num ber
of nuclei-m ode particles and a decrease in the num ber of accum u lation-m ode particles as injection pressure was increased. This shift did not occur, however, at high engine speeds and loads, but only at low to interm ediate speeds and loads. The increase in num ber concentration of
nuclei-m ode particles was m u ch lower than the two orders of m a gnitude increase reported by Kreso et al. (1998) or Bagley et al. (1996). One m u st use caution in applying the results of Pagan to m odern 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

com m on rail injection was applied. This would likely preclude this particular engine from

m eeting current on-highway PM or NO

x standards. Although som e studies have suggested that increased injection pressure can lead to elevated ultrafine DPM num ber counts, Kittelson et al.

2-83(1999) cite a Germ an study that reported a decrease in ultrafine DPM num ber and m a ss with increasing injection pressure. Although the m a jority of particles in DE from m odern on-road diesel engines are in the ultrafine size range, evidence regarding a change in the size distribution over tim e is unclear. To understand the size distribution of DPM to which people are exposed will require m easurem ents under conditions that m o re closely resem b le am bient conditions.

2.3. ATMOSPHERIC TRANSFORMATION OF DIESEL EXHAUST
Prim ary diesel em issions are a com p lex m i xture containing hundreds of organic and inorganic constituents in the gas and particle phases, the m o st abundant of which are listed in
Table 2-19. The m o re reactive com pounds with short atm o spheric lifetim es will undergo rapid
transform a tion in the presence of the appropriate reactants, whereas m o re stable pollutants can be transported over greater distances. A know ledge of the atm o spheric transform a tions of gaseous and particulate com ponents of diesel em i ssions and their fate is im portant in assessing
environm ental exposures and risks. This section describes som e of the m a jor atm o spheric
transform a tion processes for gas-phase and particle-phase DE, focusing on the prim ary and secondary organic com pounds that are of significance for hum an health. For a m o re

com p rehensive sum m a ry of the atm o spheric transport and transform a tion of diesel em issions,
see W i ner and Busby (1995).

2.3.1. Gas-Phase Diesel Exhaust Gas-phase DE contains organic and inorga nic com pounds that undergo various chem ical and physical transform a tions in the atm o sphere, depending on the abundance of reactants and

m e teorological factors such as wind speed and direction, solar radiation, hum idity, tem p erature, and precipitation. Gaseous DE will react prim arily with the following species (Atkinson, 1988):

• S unlight, during daylight hours

• H ydroxyl (OH) radical, during daylight hours

• O zone (O 3 ), during daytim e and nighttim e

• H ydroperoxyl (HO 2 ) radical, typically during afternoon/evening hours

• G aseous nitrate (NO 3 ) radicals or dinitrogen pentoxide (N 2 O 5 ), during nighttim e hours

• G aseous nitric acid (HNO 3 ) and other species such as nitrous acid (HONO) and sulfuric acid (H 2 SO 4 ).

2-84Table 2-19. Classes of compounds in diesel exhaust

Particulate phase Gas phase
Het e rocy cl i c s, hy drocarbons (C 14 -C 35 ), and PAHs and deri vat i v es: Acids Cycloalkanes Al cohol s Est e rs Al kanoi c aci ds Hal ogenat e d cm pds. n-Al kanes Ket ones Anhydrides Nitrated cm pds. Arom at i c aci ds Sul f onat e s Qui nones Het e rocy cl i c s, hy drocarbons (C 1 -C 10 ), and deri vat i v es: Aci d s C y c l o al kanes, C y cl oakenes Al dehy des Di carbony l s Al kanoi c aci ds Et hy ne n-Al kanes Hal ogenat e d cm pds. n-Alkenes Ketones Anhydrides Nitrated cm pds. Arom at i c aci ds Sul f onat e s Qui nones Elem en tal carb o n Acro lein Inorgani c sul f at es and ni t r at es Am m oni a M e t a l s C a rbon di oxi de, carbon m onoxi de W a ter B enzene 1,3-B u t a di ene Form al dehy de Form ic acid Hy drogen cy ani d e, hy drogen sul f i d e Methane, m e thanol Ni t r i c and ni t r ous aci ds Ni t r ogen oxi des, ni t r ous oxi de Sul f ur di oxi de Toluene Wa t e r Sources: M a uderl y , 1992, whi c h sum m a ri zed t h e work of Li es et al ., 1986; Schuet z l e and Frazi er, 1986; Carey, 1987; Zaebst et al., 1988, updated from recent work by Johnson, 1993; McDonald, 1997; Schauer et al ., 1999.

The m a jor loss process for m o st of the DE em ission constituents is oxidation, which occurs prim arily by daytim e reaction with OH radical (Table 2-20). For som e pollutants, photolysis, reaction with O 3 , and reactions with NO 3 radicals during nighttim e hours are also
im portant rem oval processes. The atm o spheric lifetim es do not take into consideration the potential chem ical or biological im portance of the products of these various reactions. For

exam ple, the reaction of gas-phase PAHs with NO 3 appears to be of m i nor significance as a PAH loss process, but it is m o re im portant as a route of form ation of m u tagenic nitro-PAHs. The reaction products for som e of the m a jor gaseous DE com pounds are listed in Table 2-21 and are discussed briefly below.

2.3.1.1. Organic Compounds The organic fraction of diesel is a com p lex m i xture of com pounds, very few of which
have been characterized. The atm o spheric chem istry of several organic constituents of DE (which are also produced by other com bustion sources) has been studied. A few of these

2-85Table 2-20. Calculated atmospheric lifetimes for gas-phase reactions of selected compounds present in automotive emi ssions w i th important reactive species

Compound Atmospheric lifetime resulti ng from reaction with: OH a O 3 b NO 3 c HO 2 d hv e NO 2 1.3 day s 12 h 24 m i n 2 h 2 m i n NO 2.5 day s 1 m i n 1 .2 m i n 20 m i n — HNO 3 110 day s — — — — SO 2 16 day s >200 y ears > 1.4×10 4 y ears > 600 y ears — NH 3 90 day s — — — — Propane 12 day s >7,000 y ears — — — n-B u t a ne 5.6 day s >4,500 y ears 3 .6 y ears — — n-Oct a ne 1.9 day s — 1 .2 y ears — — Et hy l e ne 1.9 day s 9 day s 1.2 y ears — — Propy l e ne 7 h 1 .5 day s 6 day s — — Acet y l ene 19 day s 6 y ears > 5.6 y ears — — Form al dehy de 1.9 day s >2 - 104 y ears 84 day s 23 day s 4 h Acet al dehy de 0.6 day > 7 y ears 20 day s — 60 h B e nzal dehy de 1.2 day s — 24 day s — — Acrol e i n 0.6 day 60 day s — — — Form i c aci d 31 day s — — — — B e nzene 11 day s 600 y ears > 6.4 y ears — — Tol u ene 2 .5 day s 300 y ears 3 .6 y ears — —

m

- Xy l e ne 7 h 75 y ears 0 .8 y ears — — Phenol 6 h — 8 m i n — — Napht hal e ne f 6.8 h > 80 day s 1.5 y ears — — 2-M e t h y l napht hal e ne f 2.8 h > 40 day s 180 day s — 1-Ni t r onapht hal e ne f 2.3 day s >28 day s 1 8 y ears 1 .7 h Acenapht hene f 1.5 h > 30 day s 1.2 h — — Acenapht hy l e ne f 1.3 h

~ 43 m i n 6 m i n — — Phenanthrene f 11.2 h 41 day s 4.6 h — — Anthracene f 8.6 h — — — — Fluoranthene f ~2 .9 h — ~1 year — — Pyrene f

~ 2 .9 h — ~ 120 day s — —

a For 12-h average concent r at i on of OH radi cal of 1.6×10 6 m o lecu le/cm 3 (Pri nn et al ., 1992).

b For 24-h average O 3 concent r at i on of 7×10 11 m o lecu le/cm 3

.
c For 12-h average NO 3 concent r at i on of 5×10 8 m o lecu le/cm 3 (At k i n son, 1991).

d For 12-h average HO 2 concent r at i on of 10 8 m o lecu le/cm 3

.
e For solar zenith angle of 0 E

. f Li fet i m es from Arey (1998), for 12-h concent r at i on of OH radi cal of 1.9×10 6 m o lecu le/cm 3

.

Source: W i ner and B u sby
, 1995, unl ess not ed ot herwi s e.

2-86Table 2-21. Major components of gas-phase diesel engine emissions, their know n atmospheric transformation products, and the biological impact of the reactants and products Gas-phase emission component Atmospheric reaction products Biological impact

C a rbon di oxi de

— M a jor cont ri but or t o gl obal

warm i ng. C a rbon m onoxi de

— Hi ghl y t oxi c t o hum ans; bl ocks
oxy gen upt ake. Oxi d es of ni t r ogen Ni t r i c aci d, ozone Ni t r ogen di oxi de i s a respi r at ory t r act i rri t a nt and m a jor ozone
precursor. Ni t r i c aci d cont ri but es t o acid rain. Su lfu r d i o x i d e Su lfu r ic acid R esp i rato ry tract irritatio n

. C ont ri but or t o aci d rai n

. Hydrocarbons: Al kanes ( # C 18 ) A ldehydes, alkyl nitrates, ket ones Resp irato r y tract irritatio n

. Reactio n product s are ozone precursors (i n t h e presence of NO x ). Al kenes ( # C 4 ) (e.g., 1,3-but adi e ne) Al dehy des, ket ones R espi rat o ry t r act i rri t a t i on. Som e alkenes are m u tagenic and carcinogenic. Reaction products are ozone precursors (i n t h e presence of NO x ). Aldehydes: Form al dehy de C a rbon m onoxi de, hy droperoxy l radi cal s Form al dehy de i s a probabl e hum an carci nogen and an ozone precursor (in the presence of NO x ). Hi gher al dehy des (e.g., acetaldehyde, acrolein) Peroxyacyl nitrates Respirator y tract and eye irritation; causes pl ant dam a ge. M onocy c l i c arom at i c

com pounds (e.g., benzene,
to lu en e) Hy droxy l a t e d and hy droxy l a t e d-ni t r o deri vat i v es a B e nzene i s t oxi c and carci nogeni c i n hum ans. Som e react i on product s are

m

u tagenic in bacteria (Am e s assay). PAHs ( # 4 rings) (e.g., phenant hrene, fl uorant h ene) b Ni t r o-PAHs (4 ri ngs) c Som e of t h ese PAHs and ni t r oPAHs are known m u t a gens and carci nogens. Ni t r o-PAHs (2 and 3 ri ngs) (e.g., ni t r onapht hal e nes) Qui nones and hy droxy l a t e dni t r o deri vat i v es Som e react i on product s are

m

u tagenic in bacteria (Am e s assay).

a Som e reaction products expected to partition into the particle phase.

b Nitro

- PAHs with m o re th an two rin g s will p a rtitio n in to th e p a rticle p h a se.

c PAHs cont ai ni ng four ri ngs are usual l y pres ent i n bot h t h e vapor and part i c l e phases.

Source: Health Effects Institute, 1995.

2-872.3.1.1. Organic Compounds The organic fraction of diesel is a com p lex m i xture of com pounds, very few of which
have been characterized. The atm o spheric chem istry of several organic constituents of DE (which are also produced by other com bustion sources) has been studied. A few of these reactions and their products are discussed below. For a com p lete sum m a ry of the atm o spheric
chem istry of organic com bustion products, see Seinfeld and Pandis (1998). Acetaldehyde form s peroxyacetyl nitrate (via form ation of peroxyl radicals and reaction
with NO 2 ), which has been shown to be a direct-acting m u tagen toward S. typhimurium strain TA100 (Kleindienst et al., 1985) and is phytot oxic. Benzaldehyde, the sim p lest arom atic aldehyde, form s peroxybenzoyl nitrate or nitrophe nols following reaction with oxides of nitrogen (Table 2-21). For those PAHs present in the gas phase, reaction with the OH radical is the m a jor

rem oval route, leading to atm o spheric lifetim es of a few hours in daylight. The gas-phase reaction of PAHs containing a cyclopenta-fused ring such as acenaphthene, acenaphthylene, and acephenanthrylene with the nitrate radical m a y be an im portant loss process during nighttim e hours. Relatively few data are available concerni ng the products of these gas-phase reactions. It has been shown that in the presence of NO

x , the OH radical reactions with naphthalene, 1- and 2-

m e thylnaphthalene, acenaphthylene, biphenyl, fluor anthene, pyrene, and acephenanthrylene lead
to the form ation of nitroarenes (Arey et al., 1986; Atkinson, 1986; Atkinson et al., 1990; Zielinska et al., 1988, 1989a; Arey, 1998). In a ddition, in a two-step process involving OH
radical reaction and NO 2 addition, 2-nitrofluoranthene and 2-nitropyrene can be form ed and eventually partition to the particle phase, as will other nitro-PAHs.
The addition of the NO 3 radical to the PAH arom atic ring leads to nitroarene form ation (Sweetm a n et al., 1986; Atkinson et al., 1987, 1990; Zielinska et al., 1989a). The gas-phase
reactions of NO 3 radical with naphthalene, 1- and 2-m e thylnaphthalene, acenaphthene, phenanthrene, anthracene, fluoranthene, and pyrene produce, in general, the sam e nitro-PAH
isom ers as the OH radical reaction, but with diffe rent yields (Arey et al., 1989; Sweetm a n et al., 1986; Atkinson et al., 1987, 1990; Zielinska et al., 1986, 1989a). For exam ple, the sam e 2- nitrofluoranthene is produced from both OH radical and NO 3 gas-phase reactions, but the
reaction with NO 3 produces a m u ch higher yield. The production of several nitroarene

com pounds has been studied in environm ental cham bers (Arey et al., 1989; Zielinska et al., 1990; Atkinson and Arey, 1994; Arey, 1998; Feilberg et al., 1999), and generally the sam e nitroPAH isom ers form ed from reaction with OH and NO 3 radicals are observed in am bient air

sam p les. Secondary form ation of nitroarenes through the gas-phase reactions of the 2-, 3-, and 4-ring PAHs is the m a jor source f o r m a ny of the nitroarenes observed in am bient air (Pitts et al., 1985a-c; Arey et al., 1986; Zielinska et al., 1988). Photolysis is the m a jor rem oval pathway for

2-88nitroarenes with lifetim es of approxim a tely 2 hours (Feilberg et al., 1999; Nielsen and Ram d ahl, 1986).

2.3.1.2. Inorganic Compounds
SO 2 and oxides of nitrogen (prim a rily NO) are em itted f r om diesel engines. SO 2 is readily oxidized by the OH radical in the atm o sphere, followed by form ation of the HO 2 radical
and HSO 3 , which rapidly reacts with water to form H 2 SO 4 aerosols. Because SO 2 is soluble in water, it is scavenged by fog, cloud water, and raindrops. In aqueous system s, SO 2 is readily oxidized to sulfate by reaction with hydrogen peroxide (H 2 O 2 ), O 3 , or O 2 in the presence of a

m e tal catalyst (Calvert and Stockwell, 1983). Sulfur em itted from diesel engines is predom inantly (~98%) in the form of SO 2 , a portion of which will f o rm sulf ate aerosols by the
reaction described above. Nonroad equipm ent, which typically uses fuel containing 3,300 ppm sulfate, em its m o re SO 2 than on-road diesel engines, which use fuels currently containing an average of 340 ppm sulfur because of EPA regulations effective in 1993 decreasing diesel fuel
sulfur levels. EPA estim ates that m obile sources are responsible for about 7% of nationwide SO 2

em issions, with diesel engines contributing 74% of the m obile source total (the m a jority of the
diesel SO 2 em issions originate from nonroad engines) (U.S. EPA, 1998b).
NO is also oxidized in the atm o sphere to form NO 2 and particulate nitrate. The fraction
of m o tor vehicle NO

x exhaust converted to particulate nitrate in a 24-hour period has been calculated using a box m odel to be approxim a tely 3.5% nationwide, a portion of which can be attributed to DE (Gray and Kuklin, 1996). EPA estim ates that in 1997, m obile sources were responsible for about 50% of nationwide NO

x em issions, with diesel engines being responsible for approxim a tely one-half of the m obile source total (U.S. EPA, 1998b).

2.3.1.3. Atmospheric Transport of Gas-Phase DE Gas-phase DE can be dry deposited, depending on the deposition surface, atm o spheric stability, and the solubility and other chem ical properties of the com pound. Dry deposition of organic species is typically on the order of weeks to m onths, with dry deposition velocities of approxim a tely 10 -4 cm /sec (W iner and Busby, 1995). In contrast, inorganic species such as SO 2 and nitric acid have relatively fast deposition rates (0.1–2.5 cm /sec) and will rem a in in the

atm o sphere for shorter tim e periods com p ared with the organic exhaust com ponents. Som e gasphase species will also be scavenged by aqueous aerosols and potentially deposited via precipitation. These processes can greatly reduce the atm o spheric concentration of som e vaporphase species. Atm o spheric lifetim es for several gas-phase com ponents of DE are on the order of hours or days, during which tim e atm o spheric turbulence and advection can disperse these pollutants widely.

2-892.3.2. Particle-Phase Diesel Exhaust
Particle-associated DE is com posed of prim arily carbonaceous m a terial (organic and EC) with a very sm all fraction com posed of inorganic com pounds and m e tals. The OC fraction
adsorbed on DPM is com posed of high-m o lecular-weight com pounds, such as PAHs, which are generally m o re resistant to atm o spheric reactions than PAHs in the gas phase. The EC

com ponent of DE is inert to atm o spheric degradation, whereas the PAH com pounds are degraded by reaction with the following species:

• S unlight, during daytim e hours

• O 3 , during daytim e and nighttim e

• N O 3 and N 2 O 5 , during nighttim e hours

• OH and HO 2

• N O 2 , during nighttim e and daytim e hours

• H 2 O 2

• HNO 3 and other species such as HONO and H 2 SO 4 .

Because m a ny of the PAH derivatives form ed by reaction with som e of the reactants listed above have been found to be highly m u ta genic, a brief discussion of PAH photolysis, nitration, and oxidation follows. Som e of the m a jor degradation products from particulate DE and their biological im pact are listed in Table 2-22.

2.3.2.1. Particle-Associated PAH Photooxidation Laboratory studies of photolysis of PAHs adsorbed on 18 different fly ashes, carbon black, silica gel, and alum ina (Behym er and Hites, 1985, 1988) and several coal stack ashes (Yokely et al., 1986; Dunstan et al., 1989) have shown that the extent of photodegradation of PAHs depends very m u ch on the nature of the substrate to which they are adsorbed. The
dom inant factor in the stabilization of PAHs adsorbed on fly ash was the color of the fly ash,
which is related to the am ount of carbon black present. It appears that PAHs were stabilized if the carbon black content of the fly ash was greater than approxim a tely 5%. On black substrates, half-lives of PAHs studied were on the order of several days (Behym er and Hites, 1988). The
environm ental cham ber studies of Kam e ns et al. (1988) on the daytim e decay of PAHs present
on residential wood sm oke particles and on gasoline internal com bustion em ission particles showed PAH half-lives of approxim a tely 1 hour at m oderate hum idities and

tem p eratures. At very low angle sunlight, ve ry low water vapor concentration, or very low

tem p eratures, PAH daytim e half-lives increased to a period of days. The presence and

2-90Table 2-22. Major components of particle-phase diesel engine emissions, their know n atmospheric transformation products, and the biological impact of the reactants and products
Particle-phase emission component Atmospheric reaction products B i ol ogi cal i m pact

El em ent a l carbon — N ucl e i adsorb organi c com pounds; size p e rm its tran sp o r t d eep in to th e l ungs (al v eol i )
Inorganic sulfate and

n itrate — R esp i rato ry tract irritatio n

Hy drocarbons (C 14 -C 35 ) L ittle in fo rm atio n ; possi bl y al dehy des, ket ones, and alkyl nitrates Unknown

PAHs ( $ 4 rings) (e.g.,
py rene, benzo[a] py rene) Nitro-PAHs ( $ 4 ri ngs) a Ni t r o-PAH l act ones Larger PAHs are m a jor cont ri but ors of carci nogens i n com bust i o n

em i ssi ons. M a ny ni t r o-PAHs are pot ent m u t a gens and carci nogens. Nitro-PAHs ( $ 3 ri ngs) (e.g., ni t r opy renes) Hy droxy l a t e d-ni t r o deri vat i v es M a ny ni t r o-PAHs are pot ent

m

u t a gens and carci nogens. Som e react i on product s are m u t a geni c i n bacteria (Am e s assay).

a Nitro

- PAHs with m o re th an two rin g s will p a rtitio n in to th e p a rticle p h a se. Source: Health Effects Institute, 1995.

com position of an organic layer on the aerosol seem s to influence the rate of PAH photolysis (Jang and McDow, 1995; McDow et al., 1994; Odum et al., 1994).
Because of lim ited understanding of the m echanism s of these com p lex heterogeneous reactions, it is currently im possible to draw any firm conclusion concerning the photostability of particle-bound PAHs in the atm o sphere. Because DPM contains a relatively high quantity of EC, it is reasonable to speculate that PAHs adsorbed onto these particles m i ght be relatively
stable under standard atm o spheric conditions, leading to an anticipated half-life of 1 or m o re days.

2.3.2.2. Particle-Associated PAH Nitration Since 1978, when Pitts et al. (1978) first dem onstrated that B[ a ]P deposited on glassfiber filters exposed to air containing 0.25 ppm NO 2 with traces of HNO 3 form ed nitro-B[ a ]P,

num erous studies of the heterogeneous nitrati on reactions of PAHs adsorbed on a variety of
substrates in dif f e rent sim u lated atm o spheres have been carried out (Finlayson-Pitts and Pitts, 1986). PAHs deposited on glass-fiber and Teflon- im pregnated glass-fiber filters react with gaseous N 2 O 5 , yielding their nitro derivatives (Pitts et al., 1985b,c). The m o st abundant isom ers form ed were 1-NP from pyrene, 6-nitro-B[ a ]P from B[ a ]P, and 3-nitroperylene from perylene.

2-91The form ation of nitro-PAHs during sam p ling m a y be an im portant consideration for DPM collection because of the presence of NO 2 and HNO 3 (Feilberg et al., 1999). However, Schuetzle (1983) concluded that the artifact form a tion of 1-NP was less than 10% to 20% of the 1-NP present in the diesel particles if the sam p ling tim e was less than 23 m i n (one FTP cycle)
and if the sam p ling tem p erature was not higher than 43 o C. The form ation of nitroarenes during

am bient high-volum e sam p ling conditions has been reported to be m i nim a l, at least for the m o st abundant nitropyrene and nitrofluoranthene isom ers (Arey et al., 1988). DPM contains a variety of nitroarenes, with 1-NP being the m o st abundant am ong identified nitro-PAHs. The concentration of 1-NP was m easured in the extract of particulate

sam p les collected at the Allegheny Mountain Tunnel on the Pennsylvania Turnpike as 2.1 ppm and ~5 ppm by m a ss of the extractable m a terial from diesel and SI vehicle PM, respectively.
These values are m u ch lower than would be predicted on the basis of laboratory m easurem ents for either diesel or SI engines (Gorse et al., 1983). Several nitroarene m easurem ents have been conducted in airsheds heavily affected by m o to r vehicle em issions (Arey et al., 1987; Atkinson et al., 1988; Zielinska et al., 1989a,b; Ciccio li et al., 1989, 1993). Am bient PM sam p les were collected at three sites in the Los Angeles Basin during two sum m e rtim e periods and one
wintertim e period. Concentrations of 1-NP ranged from 3 pg/m 3 , to 60 pg/m 3 , and 3- nitrofluoranthene was also present in DPM at c oncentrations ranging from not detectable to 70 pg/m 3

.

2.3.2.3. Particle-Associated PAH Ozonolysis
Num e rous laboratory studies have shown th at PAHs deposited on com bustion-generated fine particles and on m odel substrates undergo reaction with O 3 (Katz et al., 1979; Pitts et al., 1980, 1986; Van Vaeck and Van Cauwenberghe, 1984; Finlayson-Pitts and Pitts, 1986). The
dark reaction toward O 3 of several PAHs deposited on m odel substrates has been shown to be relatively fast under sim u lated atm o spheric cond itions (Katz et al., 1979; Pitts et al., 1980, 1986). Half-lives on the order of 1 to several hours were reported for the m o re reactive PAHs, such as B[ a ]P, anthracene, and benz[a]anthracene (Katz et al., 1979). The reaction of PAHs deposited on diesel particles with 1.5 ppm O 3 under high-volum e

sam p ling conditions has been shown to be relatively fast, and half-lives on the order of 0.5 to 1 hour have been reported for m o st PAHs studi ed (Van Vaeck and Van Cauwenberghe, 1984).
The m o st reactive PAHs include B[ a ]P, perylene, benz[a]anthracene, cyclopenta[cd]pyrene, and benzo[ghi]perylene. The benzofluoranthene isom ers are the least reactive of the PAHs studied, and benzo[e]perylene is less reactive than its isom er B[ a ]P. The im plications of this study for the high-volum e sam p ling am bient POM are im portant: reaction of PAHs with O 3 could possibly occur under high-volum e sam p ling conditions during severe photochem i cal sm og

2-92episodes, when the am bient level of O 3 is high. However, the m a gnitude of this artifact is
difficult to assess from available data.

2.3.2.4. Atmospheric Transport of DE Particulate Matter Ultrafine particles em itted by diesel engines undergo nucleation, coagulation, and
condensation to form fine particles. DPM can be rem oved from the atm o sphere by dry and wet deposition. Particles of sm all diam eter (<1 μm ), such as DPM, are rem oved less efficiently than larger particles by wet and dry deposition and thus have longer atm o spheric residence tim es. Dry deposition rates vary depending on the particle size. Because of their sm all size, DE particles have residence tim es of several days (dry deposition velocities of approxim a tely 0.01

cm /sec) (W iner and Busby, 1995). Diesel particulates m a y be rem oved by wet deposition if they serve as condensation nuclei for water vapor deposition or are scavenged by precipitation in- or
below-cloud. In a study designed to assess the atm o spheric concentrations and transport of DE particles, Horvath et al. (1988) doped the sole source of diesel fuel in Vienna with an

organom etallic com pound of the heavy earth elem ent dysprosium

. The authors found that in

som e of the m o re rem o te sam p ling areas, DPM com posed m o re than 30% of the particulate

m a ss, indicating that DPM can be dispersed widely.

2.3.3. Diesel Exhaust Aging
Prim ary DE is considered “fresh,” whereas “aged” DE is considered to have undergone
chem ical and physical transform a tion and disper sion over a period of a day or two. Laboratory
dilution tunnel m easurem ents represent a hom ogeneous environm ent com p ared to the com p lex and dynam i c system into which real-world DE is em itted. The physical and chem ical
transf orm a tion of DE will vary depending on the environm ent into which it is em itted. In an
urban or industrial environm ent, DE m a y enter an atm o sphere with high concentrations of oxidizing and nitrating radicals, as well as nondi esel organic and inorganic com pounds that m a y inf l uence the toxicity, chem ical stability, and atm o spheric residence tim e. In general, secondary pollutants form ed in an aged aerosol m a ss are m o re oxidized, and therefore have increased polarity and water solubility (Finlayson-Pitts and Pitts, 1986). Kam e ns et al. (1988) reported that photooxidation of particle-bound PAH is enhanced as relative
hum idity is increased. W e ingartner et al. (1997a) and Dua et al. (1999) have reported that unlike

m a ny other types of particles, diesel particle s do not appear to undergo hygroscopic growth once

em itted to the atm o sphere and m a y even shrink in size to som e extent under increasing relative
hum idity conditions. W e ingartner et al. (1997a) evaluated the hygroscopic growth of diesel particles and found that freshly em itted diesel particles dem onstrated m i nim a l hygroscopic

2-93growth (2.5%), whereas aged particles subjected to UV radiation and ozonolysis exhibited

som e what greater but still m i nim a l hygroscopic growth. An increase in the sulfur content of
diesel fuel has also been observed to result in som e what greater water condensation onto diesel particles. To the extent that DE com ponents are oxidized or nitrated in the atm o sphere, they m a y be rem oved at rates different from their precursor com pounds and m a y exhibit different biological reactivities. Data suggesting that m i nim a l hygroscopic growth of DPM occurs also
has im plications for the dosim etry of these particles in the lung because the sm aller particles will reach the lowest airways of the lung, whereas grow th of the particle would result in deposition in the upper airways. The dosim etry of DPM is discussed in Chapter 3. In a recent experim e nt, the biological activity of DPM exposed to 0.1 ppm ozone for 48
hours was com p ared with that of DPM not exposed to ozone (Ghio et al., 2000). Instillations of the ozonated DPM in rat lung resulted in an increase in biological activity (neutrophil influx, increased protein, and lactate dehydrogenase activity) com p ared with DPM that had not been
treated with ozone. These data suggest that am bient levels of ozone can alter DPM constituents causing an increase in toxicity com p ared with nonozonated DPM. In addition to changes in particle com position with aging, particle size distributions m a y vary depending on aggregation and coagulation phenom ena in the aging process. People in vehicles, near roadways (e.g., cyclists, pedestrians, people in nearby buildings), and on

m o torcycles will be exposed to m o re fresh exhaust than the general population. In som e settings
where em issions are entrained for long peri ods through m e teorological or other factors, exposures would be expected to include both fresh and aged DE. The com p lexities of transport and dispersion of em ission arising from m o tor ve hicles have been the subject of extensive

m odeling and experim e ntal studies over the past decades and have been sum m a rized by
Sam p son (1988); exposures to DPM are discussed in the next section of this chapter.
The m a jor organic constituents of DE and their potential degradation pathways described above provide evidence for (1) direct em ission of PAHs, (2) secondary form ation of nitroarenes, and (3) secondary sulfate and nitrate form ation. Because nitro-PAH products are often m o re

m

u tagenic than their precursors, the form ation, transport, and concentrations of these com pounds in an aged aerosol m a ss are of significant interest.

2.4. AMBIENT DIESEL EXHAUST CONCENTRATIONS AND EXPOSURES
2.4.1. Diesel Exhaust Gases in the Ambient Atmosphere Although em issions of several DE com ponents have been m easured, few studies have

attem p ted to elucidate the contribution of diesel -powered engines to atm o spheric concentrations
of these com ponents. The em ission profile of gaseous organic com pounds is different for diesel and SI vehicles; the low-m o lecular-weight arom atic HCs and alkanes (<C 9 ) are m o re

2-94characteristic of SI engine em issions, whereas the heavier alkanes (>C 10 ) and arom atic HCs (such as naphthalene, m e thyl- and dim e thyl- na phthalenes, m e thyl- and dim e thyl-indans) are

m o re characteristic of diesel engine em issions. These differences were the basis for apportionm ent of gasoline- and diesel-powered vehicle em issions to am bient nonm ethane hydrocarbon (NMHC) concentrations in the Boston and Los Angeles (South Coast Air Basin)
urban areas.
The chem ical m a ss balance receptor m odel (described below) was applied to am bient

sam p les collected in these areas, along with appropriate fuel, stationary, and area source profiles (Fujita et al., 1997). The average of the sum of NMHC attributed to DE, gasoline-vehicle exhaust, liquid gasoline, and gasoline vapor was 73% and 76% for Boston and the South Coast Air Basin (SoCAB), respectively. The average source contributions of DE to NMHC
concentrations were 22% and 13% for Boston a nd the SoCAB, respectively. Diesel vehicles

em it lower levels of NMHC in the exhaust com p ared with gasoline vehicles. The relative contribution of DE clearly depends on severa l factors, including fleet com position, sam p ling location (e.g., near a bus station vs. near a highw ay or other sources), and the contribution from point and area sources. The contribution of DE to am bient NMHC showed large variations

am ong sam p ling sites in the Boston area. The source apportionm ent in the Fujita et al. (1997) study indicates that m obile vehicle-related em issions account for the m a jority of am bient NMHC in the two urban areas studied, and the results can likely be extrapolated to other urban areas
with sim ilar source com positions. Other source apportionm ent m e thods such as those used by Henry et al. (1994) have been applied to speciated HC data to separate the m obile source direct

em ission from gasoline evaporative em issions. This m e thod uses a com b ination of graphical analysis (Graphical Ratio Analysis f o r Com position Estim ates, GRACE) and m u ltivariate receptor m odeling m e thods (Source Apportionm ent by Factors with Explicit Restrictions, SAFER) and was not used to identify the diesel engine contribution to the HCs m easured.

2.4.2. Ambient Concentrations of DPM
Because DPM is chem ically com p lex, an assessm ent of am bient DPM concentrations relies prim arily on (1) studies that collect am bient sam p les and adequately characterize their
chem ical com position, or (2) m odeling studies that attem p t to recreate em issions and

atm o spheric conditions. Am bient concentrations of DPM also have been reported from studies using surrogate species. The results of these studies are sum m a rized below. Studies conducted in Europe and Japan were reviewed, but for the m o st part were not included because of questions surrounding the applicability of m easurem ents in locations that use different diesel technology
and control m easures from those in the United States.

2-952.4.2.1. Source Apportionment Studies Receptor m odels are used to infer the types and relative contributions of sources to pollutant m easurem ents m a de at a receptor site. Receptor m odels assum e that the m a ss is conserved between the source and receptor site and that the m easured m a ss of each pollutant is a

sum of the contributions from each source. Receptor m odels are referred to as “top-down” in
contrast to “bottom

- up” m e thods, which use em ission inventory data, activity patterns, and dispersion m odeling from the source to predict concentrations at a receptor site.
The m o st com m only used receptor m odel for quantifying concentrations of DPM at a receptor site is the chem ical m a ss balance (CMB) m odel. Input to the CMB m odel includes

m easurem ents of PM m a ss and chem istry m a de at the receptor site as well as m easurem ents

m a de of each of the source types suspected to im pact the site. Because of problem s involving
the elem ental sim ilarity between diesel and gasoline em ission profiles and their co-em i ssion in
tim e and space, chem ical m o lecular species that provide m a rkers for separation of these sources have been identified (Lowenthal et al., 1992). Recent advances in chem ical analytical techniques have facilitated the developm ent of sophisticated m o lecular source profiles, including detailed speciation of PM-associated organic com pounds that allow the apportionm ent of PM to gasoline and diesel sources with increased confidence. CMB analysis that uses speciation of organic com pounds in the source profiles is typically referred to as extended species CMB.
Older studies that m a de use of only EC, total OC, trace elem ents, and m a jor ions in the source profiles (conventional CMB) have been published and are sum m a rized here, but they are subject

to m o re uncertainty. It should be noted that because receptor m odeling is based on the application of source profiles to am bient m easurem ents, estim ates of DPM concentration generated by this m e thod include the contribution from on-road and nonroad sources to the extent the source prof iles are sim ilar (which would include m ilitary sources depending on the

sam p ling locations and fleet com position). In addition, this m e thod identifies sources of prim ary

em issions of DPM only, and the contribution of s econdary aerosols is not attributed to sources.
The CMB m odel has been used to assess concentrations of DPM in areas of California, Phoenix, Denver, and Manhattan (Table 2-23). DP M concentrations reported by Schauer et al. (1996) for sam p les collected in California in 1982 ranged from 4.4 μg/m 3 in west Los Angeles to 11.6 μg/m 3 in downtown Los Angeles. The average contribution of DPM to total PM 2.5 ma s s ranged from 13% in Rubidoux to 36% in downtown Los Angeles. As m e ntioned above, this

m odel accounts for prim ary em issions of DPM only; the contribution of secondary aerosol
form ation (both acid and organic aerosols) is not included. In sites downwind from urban areas, such as Rubidoux in this study, secondary nitrate form ation can account for a substantial fraction
of the m a ss (25% of the fine m a ss m easured in Rubidoux was attributed to secondary nitrate), a portion of which com e s from DE (Gray and Kuklin, 1996).

2-96Table 2-23.Ambient DPM concentrations reported from chemical mass balance modelingReferenceLocationYear of samplingLocation typeDiesel PM2.5Fg/m3mean,(range)AverageDPM%oftotal PM(range)Source profileusedSchauer etal., 1996WestLA, CA1982, annualaverage (~60samples at each site)Urban4.418EC, OCS,elementsPasadena, CAUrban5.319Rubidoux, CAUrban5.413Los Angeles, CAUrban11.636Chow etal., 1991WestPhoenix, AZ1989-90, winter11 days at each siteUrban13 (max. 22)18EC, OCT, MI,elementsCentralPhoenix, AZUrban13 (max. 16)20South Scottsdale, AZUrban10 (max. 12)17EstrellaPark, AZNonurban59GunneryPark, AZNonurban310Pinnacle Peak, AZNonurban212bCalifornia EPA, 1998aCalifornia, 6 air basins1988-92, annualUrbanc1.8–3.6abEC, OCT, MI,elementsCalifornia, 9 air basinsNonurbanc0.2–2.6aEC, OCT, MI,elementsWittorff et al., 1994Manhattan, NY1993,spring 3 daysUrban29.2(13.2–46.7)a53 (31–68)EC, OCT, MI,elementsMaricopa Association ofGovernments, 1999Phoenix, AZ1994–95, winter 12 daysUrban2.4 (0–5.3)15 (0–27)EC, OCS, MI,elementsFujitaetal., 1998Welby,CO1996–97, winter 60 daysUrban1.7 (0–7.3)10 (0–26)EC, OCS, MI,elementsBrighton, COSuburban1.2 (0–3.4)10 (0–38)

2-97aPM10.bNot available.cUrban air basins are qualitatively defined asthose areas that are moderately or largely urbanized, and nonurban air basins arethose areas that are largelynonurban, butmayhave one or more denselypopulated areas.Abbreviations:EC:Elementalcarbon;OCT:OCtotal;OCS:OCspecies;MI:Major ions including nitrate, sulfate, chloride and,in somecases, ammonium,sodium, potassium.The California Environm ental Protection Agency (Cal EPA) reported am bient DPM
concentrations for 15 air basins in California based on am bient m easurem ents taken statewide
from 1988 to 1992 (Cal EPA, 1998a). Cal EPA used CMB analysis of am bient m easurem ents from the San Joaquin Valley (1988-89), South Coast (1986), and San Jose (winters for 1991–92 and 1992–93) to determ ine m obile source contributions and then applied the California 1990
PM 10 em issions inventory to determ ine the fraction of m obile source PM 10 attributable to diesel

em issions. The results of this analysis indicate that annual average basin-wide levels of direct
DPM m a y be as low as 0.2 μg/m 3 and m a y range up to 2.6 μg/m 3 for basins that are largely
nonurban but m a y have one or m o re densely populated areas (such as Palm Springs in the Salton Sea basin). DPM concentrations for air basins that are m oderately or largely urbanized ranged
from 1.8 μg/m 3 to 3.6 μg/m 3

. Two studies using CMB analysis that report DPM concentrations have been conducted in
the Phoenix area. A wintertim e study in 1989–90 reported DPM concentrations for nonurban areas ranging from 2 μ g/m 3 to 5 μ g/m 3 and DPM concentrations for central and south Phoenix urban areas ranging from 10 μ g/m 3 to 13 μg/m 3 (Chow et al., 1991). Chow et al. (1991) reported that DPM levels on single days can range up to 22 μg/m 3 at the central Phoenix site. A m o re recent study conducted from Novem b er 1994 through March 1995 reported DPM concentrations for Phoenix averaging 2.4 μg/m 3 and reaching 5.3 μg/m 3 (Maricopa Association of Governm e nts, 1999). The extended species CMB was used for this study, providing a m o re confident identification of DPM separate from gasoline PM em issions than the earlier Phoenix study. DPM accounted for an average 15% of am bient PM 2.5 , and gasoline PM accounted for an average of 52% of am bient PM 2.5 in the 1994–95 Phoenix study. In a recently published study designed to investigate the ability of a new type of factor analysis, positive m a trix factorization, to separate sources contributing to the urban aerosol in Phoenix, Ram a dan et al. (2000) report their success in separating the DE PM from other m o tor vehicle PM. Fine PM sam p les were collected by two different types of sam p lers in Phoenix, one
set collected from March 1995 through June 1998 and a second set from June 1996 through June
1998. Elem ental and OC were analyzed using TOT. Particles of DE origin were identified by their high EC content in addition to specific trace elem ents, including m a nganese, sulfur, and
iron. DPM concentrations exceeding 5 μg/m 3 were reported for winter m onths during the study period. The investigators concluded that m o tor vehicles, vegetative burning, and HD DE were
the three m a jor sources contributing to am bient fine PM in Phoenix, with higher contributions in
the winter than in sum m e r. During the winter of 1997, a study assessed DPM concentrations at two urban sites in the Denver area (Fujita et al., 1998). The Northern Front Range Air Quality Study (NFRAQS), initiated to assess the sources of the “brown cloud” observed along Colorado’s Front Range,

2-98conducted air quality sam p ling during the wint er of 1996, sum m e r of 1996, and winter of 1997. For a 60-day period from Decem ber 1996 through January 1997, am bient sam p les collected at two urban Denver sites were analyzed for OC species for use in the extended-species CMB. The average DPM concentrations reported for the urban site at W e lby, CO, and the suburban site at Brighton, CO, were 1.7 μg/m 3 and 1.2 μg/m 3 , respectively. During the study period, DPM concentrations exceeded 5 μg/m 3 on two occasions in W e lby, with reported DPM concentrations of 5.7 μg/m 3 and 7.3 μg/m 3

. DPM accounted for an average of 10% of am bient PM 2.5 , and gasoline PM accounted for an average of 27% of am bient PM 2.5 . One of the m a jor claim s from the NFRAQS was a substantial contribution of EC from gasoline-powered vehicles, m a inly f r om cold-start and high-em itting vehicles. At the W e lby site, the contribution of diesel and gasoline em issions to EC m easurem ents was 52% and 42%, respectively. At the Brighton site, the contribution of diesel and gasoline em issions to EC

m easurem ents was 71% and 26%, respectively. The f i ndings f r om the NFRAQS are com p elling and suggest the need for further investigations to quantify the contribution from cold-start and high-em itting vehicle em issions f o r both gasoline a nd diesel vehicles. Geographical, tem poral, and other site-specif i c param e ters that inf l uence PM concentrations, such as altitude, m u st be considered when extrapolating the NFRAQS findings to other locations. In addition to the need for urban and rural average DPM concentrations, an assessm ent of potential health effects resulting from DPM exposure includes an assessm ent of people in
environm ents with potentially elevated levels of DPM. Lim ited data are available to allow a characterization of DPM concentrations in “hotspots” such as near heavily traveled roadways, bus stations, train stations, and m a rinas. Only one CMB study has attem p ted to apportion PM

m easured in an urban hotspot. W ittorff et al. (1994) reported results of conventional CMB perform e d on PM sam p les collected in the spring of 1993 over a 3-day period at a site adjacent

to a m a jor bus stop on Madison Avenue in m i dtown Manhattan. Buses in this area idle for as long as 10 m i nutes, and PM em issions are augm ented by the elevated levels of DPM em itted during acceleration away from the bus stop (dis cussed in Section 2.2.5). DPM concentrations reported from this study ranged from 13.0 μg/m 3 to 46.7 μg/m 3

. This study attributed, on average, 53% of the PM 10 to DE. The DPM concentrations resulting from the source apportionm ent m e thod used in this study require som e caution because the CMB m odel overpredicted PM 10 concentrations by an average 30%, which suggests that additional sources of
the m a ss were not accounted for in the m odel. The relevance of the Manhattan bus stop concentrations and potential exposure for large urban populations provide strong m o tivation for further studies in the vicinity of such hotspots.
In sum m a ry, source apportionm ent studies of am bient sam p les collected before 1990 suggest that seasonal and annual average DPM concentrations for nonurban areas ranged from 2

2-99μ g/m 3 to 5 μ g/m 3

. DPM concentrations reported fromCMB studies for urban areas in the pre1990 tim efram e ranged from 4.4 μ g/m 3 to 13 μg/m 3 , with concentrations on individual days ranging up to 22 μ g/m 3

. Source apportionm ent applied to am bient m easurem ents taken in 1990 or later suggest that seasonal or annual average DPM levels in suburban/nonurban locations can range from 0.2 μ g/m 3 to 2.6 μg/m 3 , with m a xim u m reported values ranging up to 3.4 μg/m 3

. DPM concentrations reported from CMB studies in urban areas during 1990 or later range from
1.7 μ g/m 3 to 3.6 μg/m 3 , with m a xim u m concentrations up to 7.3 μg/m 3

. The highest DPM concentrations reported from CMB analysis of am bient m easurem ents were those in the vicinity of a bus stop in m i dtown Manhattan, which ranged from 13.2 μ g/m 3 to 46.7 μg/m 3

.

2.4.2.2. EC Surrogate for DPM
EC is a m a jor com ponent of DE, contributing approxim a tely 50% to 85% of diesel particulate m a ss, depending on engine technology, fu el type, duty cycle, engine lubrication oil

consum ption, and state of engine m a intenan ce (Graboski et al., 1998b; Zaebst et al., 1991; Pierson and Brachaczek, 1983; W a rner-Selph and Dietzm ann, 1984). In urban am bient
environm ents, DE is one of the m a jor contributors to EC, with other potential sources including spark-engine exhaust; com bustion of coal, oil, or wood; charbroiling; cigarette sm oke; and road
dust. Although cold-start em issions from gasoline com bustion vehicles were reported to be an
im portant source of EC in wintertim e sam p les collected in two cities in the Denver area (Fujita et al., 1998), it is currently unclear to what extent these results are transferable to other locations. It is noteworthy that the EC content of the cold-start em issions from gasoline com bustion vehicles
was lower than that from diesel com bustion engi nes in the sam e study by alm o st a factor of 2. Fowler (1985) evaluated several com ponents of DE and concluded that EC is the m o st reliable overall m easure of am bient DE exposure. Because of the large portion of EC in DPM,
and the fact that DE is one of the m a jor contributors to EC in m a ny am bient environm ents, DPM concentrations can be bounded using EC m easurem ents. Surrogate calculations of DPM have
been based on the fraction of am bient EC m easured in a sam p le that is attributable to diesel engine exhaust and the fraction of the diesel particle m a ss accounted for by EC. In the recent Multiple Air Toxics Exposure Study in the South Coast Air Basin (MATES-II, SCAQMD, 2000), EC m easurem ents were used to estim ate DPM concentrations by the following relationship: approxim a tely 67% of fine EC in the am bient air in the Los Angeles area originates
from diesel engine exhaust (Gray, 1986), and th e average EC fraction of diesel particles

m easured was 64%. Therefore, in the MATES-II study, the South Coast Air Quality Managem e nt District calculated DPM concentrations f r om EC m easurem ents by m u ltiplying a

m easured EC concentration by 67% and dividing by the fraction of DPM m a ss accounted for by EC of 64%, for exam ple, DPM concentration = (EC * 0.67)/0.64, or DPM = EC * 1.04 (not

2-100appreciably different from EC . DPM). This calculation, used in the MATES-II study, relies on
data collected in the 1982 tim efram e and m a y not accurately represent the current day contributions of diesel engines to the am bien t EC inventory. Using a 1998 em issions inventory for the South Coast Air Basin, it is now estim ated that a m o re appropriate conversion from EC to
DPM is to m u ltiply EC by 1.24 (MATES-II, SCAQMD, 2000). An alternative calculation can be derived using data from recent studies in Colorado and Arizona (Fujita et al., 1998; Maricopa Associa tion of Governm e nts, 1999). The fraction of EC
attributable to DE can be estim ated from detailed source profiles applied to a CMB m odel as
discussed above. The contribution of diesel e ngines to EC averaged 68% ± 20% for Brighton, CO, and 49% ± 26% at W e lby, CO, as part of the winter 1996-1997 NFRAQS. In Phoenix, diesel engine exhaust was estim ated to account for approxim a tely 46% ± 22% of the am bient
EC. For som e environm ents, such as certain occupational settings in which diesel engines are in proxim ity to workers, all the EC m a y realistically be attributed to DE as a reasonable upper
bound estim ate of DPM concentrations. As discussed in Section 2.2, the EC content of DPM can vary widely depending on engine type, load conditions, and the test cycle. However, typical profiles for HD and LD diesel engines have been determ ined and the typical EC fraction of DPM ranges from approxim a tely
52% to 75%.
Am bient EC attributed to DE in the studies described above ranges from 46% to 68%. A
lower-bound estim ate of DPM from am bient EC m easurem ents in areas with sim ilar source contributions to those in the Phoenix and Denver areas can be derived using the equation:

DPM = (EC * 0.46)/0.75 or DPM = EC * 0.62

An upper-bound estim ate uses the equation:

DPM = (EC * 0.68)/0.52 or DPM = EC * 1.31

Using the average of the ranges provides the equation:

DPM = EC * 0.89.

Clearly the choice of a point estim ate can provide a surrogate calculation of DPM that can vary by at least a factor of two. Although a recom m e nded surrogate DPM calculation

m e thod is not provided here, the surrogate DPM calculation is used to illustrate the usefulness of

2-101this approach for estim ating DPM in the absence of a m o re sophisticated receptor m odeling analysis for locations where fine PM EC concentrations are available.
One source of variability in EC concentrations reported f o r am bient studies is the

m easurem ent m e thod used to quantify EC. As discussed in Section 2.2.8.1, EC and OC are operationally defined. Am bient sam p les are typically analyzed for EC using therm a l optical
ref l ectance or therm a l optical transm ittance. The m easurem ent technique used in the NFRAQS and Phoenix studies was TOR, which, as discussed in Section 2.2.8.2, often results in higher EC
levels com p ared to TOT analyses. Table 2-24 provides a lower- and upper-bound DPM estim ate from annual average EC concentrations for three urban areas, in addition to DPM concentrations reported from EC

m easurem ents for the MATES- II (SCAQMD, 2000). Under an EPA research grant with the Northeastern States for Coordinated Air Use Managem e nt (NESCAUM), PM 2.5 sam p les were collected every 6 days for 1 year (1995) in Boston (Kenm o re Square), MA, and Rochester, NY, and were analyzed for EC using TOT (Salm on et al., 1997). DPM concentrations were estim ated to be in the range from 0.8 μg/m 3 to 1.7 μg/m 3 in Boston, and from 0.4 μg/m 3 to 0.8 μg/m 3 in Rochester (Table 2-24).

Table 2-24. Ambient diesel particulate matter concentrations from elemental carbon measurements in urban locations

Reference Year of

sa mpling Lo ca tio n DPM 2.5 F g/ m 3 lower-upper bound range (point estimate) a DPM % of total

PM Sal m on et al ., 1997 1995, annual B ost on, M A 0.8–1.7 (1.1) 6 -12 R o chest e r, NY 0.4-0.8 (0.5) 3 -6 Si sl er, 1996 1992-1995, annual W a shi ngt on, DC 0.9-2.2 (1.5) 4 -12

MATES II c Diesel PM 2.5 F g/ m 3 avg± st d dev. Sout h C o ast Air Qu ality M a nagem e nt Di st ri ct , 1999 1995-6, annual A nahei m , C A 2.4 ± 1.8 b

B u rbank, C A 3.3 ± 1.9 b Los Angel e s, C A 3.5 ± 1.9 b

Font ant a , C A 3.4 ± 2.3 b

Hunt i ngt on Park, C A 4.5 ± 2.4 b Long Beach, CA 2.5 ± 1.7 b

Pi co R i vera, C A 4.4 ± 2.2 b

R

ubi doux, C A 3.4 ± 2.0 b

a Lower-bound range: DPM = EC *0.62; upper-bound range: DPM = EC *1.31; poi nt est i m at e: DPM = EC *0.89.

b No t av ailab l e.

c The Multiple Air Toxics Exposure Study in the S outh Coast Air Basin reported DPM calculated from EC
concent r at i ons as DPM = EC *1.04. St andard devi at i ons are report e d.

2-102The Interagency Monitoring of Protected Visual Environm ents (IMPROVE) project being conducted by the National Park Service includes an extensive aerosol m onitoring network

m a inly in rural or rem o te areas of the country (national parks, national m onum ents, wilderness areas, national wildlife refuges, and national seashores), and also in W a shington, DC (Sisler, 1996). PM 2.5 sam p les, collected from March 1992 through February 1995 twice weekly for 24- hour duration at 43 sites (som e co-located in the sam e rural park area), were analyzed for a suite
of

chem ical constituents, including EC (using TOR). EC concentrations in these rural locations

m a y have EC source contributions quite different from those in the urban areas in which the
fraction of EC attributable to DE has been reported. The lack of inform ation regarding EC
sources in these rural locations m a kes the application of the EC surrogate highly uncertain. It is noteworthy that annual average EC concentrations in the rural and rem o te regions reported as part of the IMPROVE network range from 0.1 μg/m 3 for Denali National Park, AK, to 0.9 μg/m 3 for the Lake Tahoe, CA, area. In W a shington, DC, the annual average EC concentration of 1.7 μg/m 3 is estim ated as an annual average DPM concentration of 1.4 μg/m 3. The annual average EC m easurem ents in W a shington, DC, suggest that the DPM concentrations are in the range from 1.0 μg/m 3 to 2.2 μg/m 3, accounting for 5% to 12% of

am bient PM 2.5 . Seasonally averaged data for the W a shington, DC, site indicate that EC concentrations and, by extension, DPM concentrations peak in the autum n and winter (2.0 μg/m 3 and 0.9 μg/m 3 EC, respectively). DPM concentrations reported recently as part of the MATES-II study at eight locations ranged from 2.4 μg/m 3 to 4.5 μg/m 3. DPM concentrations at Huntington Park and Pico Rivera, CA, were higher than other DPM concentrations in the South Coast Air Basin, perhaps because of higher diesel truck traffic, proxim ity to nonr oad diesel sources, or nondiesel sources of EC, including gasoline vehicle traffic. In a recent study of the trends in fine particle and EC concentrations in Southern California, Christoforou et al. (2000) report that EC concentrations m e asured in 1993 were 29%-
40% of EC concentrations m easured in 1982 at four urban Los Angeles sites. The authors credit
lower PM em ission rates from on-road diesel engi nes as well as cleaner-burning diesel fuel for
the observed EC decrease. The extent to which nonroad diesel equipm ent im pacts a given site
will influence the trend in am bient EC concentrations because fewer regulations have been

prom ulgated to control the PM em issions from these engines.

2.4.2.3. Dispersion Modeling Results Dispersion m odels estim ate am bient levels of PM at a receptor site on the basis of

em ission factors for the relevant sources and param e ters that sim u late atm o spheric processes such as the advection, m i xing, deposition, and chem ical transform a tion of com pounds as they are

2-103transported from the source to the receptor site(s). Cass and Gray (1995), Gray and Cass (1998),
and Kleem an and Cass (1998) have applied dispersion m odels to the South Coast Air Basin to
estim ate DPM concentrations. The m odels used by these investigators applied em ission factors
from 1982 and consequently are representative of concentrations prior to the im plem entation of
DPM em ission controls. In addition to offering another approach for estim ating am bient DPM concentrations, dispersion m odels can provide the ability to distinguish on-highway from nonroad diesel source contributions and have presented an approach for quantifying the concentrations of secondary aerosols from DE. Cass and Gray (1995) used a Lagrangian particle-in-cell m odel to estim ate the source
contributions to atm o spheric fine carbon particle concentrations in the Los Angeles area, including diesel em ission factors from on-highway and off-highway sources. Their dispersion

m odel indicates that for 1982, the annual average am bient concentrations of DPM ranged from 1.9 μg/m 3 in Azusa, CA, to 5.6 μg/m 3 in downtown Los Angeles (Table 2-25). The contribution of on-highway sources to DPM ranged from 63.3% in downtown Los Angeles to 89% in west Los Angeles. Of the on-highway diesel contribution, the m odel predicted that for southern California, HD trucks m a de up the m a jority (85%) of the DPM inventory, and overall they
contributed 66% of the DPM in the am bient air. Nonroad sources of DE include pum ping stations, construction sites, shipping docks, railroad yards, and heavy equipm ent repair f acilities. Cass and Gray (1995) also report that wintertim e peaks in DPM concentrations can reach 10 μg/m 3

.

Table 2-25. Ambient diesel particulate matt er concentrations from dispersion modeling

Reference Location Year of sampling L ocation type DPM 2.5 g/

m 3 (mean) DPM % of
total PM Cass and Gray,
1995 Azusa, C A 1982, annual N onurban 1 .4 a 5 Lennox, C A 1982, annual N onurban 3 .8 a 13 Anahei m , C A 1982, annual U rban 2.7 a 12 Pasadena, C A 1982, annual U rban 2.0 a 7 Long Beach, CA 1982, annual U rban 3.5 a 13 Downt o wn LA, C A 1982, annual U rban 3.5 a 11 W e st LA, C A 1982, annual U rban 3.8 a 16 Kleem an and Cass,
1998 C l arem ont , C A 18-19 Aug 1987 Nonurban 2 .4 (4.0) a,b 8 (6) b

Kleem an et al.,
2000 Long Beach, CA 24 Sept 1996 Urban 1 .9(2.6) b 8 (7) b Ful l e rt on, C A 24 Sept 1996 Nonurban 2 .4(3.9) b 9 (8) b
R i versi d e, C A 25 Sept 1996 Suburban 4 .4(13.3) b 12 (13) b

a On-road diesel vehicles only; all other values are for on-road plus nonroad diesel em issions.

b Value in parentheses includes secondary DPM (nitrate, am m onium , sulfate and hydr ocarbons) attributable to atm o spheric reaction s of prim ary diesel em issions of NO x , SO 2 and hydrocarbons. For the fraction of am bient PM attributab le to DPM, the value in pa renthesis reports total DPM (prim a ry plus secondary) as a fraction of total am bient PM (prim a ry plus secondary).

2-104Kleem an and Cass (1998) developed a Lagrangian m odel that exam ines the size and
chem ical evolution of aerosols, including gas-to -particle conversion processes during transport.
This m odel was applied to one well-characterized episode in Clarem ont, CA, on August 27-28,
1987. The m odel provided reasonable predictions of PM 10 (overpredicting PM 10 by 13%), EC, and OC, and it adequately reconstructed the size distribution of the aerosols. The m odel indicated that on August 27-28, 1987, the PM 2.5 concentration was 76.7 μg/m 3 , 13.2% (10.1 μg/m 3 ) of which was attributable to diesel engine em issions. This estim ate includes secondary
aerosol form ation for sulfate, am m onium , nitrate, and organic com pounds, which accounted for 4.9 μg/m 3 of the total estim ated DPM m a ss. The secondary organic aerosol was estim ated to be 1.1 μg/m 3 , or 31% of the total secondary aerosol m a ss, with the rem a inder com posed of nitrate,

am m onium , and sulfate aerosols. Dispersion m odeling estim ates of diesel PM concentrations from on-highway and nonroad sources have recently been developed as part of the EPA National Air Toxics
Assessm ent (NATA) National Scale Assessm ent. This assessm ent uses the Assessm ent System for Population Exposure Nationwide (ASPEN) dispersion m odel to estim ate am bient concentrations for the year 1996. The NATA national scale assessm ent reports concentrations of
DPM and 32 additional urban air toxic com pounds at the county, State, and national level (NATA, 2001).
ASPEN m a kes a num ber of sim p lifying assum p tions in order to m odel concentrations on
a nationwide scale. For instance, concentration estim ates at the census tract level were estim ated using m odeling assum p tions to allocate em issions from the county level, and the m odel is very
sensitive to the assum p tions used. In addition, dispersion of em issions from nonpoint sources (e.g., on-highway and nonroad vehicles) was treated sim p listically. For resident tracts that have radii greater than 0.3 km , non-point-source am bient concentrations are estim ated on the basis of five pseudo-point sources. The average concentration for the census tract is determ ined by spatially averaging the am bient concentrations associated with the receptors defined for the five pseudosources that fall within the bounds of the tract. Other lim itations include the following:
terrain im pacts on dispersion were not included; the study relied on long-term clim ate sum m a ry data, and no long-range transport was included for DPM (m edium -range transport for DPM,
within 300 km , was included). Because of the lim itations, the results are m o st m e aningfully interpreted when viewed over large geographic areas. The 1996 results from ASPEN com p are well (generally within a factor of 1.5) with estim ated concentrations from EC m easurem ents and receptor m odeling, as well as data from other dispersion m odeling studies. The com p lete results
of the assessm ent are available at http://www.epa.gov/ttn/uatw/nata.

2-105Table 2-26 presents 25 th percentile, average, and 75 th percentile nationwide
concentrations from the 1996 National-Scale Assessm ent as well as the contribution of on-road
and nonroad DE the sources to the nationwide average. The national average DPM concentration reported in the National-Scale Assessm ent is 2.1 μg/m 3 , of which nonroad sources
are estim ated to contribute 67% and on-road sources contribute the rem a inder. Less than 2% of the nationwide DE inventory is attributed to poi nt sources, and these were not included in the

m odeling as part of National-Scale Assessm ent. A wide range in average State-specific am bient DPM concentrations was reported by the National-Scale Assessm ent with the lowest values f o r

m a inly rural States with few DE sources, such as W yom ing (annual average of 0.2 μg/m 3 ), and the highest values for States with large urban centers such as New York (annual average of 5.4 μg/m 3 ).

Table 2-26. Nationw i de ambient diesel partic ulate matter concentrations for 1996 from the National Air Toxics Assessment Nation al-Scale Assessment dispersion modeling

Lo ca tio n 25th percentile,

DPM 10 mg/m 3 Average, DPM 10 mg/m 3 75th percentile,

DPM 10 mg/m 3 Contribution toaverage from onroad sources, DPM 10 mg/m 3 Contribution to average from nonroad sources, DPM 10 mg/m 3 Nat i onwi d e 0 .9 2.1 2 .5 0.6 1 .4

Al l urban
count i e s 1.2 2 .4 2.7 0 .7 1.7

All rural
count i e s 0.4 0 .7 1.0 0 .3 0.5

Source: NATA, 2001. Data available at http://www.epa.gov/ttn/uatw/nata

.

2.4.3. Exposures to Diesel Exhaust
Ultim ately, it is personal exposure that determ ines health im pacts. In the following sections, m odeled average exposures and som e inform ation reflecting potential exposures for those who spend a large portion of their tim e outdoors are presented. Occupational exposures to
DPM are sum m a rized for the variety of workplaces in which diesel engines are used. These occupational exposures are placed into context with equivalent environm ental exposures to understand the potential for overlap in average occupational and average am bient exposures.
Because DE is a m i xture of particles and gases, one m u st choose a m easure of exposure (i.e.,
dosim eter); μg/m 3 of DPM has historically been used in m a ny studies as the dosim eter for the
entire DE m i xture.

2-1062.4.3.1. Occupational Exposure to DE The National Institute for Occupational Safety and Health (NIOSH, 1988) estim ates that approxim a tely 1.35 m illion workers are occupationally exposed to DE em issions. Such workers

em issions include m i ne workers, railroad workers, bus and truck drivers, truck and bus

m a intenance garage workers, loading dock workers, firefighters, heavy equipm ent operators, and
farm workers.
Measurem ents of DPM exposure in occupational environm ents have included respirable particulate (<3.5 μm ), sm oking-corrected respirable particulate, com bustible respirable particulate, and EC, am ong other m e thods. The m easurem ent m e thod used in each of the studies
discussed below is listed in Table 2-27. Occupational exposures to DPM as well as breathing
zone concentrations of DPM have been described in som e detail by W a tts (1995), Groves and Cain (2000), Ham m ond (1998), the W o rld Health Organization (1996), and Birch and Cary (1996) and are briefly, but not com p rehensively, sum m a rized here. The highest occupational exposures to DPM are for workers in coal m i nes and noncoal

m i nes using diesel-powered equipm ent. These exposures, reported by several investigators, range from approxim a tely 10 μg/m 3 to 1,280 μg/m 3 (Table 2-27). Rogers and W h elan (1999) report exposures to specific DPM-associat ed PAHs (including naphthalene, fluorene, phenanthrene, pyrene, and benz[a]anthracene) fo r m i ne workers using diesel fuels containing low and high levels of sulfur, aliphatic, and arom atic com pounds. Results of this study indicate
that the com position of DPM to which workers were exposed varies considerably based on engine condition, fuel, and other operating para m e ters. Mine worker exposures to PAH

com pounds were highest for naphthalene, rangi ng from 1,312 μg/g to 3,228 μg/g of organics, and exposures were lowest for benz[a]anthracene, ranging from less than 3 μg/g up to 18 μg/g of organics. Other investigators have reported DPM-a ssociated PAH concentrations that do not necessarily represent personal exposures but are a snapshot of short periods of elevated
concentration that m a ke up a portion of a worker’s daily exposure. Bagley et al. (1991, 1992) reported levels of B[ a ]P ranging from below the detection lim it of 0.05 ng/m 3 to 61 ng/m 3 collected only during periods of m i ning activity. W a tts (1995) reported DPM concentrations in
four m i nes collected during significan t diesel activity, ranging from 850 μg/m 3 to 3,260 μg/m 3

. Heino (1978) reports DPM concentrations for locom o tive engineers reaching 2,000 μg/m 3

. In a study of four railroads, W o skie et al. (1988) reported concentrations of respirable dust (corrected for cigarette sm oke particulate) that ranged from 39 μg/m 3 for engineers/firers to 134 μg/m 3 for locom o tive shop workers and 191 μg/m 3 for hostlers. W o skie et al. (1988) also reported sm oking-corrected respirable dust for railroad clerks (17 μg/m 3 ), who are considered to be not exposed to DE. Although these exposures m a y have included nondiesel PM (background

2-107Table 2-27. Occupational exposure to DPM

Author Year of sample L o cati on/ j ob type, typical work schedule of 8 hours n Sample type Range in DPM, μg/m 3 Gangal and Dai n t y , 1993 a NA Noncoal m i ne workers

~ 200 R C D 100–900

Säveri n, 1999 1992 Noncoal m i ne workers 255 b R T C 38–1,280

R ogers and W h el an, 1999 1990-99 C o al m i ne workers >1,300 DPSM M 10–640

Haney
, 1990 a 1980s C o al m i ne workers (fi ve m i nes) NA SJI 180–1,000

Am bs, 1991a a NA Coal m i ne workers (four m i nes) NA PDEAS 750–780

W o ski e et al ., 1988 3-y ears i n m i d-1980s R a i l r oad engi neer/ fri er 128 AR P 39–73 R a i l r oad braker/ c onduct o r 158 AR P 52–191 R a i l r oad shop workers 176 AR P 114–134

Groves and C a i n , 2000 NA Railway repair 64 EC(U) 7 -50

Froi nes et al ., 1987 1985 Fi refi ght ers (t wo st at i ons) 238 TSP 63–748

NIOSH, 1992 a NA Firefighters (t hree st at i ons) 18 EC (T) 6–70

B i rch and C a ry , 1996 NA Fi refi ght ers N A E C ( U) 20–79

NA Fi re st at i on em pl oy ees (four
st at i ons) NA EC (U) 4–52

B i rch and C a ry , 1996 NA Ai rport ground crew NA EC (U) 7–15

NA Publ i c t r ansi t workers N A E C ( U) 15–98

NIOSH, 1990 1990 Di esel forkl i f t dockworkers 24 EC (T) 12–61 Zaebst et

al ., 1991 1990 Dockworkers 7 5 E C ( T) 9–20

M echani c s 8 0 E C ( T) 5–28
Long- and short

- haul t r uckers 128 EC (T) 2–7 Groves and C a i n , 2000 NA B u s garage/ r epai r

Forkl i f t t r

ucks 5 3

27 E C ( U)

EC (U) 7-217

7 -403
Kittelson et al., 2000 1999-2000 Bus drivers 3 9 E C(T) 1–3

Parki ng ram p at t e ndant s 1 2 E C ( T) 2 ± 0.4

a C i t e d i n W a t t s (1995). NA: not avai l a bl e. b Personal exposure and area sam p l e s were not report e d separat e l y for t h i s st udy

. R C D : respi r abl e com bust i b l e dust ; R T C : respi r abl e t o t a l carbon SPM : subm i c rom e t e r PM ; DPSM M : di esel part i c ul at e subm i c ron m a ss (t wo-st a ge i m pact i on sam p l e r used t o separat e PM by si ze); EC (T): el em ent a l carbon anal y zed by TOT; EC (R ) el em ent a l carbon anal y zed by TOR ; EC (U) el em ent a l carbon anal y zed by col ourem et ri c

m e t hod or m e t hod not report e d; SJI: si ngl e-jet i m pact or agreed wi t h i n 10% wi t h si m u l t a neous PDEAS

m easurem ents; PDEAS: personal DE aerosol sam p ler collects DPM <0.8 F m , SPM: p a rticu l ate m a tter; ARP:
resp irab le p a rticu l ate ad j u sted to rem o v e th e in flu e n ce o f cig a rette sm o k e ; TSP: to tal su sp en d e d p a rticu l ate m a tter.

2-108respirable dust levels have been estim ated to have contributed approxim a tely 10 μg/m 3 to 33 μg/m 3 for this study), the m a jority of the respirable PM is believed to have originated from diesel
locom o tive em issions. Groves and Cain (2000) reported EC exposures am ong railway repair workers averaging 21 μg/m 3 with a range from 7-50 μg/m 3

. DPM exposures reported for firefighters operating diesel engi ne vehicles range from 4 μg/m 3 to 748 μg/m 3 , which also

encom p asses the range of DPM exposures reported for airport ground crew and public transportation system personnel (7 μg/m 3 to 98 μg/m 3 ). Studies reporting DE exposure am ong fire sta tion em ployees typically report particulate levels below 100 μg/m 3 (ranging from 4 μg/m 3 to 79 μg/m 3 ) (NIOSH, 1992; Birch and Cary, 1996). In a study by Froines et al. (1987), DPM expos ures for firefighters in two stations ranged
from 39 μg/m 3 to 73 μg/m 3

. Birch and Cary (1996) also reported DPM exposures for airport ground crew and public transit workers, ranging from 7 μg/m 3 to 15 μg/m 3 for airport ground crews and 15 μg/m 3 to 98 μg/m 3 for public transit workers. Dock workers using diesel-powered forklifts have been reported to have DPM exposures ranging from 6 μg/m 3 to 403 μg/m 3 (NIOSH, 1990; Zaebst et al., 1991; Groves and Cain, 2000). In studies by NIOSH (1990) and Fowler (1985), the organic m a terial m easured acc ounted for about one-half to alm o st all of the carbonaceous DPM exposures, providing evidence that som e pieces of nonroad equipm ent (forklifts and construction equipm ent) em itted DPM with a significant OC fraction in the 1980s and early 1990s. Zaebst et al. (1991) also reported DPM exposures for m echanics, road drivers, and local drivers for 8-hour shifts at each of six large hub truck term inals. Residential background and highway background sam p les at fixed sites were also collected during warm -weather and coldweather periods, and the geom etric m ean for DPM concentrations ranged from 1 μg/m 3 to 5 μg/m 3

. DPM exposures for road and local truckers in warm - and cold-weather periods ranged
from 2 μg/m 3 to 7 μg/m 3 , whereas exposure levels for m echanics were reported between 5 μg/m 3 and 28 μg/m 3 (geom e tric m eans). Kittelson et al. (2000) are m easuring DPM exposures for bus drivers, parking garage attendants, and m echanics using TOT to quantify EC. Personal exposures for bus drivers on four different routes range from 1 μg/m 3 to 3 μg/m 3 and exposure am ong parking ram p attendants averaged 2 μg/m 3

. These results are prelim inary, and data for the m echanics have not yet been analyzed. This study will also characterize PAH com pounds to which these workers are exposed. Bus garage workers have also been assessed for exposure to DE using urinary excretion
of 8-oxo-2' -deoxyguanosine (Loft et al., 1999). Other biom arkers of DE exposure in occupational workers have included m easurem en ts of urinary 1-hydroxypyrene, adducts of DNA

2-109and hem oglobin, and 8-hydroxyguanosine in lung tissu e (Nielsen et al., 1996; Tokiwa et al., 1999; Zwirner-Baier and Neum ann, 1999; Hara et al., 1997).
To estim ate an environm ental exposure that is equivalent to an occupational lifetim e exposure, the fraction of lifetim e worker inhalation exposure (calculated as the am ount of air breathed on the job m u ltiplied by the typical am ount of tim e spent on the job) is calculated relative to 70-year lifetim e inhalation exposure: (10 m 3 /shift/20 m 3 /day) * (5 days/7 days) * (48 weeks/52 weeks) * (45-year career/70-year life tim e) = 0.21. Using this calculation, 21% of an annual average occupational lifetim e exposure is roughly equivalent to a 70-year annual average
lifetim e environm ental exposure. The equivalent environm ental exposures for the occupational exposures presented in Table 2-28 range from 0.6 μg/m 3 to 14 μg/m 3 for truckers, dock workers,
and m echanics, and from 2 μg/m 3 to 269 μg/m 3 for m i ners. The low end of the range of
environm ental equivalent exposures for several of the occupational settings overlaps with

average m odeled exposures and with am bient concentrations of DPM in urban areas in the
1990–1996 tim efram e

. The overlap between som e occupational exposures and environm ental exposures, as well as the sm all difference between occupational environm ental equivalent exposures and environm ental exposures, is a significant concern and suggests the potential for significant risk in the general population. The possible m a gnitude of the cancer risk in the general population is discussed in Chapter 8, Section 8.3.

Table 2-28. Ranges of occupational exposure to DPM by job category w i th estimates of equivalent environmental exposures Year of sampl i n g O ccupati ons O ccupati onal DPM, F g/ m 3 Environmental equivalent a

exposure, F g/ m 3

1980s and 1990s M i ners 10–1,280 2–269
1980s R a i l r oad workers 39–191 8–40
1985 and l a t e r F i r efi ght ers 4–748 1–157
NA Ai rport crew, publ i c t r ansi t workers 7–98 2–21
1990 Dockworkers, m echani c s 5–61 1–13
1990 Long- and short

- haul t r uckers 2–7 0.4–2

a Envi ronm ent a l equi val e nt exposure i s cal cu l a t e d as t h e occupat i onal exposure * (10 m 3 / s hi ft /

20 m 3 /d ay)* (5 day s / 7day s ) * (48 weeks / 52 weeks) * (45 y ear career / 70 y ear l i f et i m e), or occupat i onal exposure * 0.21 (di s cussed i n sect i on 2.4.3.1.

2.4.3.2. Ambient Exposure to DE
Modeled estim ates of population exposures to DPM integrate exposure in various indoor
and outdoor environm ents and also account for the dem ographic distribution, tim e-activity

2-110patterns, and DPM concentrations in various environm ents, including job-related exposures.
Two m odeling efforts have been developed to determ ine DPM exposures in the general population: the Hazardous Air Pollutant Exposure Model for Mobile Sources, version 3 (HAPEM-MS3) and the California Population I ndoor Exposure Model (CPIEM). EPA has also developed version 4 of the HAPEM, which provi des State-specific average exposures for DPM
and 32 other urban air toxic com pounds. The draft exposure assessm ent using HAPEM version 4 (HAPEM4) has been conducted as part of th e National Air Toxics Assessm ent National-Scale Analysis described in Section 2.4.2.3 above and results are provided here.

2.4.3.2.1. The Hazardous Air Pollutant Exposure Model

. To estim ate population exposures to DPM, EPA has used HAPEM-MS3 (U.S. EPA, 1999b). This m odel provides national and urban-area-specific exposures to DPM from on-ro ad sources only. HAPEM-MS3 is based on the CO probabilistic National Am bient Air Quality Standards (NAAQS) exposure m odel (pNEM/CO), which is used to estim ate the frequency distribution of population exposure to CO and the resulting carboxyhem a globin levels (Law et al., 1997). HAPEM sim u lates the CO exposure scenario of individuals in 22 dem ographic groups for 37 m i croenvironm ents. CO
concentrations are based on am bient m easurem ents m a de in 1990 and are related to exposures of
individuals in a 10-km radius around the sam p ling site. DPM exposures are calculated as in Equation 2-5, using a ratiom e tric approach to CO.

DPM μ g m 3 = ( C O μ

g m 3 / C O g m i ) × D PM g / m i (2-5)
/ / /

Data provided to the m odel include CO m onitoring data for 1990; tim e-activity data collected in Denver, W a shington, DC, and Ci ncinnati from 1982 to 1985; m i croenvironm ental data; and 1990 census population data. Motor vehicle DPM and CO em ission rates reported by EPA (1999c) are used to calculate m obile-source DPM exposures, and exposures in future years are projected based on the increase in vehicle m iles traveled. EPA’s PART5 m odel is used to
estim ate DPM em ission rates (g/m i) for the fleet as a whole in any given calendar year. PART5 is currently being m odified to account for deterioration, actual in-use em issions, poor

m a intenance, and tam p ering effects, all of which increase em ission factors. As a result, HAPEM-MS3 exposure estim ates based on PART5 em ission factors m a y underestim ate true exposures from on-road sources. A com p arison of PART5 HD diesel vehicle em ission factors with those presented earlier in this chapter suggests that PART5 m a y underestim ate HD diesel
vehicle em issions by up to 50%.
HAPEM-MS3 assum e s that the highway fleet (gasoline plus diesel) em issions ratio of CO to DPM can be used as an adjustm e nt factor to convert estim ated CO personal exposure to

2-111DPM exposure estim ates. This assum p tion is supported by the observation that even though gasoline vehicles em it the large m a jority of CO, gasoline and diesel highway vehicles travel on
the sam e roadways. DPM and CO are both relatively long-lived atm o spheric species (1–3 days) except under certain conditions (Seinfeld and Pandis, 1998); therefore, the m odel does not
account for chem ical and physical differences between the DPM and CO, and the m odel assum e s that for the average person in a m odeled air district, CO and DPM are well m i xed. Exposure in

m i croscale environm ents in which these assum p tions m a y not be valid were not m odeled. A validation study conducted for the pNEM/CO m odel on which HAPEM-MS3 is based indicates that CO exposures for the population in the 5 th percentile were overestim ated by approxim a tely 33%, whereas those with exposures in the 98 th percentile were underestim ated by
about 30%. This validation study is considered applicable to the HAPEM-MS3 m odel. To
address the underestim ate of exposures for th e m o st highly exposed, Brodowicz (1999) used CO
concentrations relevant to the m o st highly exposed populations to determ ine DPM exposures for
different dem ographic groups within this population; the results are discussed below. Annual average DPM exposures from on-road vehicles and nonroad sources nationwide for the general population, rural and urban population, outdoor workers, and urban children are reported in Tables 2-29 and 2-30. The m odeled annual average DPM exposure nationwide (urban and rural areas) in 1996 from on-road sources only was 0.8 μg/m 3

. The m odeled annual average exposure in urban areas for the sam e year was 0.8 μ g/m 3 , and the m odeled exposure for rural areas was 0.4 μg/m 3

. Am ong the dem ographic groups m odeled, urban outdoor workers in general were found to have the highest average exposure to DPM, averaging 1.0 μg/m 3 fromonroad sources in 1996. DPM exposures attributable to on-road sources are projected to decrease until approxim a tely 2007 because of fleet turnover and the full im plem entation of Federal regulations that are currently in place. Full im plem entation of the recently finalized Heavy-Duty Engine and Vehicle Standards and Highway Dies el Fuel Sulfur Control Requirem e nts would significantly lower DPM exposures from on-road sources in the post-2007 tim efram e (U.S. EPA, 2000b). Because diesel vehicle traffic, and therefore exposure to DPM, varies for different urban areas, HAPEM-MS3 was used to estim ate annual average population exposures for 10 urban

areas. Modeled 1996 DPM exposures in the cities ranged from 0.6 μg/m 3 in Chicago and St. Louis to 1.3 μg/m 3 in Phoenix (Table 2-31). In 1996, estim ated average DPM exposure fromonroad sources was higher than the national average in f i ve cities: Atlanta, Minneapolis, New York, Phoenix, and Spokane. Nationally in 1996, 97% of DPM exposure from on-road vehicles was attributable to HD diesel vehicles, and the rest was generated m a inly by LD diesel trucks.

2-112Table 2-29. Annual average nationw i de DPM exposure estimates (μg/m 3 ) from on-road sources for rural and urban demographic groups in 1990, 1996, and 2007 using HAPEM-MS3

Demographi c group 1990 1996 2007

50-St at e popul at i o n 0 .8 0.8 0 .4

R

u ral popul at i o n 0 .5 0.4 0 .2
Urban popul at i o n 0 .9 0.8 0 .4

Urban out door workers 1 .1 1.0 0 .5

Urban chi l d ren (0-17) 0.9 0 .8 0.4
Source : U.S. EPA, 1999b, adjust ed t o refl ect HDDV VM T descri bed i n U.S. EPA, 2000b.

Table 2-30. Draft annual average, 25 th , and 75 th percentile nationw i de DPM exposure estimates (μg/m 3 ) from on-road and nonroad sources for rural and urban counties in 1996 using HAPEM4

Demographi c group 25 th Percentile, DPM 10 mg/m 3 Average, DPM 10 mg/m 3 75 th Percentile, DPM 10 mg/m 3 Contribution toaverage from onroad sources, DPM 10 mg/m 3 Contribution to average from nonroad sources, DPM 10 mg/m 3
Nat i onwi d e 0 .6 1.4 1 .8 0.5 0 .9

R

u ral popul at i o n 0 .3 0.6 0 .7 0.2 0 .3 Urban popul at i o n 0 .8 1.6 2 .0 0.5 1 .1 Source: NATA, 2001. Data availabl e at http://www.ep a.gov/ttn/uatw/nata

.

Because HAPEM-MS3 is suspected to underestim ate exposures in highly exposed populations, 1990 CO concentrations relevant to the m o st highly exposed populations were used
to determ ine 1990 DPM exposures for different dem ographic groups in this population. The highest DPM exposures ranged from 0.8 μg/m 3 for outdoor workers in St. Louis to 2.0 μg/m 3 for outdoor workers in Spokane and up to 4.0 μg/m 3 for outdoor children in New York (Table 2-31). The highest exposed dem ographic groups were those who spend a large portion of their tim e outdoors. It is im portant to note that these exposure estim ates are lower than the total exposure to DPM because they reflect only DPM from on-road sources and not exposure to nonroad DPM

em issions.

2-113Table 2-31. Annual average DPM exposures for 1990 and 1996 in the general population and among the highest exposed de mographic groups in nine urban areas and nationw i de from on-road sources only using HAPEM-MS3

Urban area 1990 Popul ati o n average

exposure, μg/m 3 1996 Popul ati o n average

exposure, μg/m 3 Hi ghest DPM exposure i n 1990, μg/m 3 (demographi c group experiencing this exposure)

N

a t i onw i d e 0 .8 0.8 N A
Atlanta, GA 0.8 0 .9 NA

C h i cago, IL 0.8 0 .6 1.3 (out door workers)

Denver, C O 0.7 0 .8 1.2 (out door workers)

Houst on, TX 0.6 0 .9 0.8 (out door workers)

M i nneapol i s , M N 1.0 1 .0 1.5 (out door workers)

New York, NY 1.6 1 .2 4.0 (out door chi l d ren)

Phi l a del phi a, PA 0.7 0 .7 1.2 (out door chi l d ren)

Phoeni x, AZ 1.4 1 .3 2.4 (nonworki ng m e n 18-44)
Spokane, W A 1.3 1 .1 2.0 (out door workers)

St . Loui s, M O 0.6 0 .6 0.8 (out door workers)
NA - Not available. Source : U.S. EPA, 1999b, adjust ed t o refl ect HDDV VM T descri bed i n U.S. EPA, 2000b.

The HAPEM4 m odeling approach provides exposure estim ates from on-road and nonroad sources as well as point and area sources for pollutants other than DPM. In addition, HAPEM4 incorporates technical advancem ents over previous Agency exposure assessm ents. Instead of using a surrogate pollutant such as CO to estim ate exposure, HAPEM4 uses census tract DPM concentrations provided by the ASPE N dispersion m odel described in Section 2.4.2.3
to estim ate DPM exposure for individuals in each census tract in the United States. The

exposure m odeling results are aggregated to provide county, State, and nationwide exposure
estim ates. HAPEM4 also incorporates the latest data regarding tim e-activity patterns f r om the
Consolidated Hum a n Activity Database and the latest data available regarding penetration of PM
to indoor environm ents. The results of this m odeling approach are currently undergoing peer review and are therefore considered a draft and subject to change. Nationwide exposure estim ates from HAPEM4 are provided in Table 2-30. The draft
National-Scale Assessm ent 1996 national average estim ate of DPM exposure attributable to onroad and nonroad sources is 1.4 μg/m 3

. On-road sources are estim ated to account for 0.5 μg/m 3 and nonroad sources 0.9 μg/m 3

. The HAPEM-MS3 1996 exposure value of 0.8 μg/m 3 and the

2-114m o st recent draft National-Scale Assessm ent value of 0.5 μg/m 3 differ slightly as a result of the
different m odeling approaches. Both the HAPEM-MS3 and HAPEM4 exposure results support the risk perspective provided in Chapter 8, Section 8.3.

2.4.3.2.2. Personal exposures: microenvironments/hotspots. Personal m onitoring for DPM exposure has focused on occupationally exposed groups, including railroad workers, m i ne workers, m echanics, and truck drivers. Although som e studies have m easured personal exposures to am bient PM, none have conducted detailed chem ical analysis to quantify the portion of PM attributable to DE (e.g., using extended species CMB, discussed above). EC concentrations have been reported for som e m i croenvironm ents and are discussed in this section.
Microenvironm ental exposures of significant concern include in-vehicle exposures such as school buses and passenger cars as well as near highways and in urban canyons. Because DPM
f r om m obile sources is em itted into the breathing zone of hum ans, this source has a greater potential for hum an exposure (per kg of em issions) com p ared to com bustion particulates em itted
from point sources.
Recent EC m easurem ents reported for enclosed vehicles driving on Sacram ento roadways ranged from below detection lim its up to 10 μg/m 3 and from 3 μg/m 3 to 40 μg/m 3 on Los Angeles roadways. Elevated levels of PM 2.5 and EC were observed when the vehicle being followed was powered by a HD diesel truck or bus (Cal EPA, 1998b). EC is also present in the exhaust of gasoline vehicles, so these m easurem ents are likely to include som e EC from gasoline
vehicles. The SHEDS (Stochastic Hum a n Exposure and Dose Sim u lation) m odel for PM predicts that although the typical person spends only about 5% of his or her tim e in a vehicle,
this m i croenvironm ent can contribute on average 20% and as m u ch as 40% of a person’s total PM exposure (Burke et al., 2000). The California Air Resources Board also collected EC near the Long Beach Freeway for 4 days in May 1993 and 3 days in Decem ber 1993 (Cal EPA, 1998a). Using em ission estim ates from their EMFAC7G m odel and EC and OC com position profiles for diesel and gasoline exhaust, tire wear, and road dust, CARB estim ated the contribution of the freeway to DPM
concentrations. For the 2 days of sam p ling in Decem ber 1993, DE from vehicles on the nearby freeway was estim ated to contribute from 0.7 μg/m 3 to 4.0 μg/m 3 excess DPM above background concentrations, with a m a xim u m of 7.5 μg/m 3

. In 1986, EC concentrations were m easured in Glendora, CA, during a carbonaceous
aerosol intercom parison study (Cadle and Mulawa, 1990; Hansen and Novakov, 1990). One technique used during the study reported EC concen trations in 1-m i nute intervals, reflecting the
im pact from diesel vehicles 50 m from the study site. The diesel vehicles were estim ated to contribute up to 5 μg/m 3 EC above the background concentration.

2-115In a study designed to investigate relationships between DE exposure and respiratory health of children in the Netherlands, EC m easurem ents were collected in 23 schools located
from 47 m to 377 m from a freeway and in 8 schools located at a distance greater than 400 m
from a freeway (Brunekreef, 1999). EC concentr ations in schools near freeways ranged from 1.1 μg/m 3 to 6.3 μg/m 3 , with a m ean of 3.4 μg/m 3 , and EC concentrations in schools m o re than 400

m

from freeways ranged from 0.8 μg/m 3 to 2.1 μg/m 3 , with a m ean of 1.4 μg/m 3

. Brunekreef et al. (2000), using a reflectance m e thod to report “s oot” or carbonaceous particulate concentrations as a surrogate for EC, found a statistically significant increase in carbonaceous particle concentrations inside and outside of the schools with increasing truck traffic (predom inantly diesel), with decreasing distance between the school and the highway, and with an increase in the percent of tim e the school was downwind of th e highway. In additional studies in elderly subjects in Helsinki and Am sterdam
, Janssen et al. (2000) reported that outdoor m easurem ents of EC were highly correlated with indoor and personal exposure m easurem ents of EC, supporting the position that short-term increases in outdoor EC concentrations are reflected in increased personal exposures even for those who spend m u ch of their tim e indoors. Although there is little quantitative inform ation regarding personal exposure to DPM, certain exposure situations are expected to result in higher than average exposures. Those in the

m o re highly exposed categories would generally include people living in urban areas in which diesel delivery trucks, buses, and garbage trucks frequent the roadways, but also included would be people living near freeways, bus stations, cons truction sites, train stations, m a rinas frequented by diesel-powered vessels, and distribution hubs us ing diesel truck transport. One study using the 1-hydroxypyrene biom arker of DE exposure reported exposure am ong m o st (76%) of the 26
adolescents sam p led in Harlem (Northridge et al., 1999). In a follow-on study, Kinney et al. (2000) reported EC concentrations from persona l m onitors worn by study staff on sidewalks at
four Harlem intersections that ranged from 1.5 μg/m 3 to 6 μg/m 3

. The EC concentrations were found to be associated with diesel bus and truck counts such that spatial variations in sidewalk
concentrations of EC were attributed to local diesel sources in Harlem

. In any situation in which diesel engines operate and a m a jority of tim e is spent outdoors, personal exposures to DE are expected to exceed average exposures. Because a large but currently undefined portion of DPM is em itted during acceleration, those living and working in the vicinity of sources operating in this transient m ode could experience highly elevated levels of
DPM. DPM enriched in soluble organic m a terial (as opposed to EC) is em itted from LD vehicles, som e nonroad equipm ent, on-road diesel engines during cold-start and m o toring conditions, and poorly m a intained vehicles. The potential health effects of acute exposures to elevated DPM levels as well as health effects resulting from chronic exposures are discussed in subsequent chapters in this docum ent.

2-1162.4.3.2.3. The California Population Indoor Exposure Model. CPIEM, developed under contract to the CARB, estim ates Californians’ e xposure to DPM using distributions of input data and a Monte Carlo approach (Cal EPA, 1998a). This m odel uses population-weighted outdoor
DPM concentrations in a m a ss balance m odel to estim ate DPM concentrations in four indoor
environm ents: residences, office buildings, schools, and stores/retail buildings. The m odel takes into account air exchange rates, penetration factors, and a net loss factor for deposition/rem oval.
In four additional environm ents (industrial plants, restaurants/lounges, other indoor places, and enclosed vehicles), assum p tions were m a de about the sim ilarity of each of these spaces to
environm ents for which DPM exposures had been calculated. Industrial plants and enclosed
vehicles were assum e d to have DPM exposures sim ilar to those in the outdoor environm ent; restaurants/lounges were assum e d to have DPM concentrations sim ilar to stores; and other indoor places were assum e d to have DPM concentrations sim ilar to offices. The estim ated DPM
concentrations in the indoor and outdoor environm ents range from 1.6 μg/m 3 to 3.0 μg/m 3 (Table 2-32).

Table 2-32. Modeled and estimated concentrations of DPM in microenvironments for California for all sources

Microenvironment E stimated mean DPM (stdev), μg/m 3

Residences 1.9 (0.9)
Offices 1.6 (0.7)
Schools 1 .9 (0.8) Stores/public/retail bldgs 2.1 (0.9) Outdoor places 3.0 (1.1) Industrial plants a 3.0 (1.1) Restaurants/lounges a 2.1 (0.9) Other indoor places a 1.6 (0.7)

Enclosed vehicles a 3.0 (1.1)

a C oncent r at i ons assum e d based on si m i l a ri t y wi t h m odel e d envi ronm ent s

. Source: C a l i f orni a EPA, 1998a.

2-117The DPM concentrations reported in Table 2-32 were used as input to CPIEM, and tim eactivity patterns for children and adults were used to estim ate total indoor and total air exposures
to DPM. Overall, total indoor exposures were estim ated to be 2.0 ± 0.7 μ g/m 3 , and total air exposures (indoor and outdoor exposures) were 2.1 ± 0.7 μg/m 3 (Table 2-33). The South Coast Air Basin and the San Francisco Bay Area were also m odeled using CPIEM, where total air exposures to DPM were estim ated to be 2.5 ± 0.9 μg/m 3 and 1.7 ± 0.9 μg/m 3 , respectively.

Table 2-33. Estimated indoor air and total air exposures to DPM in California in 1990

Exposed population Total indoor exposure (stdev),
μg /m 3 Total air exposure, (stdev), μg/m 3

All Californians 2 .0 (0.7) 2 .1 (0.8)

South Coast Air Basin 2 .4 (0.9) 2 .5 (0.9)
San Francisco Bay Area 1.7 (0.9) 1 .7 (0.9)
Source: C a l i f orni a EPA, 1998a.

Exposure estim ates were also m a de by Cal EPA (1998a) for 1995, 2000, and 2010 using a
ratiom e tric approach to 1990 exposures. Total air exposures reported for 1995 and projected for 2000 and 2010 were 1.5 μg/m 3 , 1.3 μg/m 3 , and 1.2 μg/m 3 , respectively.

2.5. SUMMARY AND DISCUSSION This chapter sum m a rizes inform ation regarding the history of the use of diesel engines, technological developm ents and their im pact on em issions over tim e, Federal standards on DE,
the chem ical and physical character of DE, atm o spheric transform a tions of DE, and am bient DE concentrations and exposures. The aspects of each of these topics that are m o st relevant to the discussion of health effects in later chapters of this docum ent are sum m a rized here. Because the

m a jority of inform ation regarding the chem ical com position and historical changes in DE pertains to on-road diesel engines, these data are discussed in greater detail than diesel em issions
from nonroad equipm ent. W h ere possible, nonroa d em issions were discussed in Chapter 2 and are briefly sum m a rized here.

2-1182.5.1. 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 am ong HD truc ks, accounting for 36% of new HD truck sales in 1960, 85% of sales in 1970, and alm o st 100% of sales in 1997. It is estim ated that in 2000,
HD diesel vehicles will travel m o re than 224 billion m iles (U.S. EPA, 2000b). In 1997, onhighway HD diesel engines contributed 66% of the PM 2.5 em itted by on-highway vehicles. To understand changes in em issions over tim e, it is im portant to note the difference
between m odel year em ission trends and calendar year em ission trends. Em ission trends by

m odel year refer to the year in which an engine was m a de; the em ission rate is specific to the technology and regulations in effect for that year

. Em issions in a specific calendar year refer to aggregate em issions due to the m i x of m odel year engines on the road. Because of the tim e required for fleet turnover, em ission rates for the on-road fleet in any calendar year are not as
low as the m o st recent m odel year em ission rate. In 1997, 40% of the HD vehicles on the road were at least 10 years old and traveled approxim a tely 17% of total HD vehicle m iles.
EPA set a sm oke standard for on-road HD diesel engines beginning with the 1970 m odel

year. In the ensuing years, standards for PM from diesel engines for on-road applications
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 em ission c ontributions of PM from diesel engines to
national PM 10 inventories reflect decreases expected to result from Federal regulations, because
the em ission factor m odels (MOBILE5 and PART5) used to provide em ission estim ates for

m obile sources largely use engine test data required for certification. The U.S. EPA Trends Report estim ates that PM 10 em issions attributable to on-road diesel vehicles decreased 27%
between 1980 and 1998. DPM em ission factors (g/m i by m odel year) m easured from in-use vehicles decreased on average by a factor of six from the m i d-1970s to the m i d-1990s.
It is im portant to note that in spite of the decreasing trend in DPM em ission factors by

m odel year, a wide range in em ission factors from in-use testing is reported, even for newer

m odel year HD vehicles (from less than 0.1 g/m i to m o re than 1 g/m i for m odel year 1996 vehicles). The high variability in DPM em issions within one m odel year has been attributed to
deterioration 3 and differences in m easurem ent m e thods and test conditions at the various testing
facilities. Studies in which consistent testing m e thods were used suggest that deterioration (even
f o r newer m odel year engines) causes som e of the variability in em ission f actors, whereas other

3 Det e ri orat i on i n cl udes i n creases i n em i ssi on rat e s (g/ bhp hr) due t o norm a l wear as wel l as m a nufact uri n g defect s and m a l f unct i ons such as ret a rded t i m i ng, fuel i n ject or m a l f unct i on, sm oke l i m i t i ng m echani s m probl em s, clogged air filter, wrong or worn tu rbocharger, clogged intercooler, engi ne m echanical failure, excess oil

consum pt i on, and el ect roni cs t h at have been t a m p ered wi t h or have fai l e d.

2-119studies clearly dem onstrate the im portant influen ce of test conditions and driving protocols (e.g., aggressive driving) on DPM em ission factors. Even though significant reductions in DPM from diesel vehicle em issions for on-road applications have been realized, diesel engines (nonroad and on-road com b ined) are still significant contributors to 1998 inventories of particulate m a tter, contributing approxim a tely
23% of PM 2.5 em issions (not including the contribution from natural and m i scellaneous sources). Technology innovations that im pact diesel engine em issions have occurred in the years since 1960, in particular the advent of turbocharg ing with charge air cooling and direct-injection engines. The use of these new technologies tends to lower em issions from on-road diesel engines; until the late 1970s, however, engines were optim ized for perform a nce rather than

em issions, so the effect on em issions prior to this tim e was sm all. The lim ited am ount of data available indicates that on-road engines in the 1950 to 1975 tim efram e had DPM em issions
sim ilar to, and in som e cases higher than, those of the m i d-1970 engines that were not yet controlled for particulates. Few data are available to assess the changes in em ission rates from locom o tive, m a rine, or other nonroad diesel engine sources over tim e. It is expected that because the typical lifespan
of a locom o tive engine is at least 40 years and PM regulations for these engines do not take effect until 2000, PM em ission rates by m odel year from locom o tives are not likely to have changed substantially since the introduction of the diesel engine into the railroad industry in the early 1950s.
Particulate m a tter regulations for nonroad diesel equipm ent are not as stringent as PM regulations for on-road diesel engines. Although PM em issions have declined for on-road trucks, it is estim ated that PM 10 em issions from nonroad diesel engines increased 17% between
1980 and 1998. DPM em issions from nonroad diesel engines are expected to continue to
increase from current levels in the absence of new regulations. No inform ation is available regarding changes in the chem ical com position of nonroad engine em issions over tim e.

2.5.2. Physical and Chemical Composition of Diesel Exhaust
Com p lete and incom p lete com bustion of fuel in the diesel engine results in the form ation
of a com p lex m i xture of hundreds of organic and inorganic com pounds in the gas and particle phases. Am ong the gaseous com ponents of DE, the aldehydes are particularly im portant because of their health effects and because they are an im portant fraction of the gaseous em issions.
Form aldehyde m a kes up a m a jority of the aldehyde em issions (65%-80%) from diesel engines,
with the next m o st abundant aldehydes be ing acetaldehyde and acrolein. Other gaseous

com ponents of DE that are notable for their h ealth effects include benzene, 1,3-butadiene, PAH,
and nitro-PAH. Dioxin com pounds have also been detected in trace quantities in DE and

2-120currently account for 1.2% of the national inventory. Dioxin com pounds are known to

accum u late in certain foods, such as beef, poultry, and dairy products. It is unknown whether deposition of DE em issions has an im pact on food chains in local areas. DPM contains EC, OC, and sm all am ounts of sulfate, nitrate, m e tals, trace elem ents, water, and unidentified com pounds. DPM is typically com posed of m o re than 50% to approxim a tely 75% EC depending on the age of the engine, deterioration, HD versus LD, fuel characteristics, and driving conditions. The OC portion of DPM originates from unburned fuel, engine lubrication oil, and low levels of par tial com bustion and pyrolysis products and typically ranges from approxim a tely 19% to 43%, although the range can be broader depending on m a ny of the sam e factors that influence the EC cont ent of DPM. Polyarom atic hydrocarbons generally
constitute less than 1% of the DPM m a ss. Metal com pounds and other elem ents in the fuel and engine lubrication oil are exhausted as ash and typically m a ke up 1%-5% of the DPM m a ss. Elem ents and m e tals detected in DE include barium , calcium , chlorine, chrom i um , copper, iron, lead, m a nganese, m e rcury, nickel, phosphorus, sodium , silicon, and zinc. The com position of DPM contrasts strongly with the typical chem ical com position of am bient DPM 2.5 that is
dom inated by sulfate for aerosols m easured in the eastern United States and by nitrate,

am m onium , and OC in the western United States. Approxim a tely 1% to 20% of the m a ss of DPM in DE is in the ultraf ine size range (nuclei-m ode), with the m a jority of particles ranging in size from 0.005 to 0.05 m i crons and having a m ean diam eter of about 0.02 m i crons. These particles account for 50%-90% of the

num ber of particles. These ultrafine particles ar e largely com posed of sulfate and/or sulfate with
condensed OC. Evidence regarding an increase in the num ber of ultrafine particles from new HD engines
is inconclusive. The dilution conditions used to m easure the size distribution of DE have a large
im pact on the num ber of ultrafine particles quantified. To understand the size distribution of DPM to which people are exposed will require m easurem ents under conditions that m o re closely

resem b le am bient conditions. Approxim a tely 80%-95% of the m a ss of particles in DE is in the size range from 0.05-1.0

m i crons, with a m ean particle diam eter of about 0.2 m i crons, and therefore in the fine PM size

range. Diesel particles in the 0.05-1.0 m i cron range are aggregates of prim ary spherical particles consisting of an EC core, adsorbed organic com pounds, sulfate, nitrate, and trace elem ents. These particles have a very large surface area per gram of m a ss, which m a kes them an excellent carrier for adsorbed inorganic and organic com pounds and, due to their sm all size, they can effectively reach the lower portions of the resp iratory tract. The EC core has a high specific surface area of approxim a tely 30-90 m 2 /g.

2-121Because of the potential toxicological signi ficance of the organic com ponents associated with DPM, it is im portant to understand, to the extent possible, the historical changes in the

am ount and com position of the DPM-associated organic fraction. The organic com ponent of DPM has typically been characterized by extraction with organic solvents, although other techniques such as therm ogravim etric m e thods have also been used. Results from studies using
sim ilar extraction m e thods were com p ared to characterize historical changes in the SOF

em ission rates, the percentage of DPM com p rised by SOF, and the com position of SOF. Data
from both engine and chassis dynam o m e ter tests s uggest that SOF em ission rates have decreased by m odel year from 1975 to 1995. W h en expressed as a percentage of total DPM, the
contribution of SOF to total DPM dem onstrates a wide range of variability that m a y be attributed to different test cycles, different engine t ypes, and different deterioration rates am ong the
vehicles tested. Currently, LD diesel engine s em it DPM with a higher fraction of SOF than do HD engines. Chassis dynam o m e ter tests dem onstrate an overall decrease in the m a ss percentage contribution of SOF to DPM, ranging from 10% to 60% in the 1980s and ~5% to 20% in the
1990s. In contrast, engine dynam o m e ter tests dem onstrate that typically 10%-50% of DPM m a ss is soluble organic m a tter for engines in m odel years 1980-1995. The higher SOF fraction of
DPM from 1990s m odel year engine dynam o m e ter tests is attributed prim arily to the differences in the engine and chassis dynam o m e ter driving cycles. The engine dynam o m e ter testing includes high- speed and low-load or low-speed lugging test m odes in the engine Federal Test Procedure that produce DPM with a high SOF fraction. The chassis dynam o m e ter data are considered to reflect real-world trends in em issions
from heavy HD vehicles by m odel year because vehicles from different m odel years, with
different m ileage and different levels of deterioration, are represented. Thus, it is expected that the percentage of SOF from new (1990 or later) m odel year heavy HD diesel vehicles is lower
than that from older vehicles. This expectation is supported by data dem onstrating an overall
increase in the fraction of EC in the carbonaceous com ponent of DPM. The im portant observation from the engine test data is that som e driving m odes occurring in real-world applications even with new (post-1990) engines m a y produce DPM with a high SOF com ponent (up to 50%). PAH and nitro-PAH are present in DPM from both new and older engine exhaust. There
is no inform ation to suggest that the overall PAH com position profile for DPM has changed. There are too few data to speculate on the changes in em issions of total PAH, nitro-PAH, or
PAH and nitro-PAH com ponents such as BaP and 1- NP. The data suggest that differences in a vehicle’s engine type and m a ke, general engi ne condition, fuel com position, and test conditions
can influence the em issions levels of PAH. Som e studies suggest that fuel com position is the

m o st im portant determ inant of PAH em issions. There is lim ited evidence that gas-phase PAH

2-122em ission rates increase with higher fuel PAH content and that som e particle-phase PAH

em ission rates increase with higher fuel PAH conten t. These data suggest that during the period
from 1960 to 1986, when the arom atic content of fuel increased, PAH em issions m a y have
increased until the arom atic content of diesel fuel was capped in 1993. The arom atic content of nonroad diesel fuel is not federally regulated a nd is typically greater than 30%. PAH em issions
from nonroad equipm ent would also be expected to vary with the PAH content of the fuel. Currently, inform ation regarding em ission rates, chem ical com position, and relative
contribution of DPM f r om high-em itting HD dies el vehicles is not available and m a y significantly change the understanding of DPM com position to which people are exposed. Som e studies have reported a substantial num ber of sm oking diesel trucks in the in-use fleet. Although
the correlation between sm oke and particulate concentration varies with the driving cycle and

m easurem ent m e thod, the results of sm oke opacity tests suggest that high-em itting HD diesel
vehicles m a y be im portant contributors to am bient DE and DPM concentrations.
The chem ical com position of DPM to which people are currently exposed is determ ined by a com b ination of older and newer technology on-road and nonroad engines. Consequently, the decrease in the SOF of DPM by m odel year does not directly translate into a proportional decrease in DPM-associated organic m a terial to which people are currently exposed. In addition, the im pact from high-em itting and/or sm oking diesel engines is not quantified at this
tim e. Because of these uncertainties, the changes in DPM com position over tim e cannot presently be quantified. The data clearly indi cate that toxicologically significant organic

com ponents of DE (e.g., PAHs, PAH derivatives, nitro-PAHs) were present in DPM and DE in the 1970s and are still present in DPM and DE as a whole. Although a significant fraction of am bient DPM (over 50% is possible) is also em itted by nonroad equipm ent, there are no data available to characterize changes in the chem ical

com position of DPM from nonroad equipm ent over tim e. Som e analysts project that diesel engines will increase substantially in the LD f l eet in

com i ng years. Although LD engines currently em it DPM with higher SOF than HD engines of
the sam e m odel year, recently prom ulgated Tier 2 standards will require control m easures in the
2004-2007 tim efram e that will reduce PM em issions from these vehicles. These control

m easures provide som e assurance that even if LD diesel use increases, DPM em itted f r om these vehicles will likely have a sm aller SOF com ponent than such engines currently em it.

2.5.3. Atmospheric Transformation of Diesel Exhaust An understanding of the physical as well as chem ical transform a tions of DE in the

atm o sphere is necessary to fully understand the im pact of this com p lex chem ical m i xture on hum an health. In the past two decades, data acquired from laboratory and am bient experim e nts

2-123have provided inform ation regarding the atm o spheric loss processes and transform a tion of DE, but knowledge concerning the products of these chem ical transform a tions is still lim ited. A recent study has suggested that DPM exposed to am bient levels of ozone is sufficiently altered to increase the rat lung inflam m a tory effect com p ared with DPM not exposed to ozone. Studies investigating the chem ical and physical changes of DE em issions suggest that there is little or no hygroscopic growth of prim ary diesel particles. This observation suggests
that the sm all size of DPM particles m i ght be m a intained upon inhalation, particularly near the

em ission source, allowing these particles to reach the lower portions of the respiratory tract. Increased solubility can increase the rem oval efficiency of secondary diesel particles com p ared with their precursor com pounds. Secondary aerosols from DE m a y also exhibit different biological reactivities f r om the prim ary particles. For exam ple, there is evidence f o r nitration of

som e PAH com pounds resulting in the form ation of nitroarenes that are often m o re m u tagenic than their precursors.

2.5.4. Ambient Concentrations and Exposure to Diesel Exhaust Because of changes in engine technology and DPM em issions over tim e, am bient concentrations reported from studies before 1990 are com p ared here to those reported after 1990.
There are no studies in which direct com p arisons can be m a de because of different analytical and

m odeling tools used to assess DPM am bient levels. DPM concentrations reported from CMB and dispersion m odeling studies in the 1980s suggest that in urban and suburban areas (Phoeni x, AZ, and Southern California), annual average DPM concentrations ranged from 2 to 13 μg/m 3 , with possible m a xim u m daily values in Phoenix of 22 μg/m 3

. In these studies, the average contribution of DPM in urban areas to total am bient PM ranged from 7% in Pasadena, CA, to 36% in Los Angeles.
In the 1990 tim efram e , annual or seasonal average DPM concentrations reported in CMB
studies and from EC m easurem ents for urban and suburban areas range from 1.2 to 4.5 μg/m 3

. The contribution of DPM to am bient PM at these sites averaged 10%-15% on a seasonal or annual basis, with contributions up to 38% on individual days (Brighton, CO). Dispersion

m odeling on individual days in Southern California in the 1990s predicts DPM concentrations ranging from 1.9 to 4.4 μg/m 3 (8%-12% of am bient PM). On individual days at a m a jor bus stop in New York City, DPM concentrations were reported to reach 46.7 μg/m 3 and averaged 53% of

am bient PM, highlighting the im portant influence of diesel bus traffic in an urban street canyon. In nonurban and rural areas in the 1980s, DPM concentrations reported range from 1.4 to 5 μg/m 3 and on average com p rised 5%-12% of the am bient aerosol. In the 1990s, nonurban air basins in California were reported to have DPM concentrations ranging from 0.2-2.6 μg/m 3

.

2-124Although estim ates from em issions m odels suggest that DPM em issions from on-road sources decreased during the 1990s, the atm o spheric data available do not provide a clear
indication of trends in DPM concentrations but ar e likely to be m o re a reflection of the choice in

sam p ling sites, source apportionm ent m e thods, and m odeling techniques. In general, from the
lim ited num ber of studies available it appears that DPM concentrations averaged over at least a season in the 1990s typically ranged from 1-4 μg/m 3

. These data can be used in m odel-m onitor

com p arisons and to provide an indication of l ong-term average exposures in som e urban areas.
Additional work is needed to assess am bient DPM and DE concentrations in several urban
environm ents, to assess m i croenvironm ents, and to evaluate the relative im pact of nonroad and
on-road sources on concentrations.
A com p rehensive exposure assessm ent cannot currently be conducted because of the lack
of data. Inform ation regarding DPM in occupational environm ents suggests that exposure ranges up to approxim a tely 1,280 μg/m 3 for m i ners, with lower exposure m easured for railroad workers (39-191 μg/m 3 ), firefighters (4-748 μg/m 3 ), public transit personnel who work with diesel equipm ent (7-98 μg/m 3 ), m echanics and dockworkers (5-65 μg/m 3 ), truck drivers (2-7 μg/m 3 ), and bus drivers (1-3 μg/m 3 ). W o rk area concentrations at fixed sites are often higher
than m easured exposures, especially for m i ning operations or other enclosed spaces. For several occupations involving DE exposure, an increased risk of lung cancer has been reported by epidem iologic studies (discussed in Chapter 7). An estim ate of the 70-year lifetim e environm ental exposure equivalent to these occupational exposures provides one m e ans of

com p aring the potential overlap between occupational exposures and exposures m odeled for the general public. The estim ated 70-year lifetim e exposures equivalent to those for the occupational groups discussed above range from 0.4-2 μg/m 3 on the low end to 2-269 μg/m 3 on the high end. The EPA has perform e d a national-scale exposure assessm ent for DPM from on-road

sources. Current national exposure m odeling using the HAPEM-MS3 m odel suggests that in 1996, annual average DPM exposure from on-road DE sources in urban areas was 0.8 μg/m 3

, whereas in rural areas, exposures were 0.4 μg/m 3

. Am ong 10 urban areas in which DPM

exposures were m odeled, 1996 annual average exposure from on-road DE sources ranged from 0.6 μg/m 3 to 1.2 μg/m 3

. Outdoor workers and children who spent a large am ount of tim e outdoors were estim ated to have elevated DPM exposures in 1990, ranging up to 4.0 μg/m 3 from on-road sources only. Based on the national inventory, nonroad em ission sources could
contribute at least twof old m o re DPM than that em itted by on-road sources. Results of the draf t National-Scale Assessm ent for 1996 indicate that national average exposure to DPM, including nonroad sources, is 1.4 μg/m 3 , with 0.9 μg/m 3 of that average attributed to em issions from
nonroad sources.

2-125Low-end exposures for m a ny of the occupational groups overlap 1990 and 1996 exposures from on-road sources m odeled for the general population (0.8 μg/m 3 ) and for the m o re highly exposed groups. This potential overlap, or sm all difference between occupational and

am bient exposures, presents a concern that health effects observed in occupational groups m a y also be evidenced in the general population. The potential m a gnitude of this risk is discussed in Chapter 8. In different exposure environm ents, the types of diesel vehicles, their m ode of operation,

m a intenance, atm o spheric transform a tion, and m a ny additional factors influence the chem ical nature and quantity of DPM to which people are exposed. The potential health consequences of both short- and long-term exposures to DE are discussed in the following chapters of this
docum ent.

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2-1433-1 3. DOSIMETRY OF DIESEL PARTICULATE MATTER

3.1. INTRODUCTION Animals and humans receive different internal doses when breathing the same external
concentrations of airbor ne materials such as diesel partic ulate matter (DPM) (Brain and Mensah, 1983; Schlesinger, 1985). The dose received in diffe rent species differs fro m the aspects of the total amount deposited within the re spiratory tract, the relative dist ribution of the dose to specific regions in the respiratory tract, and the residence time of these materials within the respiratory tract, i.e., clearance. Using an external concentration breathed by laboratory animals as a basis for any guidance for human exposur e to DPM would then be an in adequate approx imation of the total and regional dose that humans may receive. The reason for the existence of this chapte r and for consideration about interspecies dosimetry is the lack of human health effect data on DPM and the concomitant need to be able to evaluate existing animal data from the aspect of an equivalent human dose. The objective of this chapter is to evaluate and address this issue of interspecies dosimetr ic differences through:

• A general overview of what is known a bout how particles lik e DPM are deposited, transported to, and cleared from the re spiratory tract. Information on both laboratory animals (mainly rodents) and humans will be considered and interspecies similarities and differences highlighted.
• An overview of what is known about th e bioavailability of the organic compounds
adsorbed onto DPM from information in hum ans, animals, and in vitro studies, and from model predictions. • An evaluation of the suitability of available dosimetric models and procedures for DPM to estimate interspecies extrapolations whereby an exposure scenario, conditions, and outcome in laboratory an imals are adjusted to an equivalent
outcome in humans via calcula tion of an internal dose.

The focus in this chapter will be on the particulate fraction of diesel emissions, i.e, DPM. Although diesel engine exhaust consists of a complex mixture of typical combustion gases, vapors, low-molecular-weight hydrocarbons, and pa rticles, it is the pa rticle phase that is considered to be of major health concern. The major constituents of diesel engine exhaust (DE) and their atmospheric reaction produc ts are described in Chapter 2. As will be deduced in Chapter 5, pulmonary toxicity and carcinogenicity are the major focal points of diesel toxicity and of DPM depos ition. Therefore, dosimetr ic considerations are limited to the lung although DPM deposition would occur throughout the respiratory tract, from3-2 the nares to the alveoli. Aspects of respiratory tract dosimetry to be considered in this chapter include the characteristics of DPM, deposition of DPM throughout the respiratory tract, the conducting airways and alveolar regions, nor mal DPM clearance mechanisms and rates of clearance in both these regions, clearance rates durin g lung overload (in rats), elution of organics from DPM, transport of DPM to extra-alveolar site s, and the interrelationships of these factors. The overall goal in this chapter follows from the objective—to judge the feasibility and suitability of procedures allowing for derivation of an internal dose estimate of DPM for humans, i.e., of a human equivalent c oncentration to exposure concentr ations and conditions used in animal studies. This goal is of significance es pecially in the quantitative dose-response analysis of DPM effects in laboratory animals proposed in Chapter 6.

3.2. CHARACTERISTICS OF INHALED DIESEL PARTICULATE MATTER The formation, transport, and characteristic s of DPM are among the subjects considered in detail in Chapter 2. DPM consists of aggreg ates of spherical carbonace ous particles (typically
about 0.2 : m mass median aerodynamic diameter [MM AD] or, more appropr iately, mass median thermodynamic diameter [MMTD]) to which si gnificant amounts of hi gher-molecular-weight organic compounds are adsorbed. DPM has an ex tremely large surface area that allows for the adsorption of organic compounds (see Chapter 2, Section 2.2.2). The organic carbon portion of DPM can range from at least 19% to 43% from high way diesel engines; no data are available to characterize the organic content of DPM from nonr oad engines. The toxicologically relevant organic chemicals include high -molecular-weight hydrocarbons su ch as the polycyclic aromatic hydrocarbons (PAHs) and their deri vatives (Chapter 2, Section 2.2.8).

3.3. REGIONAL DEPOSITION OF INHALED DIESEL PARTICULATE MATTER This section discusses the major factors controlling the disposition of inhaled particles. Note that disposition is defined as encomp assing the processes of deposition, absorption, distribution, metabolism, and elimination. The re gional deposition of par ticulate matter in the respiratory tract is dependent on the interaction of a number of f actors, including respiratory tract anatomy (airway dimensions and branching configurations), ventilatory characteristics (breathing mode and rate, ventilatory volum es and capacities), physical pro cesses (diffusion, sedimentation, impaction, and interception), and the physicoche mical characteristics (particle size, shape, density, and electrostatic attracti on) of the inhaled pa rticles. Regional de position of particulate material is usually expressed as deposition fraction of the total pa rticles or mass inhaled and may be represented by the ratio of the particles or mass deposited in a specific region to the number or mass of particles inspired. The factors affec ting deposition in these various regions and their importance in understanding the fa te of inhaled DPM are discusse d in the following sections.3-3 It is beyond the scope of this document to present a comprehensive account of the complexities of respiratory mechanics, physiology, and toxicology, and only a brief review will be presented here. The reader is referred to publications that provide a more in-depth treatment of these topics (Weibel, 1963; Brain and Me nsah, 1983; Raabe et al ., 1988; Stöber et al., 1993; U.S. EPA, 1996). The respiratory tract in both humans and exper imental mammals can be divided into three general regions on the basis of structure, si ze, and function: the extrathoracic (ET), the tracheobronchial (TB), and the alveolar (A). In humans, inhalation can occur through the nose or mouth or both (oronasal breathing). Animal mode ls used in respiratory toxicology studies, particularly the rat, however, are obligate nose breathers.

3.3.1. Deposition Mechanisms This section provides an overview of the ba sic mechanisms by which inhaled particles deposit within the respiratory tract. Details con cerning the aerosol physi cs that explain both how and why particle deposition occurs as well as data on total human respiratory tract deposition are presented in detail in the earlie r PM Criteria Document (U.S. EP A, 1996) and will only be briefly
summarized here. For more extensive discussions of deposition processes, refer to reviews by Morrow (1966), Raabe (1982), U.S. EPA (1982) , Phalen and Oldham (1983), Lippmann and Schlesinger (1984), Raabe et al. (1988), and Stöber et al. (1993). As pictorially represented in Figure 3-1, pa rticles may deposit by five major mechanisms (inertial impaction, gravitationa l 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 respiratory tract. It is important to appreciate that these pro cesses are not necessarily independent but may, in some instances, interact with one another such that total deposition in the respiratory tract may be less than the calculated pr obabilities for deposition by the i ndividual processes (Raabe, 1982). Depending on the particle size and mass, varying degrees of de position may occur in the ET (or nasopharyngeal), TB, and A regi ons of the respiratory tract. Upon inhalation of particulate matter such as that found in DE, particle deposition will occur throughout the respiratory tract. Because of high airflow velocities and abrupt directional changes in the ET and TB regions, inertial impaction is a primary deposition mechanism, especially for particles $ 2.5 : m d
ae (aerodynamic equivalent di ameter). Although inertial impaction is a prominent process for deposition of larger particles in th e tracheobronchial region, it is of considerably less significance as a determinant of regional deposition patterns for3-4 Figure 3-1. Schematic representation of major mechanisms, including diffusion, involved in particle deposition. Airflow is signified by the arrows and particle trajectories by the dashed line.3-5 DPM, which have a d
ae # 0.2 : m and may be considered a rath er polydisperse distribution with sigma g values of 2.4 and greater. All aerosol particles are c ontinuously influenced by gr avity, but particles with a
d
ae > 0.5 : m are affected to the greatest extent. A spherical compact particle will acquire a terminal settling velocity when a balance is achieved between th e acceleration of gravity acting on the particle and the viscous resistance of the air; it is this velocity that brings the particle into contact with airway surfaces. Both sediment ation and inertial impacti on cause the deposition of many particles within the same size range. These deposition processes act together in the ET and TB regions, with inertial impac tion dominating in the upper airw ays and sedimentation becoming increasingly dominant in the lower conducting airway s, especially for the largest particles that can penetrate into the smaller bronchial airways. As particle diameters become <1 : m, the particles are increasingly subjected to diffusive deposition because of random bombardment by ai r molecules, which results in contact with airway surfaces. A d
ae of 0.5 : m is often considered a boundary between diffusion and aerodynamic (sedimentation and imp action) mechanisms of deposition. Thus, instead of having
a d
ae , diffusive particles of different shapes can be related to the diffusivity of a thermodynamic equivalent size based on sphe rical particles (Heyder et al ., 1986). Diffusive deposition of particles is favored in the A region of the respiratory tract as particles of this size are likely to penetrate past the ET and TB regions. Electrostatic precipitation is deposition related to particle charge. The electrical charge on some particles may result in an enhanced deposition over what would be expected from size alone. This is due to image charges induced on the surface of the airway by these particles, or to space-charge effects whereby repulsion of particle s containing like charges results in increased migration toward the airway wall. The effect of charge on deposition is inversely proportional to particle size and airflow rate. A recent study employing hollow airway casts of the human
tracheobronchial tree that asse ssed deposition of ultrafine (0.02 : m) and fine (0.125 : m) particles found that depos ition of singly charged particles wa s 5-6 times that of particles having no charge, and 2-3 times that of particles at Boltz mann equilibrium (Cohen et al., 1998). This suggests that within the TB region of humans, electrostatic precipitation may be a significant deposition mechanism for ultrafine and some fine particles, the latter of which are inclusive of DPM. Thus, although electrostatic precipitati on is generally a minor contributor to overall particle depositi on, it may be important for DPM. Interception is deposition by physical contact with airway surfaces and is most important for fiber deposition (U.S. EPA, 1996). Figure 3-2 shows the regional (ET, TB, A) deposition in the human respiratory tract as influenced by particle size. Keeping in mind that DPM is a polydisperse distribution with 0.2 : m3-6 Figure 3-2. Generalized regional deposition fr actions of various sized particles in the human respiratory tract. (Adapted from the International Commission on Radiological Protection (ICRP) Publicatio n 66 (1994) model. For unit density, spherical particles inhaled through the nose by an adult male with a tidal volume of 1250 mL, respiratory frequency of 20 min -1 , and functional residual capacity (FRC) of 3300 mL.) ET, extrathoracic; TB, tracheobronchial; A, alveolar.

being only the median diameter, it can be seen that principal fraction particles sized from < 0.2
down to around 0.002 : m would, as predicted based on their size and the expected mechanism of diffusion, deposit in the alveolar region. Partic les below this size range (and above around 4 : m) tend to deposit in the ET region. Specific mode ling results for depositi on of DPM particles inclusive of their distribution (i.e., F g ) are presented in Section 3.6.

3.3.1.1. Biological Factors Modifying Deposition The available experimental deposition data in humans are commonly derived using healthy adult Caucasian males. Various factors can act to a lter deposition patterns from those obtained in this group. The effects of different biological factors, incl uding gender, age, and respiratory tract disease, on pa rticle deposition have been reviewed previously (U.S. EPA, 1996, Section 10.4.1.6). In general, there appears to be an inverse relationship between airway resistance and total deposition.3-7 Differences in patterns of deposition between humans and an imals have been summarized (U.S. EPA, 1996; Schlesinger, 1985) and show clear ly that when exposed to the same aerosol or gas, humans and animals receive doses that may differ in both total and regional (i.e., ET, TB, or A) deposition from a number of variables including particle size, especially for larger sized particles, i.e. d
ae $ 1 : m. Such interspecies differences ar e important because the adverse toxic effect is likely more related to the quantitative pattern of deposition within the respiratory tract than to the exposure concentration; this pattern determines not only the initial respiratory tract tissue dose but also the specific pathways by which the inhaled material is cleared and redistributed (Schlesinger, 1985). Such differences in initial de position must be considered when relating biological responses obtained in la boratory animal studies to effects in humans. The deposition patterns of inhale d diesel particles in the re spiratory tract of humans and mammalian species has been reviewed (Health Effects Institute, 1995). Schlesinger (1985) showed that physiological diffe rences in the breathing mode for humans (nasal or oronasal breathers) and laboratory rats (obligatory nose breathers), co mbined with different airway geometries, resulted in signifi cant differences in lower respir atory tract deposition patterns for larger sized particles (>1 : m d
ae ) in that a much lower fraction of inhaled larger particles is deposited in the alveolar region of the rat compared with humans. However, alveolar deposition of the much smaller DPM (around 0.2 : m d
ae ) was not affected as much by the differences among species, as was demonstrated in mode l calculations by Xu and Yu (1987). These investigators modeled the deposition efficiency of inhaled DPM in rats, hamsters, and humans on
the basis of calculations of the models of Schum and Yeh (1980) and Weibel (1963). These simulations (Figure 3-3) indicate relative deposition patterns in the lower respiratory tract (trachea = generation 1; alveoli = generation 23) and are similar among hamsters, rats, and humans. Variations in alveolar deposition of DPM over one breathing cy cle in these different species were predicted to be within 30% of one another (Xu and Yu, 1987). Xu and Yu (1987) note that this similarity is concordant with the premise that deposition of the submicron diesel particles is dominated by diffusion rather than sedimentation or impac tion. Although these data assumed nose-breathing by humans, the results w ould not be very differe nt for mouth-breathing because of the low filtering capacity of the nose for particles in the 0.1 to 0.5 : m range (see Figure 3-2). The preceding discussion addr esses deposition patte rns and deposition efficiencies of DPM in the respiratory tract of various species including human s. The alveolar region was focused upon primarily because, as shown in Chapte r 5, this region is where adverse effects from long-term DPM exposure are typically observe d. For dosimetric calcu lations and modeling, however, it would be of much greater importance to consider the actual deposited dose. Table 3-1 presents the analysis of Xu and Yu (1987) on prediction of the deposited doses of DPM3-8 Figure 3-3. Modeled depositi on distribution patterns of inhaled DE particles in the airways of different species. Generation 1-18 are TB; >18 are A.

inhaled in 1 min in the lungs of humans, rats, and hamsters on three diffe rent bases: the total lung volume (M), the surface area of all lung airways (M 1 ), or the surface area of the epithelium of the alveolar region only (M 2 ). According to this analysis, the deposited dose is lower in humans than in the two rodent species regardle ss of how the deposited dose is expressed. These results are most certainly due predominately to the greater respiratory exchange rate in rodents
and smaller size of the rodent lung. Table 3-1 also indicates that the differences (between humans to animals) are less on a surface area basis ( . 3-fold) than on a lung volume basis ( . 14- fold). This is due to larger alveolar diameters and concomitant lower surface area per unit of lung volume in humans. Such differences in th e deposited dose in relevant target areas such as the alveolar region are important and have to be considered when extrapolating the results from DPM exposure studies in animals to humans. As will be discussed elsewhere in this document, procedures for dose extrapolation from animals to humans includes considering the process of clearance, with clearance measurements being in relation to surface area rather than to volume. Thus predicted doses of particulates w ould be based on surface areas, such as M 1 andM 2 in Table 3-1, rather than on volume, M.3-9 Table 3-1. Predicted doses of inhaled DPM per minute based on total lung volume (M), total airway surface area (M 1 ), or surface area in alveolar region (M 2 )

Species M (10 -3 : g/min/cm 3 ) M 1 (10 -6 : g/min/cm 2 ) M 2 (10 -6 : g/min/cm 2 )

Hamster 3.548 3.088 2.382

Fischer rat 3.434 3.463 2.608

Human 0.249 1.237 0.775

M = mass DPM deposited in lung per minute total lung volume

M 1 = mass DPM deposited in lung per minute total airway surface area

M 2 = 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 : m; geometric standard deviation ( F g ) = 1.9; packing density ( N ) = 0.3; and particle mass density ( D ) = 1.5 g/cm 2 ; (2) particle concentration = 1 mg/m 3 ; and (3) nose-breathing. For humans, total lung volume = 3200 cm 2 , total airway surface area = 633,000

cm 2 , surface area of the unciliated airways = 627,000 cm 2 . Corresponding values for Fisher rats are 418cm 2 , 412cm 2

, and 409cm 2 ; for hamsters, 282cm 2 , 262cm 2 , and 261cm 2 . Tidal volumes (in cm 2 ) and respiratory frequency (per min) used for humans were 500 and 14; for Fisher rats, 1.6 and 98; for hamsters, 67 and 1.0.

Source: Xu and Yu, 1987

.

Particle deposition will initiate particle redistribution processes (e.g., clearance mechanisms, phagocytosis) that transfer the par ticles to various subcompa rtments, including the alveolar macrophage pool, pulmonary interstitium, and lymph nodes. Over time, therefore, only small amounts of the original par ticle intake would be associated with the alveolar surface areas.

3.3.2. Particle Clearance and Translocation Mechanisms This section provides an overview of the m echanisms and pathways by which particles are cleared from the respiratory tract. The mechan isms of particle clearance as well as clearance routes from the various regions of the respiratory tract have been consider ed in the PM Criteria Document (U.S. EPA, 1996) and review ed by Schlesinger et al. (1997). Particles that deposit upon airway surfaces may be cleared from the respiratory tract completely, or be translocated to other sites within this syst em, by various regionally distinct processes. These clearance mechanisms can be cat egorized as either absorptive (i.e., dissolution) or nonabsorptive (i.e., transport of intact pa rticles) and may occur simultaneously or with3-10 Figure 3-4. Diagram of known and suspected clearance pathways for poorly soluble particles depositing in th e alveolar region. (Modifie d from Schlesinger, 1995). temporal variations. Particle solubility in terms of clearance refers to solubility within the respiratory tract fluids and cells. Thus, a poorly soluble particle is one whose rate of clearance by dissolution is insignificant compared to its rate of clearance as an intact particle (as is the case with DPM). The same clearance mechanisms act on different particles to different degrees, with their ultimate fate being a function of deposition site, physicochemical properties (including any toxicity), and sometime s deposited mass or number concentr ation. However, the duration of clearance for poorly soluble particles such as DP M as it exists between species, months for rats vs. years or even decades for humans, can make dissolution of DPM a significant contributor for humans (Kreyling, 1992). Figure 3-4 outlines many of the known a nd suspected clearance pathways for poorly soluble particles, such as DPM, that depos it in the alveolar region. Included are the representations of the translocation pathways from the alveolar epithelium through the insterstitium and on through the lymph nodes; this latter path will be refe rred to frequently later in this chapter.3-11 3.3.2.1. Extrathoracic Region The clearance of poorly soluble particles de posited in the nasal passages occurs via mucociliary transport, and the ge neral flow of mucus is backward s, i.e., towards the nasopharynx. Mucus flow in the most anterior portion of the nasal passages is fo rward, clearing deposited particles to the vestibular region where re moval is by sneezing, wiping, or blowing. Soluble material deposited on the nasal epithelium is accessi ble to underlying cells via diffusion through the mucus. Dissolved substances may be subsequently translocated into the bloodstream. The nasal passages have a rich vasculature, a nd uptake into the blood from this region may occur rapidly. Clearance of poorly soluble partic les deposited in the oral passa ges is by expectoration or by swallowing into the gastrointestinal tract.

3.3.2.2. Tracheobronchial Region The dynamic relationship between deposition and clearance is responsible for determining lung burden at any point in time. Clearance of poorly soluble particles from the TB region is mediated primarily by mucociliary transport, a more rapid process than those operating in alveolar regions. Mucociliary transport (often referred to as the mucociliary escalator) is accomplished by the rhythmic beating of cilia that line the respiratory tract from the trachea through the terminal bronchioles. This movement propels the mu cous layer containing deposited particles (or partic les within alveolar macrophages [AMs]) toward the larynx. Clearance rate by this system is determined primarily by the flow velocity of the mucus, which is greater in the proximal airways and decreases distally. These rates also exhibit interspecies and individual variability. Considerable species-dependent variability in tracheobronchial clearance has been reported, with dogs genera lly having faster clearance rates than guinea pigs, rats, or rabbits (Felicetti et al., 1981). The half-time (t 1/2 ) values for tracheobronchial clearance of relatively insoluble particles are usually on the order of hours, as compared to alveolar clearance, which is on the order of hundreds of days in humans and dogs. The clearance of particulate matter from the tracheobronchial region is generally recogn ized as being biphasic or multiphasic (Raabe, 1982). Some studies have shown that particles are cleared from large, intermediate, and small airways with t 1/2 of 0.5, 2.5, and 5 h, respectively. However, reports have indi cated that clearance from airways is biphasic and that the long-term component for humans may take much longer for a significant fraction of particle s deposited in this region, and may not be complete within 24 h as generally believed (Stahlhofen et al., 1990; ICRP, 1994). Although most of the particul ate matter will be cleared fro m the tracheobronchial region towards the larynx and ultimately swallowed, the contribution of this fraction relative to carcinogenic potential is unclear. With the exception of conditions of impaired bronchial3-12 Figure 3-5. Modeled cleara nce of poorly soluble 4- 

### : m particles deposited in tracheobronchial and alve olar regions in humans. clearance, the desorption t 1/2 for particle-associated organics is generally longer than the tracheobronchial clearance times, thereby making uncertain the importance of this fraction relative to toxicity in the respiratory tract (Pepelko, 1987). However, Gerde et al. (1991a) showed that for low-dose exposures, particle-associat ed PAHs were released rapidly at the site of deposition indicating that they would be availa ble for involvement in postulated carcinogenic processes. The relationship between the early clearance of poorly soluble particles of 4 : m aerodynamic diameter from the tracheobronchial regions and their longer-term clearance from the alveolar region is illustrated in Figure 3-5, clearly showing the rapid depuration from the TB region compared with the A region. This relationship, although demonstrated with 4 : m particles, is probabl y relevant and applicable to DPM-sized particles (i.e., 0.2 : m) as clearance
mechanisms are believed not to be particularly par ticle-sized dependent (Morrow et al., 1967a,b; Snipes et al., 1983). Cuddihy and Yeh (1986) reviewed respirator y tract clearance of particles inhaled by humans. Depending on the type of particle (ferric oxide, Teflon di scs, or albumin microspheres), the technique employed, and the anatomic region ( midtrachea, trachea, or ma in bronchi), particle3-13 Figure 3-6. 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). velocity (moved by mucocilia ry transport) ranged from 2.4 to 21.5 mm/min. The highest velocities were recorded for midtracheal transport, and the lowest were for main bronchi. Cuddihy and Yeh (1986) descri bed salient points to be considered when estimating particle clearance velocities from tracheobronchial regions: these include respiratory tract airway dimensions, calculated inhaled pa rticle deposition fractions for i ndividual airways, and thoracic (A + TB) clearance measurements. Predicted clearance velocities for the trachea and main
bronchi were found to be similar to those experimentally de termined for inhaled radiolabeled particles, but not those for intratracheally instilled particles. The velocities observed for inhalation studies were generally lower than those of instillation studies. Figure 3-6 illustrates a comparison of the short-term clearance of inha led particles by human subjects and the model predictions for this clearance. However, tracheobronchial clearance via the mucociliary escalator is of limited importance for long-term clearance.3-14 Exposure of F344 rats to whole exhaust cont aining DPM at concentrations of 0.35, 3.5, or 7.1 mg/m 3 for up to 24 mo did not significantly alter tracheal mucociliary clearance as assessed by clearance of 99m Tc-macroaggregated albumin instilled into the trachea (Wolff et al., 1987). The authors stated that measuring retention would yield estimates of clearance efficiency comparable to measuring the velocity for transport of the markers in the trachea. The results of this study were in agreement with similar findings of unaltered tracheal mucociliary clearance in rats exposed to DPM (0.21, 1.0, or 4.4 mg/m 3 ) for up to 4 mo (Wolff and Gray, 1980). However, the 1980 study by Wolff and Gray, as well as an earlier study by Battigelli et al. (1966), showed that acute exposure to high concentr ations of DE soot (1.0 and 4.4 mg/m 3 in the study by Wolff and Gray [1980] and 8 to 17 mg/m 3 in the study by Battigelli et al. [1966]) produced transient reductions in tracheal mucociliary clearance. Battigelli et al. (1966) also noted that the compromised tracheal clearance was not observe d following cessation of exhaust exposure. That tracheal clearance does not appear to be significantly impaired or is impaired only transiently following exposure to high concentra tions of DPM is consistent with the absence of pathological effects in the tr acheobronchial region of the resp iratory tract in experimental animals exposed to DPM. The apparent re tention of a fraction of the deposited dose in the airways could be cause for some concern rega rding possible effects in this region, especially in light of the results from simulation studies by Gerde et al. (1991b) suggesting that release of PAHs from particles may occur within minutes and therefore at the s ite of initial deposition. However, the absence of effects in the TB ar eas in long-term DPM st udies and experimental evidence that particle-associated PAHs are releas ed at the site of pa rticle deposition together suggest that these PAHs and othe r organics may be of lesser im portance in tumorigenic responses of rats than originally suspected. On the othe r hand, the data of Nikula et al. (1997a,b) could be interpreted to suggest that a larger fraction of particles are translocated to the interstitium of the respiratory tract in primates that are heavily exposed (and therefore presumably in humans) than in rats that are heavily exposed, including the interstitium of the respiratory bronchioles, an anatomical site absent in rats (Section 3.6). Mo reover, eluted PAHs in the TB region are retained longer than those in the alveoli (Gerde et al., 1999), allowing time for activation. Also, the results of Kreyling (1992) indicat e that appreciable dissolution of even poorly soluble particles may occur as a consequence of long absolute dur ation of clearance, such as years or decades, in humans. Thus PAHs may have a role in human re sponse to DE that cannot be evaluated with the
rat model. Also, impairment of mucociliary clearance function as a result of exposure to occupational or environmental respiratory tract toxicants or to cigarette smoke may significantly enhance the retention of par ticles in the TB region. For example, Vastag et al. (1986) demonstrated that not only smokers with clinical symptoms of bronchitis but also symptom-free3-15 smokers have significantly reduced mucoc iliary clearance rates. Although impaired
tracheobronchial clearance could conceivably have an impact on the effects of deposited DPM in the conducting airways, it does not appear to be relevant to the epigen etic mechanism likely responsible for DE-induced rat pulmonary tumors as the tumors observed in these studies were all or nearly all of A vice TB origin. Poorly soluble particles such as DPM that are deposited within the TB region are cleared predominantly by mucociliary transport towa rds the oropharynx, followed by swallowing. Poorly soluble particles may al so be cleared by traversingthe epithelium by endocytotic processes, and enter the peri bronchial region. Clearance may occur following phagocytosis by airway macrophages, located on or beneath the mucous lining throughout the bronchial tree, or via macrophages that enter the airway lume n from the bronchial or bronchiolar mucosa (Robertson, 1980).

3.3.2.3. A Region A number of investigators ha ve reported on the alveolar clearance kinetics of human subjects. Bohning et al. (1980) examined alve olar clearance in eight humans who had inhaled <0.4 mg of 85 Sr-labeled polystyrene particles (3.6 ± 1.6 : m diam.). A double-exponential model
best described the clearance of the particles and provided t 1/2 values of 29 ± 19 days and 298 ± 114 days for short-term and long-t erm phases, respectively. It wa s noted that of the particles deposited in the alveolar region, 75% ± 13% were cleared via th e long-term phase. Alveolar
retention t 1/2 values of 330 and 420 days were reported for humans who had inhaled aluminosilicate particles of MMAD 1.9 and 6.1 : m (Bailey et al., 1982). In a comprehensive study Bailey et al. (1985) followed the long-ter m retention of inhale d particles in a human respiratory tract. The retention of 1 and 4 : m fused aluminosilicate particles labeled with strontium-85 and yttrium-88, resp ectively, was followed in male volunteers for about 533 days. Approximately 7% of the initial lung deposit of 1 : m particles and 40% of the 4 : m particles were associated with a rapi d clearance phase corresponding to the calculated tracheobronchial deposits. Retention of the re maining material followed a tw o-component exponential function, with phases having half- times of the order of tens of days and several hundr ed days, respectively. Quantitative data on clearance rates in humans having large lung burdens of particulate matter are lacking. Bohning et al. (1982) and Cohen et al. (1979), however, did provide evidence for slower clearance in smokers, and Freedman and Robinson (1988) reported slower clearance rates in coal miners who had mild pneumoconiosis with presumably high lung burdens of coal dust. Although information on pa rticle burden and par ticle overload relati onships in humans is much more limited than in experimental animal models, inhibition of clearance does seem to occur. Stöber et al. ( 1967) estimated a clearance t 1/2 of 4.9 years in coal miners with nil or slight3-16 silicosis, based on postmortem lung burdens. The lung burdens and estimated exposure histories ranged from 2 to 50 mg/g of lung or more, well above the value at which clearance impairment is observed in the rat. Furthermore, impaired clearance resulting from smoking or exposure to other respiratory toxicants may increase the po ssibility of an enhanced particle accumulation effect resulting from exposure to other particle sources such as DPM. Normal alveolar clearance rates in laboratory animals exposed to DPM have been reported by a number of investigators (Table 3-2). Because the rat is, historically, the species for which experimentally induced lung cancer data are available and for which most clearance data exist, it is the species most often used for assessi ng human risk, and review s of alveolar clearance studies have been generally limited to this species. Chan et al. (1981) subjected 24 male F344 rats to nose-only inhalation of diluted DE generated from a diesel engine (6 mg/m 3 ) labeled with 131 Ba or 14 C for 40 to 45 min and assessed total lung deposition, retention, and elimination. Based on radiol abel inventory, the deposition efficiency in the respiratory tract was 15% to 17%. Measurement of 131 Ba label in the feces during the first 4 days following exposure i ndicated that 40% of the deposited DPM was eliminated via mucociliary clearance. Clearance of the particles from the lower respiratory tract followed a two-phase elimination process consisting of a rapid (t 1⁄2 of 1 day) elimination by mucociliary transport and a slower (t 1⁄2 of 62 days) macrophage-med iated alveolar clearance. This study provided data for nor mal alveolar clearance rates of DPM not affected by prolonged exposure or particle overloading.
Several studies have investigated the effects of expos ure concentration on the alveolar clearance of DPM by laboratory animals. Wol ff et al. (1986, 1987) provided clearance data (t 1⁄2 ) and lung burden values for F344 rats exposed to DE for 7 h/day, 5 days/week for 24 mo. Exposure concentrations of 0.35, 3.5, and 7.1 mg of DPM/m 3 were employed in this whole bodyinhalation exposure experiment. Inte rmediate (hours-days) clearance of 67 Ga 2 O 3 particles (30 min, nose-only inhalation) was a ssessed after 6, 12, 18, and 24 mo of exposure at all of the DPM concentrations. A two-compone nt function described the cl earance of the administered
radiolabel:

(3-1) 

### FA B

(t) 1 2 

### exp(-0.693 t / exp(-0.693 t/ = + τ τ ) ),

where F (t) was the percentage retained throughout the respiratory tract, A and B were the magnitudes of the two components (component A included nasal, lung, and gastrointestinal clearance, while component B represented intermediate lung clearance) and J 1 and J 2 were theTable 3-2. Alveolar clearance in laboratoryanimals exposed to DPM in whole exhaustSpecies/sexExposuretechniqueExposuredurationParticlesmg/m3Observed effects ReferenceRats, F-344, M Nose only;Radiolabeled DPM40-45 min 6 Four days after exposure, 40% of DPM eliminated bymucociliary clearance. Clearance from lower RT was in2 phases. Rapid mucociliary (t1/2= 1 day); slowermacrophage-mediated (t1/2= 62 days).Chan et al. (1981)Rats, F-344 Whole body;assessed effecton clearance of67Ga2O3particles7 h/day5 days/week24 mo0.353.57.1J1significantly higher with exposure to 7.1 mg/m3for24 mo;J2significantly longer after exposure to 7.1 mg/m3for 6 mo and to 3.5 mg/m3for 18 mo.Wolff et al. (1986,1987)Rats Whole body 19 h/day5 days/week2.5 years4 Estimated alveolar deposition = 60 mg; particle burdencaused lung overload. Estimated 6-15 mg particle-boundorganics deposited.Heinrich et al.(1986)Rats, F-344, MF Whole body 7 h/day5 days/week18 mo0.150.944.1Long-term clearance was 87 ± 28 and 99 ± 8 days for0.15 and 0.94 mg/m3groups, respectively; t1/2= 165 daysfor 4.1 mg/m3group.Griffis et al. (1983)Rats, F-344;Guinea pigs,HartleyNose-only;Radiolabeled14C45 min140 min45 min727Rats demonstrated 3 phases of clearance with t1/2= 1, 6,and 80 days, representing tracheobronchial, respiratorybronchioles, and alveolar clearance, respectively. Guineapigs demonstrated negligible alveolar clearance fromday 10 to 432.Lee et al. (1983)Rats, F-344 20 h/day7 days/week7-112 days0.256Monitored rats for a year. Proposed two clearance models.Clearance depends on initial particle burden; t1/2increaseswith higher exposure. Increases in t1/2indicate increasingimpairment of AM mobility and transition into overloadcondition.Chan et al. (1984)RT = respiratory tract.AM = alveolar macrophage.J1= clearance from primary, ciliated airways.J2= clearance from nonciliated passages.

3-173-18 Figure 3-7. Clearance from lungs of rats of 134 Cs-FAP fused aluminosilicate tracer particles inhaled after 24 months of DE exposure at concentrations of 0 (control), 0.35 (low), 3.5 (med ium), and 7.1 (high) mg DPM/m 3

. half-times for the A and B components, respectively. Th e early clearance half-times ( J 1 ), were similar for rats in all exposure groups at all time points except in the high-exposure (7.1 mg/m 3 ) group following 24 mo of exposure, which was fast er than the controls. Significantly longer B component retention ha lf-times, representing intermedia te clearance probably from nonciliated structures such as alveolar ducts and alveoli, were noted after as little as 6 mo exposure to DPM at 7.1 mg/m 3 and 18 mo exposure to 3.5 mg/m 3

. Nose-only exposures to 134 Cs fused aluminosilicate particle s (FAP) were used to assess long-term (weeks-months) clearance. Following 24-mo exposure to DPM, long-term clearance
of 134 Cs-FAP was significantly ( p <0.01) altered in the 3.5 (cum ulative exposure [C × T] of 11,760 mg
@ h/m 3 ) and 7.1 mg/m 3 , C × T = 23,520 mg
@ h/m 3 ) exposure groups (t 1⁄2 of 264 and 240 days, respectively) relative to the 0.35 mg/m 3 and control groups (t 1⁄2 of 81 and 79 days, respectively). Long-term clearance represents th e slow component of pa rticle removal from the alveoli. The decreased clearance correlated with the greater particle burden in the lungs of the 3.5 and 7.1 mg/m 3 exposure groups. Based on these findings, the cumulative exposure of > 11,760 mg
@ h/m 3 (or 3.5 mg/m 3 for a lifetime exposure) represented a particle overload condition resulting in compromised alveolar clearance mechanisms; the clearance rate at the lowest concentration (0.35 mg/m 3 ; cumulative exposure of 118 mg
@ h/m 3 ) was not different from control rates (Figure 3-7).3-19 Heinrich et al. (1986) exposed rats 19 h/da y, 5 days/week for 2.5 years to DPM at a particle concentra tion of about 4 mg/m 3 , equal to a “C × T” of 53,200 mg
@ h/m 3 . The deposition in the alveolar region was estimated to equal 60 mg. The lung particle burden was apparently sufficient to result in a “particle overload” condition (Section 3.4). With respect to the organic matter adsorbed onto the particle s, the authors estimated that ove r the 2.5-year period, 6-15 mg of particle-bound organic ma tter had been deposited and was pot entially available for biological effects. This estimation was based on the analysis of the DE used in the experiments, values for rat ventilatory functions, and estimates of deposition and clearance. Accumulated burden of DPM in the lungs following an 18-mo, 7 h/day, 5 days/week exposure to whole DE was reported by Griffis et al. (1983). Male and female F344 rats exposed to 0.15, 0.94, or 4.1 mg DPM/m 3 were sacrificed at 1 day and 1, 5, 15, 33, and 52 weeks after exposure, and DPM was extracted from lung tissue dissolved in tetramethylammonium hydroxide. Following centrifugation and washing of the supernatant, DPM content of the tissue was quantitated using spectrophotomet ric techniques. The analyti cal procedure was verified by comparing results to recovery studies usin g known amounts of DPM with lungs of unexposed rats. Lung burdens were 0.035, 0.220, and 1.890 mg/g lung tissue, respectively, in rats exposed to diluted whole exhaust at 0.15, 0.94, and 4.1 mg DPM/m 3 . Long-term retention for the 0.15 and 0.94 mg/m 3 groups had estimated half- times of 87 ± 28 and 99 ± 8 da ys, respectively. The
retention t 1/2 for the 4.1-mg/m 3 exposure group was 165 ± 8 days, which was significantly ( p <0.0001) greater than those of the lower ex posure groups. The 18-mo exposures to 0.15 or 0.96 mg/m 3 levels of DPM [C × T] equivalent of 378 and 2,368 mg
@ h/m 3 , respectively) did not affect clearance rates, whereas the exposure to the 4.1 mg/m 3 concentration C × T = 10,332

mg @ h/m 3 ) resulted in impaired clearance. Lee et al. (1983) described the clearance of DPM (7 mg/m 3 for 45 min or 2 mg/m 3 for 140 min) by F344 rats (24 per group) and Hartley guinea pigs expos ed by nose-only inhalation to diluted whole exhaust with no appa rent particle overloa d in the lungs as being in three distinct phases. The exposure protocols provide d comparable total doses based on a 14 C radiolabel.

14 CO 2 resulting from combustion of 14 C-labeled diesel fuel was re moved by a diffusion scrubber
to avoid erroneous assessment of 14 C intake by the animals. Re tention of the radiolabeled particles was deter mined up to 335 days after exposure a nd resulted in a three-phase clearance
with retention t 1/2 values of 1, 6, and 80 days. The three clearance phases are taken to represent removal of tracheobronchial deposits by the mucociliary escalator, removal of particles deposited in the respiratory bronchioles, and alveolar clearance, respectively. Species variability in clearance of DPM was also demonstrated because the Hartley guinea pigs exhibited negligible alveolar clearance from day 10 to day 432 followi ng a 45-min exposure to a DPM concentration3-20 of 7 mg/m 3 . Initial deposition efficiency (20% ± 2%) and short-term clearance were, however,
similar to those for rats. Lung clearance in male F344 rats preexposed to diluted whole DE containing DPM at 0.25 or 6 mg/m 3 20 h/day, 7 days/week for periods la sting from 7 to 112 days was studied by Chan et al. (1984). Following this preexposure protocol, rats were subjected to 45-min noseonly exposure to 14 C-DE, and alveolar clearance of radiol abel was monitored for up to 1 year. Two models were proposed: a normal biphasic clearance model a nd a modified lung retention
model that included a slow-clear ing residual component to account for sequestered aggregates of macrophages. The first model described a first- order clearance for two compartments: R(t) =
Ae -u1t + Be -u2t . This yielded clearance t 1/2 values of 166 and 562 days for rats preexposed to 6.0 mg/m 3 for 7 and 62 days, respectively. These values were significantly ( p <0.05) greater than the
retention t 1/2 of 77 ± 17 days for control rats. The sa me retention values for rats of the 0.25 mg/m 3 groups were 90 ± 14 and 92 ± 15 days, respec tively, for 52- and 112-day exposures and were not significantly different from controls. The two-compartment m odel represents overall clearance of the tracer particles, even if some of the particles were sequestered in particle-laden macrophages with substantially slower clearance rates. For the second model, which excluded transport of the residual fracti ons in sequestered macrophage a ggregates, slower clearance was observed in the group with a lung bur den of 6.5 mg (exposed to 6.0 mg/m 3 for 62 days), and no clearance was observed in the 11.8 mg group (exposed to 6.0 mg/m 3 for 112 days). Clearance was shown to be dependent on the initial burden of particles, and therefore the clearance t 1/2 would increase in higher exposure scenarios. This study emphasizes th e importance of particle overloading of the lung and the ramifications on clearance of particles; the significant increases
in half-times indicate an incr easing impairment of the alveolar macrophage mobility and subsequent transition into an overload condition as is disc ussed further in Section 3.4. Long-term alveolar clearance rates of partic les in various laborat ory animals and humans have been reviewed by Pepel ko (1987). Although retention t 1/2 varies both among and within species and is also dependent on the physicoche mical properties of the inhaled particles, the
retention t 1/2 for humans is much longer (>8 mo) than the average retention t 1/2 of 60 days for rats. Clearance from the A region occurs via a numbe r of mechanisms and pathways, but the relative importance of each is not always certain and may vary between species. Particle removal by macrophages comprises the main nonabsorptiv e clearance process in this region. Alveolar macrophages reside on the epithelium, where they phagocytize and trans port deposited material, which they contact by random motion or via directed migration under the influence of local chemotactic factors (Warheit et al., 1988). Particle-laden macrophages may be clear ed from the A region along a number of pathways (U.S. EPA, 1996). Uningested par ticles or macrophages in the interstitium may3-21 traverse the alveolar-capillary endothelium, directly enteri ng the blood (Raabe, 1982; Holt, 1981); endocytosis by e ndothelial cells followed by exocytos is into the vessel lumen seems, however, to be restrict ed to particles <0.1 : m diameter, and may increase with increasing lung burden (Lee et al., 1985; Oberdörster, 1988). Once in the systemic circ ulation, transmigrated macrophages, as well as uninges ted particles, can travel to extrapulmonary organs. Alveolar macrophages constitute an important first-line cellular defense mechanism against inhaled particles th at deposit in the alveolar region of th e lung. It is well established that a host of diverse materials, including DPM, are phagocytized by AMs shortly after deposition (White and Garg, 1981; Lehnert and Morrow, 1985) and that such cell-c ontained particles are generally rapidly sequestered from both the extrace llular fluid lining in the alveolar region and the potentially sensitive alveolar epithelial cells. In addition to this role in compartmentalizing particles from other lung constituents, AMs are prominently i nvolved in mediating the clearance of relatively insoluble particle s from the air spaces (Lehnert and Morrow, 1985). Although the details of the actual process ha ve not been delineated, AMs with their particle burdens gain access and become coupled to the mucociliary es calator and are subsequently transported from the lung via the conducting airways. Althoug h circumstantial, numerous lines of evidence indicate that such AM-mediated particle clearance is the predominant mechanism by which relatively insoluble particles are removed from the alveolar region of the lungs (Gibb and Morrow, 1962; Ferin, 1982; Harmsen et al., 1985; Lehnert and Morrow, 1985; Powdrill et al., 1989). The removal characteristics for particles deposite d in the alveolar region of the lung have been descriptively represented by numerous investigators as a multicompartment or multicomponent process in which each component follows simple first- order kinetics (Snipes and Clem, 1981; Snipes et al., 1988; Lee et al., 1983). Although the various compartments can be described mathematically, the actual physiological mechanisms determining these differing
clearance rates have not been well characterized. Lehnert et al. (1988, 1989) performed studi es using laboratory rats to examine particle-AM relations hips over the course of alveolar clear ance of low to high lung burdens of noncytotoxic microspheres (2.13 : m diam.) to obtain inform ation on potential AM-related mechanisms that form the underlying bases for ki netic patterns of alveolar clearance as a function of particle lung burde ns. The intratracheally instilled lung burdens varied from 1.6 × 10 7 particles (about 85 : g) for the low lung burden to 2.0 × 10 8 particles (about 1.06 mg) for the middose and 6.8 × 10 8 particles (about 3.6 mg) for the highest lung burden. The lungs were lavaged at various times postexposure and the numbers of spheres in each macrophage counted. Although such experiments provide information regarding the respons e of the lung to particulate matter, intratracheal instillation is not likely to result in the same depositional characteristics as3-22 inhalation of partic les. Therefore, it is unlikely that the response of alveolar macrophages to these different depositional characteristics will be quantitatively similar.
The t 1/2 values of both the early and later com ponents of clearance were virtually identical following deposition of the low and medium l ung burdens. For the highest lung burden, significant prolongations were found in both the early, more rapid, as well as the slower component of alveolar cl earance. The percentages of the pa rticle burden associated with the earlier and later components, howev er, were similar to those of th e lesser lung burdens. On the basis of the data, the authors concluded that tr anslocation of AMs from alveolar spaces by way of the conducting airways is fundame ntally influenced by the pa rticle burden of the cells so translocated. In the case of particle overloa d that occurred at the highest lung burden, the
translocation of AMs with the heaviest cellula r burdens of particles (i.e., greater than about 100 microspheres per AM) was definitely compromised. On the other hand, analysis of the disa ppearance of AMs with various numbers of particles indicates that the particles may not exclusively reflect the translocation of AMs from the lung. The observations are also consistent with a gradual redistribution of retained particles among the AMs in the lung concurrent with th e removal of particle-c ontaining AMs via the conducting airways. Experimental support sugges tive of potential proce sses for such particle redistribution comes from a variety of investig ations involving AMs a nd other endocytic cells (Heppleston and Young, 1973; Evans et al., 1986; Aronson, 1963; Sandusky et al., 1977; Heppleston, 1961; Riley and Dean, 1978).

3.3.3. Translocations of Particles to Extr a-Alveolar Macrophage Compartment Sites Although the phagocytosis of particles by ce lls free within the lung and the mucociliary clearance of the cells with their particulate matter burdens represent the most prominent mechanisms that govern the fate of particles deposite d in the alveolar region, other mechanisms exist that can affect both the retention characteristics of relatively insoluble particles in the lung and the lung clearance pathways for the particles. One mechanism is endoc ytosis of particles by alveolar lining (Type I) cells (Sorokin a nd Brain, 1975; Adamson and Bowden, 1978, 1981) that normally provide >90% of the cell surface of the alveoli in the lungs of a variety of mammalian species (Crapo et al., 1983). This process may be re lated to the size of th e particles that deposit in the lungs and the numbers of particles that are deposited. Adamson and Bowden (1981) found that with increasing load s of carbon particles (0.03 : m diam.) instilled in the lungs of mice, more free particles were observed in the alveoli within a few days; it should be noted, however, that this phenomenon was demonstrated with very high doses given as a bolus such that the
mechanism and relevance of this phenomenon at lower concentrati ons may be different or even unrelated to what may happen at much lower con centrations. The relativ e abundance of particles3-23 endocytosed by Type I cells also increased with increasing lung burdens of the particles, but instillation of larg e particles (1.0 : m) rarely resulted in their undergoing endocytosis. A 4 mg
burden of 0.1 : m diameter latex particles is equivalent to 8 × 10 12 particles, whereas a 4 mg
burden of 1.0 : m particles is composed of 8 × 10 9 particles. Regard less, DPM with volume
median diameters between 0.05 and 0.3 : m (Frey and Corn, 1967; Kittleson et al., 1978) would be expected to be within the size range for en gulfment by Type I cells should suitable encounters occur. Indeed, it has been demonstrated that DP M is endocytosed by Type I cells in vivo (White and Garg, 1981). Unfortunately, information on th e kinetics of partic le engulfment (endocytosis) by Type I cells relative to that by AMs is scanty. Even when relatively low burdens of particulate matter are deposited in the lungs, some fraction of the particles usually appears in the regional lymph nodes (Ferin and Feldstein, 1978; Lehnert, 1989). As will be discussed, endoc ytosis of particles by Type I cells is an initial, early step in the passage of partic les to the lymph nodes. Assuming particle phagocytosis is not su fficiently rapid or perfectly e fficient, increasing numbers of particles would be expected to gain entry into the Type I epithelial cell compartment during chronic aerosol exposures. Additionally, if par ticles are released on a continual basis by AMs that initially sequestered them after lung de position, some fraction of the “free” particles so released could also undergo passage from the alveolar space into Type I cells. The endocytosis of par ticles by Type I cells represents only the initial stage of a process that can lead to the accumulation of particles in the lung’s interstitial compartment and the subsequent translocation of particles to the regi onal lymph nodes. As suggested by the results of Adamson and Bowden (1981), a vesicular transpor t mechanism in the Type I cell can transfer particles administered at high concentrations by instillation fro m the air surface of the alveolar epithelium into the lung’s interstitium, where particles may be phagocytized by interstitial macrophages or remain in a “free” state for a poor ly defined period that may be dependent on the physicochemical characteristics of the particle. The lung’s in terstitial compartment accordingly represents an anatomical site for the retention of pa rticles in the lung, although the kinetics on
movement into and out of this site remain obscure for both human s and test species. Whether or not AMs, and perhaps polymorphonuc lear neutrophils (PMNs) that have gained access to the alveolar space compartment and phagocytize particles there, also contribute to the particle translocation process into the lung’s interstitium also remains a controversial issue. Translocation of particulate ma tter to the various interstitial spaces within the lung is a prominent phenomenon occurring at least at hi gh (occupational) exposures that has been examined extensively for both DPM and coal dust in a species comparison between rats and primates (Nikula et al., 1997a,b). Detailed pulmonary morphomet ry conducted on F344 rats and cynomolgus monkeys that had been exposed for 24 months to occupationa l levels of DPM (1.953-24 mg/m 3 ; see Lewis et al., 1989) showed major differenc es in the pulmonary sites of particulate deposition. In rats, about 73% of DPM was present in the al veolar ducts/alveoli and 27% in interstitial compartments; for monkeys the corres ponding figures were markedly different at 43% and 57%. The corresponding pulmonary histopa thology confirmed that both species were affected, although rats are more sensitive, as incidence and sever ity scores for alveolar effects ranged from 15 of 15 with sever ity scores from 1-4 (minimal to moderate), whereas for monkeys the corresponding values were only 4 of 15 at a range of 0-2 (not observed to minimal). Similarly, both species exhibited histopathology at the interstitial sites of deposition but with effects in monkeys being slightly more severe (1 of 15 graded as slight, 14 of 15 graded as minimal) than those in rats (14 of 15 graded as slight, 1 of 15 graded as minimal). The basis for this interspecies difference may be due to any number of clear contrasts that exist between rat and primate lungs, including anatomical (pr imates and humans have respiratory bronchioles whereas rats do not), kinetic (primates and human clearance processes allow more residence time of particles in the lung than do those in rats or rats may have faster interstitial to lymph node clearance rates than do humans and primates), or morphological (pr imates and humans have more interstitial tissue, more a nd thicker pleura, and wider in terstitial spaces than do rats). Aspects of the study itself that may obscure its interpretation include the relative lifespan the exposure represented between the tested species (lifetime for rat vs . about 10% lifetime of primate), that there was only the single time point at which the relative burdens were determined, and that rat lymph node burdens were not included in the analys is. The analysis of Kuempel (2000) using human occupational da ta clearly showed that models require an interstitialization process to provide adequa te fits to the empirical human (min ers’) lung deposition data discussed in that study. Hypotheses about possible mechan isms for the interstitialization process are scant, although Harmsen et al. (1985) provided some evidence in dog s that migration of AMs may contribute to the passage of par ticles to the interstitial compartme nt and also may be involved in the subsequent translocation of particles to draining lymph nodes. Translocation to the extrapulmonary regional lymph node s apparently can involve the pa ssage of free particles as well as particle-containing cells via lymphatic channels in the lungs (Harmsen et al., 1985; Ferin and Feldstein, 1978; Lee et al., 1985). Further, it has been noted that particles accumulate both more rapidly and more abundantly in lymph nodes that receive lymphatic drainage from the lung (Ferin and Feldstein, 1978; Lee et al., 1985). It should be stressed that further i nvestigation is required to confirm the character and even existence of the interstitialization process in the lungs of humans with exposures to particle s at lower environmental concen trations, or to submicrometer particles such as DPM, or to examine the kinetics and time course of the interstitialization

process.3-25 3.3.3.1. Clearance Kinetics
The clearance kinetics of PM have been revi ewed in the PM CD (U.S. EPA, 1996) and by Schlesinger et al. (1997), the results of which indicate that cl earance kinetics may be profoundly influenced by several factors. The influence of time, for example, is definitively showed by the work of Bailey et al. (1985; di scussed above), who showed that the rate of clearance from the pulmonary region to the GI tract decreased nearly fourfold from initial values to those noted at 200 days and beyond afte r particle inhalation.

3.3.3.2. Interspecies Patterns of Clearance The inability to study the retention of certain materials in humans for direct risk assessment requires the use of laboratory an imals. Adequate toxicological assessment necessitates that interspecies comparisons cons ider aspects of dosimetry including knowledge of clearance rates and routes. The basic mechanisms and overall patterns of clearance from the respiratory tract are similar in humans and most other mammals. Regional clearance rates, however, can show substantial variation between species, even for similar particles deposited under comparable exposure conditions (U.S. EPA, 1996; Schlesinger et al., 1997; Snipes et al., 1989). In general, there are species-dependent rate constants for various clearance pathways. Differences in regional and total clearance rates between some species are a reflection of differences in mechanical clearance processes. For consideration in asse ssing particle dosimetry, the end result of interspecies differences in cl earance is that the retained doses in the lower respiratory tract can differ between species, which may result in di fferences in response to similar particulate exposures.

3.3.3.3. Clearance Modifying Factors and Susceptible Populations
A number of host and environmental factor s may modify clearance kinetics and may consequently make individuals exhibiting or afflicted with these factors particularly susceptible to the effects resulting from exposure to DPM. These include age, gender, physical activity, respiratory tract disease, and inhalation of irritants (U.S. EPA, 1996, Section 10.4.2.5). Respiratory tract clearance appears to be prolong ed in a number of pathophysiological conditions in humans, including chronic sinusitis, chronic bronchitis, asthma, chronic obstructive lung disease, and various acute respiratory infections.

3.3.3.4. Respiratory Tract Disease Earlier studies reviewed in the PM CD (U.S . EPA, 1996) noted that various respiratory
tract diseases are associated with alterations in overall clearance and clear ance rates. Prolonged3-26 nasal mucociliary clearance in humans is associated with chronic sinusitis or rhinitis, and cystic fibrosis. Bronchial mucus transport may be impaired in people with bronchial carcinoma, chronic bronchitis, asthma, and various acute infections. In certain of these cases, coughing may enhance mucus clearance, but it generally is e ffective only if excess secretions are present. The rates of A region particle clearance are reduced in humans with chronic obstructive lung disease and in laboratory animals with viral infections, whereas the viability and functional activity of macrophages are impaired in human asth matics and in animals w ith viral-induced lung infections (U.S. EPA, 1996). However, an y modification of f unctional properties of macrophages appears to be injury specific, reflec ting the nature and anato mic pattern of disease.

3.4. PARTICLE “OVERLOAD”
3.4.1. Introduction Some experimental studies using labor atory rodents employed high exposure concentrations of relatively nontox ic, poorly soluble particles. These particle loads interfered with normal clearance mechanisms, producing clear ance rates different from those that would occur at lower exposure levels. Prolonged exposure to high particle concen trations is associated with what is termed particle overload. This is defined as the overwhelming of macrophagemediated clearance by the deposition of particles at a rate exceeding the capacity of that clearance pathway. Aspects and occurrence of th is phenomenon have alread y been alluded to in earlier portions of this chapter on alveolar clearance (Section 3.3.2.3). The relevance of this phenomenon for human risk assessment has long been the object of scientific inquiry. A monograph on this matter and many others relevant to DPM has appeared (ILSI, 2000), and the results, opinions, and judgments put forth therein ar e used extensively in this chapter and in this

assessment. Wolff et al. (1987) used 134 Cs-labeled fused aluminosilicate particles to measure alveolar clearance in rats following 24-mo exposure to low, medium, and high concentrations of DE (targeted concentrations of DPM of 0.35, 3.5 and 7.1 mg/m 3 ). The short-term component of the multicomponent clearance curves was similar fo r all groups, but long-term clearance was retarded in the medium- and high-exposure groups (Figure 3-7). The half times of the long-term clearance curves were 79, 81, 264, and 240 days, resp ectively, for the control, low-, medium-, and high-exposure groups. Clearance was overloaded at the high and medium but not at the low exposure level. Lung burdens of DPM were measured after 6, 12, 18, and 24 mo of exposure. The results (Figure 3-8) indicate that the l ung burden of deposited pa rticles was appreciably increased or “overloaded” compar ed with the low level of exposure in the two highest exposures post 6 months. Figure 3-8 also compares th ese observational results of lung burden with
simulated results where no overload w ould occur (McClellan, 2000). Comparison3-27 Figure 3-8. Lung burdens (in mg DPM soot/g l ung) in rats chronically exposed to DE at 0.35 (low) ( 

###  ), 3.5 (medium) ( 

### • ), and 7.1 (high) mg / m 3 ( 

###  ). The solid figures represent actual data with means and standard errors from animals sacrificed at 6, 12, and 18 months after initiation of exposures . Lines are simulated model results from these same exposure levels, assuming no effect of exposure concentration on deposition or clearance of particles (fro m Wolff et al., 1987; McClellan, 2000).

of the observed and simula ted results clearly shows that the two highest exposure levels resulted in lung burdens that were eve r-increasing and not at all concor dant with the simulated results, whereas the burdens at the low-exposure level were closely approxima ted by the simulation. Thus, at the two highest exposure levels, deposition processes were outpacing clearance mechanisms. Results from the low-exposure level indicate that clearance processes were not inhibited, the lung burden remaining the sa me throughout all time periods examined. Morrow (1988) has proposed that the condition of particle overloadi ng in the lungs is caused by a loss in the mobility of particle-engorged AMs and that such an impediment is related to the cumulative volumetric load of particles in the AMs. Morrow (1988) has further estimated that the clearance function of an AM may be comp letely impaired when the particle burden in the AM is of a volumetric size equi valent to about 60% of the nor mal volume of the AM. Morrow’s3-28 hypothesis was the initial basis for the physiology-oriented multicompartmental kinetic (POCK) model derived by Stöber et al. (1989) for estimating alveolar clear ance and retention of relatively insoluble, respirable particles in rats.
A revised version of this m odel refines the characteriza tion of the macrophage pool by including both the mobile and immobilized macr ophages (Stöber et al., 1994). Application of
the revised version of the model to experimental data suggested that lung overload does not cause
a dramatic increase in the total burden of the macrophage pool but results in a great increase in the particle burden of the inte rstitial space, a compartment that is not available for macrophagemediated clearance. The revised version of the PO CK model is discussed in greater detail in the
context of other dos imetry models below.
Oberdörster and co-workers (1992) assessed the alveolar clearance of smaller (3.3 : m diam.) and larger (10.3 : m diam.) polystyrene pa rticles, the latter of which are volumetrically equivalent to about 60% of the average normal volume of a rat AM , after intratracheal instillation into the lungs of rats. Even though both sizes of particles were found to be phagocytized by AMs within a day after deposition, a nd the smaller partic les were cleared at a normal rate, only minimal lung clearance of the larger particle s was observed over an approximately 200-day postinstillation period, thus supporting th e volumetric AM ove rload hypothesis. It has been hypothesized that when the retain ed lung burden approach es 1 mg particles/g lung tissue, overloading will begin in the rat (Morrow, 1988); at 10 mg particles/g lung tissue macrophage-mediated clearance of pa rticles would effectively cease. Overloading appears to be a nonspecific effect noted in expe rimental studies, generally in ra ts, using many different kinds of poorly soluble particles (including TiO 2 , volcanic ash, DPM, carbon black, and fly ash) and results in A region clearance slowing or stasis, with an associated inflammation and aggregation of macrophages in the lungs and increased transl ocation of particles in to the interstitium (Muhle et al., 1990a,b; Lehnert, 1990; Morrow, 1994). Following overloading, the subsequent retardation of lung clearance, accumulation of particles, chronic inflammation, and the
interaction of inflammato ry mediators with cell proliferative processes and DNA may lead to the development of fibrosis, epithe lial cell mutations, and fibrosis in rats (Mauderly, 1996). The phenomenon of overload has been di scussed in greater detail in th e previous PM CD (U.S. EPA, 1996).

3.4.2. Relevance to Humans The relevance of “lung overload” to humans, and even to species other than laboratory species (rats and mice and hamste rs; Muhle et al., 1990a,b), is not clear. Although likely to be of
little relevance for most “real world” ambien t exposures of humans, this phenomenon is of concern in interpreting some long-term expe rimental exposure data and perhaps for human3-29 occupational exposure. In addition, relevance to humans is clouded by the fact that macrophage-mediated clearance is slower and pe rhaps less important in humans than in rats (Morrow, 1994).
Particle overload appears to be an important fact or in the pulmonary carcinogenicity observed in rats exposed to DPM. A study by Gr iffis et al. (1983) demons trated that exposure (7 h/day, 5 days/week) of rats to diluted whole DE containing DPM at concentrations of 0.15, 0.94, or 4.1 mg/m 3 for 18 mo resulted in lung burdens of 0.035, 0.220, and 1.89 mg/g of lung tissue, respectively. The alveolar cl earance of those rats with the highest lung burden (1.89 mg/g of lung) was impaired, as determined by a significantly greater ( p <0.0001) retention t 1/2 for DPM. Impaired clearance was reflected in the greater lung burden/exposure concentration ratio at the highest exposure level. Similarly, in the study by Chan et al. (1984), rats exposed for 20 h/day, 7 days/week to diluted whole DE containing DPM (6 mg/m 3 ) for 112 days had an extraordinarily high lung particle bur den of 11.8 mg, with no alveolar par ticle clearance being detected over 1

year. Muhle et al. (1990a,b) indicated that overloading of rat lung s occurred when lung particle burdens reached 0.5 to 1.5 mg/g of lung tissu e and that clearance mechanisms were totally compromised at lung particle burdens $ 10 mg/g for particles with a specific density close to 1,
observations that are concordant with those of Morrow (1988). Pritchard (1989), utilizin g data from a number of DE expos ure studies, examined alveolar clearance in rats as a function of cumulative exposure. The resulting analysis noted a significant
increase in retention t 1/2 values at exposures above 10 mg/m 3

@ h/day and also showed that normal lung clearance mechanisms appeared to be compromised as the lung DPM burden approached 0.5 mg/g of lung. Animal studies have revealed that impairmen t of alveolar clearance can occur following chronic exposure to DPM (Griffis et al., 1983; Wolff et al., 1987; Vo stal et al., 1982; Lee et al., 1983) or a variety of other diverse poorly solubl e particles of low toxicity (Lee et al., 1986, 1988; Ferin and Feldstein, 1978; Muhle et al., 1990). Because high lung burdens of relatively insoluble, biochemically inert particles result in diminution of normal lung clearance kinetics or in what is now called particle overloading, this effect appears to be more related to the mass and/or volume of particles in the lung than to the nature of th e particles per se. Particle overload relates only to poorly soluble par ticles of low toxicity. It mu st 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 bur dens than DPM). Regardless, as pointed out by Morrow (1988), particle overloa ding in the lung modifies th e dosimetry for particles in the lung and thereby can alter toxicologic responses.3-30 Although quantitative data are limited regarding lung overload associated with impaired alveolar clearance in humans, impairment of clear ance mechanisms appears to occur, and at a lung burden generally in the rang e reported to impair clearance in rats, i.e., approximately 1 mg/g lung tissue. Stöber et al. (1967), in their study of coal miners, reported lung particle burdens of 2 to 50 mg/g lung tissue, for which estimated clearance t 1/2 values were very long (4.9 years). Freedman and Robinson (1988) also reported slower alveolar cl earance rates in coal miners, some of whom had a mild degree of pneumoconiosis. It must be noted, however, as has been reported even in some studies with rats ex posed lifetime to overl oad conditions (50 mg/m 3 TiO 2 ; Lee et al., 1986) that no lung cancer was reporte d among those miners w ith apparent particle
overload.
Consideration of the above information further clarifies the human relevance of noncancer effects that may be elicited from overl oad-type conditions in rats studies. Under conditions that would be most likely to elic it overload conditions in humans, such as the excessive dust burdens in the lungs of min ers, cancer is not observed although noncancer responses such as fibrosis and macrophage re sponses are documented (Freedman and Robinson, 1988; Haschek and Witschi, 1991; Oberdörster, 1994). In delibera tion on the matter of whether the rat lung nonneoplastic responses to poorly solubl e particles (such as DP M) are predictive of a similar hazard in humans, an expert panel (ILS I, 2000) opined that such responses would indeed be a useful predictor for similar responses in humans.

3.4.3. Potential Mechanisms for an AM Sequestration Compartment for Particles During
Particle Overload Several factors may be involved in the particle -load-dependent retarda tions in the rate of particle removal from the lung and the corre sponding functional app earance of an abnormally slow clearing or particle seque stration compartment. As previ ously mentioned, one potential site for particle sequestration is the containment of particles in the Type I cells. Information on the retention kinetics for particles in the Type I cells is not currently available. Also, no morphometric analyses have been performed to date to estimate wh at fraction of a retained lung burden may be contained in th e Type I cell population of the lung during lung overloading. Another anatomical region in the lung that may be a slow clearing site is the interstitial compartment (Kuempel, 2000). Little is known about the kinetics of removal of free particles or particle-containing macrophages fro m the interstitial spaces, or what fraction of a retained burden of particles is contained in the lung’s interstitium during particle overload. The gradual accumulation of particles in th e regional lymph nodes and the a ppearance of particles and cells with associated particles in lymphatic channe ls and in the peribronc hial and perivascular3-31 lymphoid tissue (Lee et al., 1985; White and Ga rg, 1981) suggest that the mobilization of particles from interstitial sites via local lymphatics is a continual process. Indeed, it is clear from histologic observations of the lungs of rodents chronically exposed to DPM that Type I cells, the interstitium, th e lymphatic channels, and pulmonary lymphoid tissues could collectively comprise subcompar tments of a more generalized slow clearing compartment. Although these sites must be considered poten tial contributors to the increased retention of particles during par ticle overload, a disturbance in par ticle-associated AM-mediated clearance is undoubtedly the predominant cause, inasmuch as, at least in rodents, the AMs are the primary reservoirs of deposited pa rticles. The factors responsible for a failure of AMs to translocate from the alveolar space compartment in lungs with high particulate matter burdens remain uncertain, although a hypothesis concerning the process i nvolving volumetric AM bur den has been offered (Morrow, 1988). Other processes also may be involved in preventing particle -laden AMs from leaving the alveolar compartment under conditions of particle overl oad in the lung. Clusters or aggregates of particle-laden AMs in the alveo li are typically found in the lungs of laboratory animals that have received large lung burdens of a variety of ty pes of particles (Lee et al., 1985), including DPM (White and Garg, 1981; McClellan et al., 1982). The aggregation of AMs may explain, in part, the reduced clearance of particle-laden AM during particle overload. The definitive mechanism(s) responsible for this clustering of AMs has not been elucidated to date. Whatever the underlying mechanism(s) for the AM aggr egation response, it is noteworthy that AMs lavaged from the lungs of DE-exposed animals con tinue to demonstrate a propensity to aggregate (Strom, 1984). This observation could result e ither from the surface characteristics of AMs being fundamentally altered or from macrophage activation by phagocytiz ed particles that then release chemotactic factors (Be llmann et al., 1990) in a manner that promotes their adherence to one another in the alveolar region. AM aggrega tion may not simply be directly caused by their
abundant accumulation as a result of immobilization by large particle loads. Furthermore, even though overloaded macrophages may redistribute particle burden to other AMs, clearance may remain 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.

3.5. BIOAVAILABILITY OF ORGANIC CO NSTITUENTS PRESENT ON DIESEL
EXHAUST PARTICLES

Because it has been shown that DPM extract is not only mutagenic but also contains known carcinogens, the organic fraction was origin ally considered to be the primary source of carcinogenicity in animal studies. Since then, evidence has been presented that carbon black,3-32 lacking an organic component, is capable of i nducing lung cancer at exposure concentrations sufficient to induce lung particle overload. Th is suggested that the relatively insoluble carbon core of the particle may be of greater importance for the pathog enic and carcinogenic processes
observed in the rat inhalation st udies conducted at high exposure concentrations. (See Chapter 7 for a discussion of this issue.) However, l ung cancer reported in epidemiologic studies was associated with diesel exposur e levels far below those inducin g particle overl oad in lifetime studies in rats. It is therefore suggested th at compounds in the organic fraction of DPM may have some role in the etiology of human lung cancers. This leads to an interest in characterizing the bioavailability of organics. The bioavailability of toxic organic compounds adsorbed to DPM can be influenced by a variety of factors. Although the agent may be ac tive while present on the pa rticle, most particles are taken up by AMs, a cell type not generally considered to be a targ et site. In order to reach the target site, elution from the particle surface is necessary followed by diffusion and uptake by the target cell. Metabolism to an active form by either the phagocytes or the target cells is also required for activity of many of the compounds present. This section describes only the various manner and mechanisms by which organics adsorbed onto DPM may become bioavailable. In vivo and in vitro results involving various biological extraction media as well as modeled scenarios of bioavailability are presented. Actual estimates of the amount of organics from DPM to which respiratory tract tissues may be exposed are discussed and pres ented in Section 3.6.2.7.

3.5.1. In Vivo Studies
3.5.1.1. Laboratory Investigations Several studies reported on th e retention of particle-a dsorbed organics following administration to various rodent species. In studies reported by Sun et al. (1982, 1984) and Bond et al. (1986), labeled organics were deposited on DPM following heating to vaporize away the organics originally present. Sun et al. (1982) compared the disposition of either pure or diesel particle-ads orbed benzo[ a ]pyrene (B[ a ]P) following nose-only inhalation by F344 rats. About 50% of partic le-adsorbed B[ a ]P was cleared with a half-time of 1 h, predominantly by mucociliary clearance. The long-term retention of particle-adsorbed 3 H-B[ a ]P at 18 days was approximately 230-fold greater than that for pure 3 H-B[ a ]P (Sun et al., 1982). At the end of exposure, about 15% of the 3 H label was found in blood, liver, a nd kidney. Similar results were reported in a companion study by Bond et al. ( 1986), and by Sun et al. (1984) with another PAH, 1-nitropyrene, except the retention half-time was 36 days. Ball and King (1985) studied the disposition and metabolism of intr atracheally instilled

14 C-labeled 1-NP (>99.9% purity) coated onto DPM. About 50% of the 14 C was excreted within3-33 the first 24 h; 20% to 30% of this appeared in the urine, and 40% to 60% was excreted in the
feces. Traces of radiolabel were detected in the trachea and esopha gus. Five percent to 12% of the radiolabel in the lung co-purified with th e protein fraction, indicatin g some protein binding. The corresponding DNA fraction contained no 14 C above background levels. Bevan and Ruggio (1991) assesse d the bioavailability of B[ a ]P adsorbed to DPM from a 5.7-L Oldsmobile diesel engine. In this study, exhaust par ticles containing 1.03 : g B[ a ]P/g particles were supplem ented with exogenous 3 H-B[ a ]P to provide 2.62 : g B[ a ]P/g of exhaust particles. In vitro analysis indicated that the supplemented B[ a ]P eluted from the particles at the same rate as the original B[ a ]P. Twenty-four hours after intr atracheal instillation in SpragueDawley rats, 68.5% of the radiolabel remained in the lungs. This is approximately a 3.5-fold greater proportion than that reported by Sun et al. (1984), possibly because smaller amounts of B[ a ]P adsorbed on the particles resulted in strong er binding or possibly because of differences between inhalation exposure and intratracheal exposure. At 3 days following administration, more than 50% of the radioactivity remained in th e lungs, nearly 30% had been excreted into the feces, and the remainder was distributed thr oughout the body. Experiments using rats with
cannulated bile ducts showed th at approximately 10% of the ad ministered radio activity appeared in the bile over a 10-h period and that less than 5% of the radioactivity entered the feces via mucociliary transport. Results of these studies showed that when organics are adsorbed to DPM the retention of organics in the lungs is increas ed considerably. Because retention time is very short following exposure to pure compounds not bound to particles, it can be concluded that the increased retention time is primarily the result of continued binding to DPM. The detection of labeled compounds in blood, systemic organs, urine, and bile as well as the trachea, however, provides evidence that at least some of the organics are eluted from the particles following deposition in the lungs and would not be availa ble as a carcinogenic dose to the lung. As discussed above, the results of Gerde (1999a,b) indicate that most of the organics eluted from particles deposite d in the alveolar region, especially P AHs, are predicted to rapidly enter the
bloodstream and thus not to contribute to potential induction of lung cancer.

3.5.1.2. Studies in Occupationally Exposed Humans DNA adducts in the lungs of experimental anima ls exposed to DE have been measured in a number of animal experiments (World Health Organization, 1996). Such studies, however, provide limited information regarding bioavailability of organics, as positive results may well
have been related to factors a ssociated with lung particle overload, a circumstance reported by Bond et al. (1990), who found carbon black, a substa nce virtually devoid of organics, to induce DNA adducts in rats at lung overload doses. Th ese authors showed that levels of DNA adducts present in pulmonary type II cells from the lungs of rats (n=15) exposed to equivalent conditions3-34 of either carbon black or DE (each at 6.2 mg/m 3 ) were nearly the same and 4- to 5-fold more than air-exposed controls. This similarity was noted despite a difference of nearly three orders of magnitude in solvent-extractable organic cont ent between DE (30%) and carbon black (0.04%).
None of the DE or carbon black adducts comigrated with B[ a ]P diol epoxide. On the other hand, DNA adduct formation a nd/or mutations in blood cells following exposure to DPM, especially at levels insufficien t to induce lung overload, can be presumed to be the result of organics diffusing into the blood. Hemminki et al. (1994) reported increased levels of DNA adducts in lymphocytes of bus maintenance and truck te rminal workers. Österholm et al. (1995) studied mutations at the hprt-locus of T-lymphocytes in bus maintenance workers. Although they were unable to identify clear-c ut exposure-related di fferences in types of mutations, adduct formation was sign ificantly increased in the expos ed workers. Nielsen et al. (1996) reported significantly increased levels of lymphoc yte DNA adducts, hydroxyvaline adducts in hemoglobin, and 1-hydroxypyrene in urine of garage workers exposed to DE.

3.5.2. In Vitro Studies
3.5.2.1. Extraction of Diesel Particle-Associ ated Organics by Biological Fluids In vitro extraction of mutagenic organics by biological fluids can be estimated by measurement of mutagenic activity in the particular fluid. Usin g this approach, Brooks et al. (1981) reported extraction efficien cies of only 3% to 10% that of dichloromethane following DPM incubation in lavage fluid, serum, saline, albumin, or dipalmitoyl lecithin. Moreover, extraction efficiency did not increase with incubation time up to 120 h. Similar findings were reported by King et al. (1981), w ho also reported that lung lavage fluid and lung cytosol fluid extracts of DPM were not mutagenic. Serum extracts of DPM did exhibit some mutagenic activity, but considerably less than that of organi c solvent extracts. Furthermore, the mutagenic activity of the solvent extract was significantly reduced when combined with serum or lung cytosol fluid, suggesting protein binding or bi otransformation of the mutagenic components. Siak et al. (1980) assessed the mutagenicity of material extr acted from DPM by bovine serum albumin in solution, simulated lung surfactant, fe tal calf serum (FCS), a nd physiological saline. Only FCS was found to extract some mutagenic activity from the DPM. Keane et al. (1991), however, reported positive effects for mutagenicity in salmonella and sister chromatid exchange in V79 cells exclusively in the supernatant fr action of DPM dispersed in aqueous mixtures of dipalmitoyl phosphatidyl choline, a major compone nt of pulmonary surfactant, indicating that pulmonary surfactant components can extract active components of DPM and result in bioavailability. The ability of biological fluids to extract organics in vitro and their effectiveness in vivo remains equivocal because of the character of the particular fluid. For example, extracellular3-35 lung fluid is a complex mixture of constitu ents that undoubtedly have a broad range of hydrophobicity (George and Hook, 1984; Wright and Clements, 1987), wh ich is fundamentally
different from serum in terms of chemical composition (Gurley et al., 1988). Moreover, assessments of the ability of lavage fluids, which actually represent substantially diluted extracellular lung fluid, to extract mutagenic activity from DPM clearly do not reflect the in vivo condition. Finally, except under very high ex posure concentrations, few particles escape phagocytosis and possible intracellu lar extraction. In this respect, Hiura et al. (1999) have shown that whole exhaust containing DPM, but not carbon black or diesel particles devoid of organics, induces apoptosis, apparently throug h generation of oxygen radicals. This study implicates organic compounds present on DPM. It also indicates the bioavailability of organics for generation of radicals from reaction with par ticle-associated organics or following elution
from DPM.

3.5.2.2. Extraction of DPM-Associated Organics by Lung Cells and Cellular Components A more likely means by which organics may be extracted from DPM and metabolized in the lung is either through par ticle dissolution or extraction of organics from the particle surface within the phagolysosomes of A Ms and other lung cells. This mechanism presupposes that the particles are internaliz ed. Specific details about the phy sicochemical conditions of the intraphagolysosomal environment, where particle dissolution in AMs presumably occurs in vivo,
have not been well characterized. It is known th at phagolysosomes constitu te an acidic (pH 4 to 5) compartment in macrophages (Nilsen et al ., 1988; Ohkuma and Poole, 1978). The relatively low pH in the phagolysosomes has been associ ated with the dissolution of some types of inorganic particles (some meta ls) by macrophages (Marafante et al., 1987; Lundborg et al., 1984), but few studies provide quantitative inform ation concerning how organics from DPM may be extracted in the phagolysosomes (Bond et al., 1983). Whatever the mechanism, assuming elution occurs, the end result is a prolonged exposure of the respiratory epithelium to DPM organics, which include low concentrations of carcinogenic agents such as PAH. Early studies by King et al. (1981) found that when pulmonary alveolar macrophages were incubated with DPM, am ounts of organic compounds and mutagenic activity decreased measurably from the amount originally associated with the particles, suggesting that organics were removed from the phagocytized particles. Leung et al. (1988) studied the ability of rat lung and liver microsomes to facilitate transfer and metabolism of B[ a ]P from diesel particles. 14 CB[ a ]P coated diesel particles, previously extr acted to remove the original organics, were incubated directly with liver or lung microsomes. About 3% of the par ticle-adsorbed B[ a ]P was transferred to the lung microso mes within 2 h. Of this amount about 1.5% was metabolized, for a total of about 0.05% of the B[ a ]P originally adsorbed to the DPM. Although transformation is3-36 slow, the long retention of pa rticles, including DPM, in hum ans may cause the fraction eluted
and metabolized to be considerab ly higher than this figure. In analyzing phagolysosomal dissolution of va rious ions from particles in the lungs of Syrian golden hamsters, however, Godleski et al . (1988) demonstrated that solubilization did not necessarily result in clearance of the ions (and therefore general bioavailability) in that binding of the solubilized components to cellular and extrace llular structures occurred. It is reasonable to assume that phagocytized DPM particles may be subject to similar processes and that these processes would be impor tant in determining th e rate of bioavailability of the particle-bound
constituents of DPM. Alveolar macrophages or macropha ge cell lines that were exposed to high concentrations
of DPM in vitro were observed to undergo apoptosis, which was a ttributed to the generation of reactive oxygen radicals (ROR) (H iura et al. 1999). Further expe rimentation showed that DPM with the organic constituents extracted was no long er able to induce apoptos is or generate ROR. The organic extracts alone, however, were able to induce apoptosis as well as the formation of stress-activated protein kinases th at play definitive roles in ce llular apoptotic pathways. The injurious effects of nonextracted DPM or of DPM extracts were obser ved to be reversible by the antioxidant radical scavenger N-acetyl cysteine. These data suggest strongly that, at least at high concentrations of DPM, the organic constituents contained on DPM play a central role in cellular toxicity and that this toxicity may be attributable to the generation of ROR.

3.5.3. Modeling Studies Gerde et al. (1991a,b) described a model simulating the effect of particle aggregation and PAH content on the rate of PAH release in the lung. According to this model, particle aggregation will occur with high exposure concentrations, resulting in a slow release of PAHs and prolonged exposure to surrounding tissues. However, large aggregates of particles are unlikely to form at doses typical of human exposur es. Inhaled particles, at low concentrations, are more likely to deposit and react with surr ounding lung medium without interference from other particles. The model pr edicts that under low-dose exposur e conditions, more typical in humans, particle-associated orga nics will be released more ra pidly from the particles because they are not aggregated. Output from this m odel suggests strongly that sustained exposure of target tissues to PAHs will result from repeated exposures, not from increased retention due to association of PAHs with carrier particles. Th is distinction is important because at low doses PAH exposure and lung tumor formation would be predicted to occur at sites of deposition rather than retention, as occurs with high doses.
The site of release of PAHs influences effective dose to the lungs because, as noted previously, at least some free organic compounds deposited in the lungs are rapidly absorbed into3-37 the bloodstream. Gerde et al. (1991b) predicted P AHs would be retained in the alveoli less than 1 min, whereas they may be retained in the c onducting airways for hours. These predictions were based on an average diffusion distan ce to capillaries of only about 0.5 : m in the alveoli, as compared to possibly greater than 50 : m in the conducting airways such as the bronchi. An experimental study by Gerde et al. (1999) provi ded support for this prediction. Beagle dogs were exposed to 3 H-B[ a ]P adsorbed on the carbonaceous core of DPM at a concentration of 15 : g B[ a ]P/gm particles. A rapidly eluting fraction from DPM deposited in the alveoli was adsorbed
into the bloodstream and metabolized in the liver , whereas the rapidly el uting fraction from DPM deposited in the conducting airways was to a large ex tent retained and metabo lized in situ in the airway epithelium. Thus, organics eluting from DPM depositing in the conducting airways (i.e., the TB region) would have a basis for a longer residence time in the tissues (and for consequent biological activity) than would organics eluting from DPM depositing in the pulmonary parenchyma. And, given the same overall deposited dose of DPM to the total pulmonary system, a deposited dose with a higher proportion in the TB region would incur a higher probability of tissue interactions with any eluted organics. This may be the case when comparing regional doses of DPM to humans as compared to rats for two reasons. First, one deposition model (Freijer et al., 1999) projects th at for air concentrations of DPM at either 0.1 or 1.0 mg/m 3 , a higher proportion of the total DPM dose to the pulmonary system would be deposited in the TB area for humans at 31% (TB/Total; 0.098 / 0.318) than for rats at only 16% (0.04 / 0.205). Second, comparative morphometry data of DPM from chronically exposed rats and primates showed higher levels of DPM adjacent to conduc ting airways in primates (i.e., the interstitium of the respiratory bronchioles) than were present in parallel regions in the rat (interstitium of the alveolar ducts) (Nikula et al., 1997a,b). The focal nature of this deposition could give rise to localized high concentrations of any organics eluted.

3.5.4. Summary and Bioavailability At present, the available data are insufficient to accurately model the effective dose of organics in the respiratory tract of humans or animals exposed to DPM. As mentioned above, though, the following Section (3.6.2.7) does presen t estimates of the actual amount of organics, including carcinogenic PAH such as B[ a ]P, that are deposited in the lung and could become
bioavailable. Overall, the results of studies presented in Section 3.6 provide evid ence that at least some of the organic matter adsorbed to DPM deposite d in the respiratory tract is eluted. The percentage taken up and metabolized to an active form by target cells is, however, uncertain. Organics eluted from particles de posited in alveoli are likely to rapidly enter the bloodstream via
translocation across endothelial cells, where they may undergo me tabolism by enzymes such as3-38 cytochromes P-450 that are capable of producing r eactive species. Organics eluted from particles deposited in the conducting airway s (the bronchioles, bronchi, and trachea) may also undergo metabolism in other cell types such as the Clara cells with constituent or inducible cytochrome P450 species. Risk of harmful effects for pa rticles deposited in th e conducting airways is predicted to be greater because solubilized or ganic compounds will be retained in the thicker tissue longer, allowing for metabolism by epithelial cells lining the airways. Furthermore, since some deposition in conducting airways occurs pr imarily at bifurcations, localized higher concentrations may occur.

3.6. MODELING THE DEPOSITION AND CLEARANCE OF PARTICLES IN THE
RESPIRATORY TRACT
3.6.1. Introduction The biological effects of inha led particles are a function of their disposition, i.e., their deposition and clearance. This, in turn, depends on their patterns of deposition (i.e., the sites within which particles initially come into contact with ai rway epithelial surfaces and the amount removed from the inhaled air at these sites) a nd clearance (i.e., the rates and routes by which deposited materials are removed from the respirat ory tract). Removal of deposited materials involves the competing processe s of macrophage-mediated clear ance and dissolution-absorption. Over the years, mathematical models fo r predicting deposition, cl earance and, ultimately, retention of particles in the respiratory tract ha ve been developed. Such models help interpret experimental data and can be us ed to make predictions of depos ition for cases where data are not
available. A review of vari ous mathematical par ticle deposition models was given by Morrow and Yu (1993) and in U.S. EPA (1996). Currently available data for long-term inha lation exposures to poor ly soluble particles (e.g., TiO 2 , carbon black, and DPM) show that pulm onary retention and clearance of these particles are not adequa tely described by simp le first-order kinetics and a single compartment representing the alveolar macropha ge particle burden. Several investigat ors have developed models for deposition, transport, and clearance of poorly soluble particulate matter in the lungs. All of these models identify various compartments and associated transpor t rates, but empirically
derived data are not available to substantiate many of the assumptions made in these models.

3.6.2. Dosimetry Models for DPM
3.6.2.1. Introduction The extrapolation of toxicological results fro m laboratory animals to humans, the goal of this chapter, requires the use of dosimetry mode ls for both species that include, first, the deposition of DPM in various regions of the re spiratory tract, and second, the transport and3-39 clearance of the particles, including adsorbed constituents, from their deposited sites. Therefore the ideal model structure would incorporat e both deposition and clearance in animals and
humans. Deposition of particles in the respiratory tract, as describe d above, can be by impaction, sedimentation, interception, a nd diffusion, with the contribution from each mechanism a function of particle size, lung structure, and size and breat hing parameters. Because of the size of diesel particles, under nor mal breathing conditions most of this deposition takes place by diffusion, and the fraction of the inhaled mass that is deposited in the thoracic region (i.e., TB plus A regions) is substantially similar for rats and humans. Among deposition models that include aspects of lung structure and breathing dynamics, the most widely used have been typical-pat h or single-path models (Yu, 1978; Yu and Diu, 1983). The single-path models are based on an idealized symmetric geometry of the lung, assuming regular dichotomous branching of the ai rways and alveolar ducts (Weibel, 1963). They lead to modeling the deposition in an average re gional sense for a given lung depth. Although the lower airways of the lung may be reasonably characterized by such a symmetric representation, there are major as ymmetries in the upper airways of the tracheobronchial tree that in turn lead to different appor tionment of airflow and particul ate burden to the different lung lobes. The rat lung structure is highly asy mmetric because of its m onopodial nature, leading to significant errors in a single-path description. Th is is rectified in the multiple-path model of the lung, which incorporates asymmetr y and heterogeneity in lung br anching structure and calculates deposition at the individual airway level. Th is model has been deve loped for the rat lung (Anjilvel and Asgharian, 1995; Freijer et al., 1999) and, in a limited fashion because of insufficient morphometric data, for the human lung (Subramania m et al., 1998; Yeh and Schum, 1980). Such models are particular ly relevant for fine and ultraf ine particles such as occur in DPM. However, models for cl earance have not yet been impleme nted in conjunction with the use of the multiple-path model. Clearance of particles in the respiratory tract takes place (1) by mechanical processes: mucociliary transport in th e ciliated conducting airways a nd macrophage phagocytosis and migration in the nonciliated airways, and (2) by disso lution. The removal of material such as the carbonaceous core of DPM is largely by mechan ical clearance, whereas the clearance of the organics adsorbed onto the carbon core is principally by dissolution. Several models currently exist that integr ate both deposition and cl earance, some specific for humans and others specific fo r laboratory animals. They differ significantly in the level of physiological detail that is captured in the model and in the uncertainties associated with the values of the parameters used. All of these models identify va rious compartments and associated transport rates, but empirically derived data are not available to validate many of the assumptions3-40 made in the models. A review of the prin cipal human and animal de position/clearance models, including candidate models for use in animal-to -human extrapolation in this assessment, are
considered below.

3.6.2.2. Human Models The International Commission on Radiologi cal Protection (ICRP) recommends specific mathematical dosimetry models as a means to calculate the mass deposition and retention by different parts of the human resp iratory tract and, if needed, tissu es beyond the respiratory tract.
The latest ICRP-recommended mode l, ICRP66 (1994), considers th e human respiratory tract as four general anatomical regions: the ET region, which is divided into two subregions; the TB region, which is also subdivided into two regi ons; and the gas-exchange tissues, which are further defined as the alveolar-interstitial (AI) region but are exactly comparable to the pulmonary or A region. The fourth region is th e lymph nodes. The deposition component of the model for the ET, TB, and A regions is semi-empirical based on equations derived from fitting experimental deposition data. The dimensional model used for the TB and A regions was adopted from several sources (Weibel, 1963; Ye h and Schum, 1980; and Phalen et al., 1985); the physical aspects of the individual airway generations for these regions were all averaged after each source was adjusted to a standard functiona l residual capacity. The equations for estimating deposition in these areas was empirical, obtained from fitting data obtained from partial human lung casts or from theoretical calculation for th ese regions. Deposition in the four regions is given as a function of pa rticle size with two different types of particle size parameters: activity median thermodynamic diameter (AMTD) for de position of particles ra nging in size from 0.0005
to 1.0 : m and the activity median aerodynamic diam eter (AMAD) for de position of particles
from 0.1 to 100 : m. Reference values of regional deposition are provi ded and guidance is given for extrapolating to specific i ndividuals and populations under diffe rent levels of activity. This model also includes consideration of particle inhalability, a measure of the degree to which particles can enter the respirator y tract and be available for de position. After deposition occurs in a given region, two different intrinsic clearan ce processes act competitively on the deposited particles: particle transport, including mucociliary clearance from the respiratory tract and physical clearance of particles to the regional lymph nodes; and absorption, including movement of material to blood and both dissolution-absorption and transport of ultrafin e particles. Rates of particle clearance derived from studies with human subjects are assumed to be the same for all types of particles. The ICRP model provides average concentra tion or average number values on a regional basis, i.e., mass or number deposited or retained in the ET, TB, or A regions. Additionally, while the ICRP66 mode l was developed primar ily for use with airborne radioactive3-41 particles and gases in humans, its use for describing the dosimetr y of inhaled mass of
nonradioactive substances in humans is also appropriate.
The National Council on Radia tion Protection (NCRP) has i ssued a human respiratory tract dosimetry model that wa s developed concurrently w ith the ICRP model (NCRP, 1997; Phalen et al.,1991). It addresse s (1) inhalability of particles, (2) new subregions of the respiratory tract, (3) dissolution-ab sorption as an important aspect of the model, and (4) body size (and age). The proposed NCRP model defines th e respiratory tract in terms of a naso-oropharyngo-laryngeal (NOPL) region, a TB region, a pulmonary (P) region, and the lung-associated lymph nodes (LN). Like the ICRP model, th e deposition component of the model for the ET region is semi-empirical, based on equations derived from fitting experimental deposition data.
The dimensional model used for the TB and A re gions was that of Yeh and Schum (1980). The data from this model were used to estimate physical processes along a typical lung path (vice multiple-path; see MPPDep model description below) on a generation-by-generation basis. The rates of dissolution-absorption of particles and their constituents are derived from clearance data from humans and laboratory animals. The effect of body growth on par ticle deposition is also considered in the model, althoug h particle clearance rates are assu med to be independent of age. The NCRP model currently available consider s deposition only within these regions of the respiratory tract. As with the ICRP model, the NCRP model can be used for evaluating inhalation exposures to all ty pes of particles. Comparison of regi onal deposition patterns estimated by the ICRP66 and the current NCRP m odels have been reported (Yeh et al., 1996). One principal difference between the models is the enhanced deposition of ultrafines in the tracheobronchial region predicted by the NCRP model compared with the ICRP model. This effect of enhanced deposition is claimed to be due to the entr ance configuration of an airway
bifurcation. The model of Freijer et al. (1999) is a mu ltiple-path particle de position model (MPPDep) for the human respiratory tract that differs fundamentally from the above two models as
described in the Introduction. Ca lculations from the model may be based on either single-path or multiple-path methods for tracking air flow and calculating aerosol deposition in the lung. The single-path method calculates deposition for a ty pical path, whereas the multiple-path method is capable of incorporating the asymmetry in l ung structure and providi ng lobar-specific and airway-specific info rmation. Two options are provided for idealizing the geometry of the human lung; one uses a symmetric geometry for the whole lung and the second option captures the asymmetry in the lobar structure, but treats the geometry within each lung lobe in a symmetric
fashion. Both models are c onstructed using morphometric da ta compiled by Yeh and Schum (1980). Within each airway, deposition is calculate d using theoretically de rived efficiencies for deposition by diffusion (most relevant to DPM), sedimentation, and impac tion within the airway3-42 or airway bifurcation. Filtration of particulate aerosols by the head is determined using empirical efficiency functions. The model calculates de position of monodisperse and polydisperse aerosols in the respiratory tract of both humans (and ra ts) for particles ranging from ultrafine (0.01 microns) to coarse (20 microns) sizes. Va rious breathing pattern s may be simulated: endotratracheal, nasal, oral, and co mbined nasal and oral (oronasa l). The exposure scenario may
be constant or variable. For the variable scenario, the user may specify different breathing patterns either on an hourly basi s during the day or activity patte rns for variable time durations. Adjustment for inhalability of th e particle is also included as an option. The software in this model provides results for the deposition fraction and mass deposited in the various regions of the respiratory tract in graphical and text formats. The combined model of Yu et al. (1991) ha s a human component that will be discussed
below.

3.6.2.3. Animal Models Strom et al. (1988) develope d a multicompartmental model for particle retention that partitioned the alveolar region into two compartments on the basis of the physiology of clearance. The alveolar region has a sepa rate compartment for sequester ed macrophages, corresponding to phagocytic macrophages that are heavily laden w ith particles and clus tered, and consequently have significantly lowered mobility. The model has the following compartments: (1) tracheobronchial tree, (2) free particulate on the alveolar surface, (3) mobile phagocytic alveolar macrophages, (4) sequest ered particle-laden alveolar macrophages, (5) regional lymph nodes, and (6) gastrointestinal tract. The mode l is based on mass-depende nt clearance (the rate
coefficients reflect this relati onship), which dictates sequestrati on of particles and their eventual transfer to the lymph nodes. The transport ra tes between various comp artments were obtained by fitting the calculated results to lung and lymph node burden experimental data for both exposure and postexposure periods. Because the number of fitted parameters was large, the model is not likely to provide unique solutions that would s imulate experimental da ta from various sources and for different exposure scenarios. For the sa me reason, it is not readily possible to use this model for extrapolating to humans. Stöber and co-workers have worked extensively in developing models for estimating retention and clearance of relatively insoluble respirable particles (as DPM) in the lung. Their most recent work (1994), a revised version of the POCK model, is a rigorous attempt to incorporate most of the physiologically known as pects of alveolar clearance and retention of inhaled relatively insoluble particles. Their multicompartmental kinetics model has five subcompartments. The transfer of particles between any of the compartments within the alveolar region is macrophage mediated. There are two co mpartments that receive particles cleared from3-43 the alveolar regions: the TB tract and the lym phatic system. The macrophage pool includes both mobile and particle-laden imm obilized macrophages. The mode l assumes a constant maximum volume capacity of the macrophages for partic le uptake and a materi al-dependent critical macrophage load that results in total loss of macrophage mobility. Sequestration of those macrophages heavily loaded with a particle burden close to a volume load capacity is treated in a sophisticated manner by approxima ting the particle load distribu tion in the macrophages. The macrophage pool is compartmentalized in terms of numbers of macrophages that are subject to discrete particle load intervals. Upon macrophage death, the phagocytized particle is released back to the alveolar surface; thus phagocytic particle collection competes to some extent with this release back to the alveolar surface. This recycled particle load is also divided into particle
clusters of size intervals defini ng a cluster size distribution on the alveolar surface. The model yields a time-dependent frequency distribution of loaded macrophages that is sensitive to both exposure and recovery peri ods in inhalation studies. The POCK model also emphasizes the importance of interstitial burden in the particle overload phenomenon and indicates th at particle overloa d (Section 3.4) is a function of a massive increase in particle burden of the interstitial space rather than total burden of the macrophage pool. The relevance of the increas ed particle burden in the inters titial space lies with the fact that this compartmental burden is not available for macrophage-mediated clearance and, therefore, persists even after cessation of exposure. Although the POCK model is the most sophis ticated in the physiological complexity it introduces, it suffers from a major disadvantage. Experimental retenti on studies provide data only on total alveolar and lymph node mass burdens of th e particles as a f unction of time. The relative fraction of the deposition between the al veolar subcompartment s in the Stöber model therefore cannot be obtained expe rimentally; the model thus uses a large number of parameters that are simultaneously fit to experimental data. Although th e model predictions are tenable, experimental data are not curren tly available to substantiate the proposed compartmental burdens or the transfer rates associated with these compartments. Thus, overparameterization in the model leads to the possibility that the model may not provide a unique solution that may be used for a variety of exposure scenario s, and for the same reason, cannot be used for extrapolation to humans. Stöber et al. have not developed an e quivalent model for humans; therefore the use of
their model in our risk assessmen t for diesel is not attempted.

3.6.2.4 . Combined Models (for interspecies extrapolation) Currently available data for long-term inha lation exposures to poor ly soluble particles (e.g., TiO 2 , carbon black, and DPM) show that pulm onary retention and clearance of these particles are not adequa tely described by simp le first-order kinetics and a single compartment3-44 representing the alveolar macr ophage particle burde n. A two-compartment lung model that could be applied to both human s and animals was developed by Smith (1985) and includes alveolar and interstitial compartments. For uptake and clearance of particles by alveolar surface macrophages and interstitial encapsu lation of particles (i.e., quartz dust), av ailable experimental data show that the rate-controlling functions followed Michaelis-Menton type kinetics, whereas other processes affecting particle transfer are assumed to be linear. The model was used in an attempt to estimate in terstitial dust and fibros is levels among a group of 171 silicon carbide workers; the levels were then compared with evidence of fibrosis from chest radiographs. A significant correlation was found between estimated fibrosis and profusion of opacities on the radiographs. This model provides as many as seven different ra te constants derived by various
estimations and under various conditions from both animal and human sources. The model was intended for estimation of generaliz ed dust described only as respir able without any other regard to sizing for establishing the va rious particle-related ra te constants. As mo st of the described functions could not be validated with experimental data, the applicability of this model, especially for particulates in the si ze range of DPM, was unclear. Yu et al. (1991; also reported as Yu and Yoon, 1990) have developed a threecompartment lung model that consists of trache obronchial (T), alveolar (A), and lymph node (L) compartments (Appendix A, Figure A-1) and, in addition, considered filtration by a nasopharyngeal or head (H) compartment. The tracheobronchial compartme nt is important for short-term considerations, whereas long-term clearance takes place via the alveolar compartment. In contrast to the Stöber and Strom approaches , the macrophage compartment in the Yu model contains all of the phagocytized particles; that is, there is no separate (and hypothetical) sequestered macrophage subcompar tment. Absorption by the blood (B) and gastrointestinal (G) compartments was also consid ered. Although the treatment of alveolar clearance is physiologically less sophisticated than that of the Stöber et al. model, the Yu model provides a more comprehensive treatment of clearance by including systemic compartments and the head, and including the clearance of the organic com ponents of DPM in addition to the relatively
insoluble carbon core. In order to progress beyond the classical human ICRP66 retention model, Yu has addressed the impairment of long-term clearance (t he overload effect) by using a set of variable transport rates for clearance from the alveolar re gion as a function of the mass of DPM in the alveolar compartment. A functiona l relationship for this was derived mathematically (Yu et al., 1989) based upon Morrow’s hypothesis for the macr ophage overload effect discussed earlier in the section on pulmonary overload. The extent of the impairment depends on the initial particle burden, with greater particulate concentration leading to slower clearance.3-45 Within this model, DPM is treated as bein g composed of three ma terial components: a relatively insoluble carbonaceous core, slowly clear ed organics (10% particle mass), and fastcleared organics (10% particle mass). Such a partitioning of organics was based on observations that the retention of particle-associated organics in lungs shows a biphasic decay curve (Sun et al., 1984; Bond et al., 1986). For any compartme nt, each of these components has a different transport rate. The total alveolar clearance ra te of each material component is the sum of
clearance rates of that materi al from the alveolar to the tracheobronchial, lymph, and blood compartments. In the Strom and Stöber models discussed above, the cl earance kinetics of DPM were assumed to be entirely dictated by those of the relatively insoluble carbonaceous core. For those organic compounds that beco me dissociated from the carbon core, clearance rates are likely to be very different, and some of these com pounds may be metabolized in the pulmonary tissue or be absorbed by blood. The transport rates for the three components we re derived from exper imental data for rats using several approx imations. The transport rates for the carbonaceous core and the two organic components were derived by fittin g to data from separate expe riments. Lung and lymph node burdens from the experiment of Strom et al. (1988) were used to determin e the transport rate of the carbonaceous core. The Yu model incorporates the impairment of clearance by including a mass dependency in the transport rate. This mass dependency is easily extracted because the animals in the experiment were sacrificed over varying periods following the end of exposure. It was assumed that the transport rates from the alveolar and lymph compartments to the blood were equal and inde pendent of the particul ate mass in the alveolar region. The clearance rates of particle-associat ed organics for rats were derived fro m the retention data of Sun et al. (1984) for B[a]P and the data of Bond et al. (1986) for nitropyrene adsorbed on diesel particles. In their model Yu et al. (1991) make tw o important assumptions to carry out the extrapolation in considera tion of inadequate human data. First, the transport rates of organics in the DPM do not change across species. This is based upon lung clear ance data of inhaled lipophilic compounds (Schanker et al., 1986), where the clearance was seen to be dependent on the lipid/water partition coefficient. In contra st, the transport rate of the carbonaceous core is considered to be significantly species depende nt (Bailey et al., 1982). DPM clearance rate is determined by two terms in the model (s ee Equation A-82 in Appendix A). The first, corresponding to macrophage-med iated clearance, is a functi on of the lung burden and is assumed to vary significantly across species. The second term, a constant, corresponding to clearance by dissolution, is assumed to be speci es independent. The ma ss-dependent term for humans is assumed to vary in the same pr oportion as in rats under the same unit surface particulate dose. The extrapolation is then achie ved by using the data of Bailey et al. (1982) for the low lung burden limit of the clearance rate. This value of 0.0017/day wa s lower (i.e., slower)3-46 than the rat value by a factor of 7.6. This is el aborated further in Appendix A. Other transport rates that have lung bur den dependence are extrapol ated in the same manner. It should be noted that the Bailey et al . (1982) experiment in humans used fused monodisperse aluminosilicat e particles of 1.9 and 6.1 : m aerodynamic diameters. Yu and coworkers have used the longer of the half-times observed in this experiment to obtain an alveolar human clearance rate ( 8 ), of 0.00169/day. In using such data for DPM 0.2 : m in diameter, they have assumed the clearance of relatively insoluble particles to be indepe ndent of size over the range in diameter from 6.1 down to 0.2 : m. This assumption is consis tent with observations and views currently in the literature indicating that clearance mechanisms are not particularly particle-size dependent (Morrow et al., 1967a,b; Snipes et al., 1983) . That the linear dimensions of an alveolar macrophage, considered to be the principal means of clearance in the A region, are significantly larger, roughly 10 : m (Yu et al., 1996), and could therefore accommodate engulfment of a range of partic le sizes also makes this assump tion reasonable. Snipes (1979), however, has reported in rats a 8 (converted here from half-time values) of 0.0022/day for 1 and
2 : m particles but a higher value of 0.0039/day for 0.4 : m particles indicating that clearance rates may indeed depend on size. In the absence of reliable data for 0.2 : m particles, the slower clearance rate pertaining to this larger partic le size, i.e., 0.00169/day, is being used. Such a choice may underestimate the actual DPM clearance rate in humans. The resulting model output (i.e., lung DPM burden) from this slower rate would predict mo re DPM in the alveolar space than may actually be present at any given time. Therefore, use of this slower 8 may be considered to be more protective of human hea lth. Long-term clearance rates for particle sizes more comparable to DPM are available, e.g., iron oxide and polystyrene spheres (Waite and Ramsden, 1971; Jammet et al., 1978), but these data show a large range in the values obtained for half-lives or are based upon a very small num ber of trials, and ther efore compare unfavorably with the quality of data from the Bailey experiment. The deposition fractions of particulate ma tter in the pulmonary and tracheobronchial regions of the human lung remain relatively unchanged over the partic le size range between
0.2 and 1.0 : m, on the basis of analysis done with the ICRP model (ICRP66, 1994). As the clearance of relatively insoluble particles is also likely to remain the same over this range, the dosimetry results in this report for the carbon aceous core component of DPM could also be extended to other particles in this size range within the PM 2.5. For respirable particles with diameters larger than this range, e.g., between 1.0 and 3.5 : m, the extent of the fraction deposited in the pulmonary region is unclear. Results from the ICRP66 (1994) model predict little change in human deposition for this diameter range, whereas the earlier model of Yu and Diu (1983) predicts a significant increase as reported in I CRP66 (1994). It is ther efore unclear if either model would be applicable for par ticles in this larger-sized rang e without changing the value for3-47 the deposition fractions. As will be presented and discussed below, regional deposition fractions of DPM-sized particles from th e MPPDep, the ICRP66 (1994) and draft NCRP models compare favorably with the human alveolar deposition in humans specific for DPM, which has been
estimated with the Yu model to be 7% to 13% (Yu and Xu, 1986). Although there was good agreement between experimental and m odeled results, this agreement follows a circular logic (as adequa tely pointed out by Yu and Yoon [1990]) because the same experimental data that figured into th e derivation of transport rates were used in the model. Nevertheless, even though this agreem ent is not a validation, it provides an important consistency check on the model. Further expe rimental data and po licy definitions on what
constitutes validation woul d be necessary for a more formal validation. The model showed that at low lung bur dens, alveolar clear ance is dominated by mucociliary transport to the tr acheobronchial region, and at hi gh lung burdens, clearance is dominated by transport to the lymphatic system . The head and tracheobronchial compartments showed quick clearance of DPM by mucociliary tr ansport and dissolution. Lung burdens of both the carbonaceous core and organics were found to be greater in humans than in rats for similar periods of exposure. The Yu and Yoon (1990) version of the m odel provides a parame tric study of the dosimetry model, examining variation over a range of exposure concentrations, breathing scenarios, and ventilation parameters; particle mass median aerodynamic diameters; and geometric standard deviations of the aerosol size distribution. It examines how lung burden varies with age for exposure over a lifespan, pr ovides dosimetry extrapola tions to children, and examines changes in lung burden with lung vol ume. The results showed that children would exhibit more diminished alveolar clearance of DPM at high lung burden than adults when exposed to equal concentrations of DPM. Thes e features make the model easy to use in risk
assessment studies. The reader is referred to A ppendix A for further details on the model and for analyses of the sensitivity of the model to change in parameter values. The Yu model presents some uncertainties in addition to those discussed earlier in the context of particle size dependence of clearance rate. The re ports of Yu and Yoon (1990) as well as Yu et al. (1991) underwent extensive peer revi ew; we list below the most important among the
model uncertainties discusse d by the review panel. The exper imental data used by the Yu model for adsorbed organics used passively adsorb ed radiolabeled compounds as surrogates for combustion-derived organics. These compounds ma y adhere differently to the carbon core than do those formed during combustion. Yu has estima ted that slowly cleared organics represent 10% of the total particle mass; the actual figure could be substantially less; the reviewers estimate that the amount of tightly bound organics is probably only 0.1% to 0.25% of the particle

mass.3-48 The model was based upon the experimental data of Strom et al. (1988), where Fischer-344 rats were exposed to DPM at a concentration of 6.0 mg/m 3 for 20 h/day and 7 days/week for periods ranging from 3 to 84 days. Such exposures lead to particle overload effects in rats, whereas human exposure patterns are usually to much lower levels at which
overload will not occur. Parame ters obtained by fitting to da ta under the conditions of the experimental scenario for rats may not be optimal for the human exposure and concentration of
interest. The extrapolation of retained dose from ra ts to humans assumed that the macrophagemediated mechanical clearance of the DPM varies with the specific particulate dose to the alveolar surface in the same proportion in hum ans and in rats, whereas clearance rates by
dissolution were assumed to be invariant across species. Th ese assumptions have not been
validated.
It should also be noted that the Yu et al. (1991) model does not possess a formal interstitial compartment although the lymph nodes, which would be the repository of particles from the interstitium, are represented. The work of Nikula et al. (1997a,b) and of Kuempel (2000) provide compelling information on the si gnificance of an extensive interstitilization process in primates and in humans. Kuempe l (2000) developed a l ung dosimetry model to
describe the kinetics of particle clearance and retention in coal miners' lungs. Models with overloading of lung clearance, as observed in rodent studies, were f ound to be inadequate to describe the end-of-life lung dus t burdens in those miners. The model that provided the best fit to the human data included a seque stration process representing the transfer of particles to the interstitium. These findings are consistent with a study showing reduced lung clearance of particles in retired coal min ers (Freedman and Robinson, 1988) and with studies showing increased retention of partic les in the lung interstitium of humans and nonhuman primates compared to rodents exposed to coal dust and/or DE (Nikula et al., 1997a ,b). These findings are
also consistent with the esta blished observation that humans a nd primates clear particles slowly from the alveolar interstitium compared with rates in rodent species such as rats and mice (Hsieh and Yu, 1998). Because several aspects of the Yu model have not been validated on human data
and because it does not include a formal inters titial compartment, it is acknowledged that this model may therefore have some uncertainty concerning the lung burdens in humans exposed to occupational levels of dust. However, it is also not known whether the model based on coal miner data (Kuempel, 2000) would also describe the clearance a nd retention processes in the lungs of humans with exposures to particles at lower envir onmental concentrations, or to submicrometer particles such as DE particulate. Further investigation of these issues is needed.3-49 3.6.2.5. Use of the Yu et al. (1991) M odel for Interspecies Extrapolation In addressing the objectives of this chapte r, i.e., consideration of what is known and applicable to DPM concerning particle disposition and the bioavailability of adsorbed organics on DPM, it is apparent that th e database is considerable fo r both the processes involved in particle dosimetry and for DPM. This information makes the goal of predicting a human internal dose from animal data through a model utilizing this database both feasible and appropriate. In their charge to EPA through “Science and Judgment in Ri sk Assessment” (NRC, 1995), the National Research Counc il opines that EPA should have principles for judging when and how to depart from default options. The ex tensive data presented in this chapter their scientific validity, and the limitations of the current default procedures provide a basis for departing from the default options currently identified by the Agency for extrapolating from animals to humans. The default option of assu ming external concentrations of DPM in animal studies as being representative of a human con centration (and an equiva lent internal dose) is clearly not adequate given the differences in the basic processes of deposition and clearance between animals and humans documented by these da ta. Use of an alternate default option, the Agency’s dosimetric adjustment procedures for inhaled particles in animal-to-human scenarios (described in U.S. EPA, 1994), is also inadequa te as only deposition is predicted and then only
down to an MMAD of 0.5 : m, whereas the MMAD of DPM is typically 0.2 : m or smaller.
Models have been described in this section that consid er both deposition and retention specifically for DPM in both laboratory anima ls and in humans. These points provide justification for moving away from default options and utilizing the best scientific informationavailable (i.e., that integrated into deposition/clearance models) in performing the animal-tohuman extrapolation.
Evaluation of the various models discussed in this chapter should be considered from the aspect of both the rat and the human. For rats it is fairly clear that the rat portion of the model of Yu et al. (1991) is the most appropriate becau se it is based on data, especially extensive information on lung burdens, from actual DP M exposures. The model provides for both deposition and integrated clearance for DPM as we ll as for two classes of adsorbed organics. The transport rates in the Yu model are deri ved directly from experiments with DPM exposed

rats. For humans, however, several models are ava ilable and discussed a bove, none of which is based on DPM-specific data. Deposition, but not cl earance, modules are ava ilable for all models, and Table 3-3 is an attempt to compare deposition projections of th e various models to the extent possible for particles in the range of characteris tics of size, distribution, and density of DPM. Intake parameters such as br eathing rates and minute volumes were also matched among the
various models. As alluded to above and shown in Table 3-3, DPM deposition is predicted to3-50 Table 3-3. Model comparison for deposi tion of DPM under equivalent conditions

Compartment Yu a ICRP66 b MPPDep1.11 c NCRP d
A (model designation) 13% (A) 14.1% (AI) 16.6% (P) 17.3% (P)

ET (model designation) 8% (H) 6% (ET 1 + ET 2 ) 8.7% (H) 6.6% (NOPL)

TB (model designation) 8% (TB) 4% (BB + bb) 7.2% (TB) 6.2% (TB)

Total 29% 24.1% 32.5% 30.1%

a Yu and Xu, 1987 (estimated from Figures 1 and 3).

b Jarvis et al., 1996.

c Freijer et al., 1999 (The Yeh-Schum 5-lobe and URT volume of 50 mL options were used.)

d NCRP, 1997. Note: Particle characteristics were set at 0.2 MMAD, 2. 4 sigma g, 1.5 shape factor (equivalent to 0.3 packing factor), density 1.5 and a concentration of 5 : g/m 3 . Lung parameters were set at 15 breaths per minute, a tidal volume of 0.926 L/hr, and a functional residual capacity (FRC) of 3300 mL.

occur in all regions of the respiratory tract but, because diffusion would be the most likely mechanism of deposition, is most prominent in the alveolar region. When run under equivalent conditions, all models show that higher deposition in the alveolar region is higher, generally by a factor of about 2, than the other regions of the re spiratory tract. The percentages projected by the different models to be deposited in the alveolar regions were all similar to one another with a range of only 13% for the Yu model to 17.3 % for the NCRP model. The total deposition of DPM-like particles predic ted by the models was also very similar at around 30%. Only the ICRP model differed appreciably from the others in to tal deposition by a factor of about 1.3 less at 22.9%. Due to its verity and co mpleteness in representation of the lung, the MPPDep model
could be considered the most th eoretically advanced of these deposition models and, presumably, the most accurate. It can be seen that, at least at the concentration tested, the Yu results and those of the MPPDep model could be judged very similar if not the same in the ET and TB regions, albeit with the MPPDep predicting slig htly more deposition in the A region. Based on this limited analysis, total and regional DPM deposition in the human respiratory tract predicted by the Yu model appear similar to other available human models. Further model comparison may be undertak en for those human models that have clearance as well as deposition modul es available; from Table 3-3, these include the Yu et al. (1991) and ICRP66 models. Therefore, the human lung burden outputs of these two models were compared under equivalent physiol ogical parameters, particle ch aracteristics, and duration (70 years) and concentrations of exposure (Figure 3-9).3-51 Figure 3-9. Modeled estimates of lung burden in humans after a simulated lifetime exposure to DPM using the Yu et al. (1991;[o]) and ICRP66 ( 

### 9 ) models. Simulations include both deposition and clearance. Simulations were run for 70 years using a respiratory frequency of 15 min -1 and a tidal volume of 0.926 L/breath for a total daily air intake of 20 m 3 /day for the various co ncentrations shown. Particle characteristics in the ICRP66 model, including MMAD, F g , density, and packing/shape factor were all matched to those used in the Yu model.

At DPM concentrations up to about 0.2 mg/m 3 , the outputs (lung burden) from these two models are essentially identical (see insert) indicating little if any difference between them in this concentration range. This observation is c onsonant with the minor differences noted in deposition (Table 3-3). Above 0.2 mg/m 3 DPM, both models continue to demonstrate a m onotonic increase in lung burden with increasing concentration. Howe ver, the output of the Yu et al. (1991) model begins to diverge markedly from the burdens pr edicted by the ICRP model such that the Yu model predicts a greater burden for a given con centration of DPM than does the ICRP66 model. This situation would be predic ted based on the assumption in th e human portion of the Yu model of a concentration-dependent m acrophage inhibition and particle overload occurring in humans; such an inhibition would result in impaired clearance processes, thereby allowing for a greater accumulation of material in the lung with increasing concentrations of DPM. This assumption is not made in the ICRP model, a nd materials are therefore not pred icted to accumulate in the lung to the extent predicted by the Yu model.3-52 Based on this limited analysis of models and the predictions from them for both deposition and clearance of DPM in humans, the model of Yu et al. (1991) can be seen to perform similarly to other available state-of-the-art models. The Yu model(s) are chosen for further analysis for the purposes of this document primarily because the animal portion of the model is based on DPM-specifi c data and the human compone nts of the model have both deposition and clearance capacities that do not appear different from other available human respiratory tract models.

3.6.2.6. Model Variability As demonstrated in Table 3-3 and Figure 3-9, there appears to be little variability among
state-of-the-art models availa ble for predicting disposition (bot h for deposition and for clearance integrated with deposition) of low levels of DPM (i.e., up to about 200 : g/m 3 ) in the respiratory
tracts of humans. Intersubject variability and its relationship to model output, however, is acknowledged in the ICRP model for deposition efficiencies (ICRP, 1994). Th is variability, recognized as substantial by ICRP, is addressed through use of scaling constants derived from estimates of the upper and lower confidence bounds for regional deposition efficiencies, with the scaling constants representing the variability in the population. It should be noted that the same philosophy is inherent in dose-response methodolog ies such as the RfC, where variability in the population is accommodated by a 10-fold uncertainty factor rather than by scaling constants. Inspection of data in ICRP66 (e.g ., Figures D-4 through D-7 in th e ICRP reference) on nasal and extrathoracic deposition in adu lt males shows that these uppe r and lower boundaries on output due to intersubject variability are considerably less than 10-fold different from one another. Thus, dividing model outputs by a factor of 10 such as is done in Rf C derivation may well be inclusive of not only intersubject variability but also of any model-to-model variability as they exist currently.

3.6.2.7. Model Comparison — Estimations of Deposition of Adsorbed Organics The data presented in Table 3-3 may be viewed as single-breath estimates of DPM deposition patterns in the various regions of the human lung under the breathing patterns and
conditions described in the table fo r the different models considered in this report. From these data it is possible to estimate the total mass of DPM deposited in the pulmonary region under a given set of conditions. Furthermore, if the fraction of organics present on DPM and their ability to be desorbed or eluted from the DPM are assumed also to be the same, then these deposition data could be used to estimate the dose of organics to pulmonary tissues. Such a comparison would not only yield an estimate of the amount of organics but also lend a further comparison3-53 between the different human models. This ex ercise was performed for humans breathing 5 : g DPM/m 3 continuously, and the results ar e presented in Table 3-4 below.

Table 3-4. Comparative model estimates of DP M deposition in human lungs from exposure to 5 : g/m 3 continuously for one year

AB C D

Human deposition model Alveolar Dep a : g DPM deposited/year b : g organics deposited/year c : g carcinogenic PAH deposited/year d

Yu et al. (1991) 13% 4745 598 1.82

ICRP66 14.1% 5147 649 1.98
MPPDep 16.6% 6059 763 2.33

NCRP 17.3% 6315 796 2.43

a Alveolar deposition fractions predicted for DPM (Yu et al ., 1991) and for particles with DPM characteristics (from Table 3-3). No clearance is included in this calculation.

b A total air intake of 20 m 3 /day is assumed. These numbers were obtained by factoring 20 m 3 × 5 : g DPM/m 3 × Alveolar deposition % (column A) × 365 days/year.

c In three samples of DPM extract, DPM-associated organics were noted as being 11.1%, 14.7%, and 12.1% wt. organics/wt. DPM (Tong and Karasek, 1984) with the average being 12.6%; column B is factored by this average to generate column C.

d Those seven PAHs identified as being carcinogenic either to humans or to animals (U.S. EPA, 1993) were summedfrom the data of Tong and Karasek (1984), where they are reported as a concentration in extract from DPMassociated organics. In three different samples, the content of these 7 PAHs was noted as 4739, 2054, and 2360 ng/mg of organic extract, with the average being 3051 ng/mg (3.051 : g/mg) organic extract. This average value was factored with Column C (in mg) to generate column D.

Note: Estimates from different human deposition models of the total amount of DPM-associated organics deposited in the pulmonary regions in humans breathing DPM at 5 : g/m 3 continuously for 1 year.

As may be expected, the relatively minor differences (17.3 % / 13% = 1.3) in the deposition of DPM among the different human models leads to similarly minor differences in projections of dose of carcinogenic PAHs to th e lung at a relatively low concentration of 5 : g/m 3 DPM. Somewhat unexpected is the small absolute quantity of carcinogenic PAH that may be delivered to the lung tissues under the conditions of exposure to DPM in this exercise. It should be noted that exercises similar to this have been carried out by others, e.g., Valberg and Watson (1999). However, the possibility that high concen trations of DPM may re sult in localized areas of deposition (such as the conducting airways) , the fact that human exposures may be3-54 considerably greater than those pr esupposed in the exercise (e.g., 5 : g/m 3 ), the nature of the assays (i.e., in vitro in Chapter 4 vs. actual inhalation exposures), and the findings that DNA adducts may result from other known noncarcinog ens such as carbon black (Bond et al., 1990) make the interpretation of such exercise s problematic and thei r meaning unclear.

3.7. SUMMARY AND DISCUSSION
The most consistent histor ical measure of exposure fo r DE is DPM in units of : g or mg particles/m 3 , with the underlying assumption that a ll components of diesel emissions (e.g., organics in the form of volatilized liquids or gases) are present in proportion to the DPM mass.
DPM is used as the basic dosimeter for effects fro m various scenarios such as chronic and acute exposures as well as for different endpoints such as irritation, fibr osis, or even cancer. There is, however, little evidence currently available to prove or refute DPM as being the most appropriate
dosimeter.
DPM dose to the tissue is related to the ex tent of the deposition and clearance of DPM. DPM may deposit throughout the resp iratory tract via sedimentati on or diffusion, with the latter being prevalent in the alveolar region. Partic les that deposit upon airway surfaces may be cleared from the respiratory tract completely or may be translocated to other sites by regionally distinct processes that can be categorized as either ab sorptive (i.e., dissolution) or nonabsorptive (i.e., transport of intact particles via mucociliary transport). Other mechanisms that can affect retention of DPM include endocyt osis by alveolar lining cells and interstitialization, which lead
to the accumulation of DPM in the interstitia l compartment of the lung and subsequent translocation of DPM to lymph nodes; inters titialization of poorly so luble particles may be prominent in primates and humans compared with rodents, althoug h different rates for this path could also explain observed results. For poorlysoluble particles such as DPM, speciesdependent rate consta nts exist for the various clearance pathwa ys that can be modified by factors such as respiratory tract disease. In rats, prolonged exposure to high concentrations of particles will result in particle overload, a condition that is defined as the overwhelming of macrophage -mediated clearance by the deposition of particles at a rate exceeding the capacity of that clearance pathway. This condition seems to begin to occur in rats wh en the pulmonary dust burden exceeds about 1 mg particles/g lung tissue. On the other hand, there is no clear evidence for particle overload in humans. Macrophage-mediated clear ance is slower in humans than in rats, and ki netics relating to interstitialization of poorly soluble particulate matter may have a greater consequence in
humans than in rats. The degree of bioavailability of the organic fraction of DPM is still somewhat uncertain. However, reports of DNA alterati ons in occupationally exposed workers, as well as results of3-55 animal studies using radiolabeled organics depos ited on DPM, indicate that at least a fraction of the organics present are eluted prior to particle clearance. Carcinogenic organics eluted in 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 interact directly with vital cellular components. The current information suggests that DPMassociated organics could be involved in a carcinogenic proce ss, although the quantitative data are far from adequate to ma ke any firm conclusions. Use of laboratory animal data in an assessment meant to be applied to humans obligates some form of interspecies extr apolation. Review and evaluati on of the considerable, specific
database in humans and animals on disposition of DPM, its adsorbed organics, and other poorly soluble particles led to the judgment that default options available for interspecies dosimetry adjustment could be set aside for more scientif ically valid, DPM-specific processes. Refinement of this process led to the evaluation of several applicable dosimetry models that in turn led to the identification and choice of the Yu et al. (1991) model to conduct inters pecies extrapolation. This model has a three-compartme nt lung consisting of trache obronchial, alveolar, and lymph node compartments. It treats DPM as being composed of the inso luble carbonaceous core, slowly cleared organics, and fast-cleared organi cs, and considers in an integrative manner the simultaneous processes of both deposition and clearance through empirical data derived from both laboratory animals and humans. Also, the model has 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 cl earance rate varying with the dose to the alveolar surface in the same propor tion in humans as in rats. Limitations of the
model include the lack of definitive information on variability and, quite possibly, the lack of a formal interstitial compartment that may be of consequence in humans. The basis of this model is to derive an internal dose from an external DPM concentration by utilizing species-specific physiological and pharmacokinetic pa rameters and, as such, is cons idered to have addressed the pharmacokinetic aspects of intersp ecies dosimetry. This aspect of the model a ddresses some of the critical data needs for the quantitative analys is of noncancer effects from DPM, the subject of Chapter 6. As parallels have been drawn between DPM and PM 2.5 in other chapters, it is perhaps appropriate to compare them also from the asp ect of dosimetry. Obvious comparisons include the nature of the particle dist ribution, defined artificially for PM 2.5 as compared with the thorough
characterization of DPM for both MMAD (which, at around 0.2 : m, is typically more than an order of magnitude less than the PM 2.5 cutoff and which, more properl y, should be termed a mass median thermodynamic diameter, an MMTD) and geometric standard de viation. It is clear that a3-56 larger portion of PM 2.5 particles than DPM w ould be above the aerodyna mic equivalent diameter (d
ae ) of 0.5 : m, which is often considered as a boundary between diffusion and aerodynamic mechanisms of deposition. This would imply th at a somewhat larger portion of DPM may pass on to the lower respiratory tract than would PM 2.5 . Alveolar deposition in humans specific for
DPM has been estimated with the Yu model to be 7%-13% (Yu and Xu, 1986), a figure that is consistent with deposition predic tions of other human models (s ee Table 3-3). This fractional deposition may be compared to one calculated for PM 2.5 and reported in U.S. EPA (1996a); assuming a MMAD of 2.25 : m and a geometric standard devi ation of 2.4, a fractional alveolar deposition of 10.2% was repor ted. This value is within the ra nge and quite comparable to that obtained by Yu and Xu (1986), indicating that little difference may exist in alveolar deposition
between DPM and PM 2.5 , at least for this assumed geometric standard deviation.

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The application of m u tagenicity data to the question of the potential carcinogenicity of
diesel em issions is based on the prem ise that genetic alterations are found in all cancers and that
several of the chem icals found in diesel em issions possess m u tagenic activity in a variety of genetic assays. These genetic alterations can be produced by gene m u tations, deletions, translocations, aneuploidy, or am plification of genes; hence no single genotoxicity assay should be expected to predict rodent carcinogenicity. Additionally, because of the inherent biological
differences of m easured endpoints, both within genotoxicity assays and between genotoxicity assays and cancer bioassays, a direct extrapol ation should not be expected (see Brusick [1987]
for a m o re detailed discussion). Indeed, m o st genotoxicity data are generated with in vitro assays that frequently em ploy concentrations of test agent that m a y be orders of m a gnitude greater than encountered in environm ental situations. W ith diesel em issions or other m i xtures, additional com p lications arise because of the com p lexity of the m a terial being tested. Since 1978, m o re than 100 publications have appeared in which genotoxicity assays were
used with diesel em issions, the volatile and particulate f r actions (including extracts), or
individual chem icals found in diesel em issions

. The Huisingh et al. (1978) report not only
identif ied m u tagenic activity in salm onella in several f r actions of diesel particular m a ter (DPM) extracts, but also indicated that the m u tagenic activity, especially quantitatively, was af f ected by
the extraction solvent as well as m e thod and length of storage. Much of the ensuing research

em ployed bioassays (m ost com m only salm onella TA98 without S9) to evaluate (1) extraction procedures, (2) f u el m odif i cations, (3) bioavailability of chem icals f r om DPM, and (4) exhaust
f ilters or other m odif i cations and other variables associated with diesel em issions. The interest
in the contribution of m u tagens to carcinogenicity was high in the early 1980s and the lack of long- term rodent carcinogenicity inform ation on diesel em issions led to the use of

sem i quantitative m u tagenicity (and in vitro cell transf orm a tion) data f r om diesel em issions and epidem iology based cancer potency estim ations to derive a com p arative potency estim ate for
diesel em issions first published by Albert et al . (1983) and m o re fully discussed in Appendix C of this report. As indicated in Chapter 2, the num ber of chem icals in diesel em issions is very large. Many of these have been determ ined to exhibit m u tagenic activity in a variety of assay system s (see Table II. in Claxton, 1983). Although a detailed discussion of those data is beyond the scope of this docum ent, som e of the m u tagenically active com pounds found in the gas phase are ethylene, benzene, 1,3-butadiene, acrolein and seve ral PAHs (see Table 2-21). Of the particleassociated chem icals, several PAHs and nitro-PAHs have been the focus of m u tagenic investigations both in bacteria and in m a m m a lian cell system s (see Table 2-22). Several review

4-1articles, som e containing m o re detailed descriptions of the available studies, are available (Claxton, 1983; Pepelko and Peirano, 1983; Interna tional Agency for Research on Cancer, 1989;
Shirnam é -Moré, 1995). Discussions of genotoxicity in the proceedings of several sym posia on
the health effects of diesel em issions (U.S. EPA, 1980; Lewtas, 1982; Ishinishi et al., 1986) are
also available.

4.1. GENE MUTATIONS Huisingh et al. (1978) dem onstrated that dichlorom e thane extracts from DPM were

m

u tagenic in strains TA1537, TA1538, TA98, and TA100 of S. typhimurium , both with and
without rat liver S9 activation. This report contained data from several fractions as well as DPM
from different vehicles and fuels. Sim ilar resu lts with diesel extracts from various engines and
f u els have been reported by a num ber of investigators using the salm onella f r am eshif t -sensitive strains TA1537, TA1538, and TA98 (Siak et al ., 1981; Claxton, 1981; Dukovich et al., 1981; Brooks et al., 1984). Sim ilarly, m u tagenic activity was observed in salm onella forward m u tation

assays m easuring 8-azaguanine resistance (Claxton and Kohan, 1981) and in E. coli m u tation assays (Lewtas, 1983). One approach to identifying significant m u tagens in chem ically com p lex environm ental

sam p les such as diesel exhaust or am bient particulate extracts is the com b ination of short-term bioassays with chem ical fractionation (Scheutzle and Lewtas, 1986). The analysis is m o st frequently carried out by sequential extraction with increasingly polar or binary solvents. Fractionation by silica-colum n chrom a tography separates com pounds by polarity or into acidic, basic, and neutral fractions. The resulting fractions are too com p lex to characterize by chem ical

m e thods, but the bioassay analysis can be used to determ ine fractions for further analysis. In

m o st applications of this concept, salm onella strain TA98 without the addition of S9 has been
used as the indicator for m u tagenic activity. Generally, a variety of nitrated polynuclear

arom atic com pounds have been found that account for a substantial portion of the m u tagenicity (Liberti et al., 1984; Schuetzle and Frazer, 1986; Sc huetzle and Perez, 1983). However, not all
bacterial m u tagenicity has been identif ied in this way, and the identity of the rem a ining

m

u tagenic com pounds rem a ins unknown. The nitrated arom atics thus far identified in diesel engine exhaust (DE) were the subject of review in the IARC m onograph on DE (International Agency for Research on Cancer, 1989). In addition to the sim p le qualitative identification of

m

u tagenic chem icals, several investigators have used num erical data to express m u tagenic activity as activity per distance driven or m a ss of fuel consum ed. These types of calculations
have been the basis for estim ates that the nitroarenes (both m ono- and dinitropyrenes) contribute a significant am ount of the total m u tagenic activity of the whole extract (Nishioka et al., 1982;
Salm een et al., 1982; Nakagawa et al., 1983). In a 1983 review, Claxton discussed a num ber of

4-2factors that affected the m u tagenic response in salm onella assays. Citing the data from the Huisingh et al. (1978) study, the author noted that the m u tagenic response could vary by a factor of 100 using different fuels in a single diesel e ngine. More recently, Crebelli et al. (1995) used
salm onella to exam ine the effects of different fuel com ponents. They reported that although

m

u tagenicity was highly dependent on arom atic content, especially di- or triarom a tics, there was
no clear effect of sulfur content of the fuel. Later, Sjögren et al. (1996) using m u ltivariate
statistical m e thods with ten diesel fuels conclude d that the m o st influential chem ical factors in
salm onella m u tagenicity were sulfur contents, certain PAHs (1-nitropyrene) and naphthenes. Matsushita et al. (1986) tested particle-free DE gas and a number of benzene nitroderivatives and polycyclic arom atic hydrocarbons (PAHs) (m any of which have been identified

as com ponents of DE gas). The particle-free exhaust gas was positive in both TA100 and TA98, but only without S9 activation. Of the 94 nitrobenzene derivatives tested, 61 were m u tagenic,
and the m a jority showed greatest activity in TA 100 without S9. Twenty-eight of 50 PAHs tested

were m u tagenic, all required the addition of S9 for detection, and m o st appeared to show a stronger response in TA100. W h en 1,6-dinitropyrene was m i xed with various PAHs or an extract of heavy-duty (HD) DE, the m u tagenic activity in TA98 was greatly reduced when S9 was absent but was increased significantly when S9 was present. These latter results suggested
that caution should be used in estim ating m u tagenicity (or other toxic effects) of com p lex

m i xtures from the specific activity of individual com ponents. Mitchell et al. (1981) reported m u tagenic activity of DPM extracts of diesel em issions in
the m ouse lym phom a L5178Y m u tation assay. Positive results were seen both with and without
S9 activation in extracts from several different vehicles, with m u tagenic activity only slightly lower in the presence of S9. These findings have been confirm e d in a num ber of other

m a m m a lian cell system s using several different genetic m a rkers. Casto et al. (1981), Chescheir et al. (1981), Li and Royer (1982), and Brooks et al. (1984) all reported positive responses at the
HPRT locus in Chinese ham s ter ovary (CHO) cells. Morim o to et al. (1986) used the APRT and
Oua r loci in CHO cells; Curren et al. (1981) used Oua r in BALB/c 3T3 cells. In all of these studies, m u tagenic activity was observed without S9 activation. Liber et al. (1981) used the thym idine kinase (TK) locus in the TK6 hum an lym phoblast cell line and observed induced

m

u tagenesis only in the presence of rat liver S9 when testing a m e thylene chloride extract of DE. Barfknecht et al. (1982) also used the TK6 assa y to identify som e of the chem icals responsible for this activation-dependent m u tagenicity. They suggested that fluoranthene, 1-

m e thylphenanthrene, and 9-m e thylphenanthren e could account for over 40% of the observed activity.
Morim o to et al. (1986) injected DPM extract s (250 to 4,000 m g /kg) into pregnant Syrian
ham s ters and m easured m u tations at the APRT locus in em bryo cells cultivated 11 days af ter i.p.

4-3injection. Although neutral fractions from both li ght-duty (LD) and HD particle extracts resulted
in increased m u tation frequency at 2,000 and 4,000 m g /kg, the response at 1,000 m g /kg was not
different from controls. Also, because the aut hors did not present data on toxicity or cloning
ef

f i ciency, the value of the apparent positive f i ndings at extrem ely high concentrations is
uncertain at best. Belisario et al. (1984) applied the Am es test to urine from Sprague-Dawley rats exposed to single applications of DPM ad m i nistered by gastric intubation, i.p. injection, or s.c. gelatin capsules. In all cases, dose-related increases were seen in TA98 (without and with S9) from urine concentrates taken 24 h after particle adm i nistration. Urine from Swiss m i ce exposed by inhalation to filtered exhaust (particle concentration 6 to 7 m g /m 3 ) for 7 weeks (Pereira et al., 1981a) or Fischer 344 rats exposed to DPM at a concentration of 1.9 m g /m 3 for 3

m onths to 2 years (Ong et al., 1985) was negative in salm onella strains.
Schuler and Niem eier (1981) exposed drosophila m a les in a stainless steel cham ber
connected to the 3 m 3 cham ber used for the chronic anim al studies at EPA (see Hinners et al., 1980 for details). Flies were exposed for 8 h and m a ted to untreated fem a les 2 days later. Although the frequency of sex-linked recessive leth als from treated m a les was not different from that of controls, the lim ited sam p le size precluded detecting less than a threefold increase over
controls. The authors noted that, because there were no signs of toxicity, the flies m i ght tolerate exposures to higher concentrations for longer tim e periods. Driscoll et al. (1996) exposed Fischer 344 m a le rats to aerosols of carbon black (1.1, 7.1,
and 52.8 m g /m 3 ) or air for 13 weeks (6 hr/day, 5 days/week) and m easured hprt m u tations in alveolar type II cells in anim als im m e diately after exposure and at 12 and 32 weeks after the end of exposure. Both of the two higher concentrations resulted in significant increases in m u tant frequency. W h ereas the m u tant frequency from the 7.1 m g /m 3 group returned to control levels by 12 weeks, the m u tant frequency of the high-exposure group was still higher than controls
even after 32 weeks. Carbon black particles have very little adsorbed PAHs, hence a direct
chem ically induced m echanism is highly unlikely. Induction of hprt m u tations were also observed in rat alveolar epithelial cells after intratracheal instillation with carbon black, 

### "

- quartz, and titanium dioxide (Driscoll et al., 1997). All three types of particles elicited an
inflam m a tory response as shown by significant increases of neutrophils in bronchoalveolar lavage (BAL) fluid. Culturing the BAL from exposed rats with a rat lung epithelial cell line
also resulted in elevation of hprt m u tational response. This response was effectively elim inated
when catalase was included in the incubation m i xture, providing evidence for cell-derived
oxidative dam a ge. Recently, Sato et al. (2000) exposed m a le Big Blue transgenic F344 rats to diluted DE (1 and 6 m g /m 3 suspended particle concentration) for 4 weeks. Mutant frequency in lung DNA was significantly elevated (4.8x control) at 6 m g /m 3 but not at 1 m g /m 3

. Lung DNA
adduct levels m easured by 32 P-postlabeling and 8-hydroxydeoxyguanosine m easured by HPLC

4-4were elevated at both particle concentrations, but to a lesser extent than m u tant frequencies. Sequence analysis of m u tants indicated that som e , but not all, of the m u tations could be explained by an oxidative dam a ge m echanism

. Specific-locus m u tations were not induced in (C3H × 101)F 1 m a le m i ce exposed to DE 8 h/day, 7 days/week for either 5 or 10 weeks (Russell et al., 1980). The exhaust was a 1:18 dilution and the average particle concentration was 6 m g /m 3

. After exposure, m a les were m a ted
to T-stock fem a les and m a tings continued for the reproductive life of the m a les. The results were unequivocally negative; no m u tants we re detected in 10,635 progeny derived from postsperm atogonial cells or in 27,917 progeny derived from sperm a togonial cells. Hou et al. (1995) m easured DNA adducts and hprt m u tations in peripheral lym phocytes
of 47 bus m a intenance workers and 22 control individuals. All were nonsm oking m e n from garages in the Stockholm area and the exposed group consisted of 16 garage workers, 25

m echanics, and 6 other garage workers. There were no exposure data, but the three groups were considered to be of higher to lower exposure to diesel engine exhaust. Levels of DNA adducts
determ ined by 32 P-postlabeling were significantly higher in workers than controls (3.2 versus 2.3
× 10 -8 ), but hprt m u tant frequencies were not different 8.6 versus 8.4 × 10 -6 ). Although group

m ean m u tant frequencies were not different, both adduct level and m u tagenicity were highest

am ong the 16 m o st exposed and m u tant frequency was significantly correlated with adduct level. All individuals were genotyped for glutathi one transferase GSTM1 and arom atic am ino transferase NAT2 polym orphism

. Neither GSTM 1 nulls nor NAT2 slow acetylators exhibited
effects on either DNA adducts or hprt m u tant frequencies.

4.2. CHROMOSOME EFFECTS Mitchell et al. (1981) and Brooks et al. (1984) reported increases in sister chrom a tid exchanges (SCE) in CHO cells exposed to DPM extracts of em issions from both LD and HD diesel engines. Morim o to et al. (1986) obs erved increased SCE from both LD and HD DPM
extracts in PAH-stim ulated hum an lym phocyte cultures. Tucker et al. (1986) exposed hum an peripheral lym phocyte cultures from four donors to direct DE for up to 3 h. Exhaust was cooled by pum ping through a plastic tube about 20 feet long; airflow was 1.5 L/m i n. Sam p les were taken at 16, 48, and 160 m i n of exposure. Cell cycle delay was observed in all cultures; significantly increased SCE levels were reported for two of the four cultures. Structural
chrom o som e aberrations were induced in CHO cells by DPM extracts from a Nissan diesel engine (Lewtas, 1983) but not by sim ilar extracts from an Oldsm obile diesel engine (Brooks et al., 1984). DPM dispersed in an aqueous m i xture containing dipalm itoyl lecithin (DPL), a

com ponent of pulm onary surfactant or extracted with dichlorom e thane (DCM) induced sim ilar

4-5responses in SCE assays in Chinese ham s ter V79 cells (Keane et al., 1991), m i cronucleus tests in V79 and CHO cells (Gu et al., 1992), and unsche duled DNA synthesis (UDS) in V79 cells (Gu et al., 1994). After separating the sam p les into supernatant and sedim e nt fractions, m u tagenic activity was confined to the sedim e nt fraction of the DPL sam p le and the supernatant of the
DCM sam p le. These findings suggest that the m u tagenic activity of DPM inhaled into the lungs
could be m a de bioavailable through solubilization a nd dispersion of pulm onary surfactants. In a later study in the sam e laboratory, Liu et al. (1996) found increased m i cronuclei in V79 cells treated with crystalline quartz and a noncrystalline silica, but response was reduced after pretreatm e nt of the particles with the sim u lated pulm onary surfactant. Pereira et al. (1981a) exposed fem a le Swiss m i ce to DE 8 h/day, 5 days/week for 1, 3,
and 7 weeks. The incidence of m i cronuclei and structural aberrations was sim ilar in bone

m a rrow cells of both control and exposed m i ce. Increased incidences of m i cronuclei, but not SCE, were observed in bone m a rrow cells of m a le Chinese ham s ters after 6 m onths of exposure to DE (Pereira et al., 1981b). Guerrero et al. (1981) observed a linear concentration-related increase in SCE in lung
cells cultured af ter intratracheal instillation of DPM at doses up to 20 m g /ham ster. However, they did not observe any increase in SCE after 3 m onths of inhalation exposure to DE particles (6 m g /m 3 ). Pereira et al. (1982) m easured SCE in em bryonic liver cells of Syrian ham s ters. Pregnant
fem a les were exposed to DE diluted with air 1:9 to contain about 12 m g /m 3 particles from days 5 to 13 of gestation or injected intraperitoneally with diesel particles or particle extracts on gestational day 13 (18 h before sacrifice). Neith er the incidence of SCE nor m itotic index was affected by exposure to DE. The injection of DPM extracts but not DPM resulted in a doserelated increase in SCE; however, the toxicity of the DPM was about twofold greater than the
DPM extract. In the only studies with m a m m a lian germ cells, Russell et al. (1980) reported no increase
in either dom inant lethals or heritable translocations in m a les of T-stock m i ce exposed by
inhalation to diesel em issions. In the dom inant lethal test, T-stock m a les were exposed for 7.5
weeks and im m e diately m a ted to fem a les of different genetic backgrounds (T-stock; [C3H × 101]; [C3H × C57BL/6]; [SEC × C 57BL/6]). There were no differences from controls in any of the param e ters m easured in this assay. For heritable translocation analysis, T-stock m a les were exposed for 4.5 weeks and m a ted to (SEC × C57BL/6) fem a les, and the F 1 m a les were tested f o r the presence of heritable translocations. Although no translocations were detected am ong 358 progeny tested, the historical control incidence is less than 1/1,000.

4-64.3. OTHER GENOTOXIC EFFECTS Pereira et al. (1981b) exposed m a le strain A m i ce to DE em issions for 31 or 39 weeks using the sam e exposure regim e n noted in the previous section. Analyses of caudal sperm for

sperm

- head abnorm a lities were conducted independently in three separate laboratories. Although the incidence of sperm abnorm a lities was not significantly above controls in any of the three laboratories, there were extrem ely large differences in scoring am ong the three (control values were 9.2%, 14.9%, and 27.8% in the three laboratories). Conversely, m a le Chinese
ham s ters exposed for 6 m o (Pereira et al., 1981c) exhibited alm o st a threefold increase in sperm

head abnorm a lities. It is noted that the control incidence in the Chinese ham s ters was less than
0.5%. Hence, it is not clear whether the differi ng responses reflect true species differences or experim e ntal artifacts.
A num ber of studies m easuring DNA adducts in anim als exposed to DPM, carbon black or other particles have been reported and are reviewed by Shirnam é -Moré (1995). Although

m odest increases in DNA adducts have been observed in lung tissue of rats after inhalation of DPM (W ong et al., 1986; Bond et al., 1990), the m a gnitude of the increases is sm all in

com p arison with those induced by chem ical carcinogens present in DE (Sm ith et al., 1993).
W h ile Gallagher et al. (1994) found no increases in total DNA adducts in lung tissue of rats exposed to DE, carbon black, or titanium dioxide they did observe an increase in an adduct with

m i gration properties sim ilar to nitrochrysene a nd nitro-benzo(a)pyrene adducts from diesel but
not carbon black or titanium dioxide exposures. The m a jority of the studies used the 32 Ppostlabeling assay to detect adducts. Although this m e thod is sensitive, chem ical identity of adducts can only be inferred if an adduct spot m i grates to the sam e location as a known prepared
adduct.
DNA adducts have also been m easured in hum ans occupationally exposed to DE. Distinct adduct patterns were found am ong garage workers occupationally exposed to DE when

com p ared to nonexposed controls (Nielsen and Autrup, 1994). Furtherm ore, the findings were concordant with the adduct patterns observed in groups exposed to low concentrations of PAHs
from com bustion processes. Hem m i nki et al. ( 1994) also reported significantly elevated levels of DNA adducts in lym phocytes from garage workers with known DE exposure com p ared with unexposed m echanics. Hou et al. (1995) found elevated adduct levels in bus m a intenance workers exposed to DE. Although no difference in m u tant frequency was observed between the groups, the adduct levels were significantly different (3.2 vs. 2.3 × 10 -8 ). Nielsen et al. (1996) reported significantly increased levels of three biom arkers (lym phocyte DNA adducts, hydroxyethylvaline adducts in hem oglobin, and 1-hydroxypyrene in urine) in DE-exposed bus garage workers.

4-7The role of oxidative dam a ge in causing m u tations has received increasing focus recently. More than 50 different chem icals have been studied in rodents usually m e asuring the
form ation of 8-hydroxydeoxyguanosine (8-OH-dG), a highly m u tagenic adduct (Loft et al., 1998). Increases in the m u tagenic DNA adduct 8-hydroxydeoxyguanosine were found in m ouse lung DNA after intratracheal instillation of dies el particles (Nagashim a et al., 1995). The response was dose dependent. Mice fed on a high-fat diet showed an increased response whereas the responses were partially reduced when the antioxidant, 

### $ -carotene, was included in the diet (Ichinose et al., 1997). Oxidative dam a ge also has been m easured in rat lung tissue after intratracheal instillation of quartz (Nehls et al., 1997) and in rat alveolar m acrophages after in
vitro treatm e nt with silica dust (Zhang et al., 2000). Arim oto et al. (1999) dem onstrated that
redissolved m e thanol extracts of DPM also induced the form ation of 8-OH-dG adducts in L120

m ouse cells. The response was dependent on both DPM concentration and P450 reductase. A detailed discussion of the potential role of oxidative dam a ge in DE carcinogenesis is presented in Chapter 7, Section 7.4.

4.4. SUMMARY AND DISCUSSION
Extensive studies with salm onella have unequivocally dem onstrated m u tagenic activity in both particulate and gaseous fractions of DE. In m o st of the studies using salm onella, DPM extracts and individual nitropyrenes exhibited the strongest responses in strain TA98 when no exogenous activation was provided. Gaseous fractions reportedly showed greater response in TA100, whereas benzo[ a ] pyrene and other unsubstituted PAHs are m u tagenic only in the presence of S9 fractions. The induction of gene m u tations has been reported in several in vitro

m a m m a lian cell lines after exposure to extracts of DPM. Note that only the TK6 hum an cell line did not give a positive response to DPM extracts in the absence of S9 activation. Mutagenic activity was recovered in urine from anim als treated with DPM by gastric intubation and i.p. and
s.c. im plants, but not by inhalation of DPM or dilu ted diesel exhaust. Dilutions of whole diesel exhaust did not induce sex-linked recessive lethals in drosophila or specific-locus m u tations in

m a le m ouse germ cells.
Structural chrom o som e aberrations and SCE in m a m m a lian cells have been induced by particles and extracts. W hole exhaust induced m i cronuclei but not SCE or structural aberrations
in bone m a rrow of m a le Chinese ham s ters exposed to whole diesel em issions for 6 m o. In a shorter exposure (7 weeks), neither m i cronuclei nor structural aberrations were increased in bone

m a rrow of fem a le Swiss m i ce. Likewise, whole DE did not induce dom inant lethals or heritable
translocations in m a le m i ce exposed for 7.5 and 4.5 weeks, respectively. The application of m u tagenicity data to the question of the potential carcinogenicity of
diesel em issions is based on the prem ise that genetic alterations are found in all cancers and that

4-8several of the chem icals found in diesel em issions possess m u tagenic activity in a variety of genetic assays. These genetic alterations can be produce by gene m u tations, deletions, translocations, aneuploidy, or am plification of genes, hence no single genotoxicity assay should be expected to either qualitatively or quantitatively predict rodent carcinogenicity. W ith diesel

em issions or other m i xtures, additional com p lications arise because of the com p lexity of the

m a terial being tested. Exercises that com b ined the salm onella m u tagenic potency with the total
concentration of m u tagenic chem icals deposited in the lungs could not account for the observed

tum o r incidence in exposed rats (Rosenkranz, 1993; Goldstein et al., 1998). However, such calculations ignored the contribution of gaseous phase chem icals which have been estim ated to
contribute from less than 50% (Rannug et al., 1983) to over 90% (Matsushita et al., 1986) of the
total m u tagenicity. This wide range is partly reflective of the differences in m a terial tested,

sem i volatile extracts in the f o rm er and whole gaseous em ission in the latter. Of greater im portance is that these calculations are based on a reverse m u tation assay in bacteria with

m e tabolic processes strikingly different from m a m m a ls. This is at least partly reflected in the observations that different nitro-PAHs give differe nt responses in bacteria and in CHO cells (Li and Dutcher, 1983) or in hum an hepatom a -derived cells (Eddy et al., 1986).

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4-135-1 5. NONCANCER HEALTH EFFECTS OF DIESEL EXHAUST

The objective of this chapter is to review and evaluate potential health effects other than cancer associated with inhalation exposure to diesel engine exhaust (DE). Data have been obtained from diverse human, laboratory animal, and in vitro test systems. The human studies comprise both occupational and human experimental exposures, the former consisting of exposure to DE in the occupational environment, and the latter consisting of exposure to diluted DE or diesel particulate matter (DPM) under controlled conditions. The laboratory animal studies consist of both acute and chronic exposures of laboratory animals to DE or DPM. Diverse in vitro test systems composed of human and laboratory animal cells treated with DPM or components of DPM have also been used to investigate the effects of DPM at the cellular and molecular levels. DPM mass (mg/m 3 ) has been used almost exclusively as a measure of DE exposure in human and experimental studies. The noncancer health effects of DPM have been reviewed previously by the Health Effects Institute (HEI, 1995) and in the Air Quality for Particulate Matter Criteria Document, the PM CD (U.S. EPA, 1996). The noncancer health effects attributable to ambient particulate matter (PM), which is composed in part of DPM, as well as the potential mechanisms underlying these effects have also been previously reviewed in the PM CD (U.S. EPA, 1996) and have been summarized in this document in Chapter 6, Section
6.4. This chapter begins with descriptions of studies that have shown various health effects occurring as a result of exposure to DE/DPM (Section 5.1). The human studies portion of this section (5.1.1) discusses results from both short-term and long-term studies as well as specialized studies such as those of populations contiguous to major highways (5.1.2). Studies using laboratory animals are ordered into various subsections under Section 5.1.3. Investigations devoted to elucidating the possible modes of action of DE/DPM are covered in Section 5.2; the mode-of-action issue of particle overload in animals is discussed elsewhere in the document (Chapter 3, Section 3.4). Section 5.3 describes evidence for the various
interactions of DPM with other conditions such as disease. Other sections address issues such as species-comparative responses to DE/DPM (Section 5.4) and influence of dose rate (Section 5.5). The summary/conclusion of this chapter, relating the totality of this information to possible human effects of DE/DPM, is in Section 5.6.5-2 5.1. HEALTH EFFECTS OF WHOLE DIESEL EXHAUST
5.1.1. Human Studies
5.1.1.1. Short-Term Exposures In a controlled human study, Rudell et al. (1990, 1994) exposed eight healthy subjects in an exposure chamber to diluted exhaust from a diesel engine for 1 h, with intermittent exercise. Dilution of the DE was controlled to provide a median NO 2 level of approximately 1.6 ppm. Median particle number was 4.3 × 10 6 /cm 3 , and median levels of NO and CO were 3.7 and 27 ppm, respectively (particle size and mass concentration were not provided). There were no effects on spirometry or on closing volume using nitrogen washout. Five of eight subjects experienced unpleasant smell, eye irritation, and nasal irritation during exposure. Bronchoalveolar lavage (BAL) was performed 18 hours after exposure and was compared with a control BAL performed 3 weeks prior to exposure; there was no control air exposure. Small but statistically significant reductions were seen in numbers of BAL mast cells, extent of AM phagocytosis of opsonized yeast particles, and lymphocyte CD4/CD8 ratios. A small increase in recovery of polymorphonuclear cells (PMNs) was also observed. These findings suggest that DE may induce mild airway inflammation in the absence of spirometric changes. This study provides an intriguing glimpse of the effect of DE exposure in humans, but only one exposure level was used, the number of subjects was low, and a limited range of endpoints was reported, so the data are inadequate to generalize about the human response. Rudell et al. (1996) exposed volunteers to DE for 1 h in an exposure chamber. Light work on a bicycle ergometer was performed during exposure. Exposures included either DE or exhaust with particle numbers reduced 46% by a particle trap. The engine used was a new Volvo model 1990, a six-cylinder direct-injection turbocharged diesel with an intercooler, which was run at a steady speed of 900 rpm during the exposures. Comparison of this study with others was difficult because neither exhaust dilution ratios nor particle concentrations were reported. Carbon monoxide concentrations of 27-30 ppm and NO of 2.6-2.7 ppm, however, suggested DPM concentrations may have equaled several mg/m 3 . The most prominent symptoms during exposure were irritation of the eyes and nose and an unpleasant smell. Both airway resistance and specific airway resistance increased significantly during the exposures. Despite the 46% reduction in particle numbers by the trap, effects on symptoms and lung function were not significantly attenuated. Nordenhall et al. (2000) had 15 healthy human subjects (13 males, 2 females) breathe in an exposure chamber diluted DE from an idling diesel engine to give a PM 10 concentration of
300 

### : g/m 3 , which was also associated with a median steady-state NO 2 concentration of 1.6 ppm.5-3 Exposures were for 1 h, with each individual serving as their own control by being exposed to filtered air, also for 1 h but at a different time. Sputum production was then induced and sputum examined at 6 and 24 hr postexposure (for both air and DPM) with differential cell counts and soluble protein counts performed. In comparing the same individual’s results after exposure to air and after exposure to DE, increases were found in the percentage of sputum neutrophiles (37.7% vs. 26.2%) after 6 hr, along with increases in concentrations of the soluble proteins interleukin-6 (12.0 vs. 6.3 pg/mL) and methylhistamine (0.11 vs. 0.12 ug/L). These differences between air and DPM were not present at 24 hr. Thus, breath exposure to DE produces early induction of an inflammatory response in healthy humans that can be detected using sputum analysis. Wade and Newman (1993) describe the situation of three railroad workers who developed persistent asthma associated with overexposure to DE from locomotives. The overexposure was a consequence of multiple hours of high levels of diesel exposure from riding in locomotive units trailing immediately behind the lead locomotive. Lines of evidence supporting railroad locomotive DE inducing asthma in these individuals include, (1) all three exhibited clear signs of asthma leading (in two of the three cases) to immediate first-time hospitalization and treatment for asthma, (2) all three developed symptoms within a few hours of the overexposure, and (3) all three experienced exacerbation of symptoms upon reexposure to locomotive DE. Although this report and that of Kahn et al. (1988) described below both provide supporting evidence for DE being able to cause asthma in humans under extreme but uncharacterized conditions, both suffer from the same limitations, including no reliable data on the concentration of diesel emissions and associated gaseous components, the duration of the exposures, or information on others that were exposed under these conditions but who did not develop asthma symptoms. Kahn et al. (1988) reported the occurrence of 13 cases of acute overexposure to DE among Utah and Colorado coal miners. Twelve miners had symptoms of mucous membrane irritation, headache, and lightheadedness. Eight individuals reported nausea; four reported a sensation of unreality; four reported heartburn; three reported weakness, numbness, and tingling in their extremities; three reported vomiting; two reported chest tightness; and two others reported wheezing. Each miner lost time from work because of these symptoms, which resolved within 24 to 48 h. No air monitoring data were presented; poor work practices were described as the predisposing conditions for overexposure. No follow-up was available for these exposed
individuals. El Batawi and Noweir (1966) reported that among 161 workers from two garages where diesel-powered buses were serviced and repaired, 42% complained of eye irritation, 37% of headaches, 30% of dizziness, 19% of throat irritation, and 11% of cough and phlegm. Ranges of5-4 mean concentrations of DE components in the two diesel bus garages were as follows: 0.4 to 1.4 ppm NO 2 , 0.13 to 0.81 ppm SO 2 , 0.6 to 44.1 ppm aldehydes, and 1.34 to 4.51 mg/m 3 of DPM; the highest concentrations were obtained close to the exhaust systems of the buses. Eye irritation was reported by Battigelli (1965) in six subjects after 40 s of chamber exposure to diluted DE containing 4.2 ppm NO 2 , 1 ppm SO 2 , 55 ppm CO, 3.2 ppm total hydrocarbons, and 1 to 2 ppm total aldehydes; after 3 min and 20 s of exposure to diluted DE containing 2.8 ppm NO 2 , 0.5 ppm SO 2 , 30 ppm CO, 2.5 ppm total hydrocarbons, and <1 to 2 ppm total aldehydes; and after 6 min of exposure to diluted DE containing 1.3 ppm NO 2 , 0.2 ppm SO 2 , <20 ppm CO, <2.0 ppm total hydrocarbons, and <1.0 ppm total aldehydes. The concentration of DPM was not reported. 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 complained on three occasions of minor eye and throat irritation, no correlation was established with concentrations of any particular component of DE. The role of radicals generated from particulate matter, including DPM, in producing toxicity has been discussed in the literature (Valavanidis et al., 2000), as has the role of antioxidant defenses in protecting against species such as radicals that may arise from acute DE exposure. Blomberg et al. (1998) investigated changes in the antioxidant defense network within the respiratory tract lining fluids of human subjects following DE exposure. Fifteen healthy, nonsmoking, asymptomatic subjects were exposed to filtered air or DE (DPM 300


### : g/m 3 ) for 1 h on two separate occasions at least 3 weeks apart. Nasal lavage fluid and blood samples were collected prior to, immediately after, and 5 1⁄2 h post exposure. Bronchoscopy was performed 6 h after the end of DE exposure. Nasal lavage ascorbic acid concentration increased tenfold during DE exposure, but returned to basal levels 5.5 h postexposure. DE had no significant effects on nasal lavage uric acid or GSH concentrations, and did not affect plasma, bronchial wash, or bronchoalveolar lavage antioxidant concentrations, nor malondialdehyde or protein carbonyl concentrations. The authors concluded that the physiological response to acute DE exposure is an acute increase in the level of the antioxidant ascorbic acid in the nasal cavity.

5.1.1.1.1. Diesel exhaust odor . The odor of DE is considered by most people to be objectionable; at high intensities, it may produce sufficient physiological and psychological effects to warrant concern for public health. The intensity of the odor of DE is an exponential function of its concentration such that a tenfold change in the concentration will alter the intensity of the odor by one unit. Two human panel rating scales have been used to measure DE odor intensity. In the first (Turk, 1967), combinations of odorous materials were selected to simulate DE odor; a set of 12 mixtures, each having twice the concentration of that of the5-5 previous mixture, is the basis of the diesel odor intensity scale (D-scale). The second method is the TIA (total intensity of aroma) scale based on seven steps, ranging from 0 to 3, with 0 being undetectable, 1⁄2 very slight, and 1 slight and increasing in one-half units up to 3, strong (Odor Panel of the CRC-APRAC Program Group on Composition of Diesel Exhaust, 1979; Levins, 1981). Surveys, utilizing volunteer panelists, have been taken to evaluate the general public’s response to the odor of DE. Hare and Springer (1971) and Hare et al. (1974) found 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, slight to moderate odor intensity), about 95% perceived the odor, and 75% objected to it, and, at a D rating of 5 (TIA = 1.8, almost moderate), about 95% objected to it. Linnell and Scott (1962) reported odor threshold measurement in six subjects and found that the dilution factor needed to reach the threshold ranged from 140 to 475 for this small sample of people. At these dilutions, the concentrations of formaldehyde ranged from 0.012 to 0.088 ppm.

5.1.1.1.2. Pulmonary and respiratory effects . Battigelli (1965) exposed 13 volunteers to three dilutions of DE obtained from a one-cylinder, four-cycle, 7-hp diesel engine (fuel type unspecified) and found that 15-min to 1-h exposures had no significant effects on pulmonary resistance. Pulmonary resistance was measured by plethysmography utilizing the simultaneous recording of esophageal pressure and airflow determined by electrical differentiation of the volume signal from a spirometer. The concentrations of the constituents in the three diluted exhausts were 1.3, 2.8, and 6.2 ppm NO 2 ; 0.2, 0.5, and 1 ppm SO 2 ; <20, 30, and 55 ppm CO; and <1.0, <1 to 2, and 1 to 2 ppm total aldehydes, respectively. DPM concentrations were not reported. A number of studies have evaluated changes in pulmonary function occurring over a workshift in workers occupationally exposed to DE (specific time period not always reported but assumed to be 8 h). In a study of coal miners, Reger (1979) found that both forced expiratory volume in 1 s (FEV 1 ) and forced vital capacity (FVC) decreased by 0.05 L in 60 diesel-exposed miners, an amount not substantially different from reductions seen in non-diesel-exposed miners (0.02 and 0.04 L, respectively). Decrements in peak expiratory flow rates were similar between diesel and non-DE-exposed miners. Although the monitoring data were not reported, the authors stated that there was no relationship between the low concentrations of measured respirable dust
or NO 2 (personal samplers) when compared with shift changes for any lung function parameter measured for the diesel-exposed miners. In summary, this study (available as an abstract only) states that no evidence was found for additional lung function effect over a shift for miners5-6 exposed to diesel emissions as compared with controls, i.e., nonexposed office workers and coal miners not exposed to diesel emissions. Ames et al. (1982) compared the pulmonary function of 60 coal miners exposed to DE with that of a control group of 90 coal miners not exposed to DE for evidence of acute respiratory effects associated with exposure to DE. Changes over the workshift in FVC, FEV 1 , and forced expiratory flow rate at 50% FVC (FEF 50 ) were the indices for acute respiratory effects. The environmental concentrations of the primary pollutants were 2.0 mg/m 3 respirable dust (<10 μm MMAD), 0.2 ppm NO 2 , 12 ppm CO, and 0.3 ppm formaldehyde. The investigators reported a statistically significant decline in FVC and FEV 1 over the workshift in both the diesel-exposed and comparison groups. Current smokers had greater decrements in FVC, FEV 1 , and FEF 50 than did ex-smokers and nonsmokers. There was a marked disparity between the ages and the time spent underground for the two study groups. Diesel-exposed miners were about 15 years younger and had worked underground for 15 fewer years (4.8 versus 20.7 years) than miners not exposed to DE. The significance to the results of these differences between the populations is difficult to ascertain. Except for the expected differences related to age, 120 underground iron ore miners exposed to DE had no workshift changes in FVC and FEV 1 when compared with 120 matched surface miners (Jörgensen and Svensson, 1970). Both groups had equal numbers (30) of smokers and nonsmokers. The frequency of bronchitis was higher among underground workers, much higher among smokers than nonsmokers, and also higher among older than younger workers. The authors reported that the underground miners had exposures of 0.5 to 1.5 ppm
NO 2 and between 3 and 9 mg/m 3 particulate matter, with 20% to 30% of the particles <5 μm MMAD. The majority of the particles were iron ore; quartz was 6% to 7% of the fraction <5 μm MMAD. Gamble et al. (1979) measured preshift FEV 1 and FVC in 187 salt miners and obtained peak flow forced expiratory flow rates at 25%, 50%, and 75% of FVC (FEF 25 , FEF 50 , or FEF 75 ). Postshift pulmonary function values were determined from total lung capacity and flows at preshift percentages of FVC. The miners were exposed to mean NO 2 levels of 1.5 ppm and mean respirable particulate levels of 0.7 mg/m 3 . No statistically significant changes were found between changes in pulmonary function and in NO 2 and respirable particles combined. Slopes of the regression of NO 2 and changes in FEV 1 , FEF 25 , FEF 50 , and FEF 75 were significantly different from zero. The authors concluded that these small reductions in pulmonary function
were attributable to variations in NO 2 within each of the five salt mines that contributed to the
cohort. Gamble et al. (1987a) investigated the acute effects of DE in 232 workers in four diesel bus garages using an acute respiratory questionnaire and before and after workshift spirometry.5-7 The prevalence of burning eyes, headaches, difficult or labored breathing, nausea, and wheeze experienced at work was higher in the diesel bus garage workers than in a comparison population of lead/acid battery workers who had not previously shown a statistically significant association of acute symptoms with acid exposure. Comparisons between the two groups were made without adjustment for age and smoking. There was no detectable association of exposure
to NO 2 (0.23 ppm ± 0.24 S.D.) or inhalable (less than 10 μm MMAD) particles (0.24 mg/m 3 ± 0.26 S.D.) and acute reductions in FVC, FEV 1 , peak flows, FEF 50 , and FEF 75 . Workers who had respiratory symptoms had slightly greater but statistically insignificant reductions in FEV 1 and
FEF 50 . Ulfvarson et al. (1987) evaluated workshift changes in the pulmonary function of 17 bus garage workers, 25 crew members of two types of car ferries, and 37 workers on roll-on/roll-off ships. The latter group was exposed primarily to DE; the first two groups were exposed to both gasoline and DE. The diesel-only exposures that averaged 8 h consisted of 0.13 to 1.0 mg/m 3 particulate matter, 0.02 to 0.8 mg/m 3 (0.016 to 0.65 ppm) NO, 0.06 to 2.3 mg/m 3 (0.03 to 1.2 ppm) NO 2 , 1.1 to 5.1 mg/m 3 (0.96 to 4.45 ppm) CO, and up to 0.5 mg/m 3 (0.4 ppm) formaldehyde. The largest decrement in pulmonary function was observed during a workshift following no exposure to DE for 10 days. Forced vital capacity and FEV 1 were significantly reduced over the workshift (0.44 L and 0.30 L, p <0.01 and p <0.001, respectively). There was no difference between smokers and nonsmokers. Maximal midexpiratory flow, closing volume expressed as the percentage of expiratory vital capacity, and alveolar plateau gradient (phase 3) were not affected. Similar but less pronounced effects on FVC (



### ! 0.16 L) were found in a second, subsequent study of stevedores (n = 24) only following 5 days of no exposure to diesel truck exhaust. Pulmonary function returned to normal after 3 days without occupational exposure to DE. No exposure-related correlation was found between the observed pulmonary effects and concentrations of NO, NO 2 , CO, or formaldehyde; however, it was suggested that
NO 2 adsorbed onto the DE particles may have contributed to the overall dose of NO 2 to the lungs. In a related study, six workers (job category not defined) were placed in an exposure chamber and exposed to diluted DE containing 0.6 mg/m 3 DPM and 3.9 mg/m 3 (2.1 ppm) NO 2 . The exhaust was generated by a 6-cylinder, 2.38-L diesel engine, operated for 3 h and 40 min at constant speed, equivalent to 60 km/h, and at about one-half full engine load. No effect on pulmonary function was observed. In a hypothesis-generating study, Kilburn (2000) examined neurobehavioral and pulmonary function of a small group of workers exposed to DE either as railroad workers (n=10) over a range of 15 to 50 years or as electricians (n=6) over a range of 0.6 to 1.5 years. Neurobehavioral and visual functions batteries showed nearly all of these individuals to be neurobehaviorally impaired in relation to a referent population in one or more areas, including5-8 reaction time, balance, blink reflex latency, verbal recall, and color vision confusion indices. Pulmonary function tests also showed that 10 of the 16 had airway obstruction and another group of 10 of the 16 had chronic bronchitis, chest pain, tightness, and hyperreactive airways. This work implies that with sufficiently sensitive methods, noncancer effects from DPM/DE exposure may be detectable in sufficiently exposed human populations.

5.1.1.1.3. Immunological effects . Salvi et al. (1999) exposed healthy human subjects to diluted DE (DPM 300 μg/m 3 ) for 1 h with intermittent exercise. Although there were no changes in pulmonary function, there were significant increases in neutrophils and B lymphocytes as well as histamine and fibronectin in airway lavage fluid. Bronchial biopsies obtained 6 h after DE exposure showed a significant increase in neutrophils, mast cells, and CD4+ and CD8+ T lymphocytes, along with upregulation of the endothelial adhesion molecules ICAM-1 and VCAM-1 and increases in the number of LFA-1+ in the bronchial tissue. Significant increases in neutrophils and platelets were observed in peripheral blood following exposure to DE. In a follow-up investigation of potential mechanisms underlying the DE-induced airway leukocyte infiltration, Salvi et al. (2000) exposed healthy human volunteers to diluted DE, on two separate occasions for 1 h each, in an exposure chamber. Fiber-optic bronchoscopy was performed 6 h after each exposure to obtain endobronchial biopsies and bronchial wash (BW) cells. These workers observed that DE exposure enhanced gene transcription of IL-8 in the bronchial tissue and BW cells and increased growth-regulated oncogene-a protein expression and IL-8 in the bronchial epithelium; there was also a trend toward an increase in IL-5 mRNA gene transcripts in the bronchial tissue.In an attempt to evaluate the potential allergenic effects of DPM in humans, DiazSanchez and associates carried out a series of clinical investigations. In the first of these (DiazSanchez et al., 1994), healthy human volunteers were challenged by spraying either saline or 0.30 mg (300 μg) DPM into their nostrils. The authors considered this dose to be equivalent to breathing the outdoor air in Los Angeles for a 24-h period on an average day. Enhanced IgE levels were noted in nasal lavage cells in as little as 24 h, with peak production observed 4 days after DPM challenge. The effects seemed to be somewhat isotype-specific, because in contrast to IgE results, DPM challenge had no effect on the levels of IgG, IgA, IgM, or albumin. Theselective enhancement of local IgE production was demonstrated by a dramatic increase in IgEsecreting cells. Although direct effects of DPM on B-cells have been demonstrated by in vitro studies, it was considered likely that other cells regulating the IgE response may also be affected. Cytokine production was therefore measured in nasal lavage cells from healthy human volunteers challenged with DPM (0 or 0.15 mg in 200 μL saline) sprayed into each nostril (Diaz-5-9 Sanchez et al., 1996). Before challenge with DPM, most subjects’ nasal lavage cells had detectable levels of only interferon- 

### ( , IL-2, and IL-13 m RNA. After challenge with DPM, the cells produced readily detectable levels of m RNA for IL-2, IL-4, IL-5, IL-6, IL-10, IL-13, and
interferon- 

### ( . Although the cells in the nasal lavage before and after challenge do not necessarily represent the same ones either in number or type, the broad increase in cytokine production was considered by the authors not to be simply the result of an increase in T cells recovered in the lavage fluid. On the basis of these findings, the authors concluded that the increase in nasal cytokine expression after exposure to DPM can be predicted to contribute to enhanced local IgE production and thus play a role in pollutant-induced airway disease. The ability of DPM to act as an adjuvant to the ragweed allergen Amb a I was also examined by nasal provocation in ragweed-allergic subjects using 0.3 mg (300 μg) DPM, Amb aI, or both (Diaz-Sanchez et al., 1997). Although allergen and DPM each enhanced ragweedspecific IgE, DPM plus allergen promoted a 16-times greater antigen-specific IgE production. Nasal challenge with DPM also influenced cytokine production. Ragweed challenge resulted in a weak response, DPM challenge caused a strong but nonspecific response, and allergen plus DPM caused a significant increase in the expression of mRNA for TH0 and TH2-type cytokines (IL-4, IL-5, IL-6, IL-10, IL-13), with a pronounced inhibitory effect on IFN- 

### ( gene expression. The author concluded that DPM can enhance B-cell differentiation and, by initiating and elevating IgE production, may be a factor in the increased incidence of allergic airway disease. In a further extension of these studies, Diaz-Sanchez et al. (1999) examined the potential for DPM to lead to primary sensitization of humans by driving a de novo mucosal IgE response to a neoantigen, keyhole limpet hemocyanin (KLH). Ten atopic subjects were given an initial nasal immunization of KLH followed by two biweekly nasal challenges with KLH. Fifteen different atopic subjects were treated identically, except that DPM was administered 24 h beforeeach KLH exposure. Intranasal administration of KLH alone led to the generation of an antiKLH IgG and IgA humoral response, which was detected in nasal fluid samples. No anti-KLH IgE was observed in any of these subjects. In contrast, when challenged with KLH preceded by DPM, 9 of the 15 subjects produced anti-KLH-specific IgE. KLH-specific IgG and IgA at levels similar to those seen with KLH alone were also detected. Subjects who received DPM and KLH had significantly increased IL-4, but not IFN-gamma, levels in nasal lavage fluid, whereas these levels were unchanged in subjects receiving KLH alone. These investigators concluded that DPM can function as a mucosal adjuvant to a de novo IgE response and may increase allergic sensitization among atopic individuals.

5.1.1.1.4. Human cell culture studies

. The potential mechanisms by which DPM may act to cause allergenic effects has been examined in human cell culture studies. Takenaka et al. (1995)5-10 reported that DPM extracts enhanced IgE production from purified human B cells. IgE production in these cells (stimulated by exogenous addition of interleukin-4 plus monoclonal antibody) was enhanced (i.e., further stimulated) 20% to 360% by the addition of DPM extracts (1-50 ng/mL) over a period of 10-14 days. DPM extracts in the absence of exogenously added IL-4 and/or monoclonal antibodies did not themselves induce IgE production or synergize with interleukin-4 alone to induce IgE from purified B cells, suggesting 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 increasing incidence of allergic airway disease. Terada et al. (1997) examined the effects of DPM and DPM extract on eosinophil adhesion, survival rate, and degranulation. Eosinophils, human mucosal microvascular endothelial cells (HMMECs), and human nasal epithelial cells (HNECs) were preincubated in the presence of DPM and DPM extract. 35S-labeled eosinophils were allowed to adhere to monolayers of HMMECs and HNECs. Although neither DPM nor DPM extract affected the adhesiveness of HMMECs and HNECs to eosinophils, DPM and DPM extract each significantly increased eosinophil adhesiveness to HNECs; neither affected eosinophil adhesiveness to HMMECs. DPM extract also induced eosinophil degranulation without changing the eosinophil survival rate. These results indicate that DPM may play an important role in promoting the nasal hypersensitivity induced by enhanced eosinophil infiltration of epithelium and eosinophil degranulation. It should also be noted that eosinophils are major components of allergic inflammatory disorders, including asthma and nasal allergy. Terada et al. (1999) examined the effects of DPM extract on the expression of histamineH1 receptor (H1R) mRNA in HNECs and HMMECs, and on the production of IL-8 and GMCSF induced by histamine. HNECs and HMMECs, isolated from human nasal mucosa specimens, were cultured with DPM extract. DPM extract increased the expression of H1R mRNA in both HNECs and HMMECs. The amount of IL-8 and GM-CSF induced by histamine was also significantly higher in HNECs and HMMECs treated with DPM extract. These results strongly suggest that DPM accelerates the inflammatory change by not only directly upregulating H1R expression but also by increasing histamine-induced IL-8 and GM-CSF production. Histamine is the most important chemical mediator in the pathogenesis of nasal allergy. Steerenberg et al. (1998) studied the effects of exposure to DPM on airway epithelial cells, the first line of defense against inhaled pollutants. Cells from a human bronchial cell line (BEAS-2B) were cultured in vitro and exposed to DPM (0.04-0.33 mg/mL) and the effects on IL-6 and IL-8 production were observed. Increases in IL-6 and IL-8 production compared to the5-11 nonexposed cells (11- and 4-fold, respectively) were found after 24 or 48 h exposure to DPM. This increase was lower (17- and 3.3-fold) compared to silica and higher compared to titanium dioxide, which showed no increase for either IL-6 or IL-8. The study was extended to observe the effects of DPM on inflammation-primed cells. BEAS-2B cells were exposed to TNF- 

### " followed by DPM. Additive effects on IL-6 and IL-8 production by BEAS-2B cells were found
after TNF- 

### " priming and subsequent exposure to DPM only at a low dose of DPM and TNF- 

### " (0.05-0.2 ng/mL). The investigators concluded that BEAS-2B phagocytized DPM and produced an increased amount of IL-6 and IL-8, and that in TNF- 

### " -primed BEAS-2B cells DPM increased interleukin production only at low concentrations of DPM and TNF- 

### "

. Ohtoshi et al. (1998) studied the effect of suspended particulate matter (SPM), obtained from high-volume air samplers, and DPM obtained from exhaust of a stationary diesel engine on the production of IL-8 and granulocyte-colony stimulating factor (GM-CSF) by human airway epithelial cells in vitro. Nontoxic doses of DPMs stimulated production of IL-8 and GM-CSF by three kinds of human epithelial cells (nasal polyp-derived upper airway, normal bronchial, and transformed bronchial epithelial cells) in a dose- and time-dependent fashion at a DPM concentration as low as 10 μg/mL. SPM applied at 250 and 2,500 μg/mL had a stimulatory effect on GM-CSF, but not on IL-8 production. The effects could be blocked with a protein synthesis inhibitor, suggesting that the process required de novo protein synthesis, and appeared to be due to an extractable component because neither charcoal nor graphite showed such stimulatory effects. The authors concluded that SPM and DPM, a component of SPM, may be important air pollutants in the activation of airway cells for the release of cytokines relevant to allergic airway inflammation. The mechanisms underlying DPM-induced injury to airway cells were investigated in human bronchial epithelial cells (HBECs) in culture (Bayram et al., 1998a). HBECs from bronchial explants obtained at surgery were cultured and exposed to DPM (10-100 μg/mL) suspended in a serum-free supplemented medium (SF-medium) or to a SF-medium filtrate of DPM. The filtrate was obtained by incubating DPM (50 μg/mL) in SF-medium for 24 h. The effects of DPM and DPM filtrate on permeability, ciliary beat frequency (CBF), and release of inflammatory mediators were observed. DPM and filtered solution of DPM significantly
increased the electrical resistance of the cultures but did not affect movement of bovine serum albumin across cell cultures. DPM and filtered DPM solution significantly attenuated the CBF of these cultures and significantly increased the release of IL-8. DPM also increased the release by these cultures of GM-CSF and soluble intercellular adhesion molecule-1 (sICAM-1). These authors also observed that activated charcoal was not able to induce changes in electrical resistance, attenuate CBF, and increase the release of inflammatory mediators from HBEC, and proposed that these effects were due most likely to the compounds adsorbed onto the DPM5-12 rather than the size of DPM. The authors concluded that exposure of airway cells to DPM may lead to functional changes and release of proinflammatory mediators and that these effects may influence the development of airway disease. Bayram et al. (1998b) investigated the sensitivity of cultured airway cells from asthmatic patients to DPM. Incubation with DPM (10-100 μg/mL) significantly attenuated the CBF in both the asthmatic and nonasthmatic bronchial epithelial cell cultures. Cultured airway cells from asthmatic patients constitutively released significantly greater amounts of IL-8, GM-CSF, and sICAM-1 than cell cultures from nonasthmatic subjects. Only cultures from asthmatic patients additionally released RANTES. The authors concluded that cultured airway cells from asthmatic subjects differ with regard to the amounts and types of proinflammatory mediators they can release and that the increased sensitivity of bronchial epithelial cells of asthmatic subjects to DPM may result in exacerbation of their disease symptoms. Devalia et al. (1999) investigated the potential sensitivity of HBECs biopsied from atopic mild asthmatic patients and non-atopic nonasthmatic subjects to DPM. HBECs from asthmatic 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 patients. Incubation of the asthmatic cultures with 10 

### : g/mL DPM significantly increased the release of IL-8, GM-CSF, and sICAM-1 after 24 h. In contrast, only higher concentrations (50-
100 

### : g/mL DPM) significantly increased the release of IL-8 and GM-CSF from HBECs of nonasthmatics. The authors conclude that the increased sensitivity of the airways of asthmatics to DPM may be, at least in part, a consequence of greater constitutive and DPM-induced release of specific pro-inflammatory mediators from bronchial epithelial cells. Abe and co-workers have demonstrated formation of increased cytokine levels in cultured human bronchial epithelial cells exposed to freshly generated DE, but not to filtered DE, i.e., particle-free DE (Abe et al., 2000). Cytokine IL-8 protein as well as transforming growth factor (TGF)- 

### $ 1 mRNAs were induced in a time-dependent manner (from 0.5 to 14 h of exposure) in BET-1A human bronchial epithelial cells in response to exposure to freshly generated, cooled, humidified DE that was diluted to 2.9 mg DPM/m 3 . The gas obtained by filtration of DE alone did not show any sustained increase in these indicators, suggesting that DE particles play a more important role in eliciting these responses than do the accompanying gases (10.6 ppm CO, 7.3 ppm NO 2 , and 3.3 ppm SO 2 ). To elucidate the intracellular signal transduction pathway regulating IL-8 and RANTES production, Hashimoto et al. (2000) examined the role of p38 mitogen-activated protein (MAP) kinase in DPM-induced (DPM = 10, 50, or 100 

### : g/mL) IL-8 and RANTES production by HBECs. They also examined the effect of a thiol-reducing agent, N-acetylcysteine (NAC), on DPM-induced p38 MAP kinase activation and cytokine production. The authors conclude that5-13 p38 MAP kinase plays an important role in the DPM-activated signaling pathway that regulatesIL-8 and RANTES production by HBECs and that the cellular redox state is critical for DPMinduced p38 MAP kinase activation leading to IL-8 and RANTES production. Boland et al. (1999) compared the biological effects of carbon black and DPM (2.5


### : g/cm 2 culture surface) 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 DPM was phagocytosed by epithelial cells and translocated through the epithelial cell sheet. The time and dose dependency of phagocytosis and its nonspecificity for different particles (DPM, carbon black, and latex particles) were established by flow cytometry. DPM also induced a time-dependent increase in interleukin-8, GM-CSF,
and interleukin-1 

### $ release. The inflammatory response occurred later than phagocytosis and, because carbon black had no effect on cytokine release, its extent appeared to depend on the content of adsorbed organic compounds. Furthermore, treatment of the exhaust gas to decrease the adsorbed organic fraction reduced the DPM-induced increase in GM-CSF factor release. These results indicate that DPM can be phagocytosed by and induce a specific inflammatory response in airway epithelial cells.

5.1.1.1.5. Summary

. In the available exposure studies, considerable variability is reported in DE detection threshold. The odor scales described in some of these studies have no general use at present because they are not objectively defined; however, the studies do clearly indicate substantial interindividual variability in the ability to detect odor and the level at which it becomes objectionable. Much of what is known about the acute effects of DE comes from case reports that lack clear measurements of exposure concentrations. The studies of pulmonary function changes in exposed humans have looked for changes occurring over a workshift or after a short-term exposure. The overall conclusion of these studies is that reversible changes in pulmonary function in humans can occur in relation to DE exposure, although it is not possible to relate these changes to specific exposure levels. Numerous studies described in this section, conducted in humans and in isolated cell systems derived from humans exposed to DPM, revealed various biochemical and pathophysiological alterations, such as IgE changes, altered levels of cytokines/chemokines, and goblet-cell hyperplasia, with nearly all these responses being key changes and markers of allergic inflammatory disorders of the airways such as asthma and nasal allergies (Nel et al., 1998). Thus, a major point of significance about these findings is that they indicate that DPM could be viewed as having the potential to elicit inflammatory and immunological responses and responses typical of asthma, and that DPM may be a likely factor in the increasing incidence of allergic hypersensitivity. These studies have also shown that effects are due primarily to the organic fraction and that DPM enhances the allergic response to5-14 known allergens. Results from these studies, including those wih laboratory animals, indicate that DPM could be viewed as having the potential to influence the development of airway inflammation and disease through its adjuvant properties and by causing the release of proinflammatory mediators.

5.1.1.2. Long-Term Exposures Several epidemiologic studies have evaluated the effects of chronic exposure to DE on occupationally exposed workers. Battigelli et al. (1964) measured several indices of pulmonary function, including vital capacity, FEV 1 , peak flow, nitrogen washout, and diffusion capacity in 210 locomotive repairmen exposed to DE in 3 engine houses. The average exposure of these locomotive repairmen to DE was 9.6 years. When compared with a control group matched for age, body size, “past extrapulmonary medical history” (no explanation given), and job status (154 railroad yard workers), no significant clinical differences were found in pulmonary function or in the prevalence of dyspnea, cough, or sputum between the DE-exposed and nonexposed groups. Exposure to DE showed marked seasonal variations because the doors of the engine house were open in the summer and closed in the winter. For the exposed group, the maximum daily workplace concentrations of air pollutants measured were 1.8 ppm NO 2 , 1.7 ppm total aldehydes, 0.15 ppm acrolein, 4.0 ppm SO 2 , and 5.0 ppm total hydrocarbons. The concentration of airborne particles was not reported. Gamble et al. (1987b) examined 283 diesel bus garage workers from four garages in two cities to determine if there was excess chronic respiratory morbidity associated with exposure to DE. Tenure of employment was used as a surrogate of exposure; mean tenure of the study population was 9 years ± 10 years S.D. Exposure-effect relationships within the study population showed no detectable associations of symptoms with tenure. Reductions in FVC,
FEV 1 , peak flow, and FEF 50 (but not FEF 75 ) were associated with increasing tenure. Compared with a control population (716 nonexposed blue-collar workers) and after indirect adjustment for age, race, and smoking, the exposed workers had a higher incidence of cough, phlegm, and wheezing; however, there was no correlation between symptoms and length of employment. Dyspnea showed an exposure-response trend but no apparent increase in prevalence. Mean
FEV 1 , FVC, FEF 50 , and peak flow were not reduced in the total cohort compared with the reference population, but were reduced in workers with 10 years or more tenure. Purdham et al. (1987) evaluated respiratory symptoms and pulmonary function in 17 stevedores employed in car ferry operations who were exposed to both diesel and gasoline exhausts and in a control group of 11 on-site office workers. Twenty-four percent of the exposed group and 36% of the controls were smokers. If a particular symptom was considered5-15 to be influenced by smoking, smoking status was used as a covariate in the logistic regression analysis; pack-years smoked was a covariate for lung function indices. The frequency of respiratory symptoms was not significantly different between the two groups; however, baseline pulmonary function measurements were significantly different. The latter comparisons were measured by multiple regression analysis using the actual (not percentage predicted) results and correcting for age, height, and pack-years smoked. The stevedores had significantly lower
FEV 1 , FEV 1 /FVC, FEF 50 , and FEF 75 ( p <0.021, p <0.023, p <0.001, and p <0.008, respectively), but not FVC. The results from the stevedores were also compared with those obtained from a study of the respiratory health status of Sydney, Nova Scotia, residents. These comparisons showed that the dock workers had higher FVC, similar FEV 1 , but lower FEV 1 /FVC and flow rates than the residents of Sydney. Based on these consistent findings, the authors concluded that the lower baseline function measurements in the stevedores provided evidence of an obstructive ventilatory defect, but caution in interpretation was warranted because of the small sample size. There were no significant changes in lung function over the workshift, nor was there a difference between the two groups. The stevedores were exposed to significantly ( p <0.04) higher concentrations of particulate matter (0.06 to 1.72 mg/m 3 , mean 0.50 mg/m 3 ) than the controls (0.13 to 0.58 mg/m 3 , mean not reported). Exposures of stevedores to SO 2 , NO 2 , aldehydes, and PAHs were very low; occasional CO concentrations in the 20 to 100 ppmrange could be detected for periods up to 1 h in areas where blockers were chaining gasolinepowered vehicles. Additional epidemiologic studies on the health hazards posed by exposure to DE have been conducted for mining operations. Reger et al. (1982) evaluated the respiratory health status of 823 male coal miners from six diesel-equipped mines compared with 823 matched coal miners not exposed to DE. The average tenure of underground work for the underground miners and their controls was only about 5 years; on average, the underground workers in diesel mines spent only 3 of those 5 years underground in diesel-use mines. Underground miners exposed to DE reported a higher incidence of symptoms 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 significant. The diesel-exposed underground miners, on the average, had lower FVC, FEV 1 , FEF 50 , FEF 75 , and FEF 90 but higher peak flow
and FEF 25 than their matched controls. These differences, however, were not statistically significant. Health indicators for surface workers and their matched controls were directionally the same as for matched underground workers. There were no consistent relationships between the findings of increased respiratory symptoms, decreased pulmonary function, smoking history, years of exposure, or monitored atmosphere pollutants (NO
x , CO, particles, and aldehydes).
Mean concentrations of NO
x at the six mines ranged from 0 to 0.6 ppm for short-term area5-16 samples, 0.13 to 0.28 ppm for full-shift personal samples, and 0.03 to 0.80 for full-shift area samples. Inhalable particles (less than 10 

### : m MMAD) averaged 0.93 to 2.73 mg/m 3 for personal samples and 0 to 16.1 mg/m 3 for full-shift area samples. Ames et al. (1984), using a portion of the miners studied by Reger, examined 280 diesel-exposed underground miners in 1977 and again in 1982. Each miner in this group had at least 1 year of underground mining work history in 1977. The control group was 838 miners with no exposure to DE. The miners were evaluated for prevalence of respiratory symptoms, chronic cough, phlegm, dyspnea, and changes in FVC, FEV 1 , and FEF 50 . No air monitoring data were reported; exposure to DE gases and mine dust particles were described as very low. These authors found no decrements in pulmonary function or increased prevalence of respiratory symptoms attributable to exposure to DE. In fact, the 5-year incidences of cough, phlegm, and dyspnea were greater in miners without exposure to DE. Attfield (1978) studied 2,659 miners from 21 mines (8 metal, 6 potash, 5 salt, and 2 trona). Diesels were employed in only 18 of the mines, but the 3 mines not using diesels were not identified. The years of diesel usage, ranging from 8 in trona mines to 16 in potash mines, were used as a surrogate for exposure to DE. Based on a questionnaire, an increased prevalence of persistent cough was associated with exposure to aldehydes; this finding, however, was not supported by the pulmonary function data. No adverse respiratory symptoms or pulmonary function impairments were related to CO 2 , CO, NO 2 , inhalable dust, or inhalable quartz. The author failed to comment on whether the prevalence of cough was related to the high incidence (70%) of smokers in the cohort. Questionnaire, chest radiograph, and spirometric data were collected by Attfield et al. (1982) on 630 potash miners from six potash mines. These miners were exposed for an average of 10 years (range 5 to 14 years) to 0.1 to 3.3 ppm NO 2 , 0.1 to 4.0 ppm aldehyde, 5 to 9 ppm CO, and total dust concentrations of 9 to 23 mg/m 3 . No attempt was made to measure dieselderived particles separately from other dusts. The ratio of total to inhalable (<10 

### : m MMAD) dust ranged from 2 to 11. An increased prevalence of respiratory symptoms was related solely to smoking. No association was found between symptoms and tenure of employment, dust exposure, NO 2 , CO, or aldehydes. A higher prevalence of symptoms of cough and phlegm was found, but no differences in pulmonary function (FVC and FEV 1 ) were found in these diesel-exposed potash miners when compared with the predicted values derived from a logistics model based on blue-collar workers working in nondusty jobs. Gamble et al. (1983) investigated respiratory morbidity in 259 miners from 5 salt mines in terms of increased respiratory symptoms, radiographic findings, and reduced pulmonary function associated with exposure to NO 2 , inhalable particles (<10 

### : m MMAD), or years worked underground. Two of the mines used diesel extensively; no diesels were used in one salt5-17 mine. Diesels were introduced into each mine in 1956, 1957, 1963, or 1963 through 1967. Several working populations were compared with the salt miner cohort. After adjustment for age and smoking, the salt miners showed no increased prevalence of cough, phlegm, dyspnea, or airway obstruction (FEV 1 /FVC) compared with aboveground coal miners, potash miners, or blue-collar workers. The underground coal miners consistently had an elevated level of symptoms. Forced expiratory volume at 1 s, FVC, FEF 50 , and FEF 75 were uniformly lower for salt miners in relation to all the comparison populations. There was, however, no association between changes in pulmonary function and years worked, estimated cumulative inhalable particles, or estimated NO 2 exposure. The highest average exposure to particulate matter was 1.4 mg/m 3 (particle size not reported, measurement includes NaCl). Mean NO 2 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 DE exposure. Average concentrations of inhalable particles and
NO 2 were 0.40, 0.60, and 0.82 mg/m 3 and 0.64, 1.77, and 2.21 ppm for the three diesel exposure categories, respectively. A statistically significant concentration-response association was found between the prevalence of phlegm in the salt miners and exposure to DE ( p <0.0001) and a similar, but nonsignificant, trend for cough and dyspnea. Changes in pulmonary function showed no association with diesel tenure. In a comparison with the control group of nonexposed, blue-collar workers, adjusted for age and smoking, the overall prevalence of cough and phlegm (but not dyspnea) was elevated in the diesel-exposed workers. Forced expiratory volumes at 1 s and FVC were within 4% of expected, which was considered to be within the normal range of variation for a nonexposed population. In a preliminary study of three subcohorts from bus company personnel (clerks [lowest exposure], bus drivers [intermediate exposure], and bus garage workers [highest exposure]) representing different levels of exposure to DE, Edling and Axelson (1984) found a fourfold higher risk ratio for cardiovascular mortality in bus garage workers, even after adjusting for smoking history and allowing for at least 10 years of exposure and 15 years or more of induction latency. Carbon monoxide was hypothesized as the etiologic agent for the increased cardiovascular disease but was not measured. However, in a more comprehensive epidemiologic study, Edling et al. (1987) evaluated mortality data covering a 32-year period for a cohort of 694 bus garage employees and found no significant differences between the observed and expected number of deaths from cardiovascular disease. Information on exposure components and their concentrations was not reported. The absence of reported noncancerous human health effects, other than infrequently occurring effects related to respiratory symptoms and pulmonary function changes, is notable. Unlike studies in laboratory animals, to be described later in this chapter, studies of the impact5-18 of DE on the defense mechanisms of the human lung have not been performed. No direct evidence is available in humans regarding doses of DE, gas phase, particulate phase, or total exhaust that lead to impaired particle clearance or enhanced susceptibility to infection. A summary of epidemiologic studies is presented in Table 5-1. Table 5-1. Human studies of exposure to diesel exhaust
Study Description Findings Acute exposures

Kahn et al. (1988) 13 cases of acute exposure, Utah and Colorado coal miners. Acute reversible sensory irritation, headache, nervous system effects, bronchoconstriction were reported at unknown exposures.

El Batawi and Noweir (1966) 161 workers, two diesel bus garages. 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. Battigelli (1965) Six subjects, eye exposure chamber, three dilutions. Time to onset was inversely related and severity of eye irritation was associated with the level of exposure to DE.

Katz et al. (1960) 14 persons monitoring DE in a train tunnel. Three occasions of minor eye and throat irritation; no correlation established with concentrations of DE components.

Hare and Springer (1971) Hare et al. (1974) Volunteer panelists who evaluated general public’s response to odor of DE. Slight odor intensity, 90% perceived, 60% objected; slight to moderate odor intensity, 95% perceived, 75% objected; moderate odor intensity, 100% perceived, almost 95% objected.

Linnell and Scott (1962) Odor panel under highly controlled conditions determined odor threshold for DE. In six panelists, the volume of air required to dilute raw DE to an odor threshold ranged from a factor of 140 to 475.

Rudell et al. (1990, 1994) Eight healthy nonsmoking subjects exposed for 60 min in chamber to DE (3.7 ppm NO, 1.5 ppm NO 2 , 27 ppm CO, 0.5 mg/m 3 formaldehyde, particles (4.3 × 10 6 /cm 3 ). Exercise, 10 of each 20 min (75 W). 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.

Rudell et al. (1996) Volunteers exposed to DE for 1 h while doing light work. Exposure concentrations uncertain. Unpleasant smell along with irritation of eyes and nose reported. Airway resistance increased. Reduction of particle concentration by trapping did not affect results. Battigelli (1965) 13 volunteers exposed to three dilutions of DE for 15 min to 1 h. No significant effects on pulmonary resistance were observed as measured by plethysmography.

Wade and Newman (1993) Three railroad workers acutely exposed to DE. The workers developed symptoms of asthma.

Diaz-Sanchez et al. (1994) Volunteers challenged by a nasal spray of 0.30 mg DPM. Enhancement of IgE production reported due to a dramatic increase in IgE-secreting cells.5-19 Table 5-1. Human studies of exposure to diesel exhaust (continued)
Study Description Findings

Takenaka et al. (1995) Volunteers challenged by a nasal spray of 0.30 mg DPM. 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.

Diaz-Sanchez et al. (1996) Volunteers challenged by a nasal spray of 0.30 mg DPM. A broad increase in cytokine expression predicted to contribute to enhanced local IgE production.

Diaz-Sanchez et al. (1997) Ragweed-sensitive volunteers challenged by a nasal spray of 0.30 mg DPM alone or in combination with ragweed allergen. Ragweed allergen plus DPM-stimulated ragweedspecific IgE to a much greater degree than ragweed alone, suggesting DPM may be a key feature in stimulating allergen-induced respiratory allergic disease.

Salvi et al. (1999) Volunteers exposed to diluted DE (DPM 300 : g/m 3 ) for 1 h with intermittent exercise. • 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 andVCAM-1; increases in the number of LFA1+ in bronchial tissue. • Significant increases in neutrophils and platelets observed in peripheral blood.

Salvi et al. (2000) Volunteers exposed to diluted DE (DPM 300 : g/m 3 ) for 1 h. • DPM enhanced gene transcription of IL-8 in bronchial tissue and bronchial wash cells • Increased expression of growth-regulated
oncogene- " and IL-8 in bronchial epithelium; trend towards increased IL-5 mRNA gene transcripts.
Nightingale et al. (2000) Volunteers exposed to resuspended DPM (200 ug/m 3 ) for 2 h at rest • DPM increased exhaled levels of CO • DPM increased sputum neutrophils and myeloperoxidase

Studies of cross-shift changes
Reger (1979) Five or more VC maneuvers by each of 60 coal miners exposed to DE at the beginning and end of a workshift. FEV 1 , FVC, and PEFR were similar between diesel and non-diesel-exposed miners. Smokers had an increased number of decrements over shift than nonsmokers.5-20 Table 5-1. Human studies of exposure to diesel exhaust (continued)
Study Description Findings

Ames et al. (1982) Pulmonary function of 60 dieselexposed compared with 90 nondiesel-exposed coal miners over workshift. Significant workshift decrements occurred in miners in both groups who smoked; no significant differences in ventilatory function changes between miners exposed to DE and those not exposed. Jörgensen and Svensson (1970) 240 iron ore miners matched for diesel exposure, smoking, and age were given bronchitis questionnaires and spirometry pre- and postworkshift. 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.

Gamble et al. (1979) 200 salt miners performed beforeand after-workshift spirometry. Personal environmental NO 2 and inhalable particle samples were collected. Smokers had greater but not significant reductions in spirometry than ex- or nonsmokers. NO 2 but not particulate levels significantly decreased FEV1, FEF 25 , FEF 50 , and FEF 75 over the workshift.

Gamble et al. (1987a) 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. Prevalence of burning eyes, headache, difficult or labored breathing, nausea, and wheeze were higher in diesel bus workers than in comparison population.

Ulfvarson et al. (1987) 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 DE, 2.1 ppm NO 2 , and 0.6 mg/m 3 particulate
matter. Pulmonary function was affected during a workshift exposure to DE, 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.

Cross-sectional and longitudinal studies Battigelli et al. (1964) 210 locomotive repairmen exposed to DE 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 DE. No significant differences in VC, FEV 1 , peak flow, nitrogen washout, or diffusion capacity or in the prevalence of dyspnea, cough, or sputum were found between the DE-exposed and nonexposed
groups.5-21 Table 5-1. Human studies of exposure to diesel exhaust (continued)
Study Description Findings

Gamble et al. (1987b) 283 male diesel bus garage workers from four garages in two cities were examined for impaired pulmonary function (FVC, FEV 1 , and flow rates). Study population with a mean tenure of 9 ± 10 years S.D. wascompared to a nonexposed bluecollar population. Analyses within the study population showed no association of respiratory symptoms with tenure. Reduced FEV 1 and FEF 50 (but not FEF 75 ) 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.
Purdham et al. (1987) Respiratory symptoms and pulmonary 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. No differences between the two groups for respiratory 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. Reger et al. (1982) Differences in respiratory symptoms and pulmonary function were assessed in 823 coal miners from 6 diesel-equipped mines compared to 823 matched coal miners not exposed to DE. 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 DE thought to be responsible.
Ames et al. (1984) Changes in respiratory symptoms and function were measured during a 5- year period in 280 diesel-exposed and 838 nonexposed U.S. underground coal miners. No decrements in pulmonary function or increased prevalence of respiratory symptoms were found attributable to DE. In fact, 5-year incidences of cough, phlegm, and dyspnea were greater in miners without exposure to DE than in miners exposed to DE. 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 DE. 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 NO 2 , CO, CO 2 , dust, or quartz.5-22 Table 5-1. Human studies of exposure to diesel exhaust (continued)
Study Description Findings

Attfield et al. (1982) Respiratory symptoms and function 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 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 1 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.

Gamble et al. (1983) 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. After adjustment for age and smoking, salt miners showed no symptoms or increased prevalence of cough, phlegm, dyspnea, or air obstruction (FEV 1 /FVC) compared with aboveground coal miners, potash workers, or blue-collar workers. FEV 1 , FVC, FEF 50 , and FEF 75 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 NO 2 . Gamble and Jones (1983) Same as above. Salt miners were grouped into low-, intermediate-, and high-exposure categories based on tenure in jobs with diesel exposure. 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 smallbut significant reductions; there was no doseresponse relationship.
Edling and Axelson (1984) Pilot study of 129 bus companyemployees classified into 3 dieselexhaust exposure categories: clerks (0), bus drivers (1), and bus garage workers. 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.
Edling et al. (1987) 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). 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.5-23 To date, no large-scale epidemiologic study has looked for effects of chronic exposure to DE on pulmonary function. In the long-term longitudinal and cross-sectional studies, a relationship was generally observed between work in a job with diesel exposure and respiratory symptoms (such as cough and phlegm), but there was no consistent effect on pulmonary function. The interpretation of these results is hampered by lack of measured DE exposure levels and the short duration of exposure in these cohorts. The studies are further limited in that only active workers were included, and it is possible that workers who have developed symptoms or severe respiratory disease are likely to have moved away from these jobs. The relationship between work in a job with diesel exposure and respiratory symptoms may be due to short-term exposure.

5.1.2. Traffic Studies The relationship between traffic density and respiratory health in children has been examined in a series of studies in Holland in children attending schools located near major freeways. Cough, wheeze, runny nose, and doctor-diagnosed asthma were reported more often for children living within 100 m of freeways carrying between 80,000 and 150,000 vehicles per day (van Vliet et al., 1997). Separate counts for truck traffic indicated a range from 8,000 to 17,500 trucks per day. Truck traffic intensity and concentration of “black smoke,” considered by the authors to be a proxy for DPM, measured in schools were found to be significantly associated with chronic respiratory symptoms, with the relationships being more pronounced in girls than in boys. Brunekreef et al. (1997) measured lung function in children in six areas located near major motorways and assessed their exposure to traffic-related air pollution using separate traffic counts for automobiles and trucks. They also measured air pollution in the children's schools. Although lung function was associated with truck traffic density, there was a lesser association with automobile traffic density. The association was stronger in those children living closest (300 m) to the roadways. Lung function was also associated with concentration of “black smoke” (source and constitution unclear from the study) measured inside the schools. The associations were stronger in girls than in boys. The authors conclude that exposure to vehicular pollution, in particular DPM, may lead to reduced lung function in children living near major motorways. In a follow-up study of traffic-related air pollution and its effect on the respiratory health of children living near roadways, Brunekreef et al. (2000) showed that the intensity of truck traffic was significantly associated with the prevalence of wheeze, phlegm, bronchitis, eye symptoms, and allergy to dust and pets. Associations with yearly averaged PM 2.5 and “soot” concentrations measured inside and outside the schools showed similar patterns. Truck traffic5-24 intensity was also significantly associated with a positive skin prick test or elevated IgE for outdoor allergens. There were no associations between traffic intensity or PM 2.5 and “soot” concentrations and lung function, bronchial responsiveness, and allergic reactions to indoor allergens. Further analysis of the data showed that the associations between traffic-related air pollution and symptoms were almost entirely related to children with bronchial hyperreactivity or sensitization to common allergens.

5.1.3. Laboratory Animal Studies Because humans and laboratory animals show similar nonneoplastic responses to inhaled particles (ILSI, 2000), animal studies have been conducted to assess the pathophysiologic effects of DPM. Because of the large number of statistical comparisons made in the laboratory animal studies, and to permit uniform, objective evaluations within and among studies, data will be reported as significantly different (i.e., p <0.05) unless otherwise specified. The exposure regimens used and the resultant exposure conditions employed in the laboratory animal inhalation studies are summarized in Tables 5-2 through 5-16. Other than the pulmonary function studies performed by Wiester et al. (1980) on guinea pigs during their exposure in inhalation chambers, the pulmonary function studies performed by other investigators, although sometimes unreported, were interpreted as being conducted on the following day or thereafter and not immediately following exposure.

5.1.3.1. Acute Exposures The acute toxicity of undiluted DE to rabbits, guinea pigs, and mice was assessed by Pattle et al. (1957). Four engine operating conditions were used, and 4 rabbits, 10 guinea pigs, and 40 mice were tested under each exposure condition for 5 h (no controls were used). Mortality was assessed up to 7 days after exposure. With the engine operating under light load, the exhaust was highly irritating but not lethal to the test species, and only mild tracheal and lung damage was observed in the exposed animals. The exhaust contained 74 mg/m 3 DPM (particle size not reported), 560 ppm CO, 23 ppm NO 2 , and 16 ppm aldehydes. Exhaust containing 5 mg/m 3 DPM, 380 ppm CO, 43 ppm NO 2 , and 6.4 ppm aldehydes resulted in low mortality rates (mostly below 10%) and moderate lung damage. Exhaust containing 122 mg/m 3 DPM, 418 ppm CO, 51 ppm NO 2 , and 6.0 ppm aldehydes produced high mortality rates (mostly above 50%) and severe lung damage. Exhaust containing 1,070 mg/m 3 DPM, 1,700 ppm CO, 12 ppm NO 2 , and 154 ppm aldehydes resulted in 100% mortality in all three species. High CO levels, which resulted in a carboxyhemoglobin value of 60% in mice and 50% in rabbits and guinea pigs, were considered to be the main cause of death in the latter case. High NO 2 levels5-25 were considered to be the main cause of lung damage and mortality seen in the other three tests. Aldehydes and NO 2 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, augmented increased airway resistance and nasal secretion induced by a histamine aerosol, increased vascular permeability in dorsal skin, and augmented vascular permeability induced by histamine. The increases in nasal airway resistance and secretion are considered typical responses of nasal mucosa against allergic stimulation. Similar results were reported for guinea pigs exposed via inhalation for 3 h to DE diluted to DPM concentrations of either 1 or 3.2 mg/m 3 (Kobayashi et al., 1997). These studies show that short-term exposure to DPM augments nasal mucosal hyperresponsiveness induced by histamine in guinea pigs.

5.1.3.2. Short-Term and Subchronic Exposures A number of inhalation studies have employed a regimen of 20 h/day, 7 days/week for varying exposure periods up to 20 weeks to differing concentrations of airborne particulate matter, vapor, and gas concentrations of diluted DE. Exposure regimens and characterization of gas-phase components for these studies are summarized in Table 5-2. Pepelko et al. (1980a) evaluated the pulmonary function of cats exposed under these conditions for 28 days to 6.4 mg/m 3 DPM. The only significant functional change observed was a decrease in maximum expiratory flow rate at 10% vital capacity. The excised lungs of the exposed cats appeared charcoal gray, with focal black spots visible on the pleural surface. Pathologic changes included a predominantly peribronchial localization of black-pigmented macrophages within the alveoli characteristic of focal pneumonitis or alveolitis. The effects of a short-term DE exposure on arterial blood gases, pH, blood buffering, body weight changes, lung volumes, and deflation pressure-volume (PV) curves of 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/m 3 DPM in the nonirradiated exhaust (RE) and 6.75 mg/m 3 in the irradiated exhaust (IE). In spite of the irradiation, levels of gaseous compounds were not substantially different between the two groups (Table 5-2). Body weight gains were significantly reduced in the RE-exposed rats and to an even greater degree in rats exposed to IE. Arterial blood gases and standard bicarbonate were unaffected, but arterial blood pH was significantly reduced in rats exposed to IE. Residual volume and wet lung weight were not affected by either exposure, but vital capacity and total lung capacity were increased significantly following exposure to RE. The shape of the deflation PV curves were nearly identical for the control, RE, and IE groups.5-26Table 5-2. Short-term effects of diesel exhaust on laboratory animalsSpecies/sex Exposure periodParticles(mg/m3)C × T(mgAh/m3)CO(ppm)NO2(ppm)SO2(ppm)Effects StudyRat, F344, M;Mouse, A/J, M; Hamster,Syrian, M20 h/day7 days/week10-13 weeks1.50.19:m MMD2,100 to 2,730 6.9 0.49—Increase in lung wt; increase inthickness of alveolar walls;minimal species differenceKaplan et al. (1982)Rat, F344, M, F; Mouse,CD-1, M, F7 h/day5 days/week19 weeks0.211.04.41406652,926—————————No effects on lung function in rats(not done in mice); increase inPMNs and proteases and AMaggregation in both speciesMauderly et al. (1981)Cat, Inbred, M20 h/day7 days/week4 weeks6.4 3,584 14.6 2.1 2.1 Feweffects on lung function; focalpneumonitis or alveolitisPepelko et al. (1980a)Rat, SpragueDawley, M20 h/day7 days/week4 weeks6.46.8a3,5843,80816.916.1a2.492.76a2.101.86aDecreased body wt; arterial bloodpH reduced; vital capacity, totallung capacities increasedPepelko (1982a)(<0.01 ppm O3)aGuinea Pig,Hartley, M, F20 h/day7 days/week4 weeks6.8a3,808 16.7 2.9 1.9 Exposure started when animalswere 4 days old; increase inpulmonary flow; bardycardiaWiester et al. (1980)(<0.01 ppm O3)aRat, F344,M20 h/day5.5 days/week4 weeks6.06.8:m MMD2,640 —— —Macrophage aggregation; increasein PMNs; Type II cellproliferation; thickened alveolarwallsWhite and Garg (1981)Guinea Pig, Hartley, M 30 min 1-2 mg DPMIntranasally— —— —Augmented increases in nasalairway resistance and vascularpermeability induced by ahistamine aerosolKobayashi and Ito(1995)Guinea Pig, Hartley, M 3 h 13.20.51.65.912.91.44.40.130.34Similar results to those reported inthe previous study using intranasalchallengeKobayashi et al. (1997)5-27Table 5-2. Short-term effects of diesel exhaust on laboratory animals (continued)Species/sex Exposure periodParticles(mg/m3)C × T(mgAh/m3)CO(ppm)NO2(ppm)SO2(ppm)Effects StudyGuinea Pig, Hartley, M,F 20 h/day7 days/week8 weeks6.3 7,056 17.4 2.32.1 Increase in relative lung wt. AMaggregation; hypertrophy of gobletcells; focal hyperplasia of alveolarepitheliumWiester et al. (1980)Mouse ICR, M 6weeks 100:g DPMintranasally— — — — DPM aggravated ovalbumininduced airway inflammation andprovided evidence that DPM canenhance manifestations of allergicasthmaTakano et al. (1997)Rat, Sprague-Dawley,M24 h 5-100:g/106AM/mL ofDPM— — — — Unchanged, but not organic-freeDPM enhanced production ofproinflammatory cytokinesYang et al. (1997)aIrradiated exhaust.PMN = Polymorphonuclear leukocyte.AM = Alveolar macrophage.5-28 In related studies, Wiester et al. (1980) evaluated pulmonary function in 4-day-old guinea pigs exposed for 20 h/day, 7 days/week for 28 days to IE having a concentration of 6.3 mg/m 3 DPM. When housed in the exposure chamber, pulmonary flow resistance increased 35%, and a small but significant sinus bradycardia occurred as compared with controls housed and measured in control air chambers ( p <0.002). Respiratory rate, tidal volume, minute volume, and dynamic compliance were unaffected, as were lead-1 electrocardiograms. 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). Heart/body weight ratios were not affected by exposure. Microscopically, there was a marked accumulation of black pigment-laden AMs throughout the lung, with a slight to moderate accumulation in bronchial and carinal lymph nodes. Hypertrophy of goblet cells in the tracheobronchial tree was frequently observed, and focal hyperplasia of alveolar lining cells was occasionally observed. No evidence of squamous metaplasia of the tracheobronchial tree, emphysema, peribronchitis, or peribronchiolitis was noted. White and Garg (1981) studied pathologic alterations in the lungs of rats (16 exposed and 8 controls) after exposure to DE containing 6 mg/m 3 DPM. Two rats from the exposed group and one rat from the control group (filtered room air) were sacrificed after each exposure interval of 6 h and 1, 3, 7, 14, 28, 42, and 63 days; daily exposures were for 20 h and were 5.5 days/week. Evidence of AM recruitment and phagocytosis of diesel particles was found at the 6- h sacrifice; after 24 h of exposure there was a focal, scattered increase in the number of Type II cells. After 4 weeks of exposure, there were morphologic changes in size, content, and shape of AM, septal thickening adjacent to clusters of AMs, and an appearance of inflammatory cells, primarily within the septa. At 9 weeks of exposure, focal aggregations of particle-laden macrophages developed near the terminal bronchi, along with an influx of PMNs, Type II cell proliferation, and thickening of alveolar walls. The affected alveoli occurred in clusters that, for the most part, were located near the terminal bronchioles, but occasionally were focally located in the lung parenchyma. Hypertrophy of goblet cells in the tracheobronchial tree was frequently observed, and focal hyperplasia of alveolar lining cells was occasionally observed. No evidence of squamous metaplasia of the tracheobronchial tree, emphysema, peribronchitis, or peribronchiolitis was noted. Mauderly et al. (1981) exposed rats and mice by inhalation to diluted DE for 545 h over a 19-week period on a regimen of 7 h/day, 5 days/week at concentrations of 0, 0.21, 1.02, or 4.38 mg/m 3 DPM. Indices of health effects were minimal following 19 weeks of exposure. There were no significant exposure-related differences in mortality or body weights of the rats or mice. There also were no significant differences in respiratory function (breathing patterns,5-29 dynamic lung mechanics, lung volumes, quasi-static PV relationships, forced expirograms, and CO-diffusing capacity) in rats; pulmonary function was not measured in mice. No effect on tracheal mucociliary or deep lung clearances were observed in the exposed groups. Rats, but not mice, had elevated immune responses in lung-associated lymph nodes at the two higher exposure levels. Inflammation in the lungs of rats exposed to 4.38 mg/m 3 DPM was indicated by increases in PMNs and lung tissue proteases. Histopathologic findings included AMs that contained DPM, an increase in Type II cells, and the presence of particles in the interstitium and tracheobronchial lymph nodes. Kaplan et al. (1982) evaluated the effects of subchronic exposure to DE on rats, hamsters, and mice. The exhaust was diluted to a concentration of 1.5 mg/m 3 DPM; exposures were 20 h/day, 7 days/week. Hamsters were exposed for 86 days, rats and mice for 90 days. There were no significant differences in mortality or growth rates between exposed and control animals. Lung weight relative to body weight of rats exposed for 90 days was significantly higher than the mean for the control group. Histological examination of tissues of all three species indicated particle accumulation in the lungs and mediastinal lymph nodes. Associated with the larger accumulations, there was a minimal increase in the thickness of the alveolar walls, but the vast majority of the particles elicited no response. After 6 mo of recovery, considerable clearance of the DPM from the lungs occurred in all three species, as evaluated by gross pathology and histopathology. However, no quantitative estimate of clearance was provided. Toxic effects in animals from acute exposure to DE appear to be primarily attributable to the gaseous components (i.e., mortality from CO intoxication and lung injury caused by cellular damage resulting from NO 2 exposure). The results from short-term exposures indicate that rats experience minimal lung function impairment even at DE levels sufficiently high to cause histological and cytological changes in the lung. In subchronic studies of durations of 4 weeks or more, frank adverse health effects are not readily apparent and, when found, are mild and result from exposure to concentrations of about 6 mg/m 3 DPM and durations of exposures of 20 h/day. There is ample evidence that subchronic exposure to lower levels of DE affects the lung, as indicated by accumulation of particles, evidence of inflammatory response, AM aggregation and accumulation near the terminal bronchioles, Type II cell proliferation, and thickening of alveolar walls adjacent to AM aggregates. Little evidence exists, however, that subchronic exposure to DE impairs lung function.

5.1.3.3. Chronic Exposures
5.1.3.3.1. Effects on growth and longevity. Changes in growth, body weight, absolute or relative organ weights, and longevity can be measurable indicators of chronic toxic effects.5-30 Such effects have been observed in some, but not all, of the long-term studies conducted on laboratory animals exposed to DE. There was limited evidence for an effect on survival in the published chronic animal studies; deaths occurred intermittently early in one study in female rats exposed to 3.7 mg/m 3 DPM; however, the death rate began to decrease after 15 mo, and the survival rate after 30 mo was slightly higher than that of the control group (Ishinishi et al., 1988). Studies of the effects of chronic exposure to DE on survival and body weight or growth
are detailed in Table 5-3. Increased lung weights and lung-to-body weight ratios have been reported in rats, mice, and hamsters. These data are summarized in Table 5-4. In rats exposed for up to 36 weeks to 0.25 or 1.5 mg/m 3 DPM, lung wet weights (normalized to body weight) were significantly higher in the 1.5 mg/m 3 exposure group than control values after 12 weeks of exposure (Misiorowski et al., 1980). Rats and Syrian hamsters were exposed for 2 years (five 16-h periods per week) to DE diluted to achieve concentrations of 0.7, 2.2, and 6.6 mg/m 3 DPM (Brightwell et al., 1986). At necropsy, a significant increase in lung weight was seen in both rats and hamsters exposed to DE compared with controls. This finding was more pronounced in the rats in which the increase was progressive with both duration of exposure and particulate matter level. The increase was greatest at 30 mo (after the end of a 6-mo observation period in the high-concentration male group where the lung weight was 2.7 times the control and at 24 mo in the high-concentration female group [3.9 times control]). Heinrich et al. (1986a,b; see also Stöber, 1986) found a significant increase in wet and dry weights of the lungs of rats and mice exposed at 4.24 mg/m 3 DPM for 1 year in comparison with controls. After 2 years, the difference was a factor of 2 (mice) or 3 (rats). After the same exposure periods, the hamsters showed increases of 50% to 75%, respectively. Exposure to equivalent filtered DE (i.e., without DPM) caused no significant effects in any of the species. Vinegar et al. (1980, 1981a,b) exposed hamsters to two levels of DE with resultant concentrations of about 6 and 12 mg/m 3 DPM for 8h/day, 7 days/week for 6 mo. Both exposures significantly increased lung weight and lungweight to body-weight ratios. The difference between lung weights of exposed and control hamsters exposed to 12 mg/m 3 DPM was approximately twice that of those exposed to 6 mg/m 3

. Heinrich et al. (1995) reported that rats exposed to 2.5 and 7 mg/m 3 DPM for 18 h/day, 5 days/week for 24 mo showed significantly lower body weights than controls starting at day 200 in the high-concentration group and at day 440 in the low-concentration group. Body weight in the low-concentration group was unaffected, as was mortality in any group. Lung weight was increased in the 7 mg/m 3 group starting at 3 mo and persisting throughout the study, while the 2.5 mg/m 3 group showed increased lung weight only at 22 and 24 mo of exposure. Mice (NMRI strain) exposed to 7 mg/m 3 in this study for 13.5 mo had no increase in mortality and insignificant decreases in body weight. Lung weights were dramatically affected, with5-31Table 5-3. Effects of chronic exposures to diesel exhaust on survival and growth of laboratory animalsSpecies/sexExposureperiodParticles(mg/m3)C × T(mgAh/m3)CO(ppm)NO2(ppm)SO2(ppm)EffectsStudyRat, F344, M, F;Monkey, Cynomolgus, M7 h/day5 days/week104 weeks2.00.23–0.36:m MMD7,280 11.5 1.5 0.8 Noeffects on growth or survivalLewis et al.(1989)Rat, F344, M;Guinea Pig, Hartley, M20 h/day5 days/week106 weeks0.250.751.50.19:m MMD2,6507,95015,9002.7a4.4a7.1a0.1b0.27b0.5b———Reduced body weight in rats at1.5 mg/m3Schreck et al.(1981)Hamster, Chinese, M 8h/day7 days/week26 weeks6.012.08,73617,472——————No effect on growth Vinegar et al.(1981a,b)Rat, Wistar, M 6h/day5 days/week87 weeks8.30.71:m MMD21,663 50.0 4.0–6.0 — Noeffect on growth or mortalityratesKaragianeset al. (1981)Rat, F344, M, F;Mouse, CD-1, M, F7 h/day5 days/week130 weeks0.353.57.10.25:m MMD1,59215,92531,8502.916.529.70.050.340.68———No effect on growth or mortalityratesMauderly et al.(1984, 1987a)Rat, Wistar, F;Mouse, MMRI, F19 h/day5 days/week104 weeks4.240.35:m MMD41,891 12.5 1.5 1.1Reduced body wts; increasedmortality in miceHeinrich et al.(1986a)Rat, F344M, F16 h/day5 days/week104 weeks0.72.26.65,82418,30454,912——32.0——————Growth reduced at 2.2 and6.6 mg/m3Brightwell et al.(1986)RatcF344/Jcl.16 h/day6 days/week130 weeks0.11d0.41d1.08d2.31d3.72e0.2–0.3:m MMD1,3735,11713,47828,82946,4261.232.123.967.1012.90.080.260.701.413.000.381.062.424.704.57Concentration-dependentdecrease in body weight; earlierdeaths in females exposed to 3.72mg/m3, stabilized by 15 moResearchCommittee forHERP Studies(1988)Rat, Wistar, F;Mouse, NMRI, F(7 mg/m3only)18 h/day5 days/week24 mo0.842.56.987,40021,80061,7002.68.321.20.31.23.80.31.13.4Reduced body weight in rats at2.5 and 6.98 mg/m3and no effectin miceHeinrich et al.(1995)5-32Table 5-3. Effects of chronic exposures to diesel exhaust on survival and growth of laboratory animals (continued)Species/sexExposureperiodParticles(mg/m3)C × T(mgAh/m3)CO(ppm)NO2(ppm)SO2(ppm)EffectsStudyMice, NMRI, F;C57BL/6N, F18 h/day5 days/week13.5 mo(NMRI)24 mo(C57BL/N)6.98 35,500 - NMRI38,300 - C5714.2 2.3 2.8 Reduced body weight in NMRImice but not in C57BL/6N miceHeinrich et al.(1995)Rats, F344, M16 h/day5 days/week23 mo2.446.3319,52050,640——————Reduced survival in 6.33 mg/m3after 300 days. Body weightsignificantly lower at 6.33 mg/m3Nikula et al.(1995)Mouse, CD-1,M,F7 h/day5 days/week104 weeks0.353.57.10.25:m MDD1,27412,74025,844317300.10.30.7———No effect on growth or mortalityratesMauderly et al.(1996)aEstimated from graphically depicted mass concentration data.bEstimated from graphically presented mass concentration data for NO2(assuming 90% NO and 10% NO2).cData for tests with light-duty engine; similar results with heavy-duty engine.dLight-duty engine.eHeavy-duty engine.5-33Table 5-4. Effects of chronic exposures to diesel exhaust on organ weights and organ-to-body-weight ratiosSpecies/sexExposureperiodParticles(mg/m3)C × T(mgAh/m3)CO(ppm)NO2(ppm)SO2(ppm)Effects StudyRat, F344, M;Mouse, A/J, M;Hamster, Syrian, M20 h/day7 days/week12-13 weeks1.50.19:m MMD2,520-2,730 — — — Noeffect on liver, kidney, spleen, orheart weightsKaplan et al. (1982)Rat, F344, M,F 7 h/day5 days/week52 weeks2.00.23–0.36:mMMD3,640 12.7 1.6 0.83 No effects onweights of lungs, liver,heart, spleen, kidneys, and testesGreen et al. (1983)Rat, F344, M20 h/day5.5 days/ week36 weeks0.251.50.19:m MMD9905,940——————Increase in relative lung weight at1.5 mg/m3only initially seen at12 weeksMisiorowski et al.(1980)Rat, F344, F 7h/day5 days/week104 weeks2.00.23–0.36:mMMD7,280 11.5 1.5 0.81 Noeffects on heart weights Vallyathan et al. (1986)Rat, F344; MGuinea Pig,Hartley, M20 h/day5.5 days/ week78 weeks0.250.751.50.19:m MMD2,1456,43512,870—————————No effects on heart mass Penney et al. (1981)Hamster, Chinese,M8 h/day7 days/week26 weeks6.012.08,73617,472——————Increase in lung weight and lung/bodyweight ratioVinegar et al. (1981a,b)Rat, Wistar, F;Hamster, Syrian,M, FMouse, NMRI, F19 h/day5 days/week120-140 weeks4.240.35:m MMD48,336-56,392 12.5 1.5 1.1Increase in rat, mouse, and hamsterlung weight and dry weightsHeinrich et al.(1986a,b)Stöber (1986)Rat, F344, M, F;Hamster, Syrian,M, F16 h/day5 days/week104 weeks0.7a2.2b6.65,82418,30454,912——32.0——————Increase in lung weight concentrationrelated in rats; heart weight/bodyweight ratio greater at 6.6 mg/m3Brightwell et al. (1986)Cat, inbred, M 8h/day7 days/week124 weeks6.0a12.0b41,66483,32820.233.22.74.42.75.0Decrease in lung and kidney weights Pepelko et al. (1980b,1981)Moorman et al. (1985)Mouse, NMRI, F(7 mg/m3only)18 h/day5 days/week24 mo0.842.56.987,40021,80061,7002.68.321.20.31.23.80.31.13.4Increased rat and mouse lung weight at7 mg/m3from 6 mo and at 2.5 mg/m3at 22 and 24 moHeinrich et al. (1995)5-34Table 5-4. Effects of chronic exposures to diesel exhaust on organ weights and organ-to-body-weight ratios (continued)Species/sexExposureperiodParticles(mg/m3)C × T(mgAh/m3)CO(ppm)NO2(ppm)SO2(ppm)Effects StudyMouse, NMRI, F;C57BL/6N, F18 h/day5 days/week13.5 mo (NMRI)24 mo(C57BL/N)6.98 35,500 - NMRI38,300 - C5714.2 2.3 2.8 Increased lung weight Heinrich et al. (1995)Rats, F344, M16 h/day5 days/week23 mo2.446.3319,52050,640——————Increase in lung weight was significantat 2 and 6 mg/m3Nikula et al. (1995)Rat 0.82.56.98Increased lung weight in rats and miceat 3.5 and 7.1 mg/m3Henderson et al.(1988a)Mouse 6.984.5a1 to 61 weeks of exposure.b62 to 124 weeks of exposure.5-35 increases progressing throughout the study from 1.5-fold at 3 mo to 3-fold at 12 mo. Mice (NMRI and C57BL/6N strains) were also exposed to 4.5 mg/m 3 for 23 mo. In NMRI mice, the body weights were reported to be significantly lower than controls, but the magnitude of the change is not reported, so biological significance cannot be assessed. Mortality was slightly increased, but statistical significance is not reported. The C57BL/6N mice showed minimal effects on body weight and mortality, which were not statistically significant. Lung weights were dramatically affected in both strains. Nikula et al. (1995) exposed male and female F344 rats to DPM concentrations of 2.4 and 6.3 mg/m 3 for 16 h/day, 5 days/week for 23 mo in a study designed to compare the effects of DPM with those of carbon black. Significantly reduced survival was observed in males exposed to 6.3 mg/m 3 but not in females or at the lower concentration. Body weights were decreased by exposure to 6.3 mg/m 3 DPM in both male and female rats throughout the exposure period. Significant increases in lung weight were first seen at 6 mo in the high-exposure group and at 12 to 18 mo in the low-exposure group. No evidence was found in the published literature that chronic exposure to DE affected the weight of body organs other than the lung and heart (e.g., liver, kidney, spleen, or testes) (Table 5-4). Morphometric analysis of hearts from rats and guinea pigs exposed to 0.25, 0.75, or 1.5 mg/m 3 DPM 20 h/day, 5.5 days/week for 78 weeks revealed no significant alteration in mass at any exposure level or duration of exposure (Penney et al., 1981). The analysis included relative wet weights of the right ventricle, left ventricle, combined atria, and ratio of right to left ventricle. Vallyathan et al. (1986) found no significant differences in heart weights and the ratio of heart weight to body weight between rats exposed to 2 mg/m 3 DPM for 7 h/day, 5 days/week for 24 mo and their respective clean-air chamber controls. No significant differences were found in the lungs, heart, liver, spleen, kidney, and testes of rats exposed for 52 weeks, 7 h/day, 5 days/week to diluted DE containing 2 mg/m 3 DPM compared with their respective controls (Green et al., 1983).

5.1.3.3.2. Effects on pulmonary function . The effect of long-term exposure to DE on pulmonary function has been evaluated in laboratory studies of rats, hamsters, cats, and monkeys. These studies are summarized in Table 5-5, along with more details on the exposure characteristics, in general order of increasing dose (C × T) of DPM. The text will be presented using the same approach. Lewis et al. (1989) evaluated functional residual capacity and airway resistance and conductance in 10 control and 10 diesel-exposed rats (2 mg/m 3 DPM, 7 h/day, 5 days/week for 52 or 104 weeks). At the 104-week evaluation, the rats were also examined for maximum flow volume impairments. No evidence of impaired pulmonary function as a result of the exposure to5-36Table 5-5. Effects of diesel exhaust on pulmonary function of laboratory animalsSpecies/sexExposureperiodParticles(mg/m3)C × T(mgAh/m3)CO(ppm)NO2(ppm)SO2(ppm)Effects StudyRat, F344, M,F 7 h/day5 days/week104 weeks2.00.23–0.36:mMMD7,280 11.5 1.5 0.8 No effect onpulmonary function Lewis et al. (1989)Monkey,Cynomolgus, M7 h/day5 days/week104 weeks2.00.23-0.36:mMMD7,280 11.5 1.50.8 Decreased expiratory flow; no effecton vital or diffusing capacitiesLewis et al. (1989)Rat, F344, M20 h/day5.5 days/week87 weeks1.50.19:m MMD14,355 7.0 0.5 —Increased functional residual capacity,expiratory volume, and flowGross (1981)Rat, Wistar, F7-8 h/day5 days/week104 weeks3.90.1:m MMD14,196-16,224 18.5 1.2 3.1 Noeffect on minute volume,compliance, or resistanceHeinrich et al. (1982)Hamster, Chinese, M 8h/day7 days/week26 weeks6.012.08,73617,472——————Decrease in vital capacity, residualvolume, and diffusing capacity;increase in static deflation lungvolumeVinegar et al. (1980,1981a,b)Rat, F344,M, F7 h/day5 days/week130 weeks0.353.57.10.23–0.26:mMMD1,59315,92531,8502.916.529.70.050.340.68———Diffusing capacity, lung compliancereduced at 3.5 and 7.1 mg/m3Mauderly et al. (1988)McClellan et al. (1986)Rat, F344, M, F;Hamster, Syrian, M, F16 h/day5 days/week104 weeks0.72.26.65,82418,30454,912—————————Large number of pulmonary functionchanges consistent with obstructiveand restrictive airway diseases at6.6 mg/m3(no specific data provided)Brightwell et al. (1986)Hamster, Syrian, M,F 19 h/day5 days/week120 weeks4.240.35:m MMD48,336 12.5 1.5 1.1Significant increase in airwayresistanceHeinrich et al. (1986a)Rat, Wistar, F 19h/day5 days/week140 weeks4.240.35:m MMD56,392 12.5 1.5 1.1 Decrease in dynamic lung compliance;increase in airway resistanceHeinrich et al. (1986a)Cat, inbred, M 8h/day7 days/week124 weeks6.0a12.0b41,66483,32820.233.32.74.42.15.0Decrease in vital capacity, total lungcapacity, and diffusing capacity after2 years; no effect on expiratory flowPepelko et al. (1980b,1981)Moorman et al. (1985)a1 to 61 weeks exposure.b62 to 124 weeks of exposure.5-37 DE was found in rats. Lewis et al. (1989) exposed male cynomolgus monkeys to DE for 7 h/day, 5 days/week for 24 mo. Groups of 15 monkeys were exposed to air, DE (2 mg/m 3 ), coal dust, or combined coal dust and DE. Pulmonary function was evaluated prior to exposure and at 6-mo intervals during the 2-year exposure, 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 restrictive disease. There was, however, evidence of obstructive airway disease as measured by low maximal flow rates in diesel-exposed monkeys. At 18 mo of exposure, forced expiratory flow at 25% of vital capacity and forced expiratory flow normalized to FVC were decreased. The measurement of forced expiratory flow at 40% of total lung capacity was significantly decreased at 12, 18, and 24 mo of exposure. The finding of an obstructive effect in monkeys contrasts with the finding of restrictive type effects in other laboratory animal species (Vinegar et al., 1980, 1981a; Mauderly et al., 1988; Pepelko et al., 1980b, 1981) and suggests a possible difference in effect between primate and small animal respiratory tracts. In these monkeys there were no specific histopathological effects reported (see next section), although particle aggregates were reported in the distal airways, suggesting more small airway deposition. Gross (1981) exposed rats for 20 h/day, 5.5 days/week for 87 weeks to DE containing 1.5 mg/m 3 DPM. When the data were normalized (e.g., indices expressed in units of airflow or volume for each animal by its own forced expiratory volume), there were no apparent functionally significant changes occurring in the lungs at 38 weeks of exposure that might be attributable to the inhalation of DE. After 87 weeks of exposure, functional residual capacity (FRC) and its component volumes (expiratory reserve [ER] and residual volume [RV]), maximum expiratory flow (MEF) at 40% FVC, MEF at 20% FVC, and FEV 0.1 were significantly greater in the diesel-exposed rats. An observed increase in airflow at the end of the forced expiratory maneuver when a decreased airflow would be expected from the increased FRC, ER, and RV data (the typical scenario of human pulmonary disease) showed these data to be inconsistent with known clinically significant health effects. Furthermore, although the lung volume changes in the diesel-exposed rats could have been indicative of emphysema or chronic obstructive lung disease, this interpretation was contradicted by the airflow data, which suggest simultaneous lowering of the resistance of the distal airways. Heinrich et al. (1982) evaluated the pulmonary function of rats exposed to a concentration of 3.9 mg/m 3 DPM for 7 to 8 h/day, 5 days/week for 2 years. When compared with a control group, no significant changes in respiratory rate, minute volume, compliance, or resistance occurred in the exposed group (number of rats per group was not stated). Chinese hamsters (eight or nine per group) were exposed 8 h/day, 7 days/week, for 6 mo to concentrations of either about 6 mg/m 3 or about 12 mg/m 3 DPM (Vinegar et al., 1980,5-38 1981a,b). Vital capacity, vital capacity/lung weight ratio, residual lung volume by water displacement, and CO 2 diffusing capacity decreased significantly in hamsters exposed to 6 mg/m 3 DPM. Static deflation volume-pressure curves showed depressed deflation volumes for diesel-exposed hamsters when volumes were corrected for body weight and even greater depressed volumes when volumes were corrected for lung weight. However, when volumes were expressed as percentage of vital capacity, the diesel-exposed hamsters had higher lung
volumes at 0 and 5 cm H 2 O. In the absence of confirmatory histopathology, the authors tentatively concluded that these elevated lung volumes and the significantly reduced diffusing capacity in the same hamsters were indicative of possible emphysematous changes in the lung. Similar lung function changes were reported in hamsters exposed at 12 mg/m 3 DPM, but detailed information was not reported. It was stated, however, that the decrease in vital capacity was 176% greater in the second experiment than in the first. Mauderly et al. (1988; see also McClellan et al., 1986) examined the impairment of respiratory function in rats exposed for 7 h/day, 5 days/week for 24 mo to diluted DE with 0.35, 3.5, or 7.1 mg/m 3 DPM. After 12 mo of exposure to the highest concentration of DE, the exposed rats (n = 22) had lower total lung capacity (TLC), dynamic lung compliance (C dyn ), FVC, and CO diffusing capacity than controls (n = 23). After 24 mo of exposure to 7.1 mg/m 3 DPM, mean TLC, C dyn , quasi-static chord compliance, and CO diffusing capacity were significantly lower than control values. Nitrogen washout and percentage of FVC expired in 0.1 s were significantly greater than control values. There was no evidence of airflow obstruction. The functional alterations were attributed to focal fibrotic and emphysematous lesions and thickened alveolar membranes observed by histological examination. Similar functional alterations and histopathologic lesions were observed in the rats exposed to 3.5 mg/m 3 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/m 3 group at any time during the study nor were there reported histopathologic changes in this group. Mauderly et al. (1989) examined the effects of DE on normal rats and on rats with experimentally induced pulmonary emphysema to see if emphysematous rats have increased susceptibility to DPM. The results from parallel lifetime exposures of these 2 groups of rats at 3.5 mg/m 3 DPM showed that only possibly 1 of 65 measured parameters gave results suggesting that rats with emphysematous lungs might be more susceptible than rats with normal lungs to the effects of DE exposure. Additional studies were conducted by Heinrich et al. (1986a,b; see also Stöber, 1986) on the effects of long-term exposure to DE on the pulmonary function of hamsters and rats. The exhaust was diluted to achieve a concentration of 4.24 mg/m 3 DPM; exposures were for 19 h/day, 5 days/week for a maximum of 120 weeks (hamsters) or 140 weeks (rats). After 1 year of exposure to the DE, the hamsters exhibited a significant increase in airway resistance and a5-39 nonsignificant reduction in lung compliance. For the same time period, rats showed increased lung weights, a significant decrease in C dyn , and a significant increase in airway resistance. These indices did not change during the second year of exposure. Syrian hamsters and rats were exposed to 0.7, 2.2, or 6.6 mg/m 3 DPM for five 16-h periods per week for 2 years (Brightwell et al., 1986). There were no treatment-related changes in pulmonary function in the hamster. Rats exposed to the highest concentration of DE exhibited changes in pulmonary function (data not presented) that were reported to be consistent
with a concentration-related obstructive and restrictive disease. Pepelko et al. (1980b; 1981; see also Pepelko, 1982b) and Moorman et al. (1985) measured the lung function of adult cats chronically exposed to DE. The cats were exposed for 8 h/day and 7 days/week for 124 weeks. Exposures were at 6 mg/m 3 for the first 61 weeks and 12 mg/m 3 from weeks 62 to 124. No definitive pattern of pulmonary function changes was observed following 61 weeks of exposure; however, a classic pattern of restrictive lung disease was found at 124 weeks. The significantly reduced lung volumes (TLC, FVC, FRC, and inspiratory capacity [IC]) and the significantly lower single-breath diffusing capacity, coupled with normal values for dynamic ventilatory function (mechanics of breathing), indicate the presence of a lesion that restricts inspiration but does not cause airway obstruction or loss of elasticity. This pulmonary physiological syndrome is consistent with an interstitial fibrotic response that was later verified by histopathology (Plopper et al., 1983). Pulmonary function impairment has been reported in rats, hamsters, cats, and monkeys chronically exposed to DE. In all species but the monkey, the pulmonary function testing results have been consistent with restrictive lung disease. The monkeys demonstrated evidence of small airway obstructive responses. The disparity between the findings in monkeys and those in rats, hamsters, and cats could be in part the result of increased particle retention in the smaller species resulting from (1) exposure to DE that has higher airborne concentrations of gases, vapors, and particles and/or (2) longer duration of exposure. The nature of the pulmonary impairment is also dependent on the site of deposition and routes of clearance, which are determined by the anatomy and physiology of the test laboratory species and the exposure regimen. The data on pulmonary function effects raise the possibility that DE produces small airway disease in primates compared with primarily alveolar effects in small animals and that similar changes might be expected in humans and monkeys. The findings of Nikula et al. (1997a,b) suggest that a larger fraction of particles are translocated to the interstitium of the respiratory tract in primates that are heavily exposed than in rats that are heavily exposed, including the interstitium of the respiratory bronchioles, an anatomical site absent in rats. Nikula and co-workers’ pulmonary histopathological findings may have a relationship to these functional findings (see Chapter 3 for a complete discussion). Unfortunately, the available data in primates are too
limited to draw clear conclusions.5-40 5.1.3.3.3. Lung morphology, biochemistry, and lung lavage analysis . Several studies have examined the morphological, histological, and histochemical changes occurring in the lungs of laboratory animals chronically exposed to DE. The histopathological effects of diesel exposure in the lungs of laboratory animals are summarized in Table 5-6, ranked in order of C × T. Table 5-6 also contains an expanded description of exposures. Kaplan et al. (1982) performed macroscopic and microscopic examinations of the lungs of rats, mice, and hamsters exposed for 20 h/day, 7 days/week for 3 mo to DE containing 1.5 mg/m 3 DPM. Gross examination revealed diffuse and focal deposition of the diesel particles that produced a grayish overall appearance of the lungs with scattered, denser black areas. There was clearance of particles via the lymphatics to regional lymph nodes. Microscopic examination revealed no anatomic changes in the upper respiratory tract; the mucociliary border was normal in appearance. Most of the particles were in macrophages, but some were free as small aggregates on alveolar and bronchiolar surfaces. The particle-laden macrophages were often in masses near the entrances of the lymphatic drainage and respiratory ducts. Associated with these masses was a minimal increase in the thickness of the alveolar walls; however, the vast majority of the particles elicited no response. After 6 mo of recovery, the lungs of all three species contained considerably less pigment, as assessed by gross pathological and histopathological
examinations. Lewis et al. (1989; see also Green et al., 1983) performed serial histological examinations of rat lung tissue exposed to DE containing 2 mg/m 3 DPM for 7 h/day, 7 days/week for 2 years. Accumulations of black-pigmented AMs were seen in the alveolar ducts adjacent to terminal bronchioles as early as 3 mo of exposure, and particles were seen within the interstitium of the alveolar ducts. These macular lesions increased in size up to 12 mo of exposure. Collagen or reticulum fibers were seen only rarely in association with deposited particles; the vast majority of lesions showed no evidence of fibrosis. There was no evidence of focal emphysema with the macules. Multifocal histiocytosis (24% of exposed rats) was observed only after 24 mo of exposure. These lesions were most commonly observed subpleurally and were composed of collections of degenerating macrophages and amorphous granular material within alveoli, together with fibrosis and chronic inflammatory cells in the
interstitium.5-41Table 5-6. Histopathological effects of diesel exhaust in the lungs of laboratory animalsSpecies/sexExposureperiodParticles(mg/m3)C × T(mgAh/m3)CO(ppm)NO2(ppm)SO2(ppm)Effects StudyRat, F344, M;Mouse, A/J, M;Hamster, Syrian, M20 h/day7 days/week12-13 weeks1.50.19:m MDD2,520-2,730 — — —Inflammatory changes, increase in lungweight, increase in thickness of alveolarwallsKaplan et al. (1982)Monkey, Cynomolgus,M7 h/day5 days/week104 weeks2.00.23–0.36:mMDD7,280 11.5 1.5 0.8AM aggregation; no fibrosis,inflammation, or emphysemaLewis et al. (1989)Rat, F344, M,F 7 h/day5 days/week104 weeks2.00.23–0.36:mMDD3,640 11.5 1.5 0.8Multifocal histiocytosis, inflammatorychanges, Type II cell proliferation,fibrosisBhatnagar et al.(1980)Pepelko (1982a)Rat, Sprague-Dawley,M; Mouse, A/HEJ, M8 h/day7 days/week39 weeks6.0 13,104 — — — Increasein lung protein content andcollagen synthesis but a decrease inoverall lung protein synthesis in bothspecies; prolylhydroxylase activityincreased in rats in uteroBhatnagar et al.(1980)Pepelko (1982a)Hamster, Chinese, M 8h/day5 days/week26 weeks6.012.06,24012,480——————Inflammatory changes, AM accumulation, thickened alveolar lining, Type IIcell hyperplasia,edema, increase incollagenPepelko (1982b)Hamster, Syrian, M,F 7-8 h/day5 days/week120 weeks3.90.1:m MDD16,380-18,720 18.5 1.2 3.1Inflammatory changes, 60%adenomatous cell proliferationHeinrich et al. (1982)Rat, Wistar, M 6h/day5 days/week87 weeks8.30.71:m MDD21,663 50.0 4.0-6.0 —Inflammatory changes, AM aggregation,alveolar cell hypertrophy, interstitialfibrosis, emphysema (diagnostic methodology not described)Karagianes et al.(1981)Rat, F344, F 8h/day7 days/week104 weeks4.9 28,538 7.0 1.8 13.1 Type II cell proliferation, inflammatorychanges, bronchial hyperplasia, fibrosisIwai et al. (1986)Rat, F344, M, F;Mouse, CD-1,M, F7 h/day5 days/week130 weeks0.353.57.10.23:m MDD1,59215,92531,8502.916.529.70.050.340.68———Alveolar and bronchiolar epithelialmetaplasia in rats at 3.5 and 7.0 mg/m3,fibrosis at 7.0 mg/m3in rats and mice,inflammatory changesMauderly et al.(1987a)Henderson et al.(1988a)5-42Table 5-6. Histopathological effects of diesel exhaust in the lungs of laboratory animals (continued)Species/sexExposureperiodParticles(mg/m3)C × T(mgAh/m3)CO(ppm)NO2(ppm)SO2(ppm)Effects StudyRats, SPF 3447 h/day5days/week104 weeks2 mg/m3coaldust (CD)2 mg/m3DPM1 mg/m3CD + 1 mg/m3DPM— — — — • Assessed pharmacologicalresponses of rat airwaysmooth muscle in vitro• Maximal contractileresponses to acetylcholine oftissues from CD-, DPM-, andCD + DPM- exposed animalssignificantly increased;effects of CD and DPM wereadditive• Maximal relaxation responseto isoproterenol increasedsignificantly by CD + DPMexposure, but not byindividual treatments• The results indicate thatchronic exposure to CD,DPM, and CD + DPMproduce differentialmodifications in the behaviorof rat airway smooth muscleFeden et al. (1985)Rat, Wistar, F;Mouse, NMRI, F(7 mg/m3only)18 h/day5 days/week24 mo0.82.56.987,40021,80061,7002.68.321.20.31.23.80.31.13.4Bronchioalveolar hyperplasia, interstitialfibrosis in all groups. Severity andincidence increase with exposureconcentrationHeinrich et al. (1995)Mouse, NMRI, F;C57BL/6N, F18 h/day5 days/week13.5 mo (NMRI)24 mo(C57BL/N)6.98 35,500 - NMRI38,300 - C5714.2 2.3 2.8 No increase in tumors. Noncancereffects not discussedMouse 4.5 No increase in tumorsNoncancer effects not discussedRat, M, F,F344/Jcl.16 h/day6 days/week130 weeks0.11a0.41a1.08a2.31a3.72b1,3735,11713,47828,82946,3361.232.123.967.1012.90.080.260.701.413.000.381.062.424.704.57Inflammatory changes Type II cellhyperplasia and lung tumors seen at>0.4 mg/m3; shortening and loss of ciliain trachea and bronchiResearch Committeefor HERP Studies(1988)Mouse, NMRI, F19 h/day5 days/week120 weeks4.24 48,336 12.5 1.51.1 Inflammatory changes, bronchioloalveolar hyperplasia, alveolar lipoproteinosis, fibrosisHeinrich et al.(1986a)Table 5-6. Histopathological effects of diesel exhaust in the lungs of laboratory animals (continued)Species/sexExposureperiodParticles(mg/m3)C × T(mgAh/m3)CO(ppm)NO2(ppm)SO2(ppm)Effects Study

5-43Rat, Wistar, F 19h/day5 days/week140 weeks4.24 56,392 12.5 1.5 1.1Thickened alveolar septa; AMaggregation; inflammatory changes;hyperplasia; lung tumorsHeinrich et al.(1986a)Guinea Pig, Hartley,M 20 h/day5.5 days/week104 weeks0.250.751.56.02,8608,58017,16068,640————————————Minimal response at 0.25 andultrastructural changes at 0.75 mg/m3;thickened alveolar membranes; cellproliferation; fibrosis at 6.0 mg/m3;increase in PMN at 0.75 mg/m3and1.5 mg/m3Barnhart et al. (1981,1982)Vostal et al. (1981)Wallace et al. (1987)Cat, inbred, M 8h/day7 days/week124 weeks6.0c12.0d41,66483,32820.233.22.74.42.15.0Inflammatory changes, AM aggregation,bronchiolar epithelial metaplasia, Type IIcell hyperplasia, peribronchiolar fibrosisPlopper et al. (1983)Hyde et al. (1985)Rat, F344, M16 h/day5 days/week23 mo2.446.3319,52050,640——————AM hyperplasia, epithelial hyperplasia,inflammation, septal fibrosis,bronchoalveolar metaplasiaNikula et al. (1995)Mouse, CD-1, M,F7 h/day5 days/week104 weeks0.353.57.10.25:m MDD1,27412,74025,844317300.10.30.7———Exposure-related increase in lung soot,pigment-laden macrophages, lunglesions.Bronchiolization in alveolar ducts at7.1 mg/m3Mauderly et al.(1996)aLight-duty engine.bHeavy-duty engine.c1 to 61 weeks exposure.d62 to 124 weeks of exposure.AM = Alveolar macrophage.PMN = Polymorphonuclear leukocyte.5-44 Epithelial lining cells adjacent to collections of pigmented macrophages showed a marked Type II cell hyperplasia; degenerative changes were not observed in Type I cells. Histological examination of lung tissue 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 cytoplasm of macrophages in the alveolar spaces as well as the interstitium. Fibrosis, focal emphysema, or inflammation was not observed. No specific histopathological lesions were reported for the monkey. Nikula et al. (1997a,b) reevaluated the lung tissue from this study. They concluded that there were no significant differences in the amount of retained particulate matter between monkeys and rats exposed under the same conditions. The rats, however, retained a greater portion of the particulate matter in lumens of the alveolar ducts and alveoli than did the monkeys. Conversely, monkeys retained a greater portion of the particulate material in the interstitium than did rats. Aggregations of particle-laden macrophages in the alveoli were rare, and there were few signs of particle-associated inflammation in the monkeys. Minimal histopathologic lesions were detected in the interstitium. Histopathological effects of DE on the lungs of rats have been investigated by the Health Effects Research Program on Diesel Exhaust (HERP) in Japan (Ishinishi et al., 1986, 1988). Both light-duty (LD) and heavy-duty (HD) diesel engines were used. The exhaust was diluted to achieve nominal concentrations of 0.1 (LD only), 0.4 (LD and HD), 1 (LD and HD), 2 (LD and HD), and 4 (HD only) mg/m 3 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/m 3 DPM or less. At concentrations above 0.4 mg/m 3 DPM, severe morphological changes were observed. These changes consisted of shortened and absent cilia in the tracheal and bronchial epithelium, marked hyperplasia of the bronchiolar epithelium, and swelling of the Type II cellular epithelium. These lesions appeared to increase in severity with increases in exhaust concentration and duration of exposure. There was no difference in the degree of changes in pulmonary pathology at the same concentrations between the LD and the HD series. Heinrich et al. (1982) investigated histological changes occurring in the respiratory tract of hamsters exposed to DE. Exposures were for 7 to 8 h/day, 5 days/week for 104 weeks to DE diluted to achieve a concentration of 3.9 mg/m 3 DPM. Significantly higher numbers of hamsters in the group exposed to DE exhibited definite proliferative changes in the lungs compared with the groups exposed to particle-free DE or clean air. Sixty percent of these changes were described as adenomatous proliferations. Heinrich et al. (1995) reported increased incidence and severity of bronchioloalveolar hyperplasia in rats exposed to 0.8, 2.5, and 7 mg/m 3 . The lesion in the lowest concentration group was described as very slight to moderate. Slight to moderate interstitial fibrosis also increased in incidence and severity in all exposed groups, but incidences were not reported. This5-45 chronic study also exposed NMRI mice to 7 mg/m 3 for 13.5 mo and both NMRI and C56BL/6N mice to 4.5 mg/m 3 for 24 mo. Noncancer histological endpoints are not discussed in any detail in the report, which is focused on the carcinogenicity of diesel as compared with titanium
dioxide and carbon black. Iwai et al. (1986) performed serial histopathology on the lungs of rats at 1, 3, 6, 12, and 24 mo of exposure to DE. Exposures were for 8 h/day, 7 days/week for 24 mo; the exposure atmosphere contained 4.9 mg/m 3 DPM. At 1 and 3 mo of exposure, there were minimal histological changes in the lungs of the exposed rats. After 6 mo of exposure, there were particle-laden macrophages distributed irregularly throughout the lung and a proliferation of Type II cells with adenomatous metaplasia in areas where the macrophages had accumulated. After 1 year of exposure, foci of heterotrophic hyperplasia of ciliated or nonciliated bronchiolar epithelium on the adjacent alveolar walls were more common, the quantity of deposited particulate matter increased, and the number of degenerative AMs and proliferative lesions of Type II or bronchiolar epithelial cells increased. After 2 years of exposure, there was a fibrous thickening of the alveolar walls, mast-cell infiltration with epithelial hyperplasia in areas where the macrophages had accumulated, and neoplasms. Heinrich et al. (1986a; see also Stöber, 1986) performed histopathologic examinations of the respiratory tract of hamsters, mice, and rats exposed to DE that had 4 mg/m 3 DPM. Exposures were for 19 h/day, 5 days/week; the maximum exposure period was 120 weeks for hamsters and mice and 140 weeks for rats. Histological examination revealed different levels of response among the three species. In hamsters, the exhaust produced thickened alveolar septa, bronchioloalveolar hyperplasia, and what were termed emphysematous lesions (diagnostic methodology not described). In mice, bronchoalveolar hyperplasia occurred in 64% of the mice exposed to the exhaust and in 5% of the controls. Multifocal alveolar lipoproteinosis occurred in 71% and multifocal interstitial fibrosis occurred in 43% of the mice exposed to exhaust but in only 4% of the controls. In exposed rats, there were severe inflammatory changes in the lungs, as well as thickened septa, foci of macrophages, and hyperplastic and metaplastic lesions. Nikula et al. (1995) reported in detail the nonneoplastic effects in male and female F344 rats exposed to 2.4 or 6.3 mg/m 3 of DPM. At 3 mo in the low-concentration group, enlarged particle-containing macrophages were found with minimal aggregation. With higher concentration and longer duration of exposure, the number and size of macrophages and aggregates increased. Alveolar epithelial hyperplasia was found starting at 3 mo and in all rats at 6 mo. These lesions progressed to chronic active inflammation, alveolar proteinosis, and septal fibrosis at 12 mo. Other lesions observed late in the study included bronchiolar-alveolar metaplasia, squamous metaplasia, and squamous cysts. This study reports in detail the progression of lesions in DE exposure and finds relatively little difference between the lesions caused by DE exposure and exposure to similar levels of carbon black particles.5-46 The effects of DE on the lungs of rats exposed to 8.3 ± 2.0 mg/m 3 DPM were investigated by Karagianes et al. (1981). Exposures were for 6 h/day, 5 days/week, for 4, 8, 16, or 20 mo. Histological examinations of lung tissue noted focal aggregation of particle-laden AMs, alveolar histiocytosis, interstitial fibrosis, and alveolar emphysema (diagnostic methodology not described). Lesion severity was related to length of exposure. No significant differences were noted in lesion severity among the DE, the DE plus coal dust (5.8 ± 3.5 mg/m 3 ), or the high-concentration (14.9 ± 6.2 mg/m 3 ) coal dust exposure groups following 20 mo of exposure. Histological changes in the lungs of guinea pigs exposed to diluted DE containing either 0.25, 0.75, 1.5, or 6.0 mg/m 3 DPM were reported by Barnhart et al. (1981; 1982). Exposures at 0.75 and 1.5 mg/m 3 for 2 weeks to 6 mo resulted in an uptake of exhaust particles by three alveolar cell types (AMs, Type I cells, and interstitial macrophages) and also by granulocytic leukocytes (eosinophils). The alveolar-capillary membrane increased in thickness as a result of an increase in the absolute tissue volume of interstitium and Type II cells. In a continuation of these studies, guinea pigs were exposed to DE (up to 6.0 mg/m 3 DPM) for 2 years (Barnhart et al., 1982). A minimal tissue response occurred at a concentration of 0.25 mg/m 3 . After 9 mo of exposure, there was a significant increase, about 30%, in Type I and II cells, endothelial cells, and interstitial cells over concurrent age-matched controls; by 24 mo only macrophages and Type II cells were significantly increased. As in the earlier study, ultrastructural evaluation showed that Type I cells, AMs, and eosinophils phagocytized the diesel particles. Exposure to 0.75 mg/m 3 for 6 mo resulted in fibrosis in regions of macrophage clusters and in focal Type II cell proliferation. No additional information was provided regarding the fibrotic changes with increasing concentration or duration of exposure. With increasing concentration/duration of DE exposure, Type II cell clusters occurred in some alveoli. Intraalveolar debris was particularly prominent after exposures at 1.5 and 6.0 mg/m 3 and consisted of secretory products from Type II
cells. In studies conducted on hamsters, Pepelko (1982b) found that the lungs of hamsters exposed for 8 h/day, 7 days/week for 6 mo to 6 or 12 mg/m 3 DPM were characterized by large numbers of black AMs in the alveolar spaces, thickening of the alveolar epithelium, hyperplasia of Type II cells, and edema. Lungs from rats and mice exposed to 0.35, 3.5, or 7.1 mg/m 3 (0.23 to 0.26 

### : m mass median diameter [MMD]) for 7 h/day and 5 days/week showed pathologic lesions (Mauderly et al., 1987a; Henderson et al., 1988a). After 1 year of exposure at 7.1 mg/m 3 , the lungs of the rats exhibited focal areas of fibrosis; fibrosis increased with increasing duration of exposure and was observable in the 3.5-mg/m 3 group of rats at 18 mo. The severity of inflammatory responses and fibrosis was directly related to the exposure level. In the 0.35 mg/m 3 group of rats, there was no inflammation or fibrosis. Although the mouse lungs contained higher burdens5-47 of diesel particles per gram of lung weight at each equivalent exposure concentration, there was substantially less inflammatory reaction and fibrosis than was the case in rats. Fibrosis was observed only in the lungs of mice exposed at 7.1 mg/m 3 and consisted of fine fibrillar thickening of occasional alveolar septa. Histological examinations were performed on the lungs of cats initially exposed to 6 mg/m 3 DPM for 61 weeks and subsequently increased to 12 mg/m 3 for Weeks 62 to 124 of exposure. Plopper et al. (1983; see also Hyde et al., 1985) concluded from the results of this study that exposure to DE produced changes in both epithelial and interstitial tissue compartments and that the focus of these lesions in the peripheral lung was the centriacinar region where the alveolar ducts join the terminal conducting airways. This conclusion was based on the following evidence. The epithelium of the terminal and respiratory bronchioles in exposed cats consisted of three cell types (ciliated, basal, and Clara cells) compared with only one type (Clara cells) in the controls. The proximal acinar region showed evidence of peribronchial fibrosis and bronchiolar epithelial metaplasia. Type II cell hyperplasia was present in the proximal interalveolar septa. The more distal alveolar ducts and the majority of the rest of the parenchyma were unchanged from controls. Peribronchial fibrosis was greater at the end of 6 mo in clean air following exposure, whereas the bronchiolar epithelial metaplasia was most severe at the end of exposure. Following an additional 6 mo in clean air, the bronchiolar epithelium more closely resembled the control epithelial cell population. Wallace et al. (1987) used transmission electron microscopy (TEM) to determine the effect of DE on the intravascular and interstitial cellular populations of the lungs of exposed rats and guinea pigs. Exposed animals and matched controls were exposed to 0.25, 0.75, 1.5, or 6.0 mg/m 3 DPM for 2, 6, or 10 weeks or 18 mo. The results inferred the following: (1) exposure to 6.0 mg/m 3 for 2 weeks was insufficient to elicit any cellular response, (2) both species demonstrated an adaptive multicellular response to DE, (3) increased numbers of fibroblasts were found in the interstitium from week 6 of exposure through month 18, and (4) there was no significant difference in either cell type or number in alveolar capillaries, but there was a significant increase at 18 mo in the mononuclear population in the interstitium of both species. Additional means for assessing the adverse effects of DE on the lung are to examine biochemical and cytological changes in bronchoalveolar lavage fluid (BALF) and in lung tissue. Fedan et al. (1985) performed studies to determine whether chronic exposure of rats affected the pharmacologic characteristics of rat airway smooth muscle. Concentration-response relationships for tension changes induced with acetylcholine, 5-hydroxytryptamine, potassium chloride, and isoproterenol were assessed in vitro on isolated preparations of airway smooth muscle (trachealis). Chronic exposure to DE significantly increased the maximal contractile responses to acetylcholine compared with control values; exposure did not alter the sensitivity5-48 (EC 50 values) of the muscles to the agonists. Exposures were to DE containing 2 mg/m 3 DPM for 7 h/day, 5 days/week for 2 years. Biochemical studies of BALF obtained from hamsters and rats revealed that exposures to DE caused significant increases in lactic dehydrogenase, alkaline phosphatase, glucose-6- phosphate dehydrogenase (G6P-DH), total protein, collagen, and protease (pH 5.1) after approximately 1 year and 2 years of exposure (Heinrich et al., 1986a). These responses were generally much greater in rats than in hamsters. Exposures were to DE containing 4.24 mg/m 3 DPM for 19 h/day, 5 days/week for 120 (hamsters) to 140 (rats) weeks. Protein, 

### $ -glucuronidase activity, and acid phosphatase activity were significantly elevated in BALF obtained from rats exposed to DE containing 0.75 or 1.5 mg/m 3 DPM for 12 mo (Strom, 1984). Exposure for 6 mo resulted in significant increases in acid phosphatase activity at 0.75 mg/m 3 and in protein, 

### $ -glucuronidase, and acid phosphatase activity at the 1.5 mg/m 3 concentration. Exposure at 0.25 mg/m 3 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 animals exposed under the same conditions and at the same site as reported on by Strom (1984). In most cases, exposures at 0.25 mg/m 3 did not cause any significant changes. The DNA content in lung tissue and the rate of collagen synthesis were significantly increased at 1.5 mg/m 3 DPM after 6 mo. Collagen deposition was not affected. Total lung collagen content increased in proportion to the increase in lung weight. The activity of prolyl hydroxylase was significantly increased at 12 weeks at 0.25 and 1.5 mg/m 3 ; it then decreased with age. Lysal oxidase activity did not change. After 9 mo of exposure, there were significant increases in lung phospholipids in rats and guinea pigs exposed to 0.75 mg/m 3 and in lung cholesterol in rats and guinea pigs exposed to 1.5 mg/m 3 . Pulmonary prostaglandin dehydrogenase activity was stimulated by an exposure at 0.25 mg/m 3 but was not affected by exposure at 1.5 mg/m 3 (Chaudhari et al., 1980, 1981). Exposures for 12 or 24 weeks resulted in a concentration-dependent lowering of this enzyme activity. Exposure of male rats and guinea pigs at 0.75 mg/m 3 for 12 weeks did not cause any changes in glutathione levels of the lung, heart, or liver. Rats exposed for 2 mo at 6 mg/m 3 showed a significant depletion of hepatic glutathione, whereas the lung showed an 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 in lung or liver tissue of exposed rats and guinea pigs. Bhatnagar et al. (1980; see also Pepelko, 1982a) evaluated changes in the biochemistry of lung connective tissue of diesel-exposed rats and mice. The mice were exposed for 8 h/day and 7 days/week for up to 9 mo to exhaust containing 6 mg/m 3 DPM. Total lung protein content was measured, as was labeled proline and labeled leucine. Leucine incorporation is an index of total protein synthesis, although collagen is very low in leucine. Proline incorporation reflects5-49 collagen synthesis. Amino acid incorporation was measured in vivo in the rat and in short-term organ culture in mice. Both rats and mice showed a large increase in total protein (41% to 47% in rats), while leucine incorporation declined and proline incorporation was unchanged. These data are consistent with an overall depression of protein synthesis in diesel-exposed animals and also with a relative increase in collagen synthesis compared to other proteins. The increase in collagen synthesis suggested proliferation of connective tissue and possible fibrosis (Pepelko, 1982a). A number of reports (McClellan et al., 1986; Mauderly et al., 1987a, 1990a; Henderson et al., 1988a) have addressed biochemical and cytological changes in lung tissue and BALF of rodents exposed for 7 h/day, 5 days/week for up to 30 mo at concentrations of 0, 0.35, 3.5, or 7.1 mg/m 3 DPM. At the lowest exposure level (0.35 mg/m 3 ), no biochemical or cytological changes occurred in the BALF or in lung tissue in either Fischer 344 rats or CD-1 mice. Henderson et al. (1988a) provide considerable time-course information on inflammatory events taking place throughout a chronic exposure. A chronic inflammatory response was seen at the two higher exposure levels in both species, as evidenced by increases in inflammatory cells (macrophages and neutrophils), cytoplasmic and lysosomal enzymes (lactate dehydrogenase, glutathione reductase, and 

### $ -glucuronidase), and protein (hydroxyproline) in BALF. Analysis of lung tissue indicated similar changes in enzyme levels as well as an increase in total lung collagen content. After 18 mo of exposure, lung tissue glutathione was depleted in a concentration-dependent fashion in rats but was slightly increased in mice. Lavage fluid levels of glutathione and glutathione reductase activity increased in a concentration-dependent manner and were higher in mice than in rats. Rats exposed for up to 17 days to diluted DE (3.5 mg/m 3 DPM) had a fivefold increase in the bronchoconstrictive prostaglandin PGF 2" and a twofold increase in the inflammatory
leukotriene LTB 4 . In similarly exposed mice, there was a twofold increase in both parameters. These investigators (Henderson et al., 1988a,b) concluded that the release of larger amounts of such mediators of inflammation from the alveolar phagocytic cells of rats accounted for the greater fibrogenic response seen in that species. Biochemical analysis of lung tissue from cats exposed for 124 weeks and held in clean air for an additional 26 weeks indicated increases of lung collagen; this finding was confirmed by an observed increase in total lung wet weight and in connective tissue fibers estimated morphometrically (Hyde et al., 1985). Exposures were for 7 h/day, 5 days/week at 6 mg/m 3 DPM for 61 weeks and at 12 mg/m 3 for weeks 62 to 124. Heinrich et al. (1995) reported on bronchoalveolar lavage in animals exposed for 24 mo and found exposure-related increases in lactate dehydrogenase, 

### $ -glucuronidase, protein, and hydroxyproline in groups exposed to 2.5 or 7 mg/m 3 , although detailed data are not presented. Lavage analyses were not carried out in concurrent studies in mice.5-50 The pathogenic sequence following the inhalation of DE as determined histopathologically and biochemically begins with the interaction of diesel particles with airway epithelial cells and phagocytosis by AMs. The airway epithelial cells and activated macrophages release chemotactic factors that attract neutrophils and additional AMs. As the lung burden of DPM increases, there is an aggregation of particle-laden AMs in alveoli adjacent to terminal 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 macrophages release mediators of inflammation and oxygen radicals that deplete a biochemical defense mechanism of the lung (i.e., glutathione). As will be described later in more detail, other defense mechanisms are affected, particularly the decreased 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 DE below which adverse structural and biochemical effects may not occur in the lung; however, differences in the anatomy and pathological responses of laboratory animals coupled with their lifespans compared with humans make a determination of human levels of exposure to DE without resultant pulmonary injury a difficult and challenging endeavor.

5.1.3.3.4. Effects on pulmonary defense mechanisms. The respiratory system has a number of defense mechanisms that negate or compensate for the effects produced by the injurious substances that repeatedly insult the upper respiratory tract, the tracheobronchial airways, and the alveoli. The effects of exposure to DE on the pulmonary defense mechanisms of laboratory animals as well as more details on exposure atmosphere are summarized in Table 5-7 and ranked by cumulative exposure (C × T). Several studies have been conducted investigating the effect of inhaled DE on the deposition and fate of inert tracer particles or diesel particles themselves. Lung clearance of deposited particles occurs in two distinct phases: a rapid phase (hours to days) from the tracheobronchial region via the mucociliary escalator and a much slower phase (weeks to months) from the nonciliated pulmonary region via, primarily but not solely, AMs. Battigelli et al. (1966) reported impaired tracheal mucociliary clearance in vitro in excised trachea from rats exposed for single or repeated exposures of 4 to 6 h at two dilutions of DE that resulted in exposures of approximately 8 and 17 mg/m 3 DPM. The exposure to 17 mg/m 3 resulted in5-51Table 5-7. Effects of exposure to diesel exhaust on the pulmonary defense mechanisms of laboratory animalsSpecies/sexExposureperiodParticles(mg/m3)C × T(mgAh/m3)CO(ppm)NO2(ppm)SO2(ppm)Effects StudyAlveolar macrophage statusGuinea Pig,Hartley20 h/day5.5days/week8 weeks0.251.50.19:m MDD2201,3202.97.5————No significant changes in absolute numbersof AMsChen et. al. (1980)Rat, F344, M 7h/day5 days/week104 weeks2.00.23–0.36:mMDD7,280 11.5 1.5 0.81Little effect on viability, cell number, oxygenconsumption, membrane integrity, lyzomalenzyme activity, or protein content of AMs;decreased cell volume and ruffling of cellmembrane and depressed luminescence ofAMCastranova et al. (1985)Rat, F344, M20 h/day5.5days/week26, 48, or52 weeks0.25a0.75a1.5b0.19:m MDD715-8,580 2.94.87.5——————AM cell counts proportional to concentrationof DPM at 0.75 and 1.5 mg/m3; AMincreased in lungs in response to rate of DPMmass entering lung rather than total DPMburden in lung; increased PMNs wereproportional to inhaled concentrations and/orduration of exposure; PMNs affiliated withclusters of aggregated AM rather than DPMStrom (1984)Vostal et al. (1982)Rat F344/Crl,M, FMouse, CD, M, F7 h/day5 days/week104 weeks(rat),78 weeks(mouse)0.353.57.00.25:m MDD1,274c12,740c25,480c2.916.529.70.050.340.68———Significant increases of AM in rats and miceexposed to 7.0 mg/m3DPM for 24 and 18mo, respectively, but not at concentrations of3.5 or 0.35 mg/m3DPM for the same exposure durations; PMNs increased in a dosedependent fashion in both rats and miceexposed to 3.5 or 7.0 mg/m3DPM and weregreater in mice than in ratsHenderson et al. (1988a)Rat, Wistar, F 18h/day5 days/week24 mo0.82.57.17,40021,80061,7002.68.321.20.31.13.4———Changes in differential cell counts in lunglavageHeinrich et al. (1995)Rat, F344/Crl, M 7h/day5 days/week24 mo3.49 12,704 9.8 1.2 —Significantly reduced AM in lavage at 24 mo Mauderly et al. (1990a)5-52Table 5-7. Effects of exposure to diesel exhaust on the pulmonary defense mechanisms of laboratory animals(continued)Species/sexExposureperiodParticles(mg/m3)C × T(mgAh/m3)CO(ppm)NO2(ppm)SO2(ppm)Effects StudyClearanceRat, M, F 7h/day5 days/week12 weeks0.21.04.50.25:m MDD844201,890—————————Evidence of apparent speeding of trachealclearance at the 4.5 mg/m3level after 1 weekof99mTc macroaggregated-albumin andreduced clearance of tracer aerosol in each ofthe three exposure levels at 12 weeks; indication of a lower percentage of ciliated cells atthe 1.0 and 4.5 mg/m3levelsWolff and Gray (1980)Rat, Wistar, F 18h/day5 days/week24 mo0.82.57.17,40021,80061,7002.68.321.20.31.23.80.31.13.4Significant increase in clearance half-time ofinhaled labeled aerosols in all groups at 3-18moHeinrich et al. (1995)Rat, F344, M,developing 0-6moadult 6-12 mo7 h/day5 days/week6 mo3.55 3,321 7.9 9.5 Clearance of 2:m, aluminosilicate particles.Half-time significantly increased in adult, notdifferent in developing ratsMauderly et al. (1987b)Rat, F344, M,F 7 h/day5 days/week18 weeks0.150.944.1<0.5:m MDD94.55922,583—————————Lung burdens of DPM were concentrationrelated; clearance half-time of DPM almostdouble in 4.1 mg/m3group compared to 0.15mg/m3groupGriffis et al. (1983)Rat, F344, M 7h/day5 days/week26-104weeks2.00.23-0.36:mMDD1,820-7,280 11.5 1.5 0.8 Nodifference in clearance of59Fe3O4particles 1 day after tracer aerosoladministration; 120 days after exposuretracer aerosol clearance was enhanced; lungburden of DPM increased significantlybetween 12 and 24 mo of exposureLewis et al. (1989)Rat, SpragueDawley, M4-6 h/day7 days/week0.1 to 14.3weeks0.98.017.02.5-10,210 ———5.02.78.00.20.61.0Impairment of tracheal mucociliary clearancein a concentration-response mannerBattigelli et al. (1966)5-53Table 5-7. Effects of exposure to diesel exhaust on the pulmonary defense mechanisms of laboratory animals(continued)Species/sexExposureperiodParticles(mg/m3)C × T(mgAh/m3)CO(ppm)NO2(ppm)SO2(ppm)Effects StudyRat, F344,M, F7 h/day5 days/week130 weeks0.353.57.10.25:m MDD1,59315,92531,8502.916.529.70.10.30.7———No changes in tracheal mucociliary clearanceafter 6, 12, 18, 24, or 30 mo of exposure;increases in lung clearance half-times asearly as 6 mo at 7.0 mg/m3level and 18 moat 3.5 mg/m3level; no changes seen at 0.35mg/m3level; after 24 mo of diesel exposure,long-term clearance half-times wereincreased in the 3.5 and 7.0 mg/m3groupsWolff et al. (1987)Rat, F344/Crl, M 7h/day5 days/week24 mo3.49 12,704 9.8 1.2 —Doubling of long-term clearance half-timefor clearance of 1.0:m aluminosilicateparticles. Less effect on clearance in animalswith experimentally induced emphysemaMauderly et al. (1990a)Microbial-induced mortalityMice CD-1, F 7h/day5 days/week4, 12, or26 weeks2.00.23–0.36:mMDD280-1,820 11.5 1.5 0.8Mortality similar at each exposure durationwhen challenged with Ao/PR/8/34 influenzavirus; in mice exposed for 3 and 6 mo, butnot 1 mo, there were increases in thepercentages of mice having lungconsolidation, higher virus growth, depressedinterferon levels, and a fourfold reduction inhemagglutinin antibody levelsHahon et al. (1985)Mice, CR/CD-1, F 8h/day7 days/week2 h up to46 weeks5.3 to 7.911-20,350 19to221.8to3.60.9to2.8Enhanced susceptibility to lethal effects ofS. pyogenesinfections at all exposuredurations (2 and 6 h; 8, 15, 16, 307, and 321days); inconclusive results withS. typhimuriumbecause of high mortalityrates in controls; no enhanced mortalitywhen challenged with A/PR8-3 influenzavirusCampbell et al. (1980, 1981)aChronic exposure lasted 52 weeks.bChronic exposure lasted 48 weeks.cCalculated for 104-week exposure.DPM = Diesel particulate matter.AM = Alveolar macrophage.PMN = Polymorphonuclear leukocyte.5-54 decreased clearance after a single exposure as well as after a cumulative exposure of 34 or 100 h. Clearance was reduced to a lesser extent and in fewer tracheas from animals exposed to 8 mg/m 3 for a cumulative exposure of 40 h. Lewis et al. (1989) found no difference in the clearance of

59 Fe 3 O 4 particles (1.5 

### : m MMAD, 

### F g 1.8) 1 day after dosing control and DE-exposed rats (2 mg/m 3 , 7 h/day, 5 days/week for 8 weeks). Wolff et al. (1987) and Wolff and Gray (1980) studied the effects of both subchronic and chronic DE exposure on the tracheal clearance of particles. Tracheal clearance assessments were made by measuring the retention of radiolabeled technetium macroaggregated-albumin remaining 1 h after instillation in the distal trachea of rats. In the subchronic studies, rats were exposed to 0.2, 1.0, or 4.5 mg/m 3 DPM on a 7 h/day, 5 days/week schedule for up to 12 weeks. After 1 week there was an apparent speeding of tracheal clearance at the 4.5 mg/m 3 exposure level ( p =0.10), which returned toward baseline after 6 weeks and was slightly below the baseline rate at 12 weeks. In the 1.0 mg/m 3 group, there was a progressive significant reduction in the clearance rate at 6 and 12 weeks of exposure. There was a trend toward reduced clearance in the 0.2 mg/m 3 group. Scanning electron micrographs indicated minimal changes in ciliary morphology; however, there was an indication of a lower percentage of ciliated cells at the 1.0 and 4.5 mg/m 3 levels. In the chronic studies, rats were exposed to 0, 0.35, 3.5, or 7.1 mg/m 3 for 7 h/day, 5 days/week for 30 mo. There were no significant differences in tracheal clearance rates between the control group and any of the exposure groups after 6, 12, 18, 24, or 30 mo of exposure. The preexposure measurements for all groups, however, were significantly lower than those during the exposure period, suggesting a possible age effect. The preexposure value for the 3.5-mg/m 3 group was also significantly lower than the control group. There is a substantial body of evidence for an impairment of particle clearance from the bronchiole-alveolar region of rats following exposure to DE. Griffis et al. (1983) exposed rats 7 h/day, 5 days/week for 18 weeks to DE at 0.15, 0.94, or 4.1 mg/m 3 DPM. Lung burdens of the 0.15, 0.94, and 4.1 mg/m 3 levels were 35, 220, and 1,890 

### : g/g lung, respectively, 1 day after the 18-week exposure. The clearance half-time of the DPM was significantly greater, almost double, for the 4.1 mg/m 3 exposure group than for those of the lower exposure groups, 165 ± 8 days versus 99 ± 8 days (0.94 mg/m 3 ) and 87 ± 28 days (0.15 mg/m 3 ), respectively. Chan et al. (1981) showed a dose-related slowing of 14 C-diesel particle clearance in rats preexposed to DE at 0.25 or 6 mg/m 3 particulate matter for 20 h/day, 7 days/week for 7 to 112 days. Clearance was inhibited in the 6 mg/m 3 group when compared by length of exposure or compared with the 0.25 mg/m 3 or control rats at the same time periods. Heinrich et al. (1982) evaluated lung clearance in rats exposed for approximately 18 mo at 3.9 mg/m 3 DPM for 7 to 8 h/day, 5 days/week. Following exposure to 59 Fe 2 O 3 -aerosol, the rats were returned to the DE exposure and the radioactivity was measured over the thoracic area5-55 at subsequent times. The biological half-life of the iron oxide deposited in the rats’ lungs was nearly twice that of controls. Heinrich also used labeled iron oxide aerosols to study clearance in rats exposed to 0.8, 2.5, or 7 mg/m 3 diesel DPM for 24 mo (Heinrich et al., 1995). Clearance measurements were carried out at 3, 12, and 18 mo of exposure. Half-times of clearance were increased in a concentration- and duration-related manner in all exposed groups, with a range of a 50% increase in the 0.8 mg/m 3 group at 3 mo to an 11-fold increase in the 7 mg/m 3 group at 19 mo.
The differential cell counts in these animals were stated to have shown clear effects in the 2.5 and 7 mg/m 3 groups, but specific information about the changes is not reported. Wolff et al. (1987) investigated alterations in DPM clearance from the lungs of rats chronically exposed to DE at 0, 0.35, 3.5, or 7.1 mg/m 3 DPM for 7 h/day, 5 days/week for up to 24 mo. Progressive increases in lung burdens were observed over time in all groups; 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/m 3 exposure levels, respectively. There were significant increases in 16-day clearance half-times of inhaled radiolabeled particles of 67 Ga 2 O 3 (0.1 

### : m MMD) as early as 6 mo at the 7.1 mg/m 3 level and 18 mo at the 3.5 mg/m 3 level; no significant changes were seen at the 0.35 mg/m 3 level at any time point examined. Rats inhaled fused aluminosilicate particles (2 

### : m MMAD) labeled with 134 Cs after 24 mo of DE exposure; long-term clearance half-times were 79, 81, 264, and 240 days for the 0, 0.35, 3.5, and 7.1 mg/m 3 groups, respectively. Differences were significant between the control and the 3.5 and 7.1 mg/m 3 groups ( p < 0.01), but not between the control and the 0.35 mg/m 3 group. Mauderly et al. (1987b) compared the effects of DE in the developing lung to the adult lung by exposing groups of male F344 rats to 3.5 mg/m 3 for 7 h/day, 5 days/week for 6 mo. One group (adult) was exposed between 6 and 12 mo of age, and the other was exposed beginning in utero and until 6 mo of age. Clearance of an inhaled monodisperse 2 

### : m aluminosilicate particle was measured after exposure for 6 mo. The clearance half-time of the slow phase was found to be doubled in the diesel-exposed adult rats compared with age-matched controls and was not significantly affected in developing rat lungs. Mauderly et al. (1990a) compared the effects of DE in normal lungs with rats in which emphysema had been induced experimentally by instillation of elastase 6 weeks before DE exposures. The rats were exposed to 3.5 mg/m 3 DPM for 7 h/day, 5 days/week for 24 mo. Measurements included histopathology, clearance, pulmonary function, lung lavage, and immune response. In the rats that were not pretreated with elastase, there was a significant reduction in the number of macrophages recovered by pulmonary lavage in contrast to the increases in macrophages reported by Strom (1984) and Henderson et al. (1988). The half-time of the slow phase of clearance of inhaled, 1 

### : m, monodisperse particles was doubled in the animals without elastase pretreatment. The elastase pretreatment did not affect clearance in5-56 unexposed animals but significantly reduced the effect of diesel. The clearance half-time was significantly less in elastase-pretreated, diesel-exposed animals than in diesel-exposed normal animals. Many other effects measured in this study were also less affected by diesel exposure in elastase-treated animals. Measurements of lung burden of DPM showed that elastase-pretreated animals accumulated less than half as much DPM mass as normal animals exposed at the same time, suggesting that the difference in effect could be explained by differences in dose to the lung. The composite results of this study indicate that, at least in a murine laboratory animal species, the presence of a pulmonary restrictive disease such as emphysema does not seem to exacerbate the effects of chronic exposure to diesel. Lewis et al. (1989) conducted lung burden and 59 Fe 3 O 4 tracer studies in rats exposed for 12 and 24 mo to 2 mg/m 3 DPM (7 h/day, 5 days/week). The slope of the Fe 3 O 4 clearance curve of the DPM-exposed animals was significantly steeper than that of the controls, indicating a more rapid alveolar clearance of the deposited 59 Fe 3 O 4 . After 120 days from the inhalation of the tracer particle, 19% and 8% of the initially deposited 59 Fe 3 O 4 were present in the lungs of control and DE-exposed rats, respectively. The lung burden of DPM, however, increased significantlybetween 12 and 24 mo of exposure (0.52 to 0.97% lung dry weight), indicating a later dosedependent inhibition of clearance. Alveolar macrophages, because of their phagocytic and digestive capabilities, are one of the prime defense mechanisms of the alveolar region of the lung against inhaled particles. Thus, characterization of the effects of DE on various properties of AMs provides information on the integrity or compromise of a key pulmonary defense mechanism. The physiological viability of AMs from diesel-exposed rats was assessed after 2 years of exposure by Castranova et al. (1985). The 7 h/day, 5 days/week exposure at 2 mg/m 3 DPM had little effect on the following: viability, cell number, oxygen consumption, membrane integrity, lysosomal enzyme activity, or protein content of the AMs. A slight decrease in cell volume, a decrease in chemiluminescence indicative of a decreased secretion of reactive oxygen species, and a decrease in ruffling of the cell membrane were observed. These latter findings could be reflective of an overall reduction in phagocytic activity. Exposure to DE has been reported both to increase the number of recoverable AMs from the lung (Strom, 1984; Vostal et al., 1982; Henderson et al., 1988a) or to produce no change in numbers (Chen et al., 1980; Castranova et al., 1985). Strom (1984) found that in rats exposed to 0.25 mg/m 3 DPM for 20 h/day, 5.5 days/week for 6 mo or 1 year, as well as in the controls, BAL cells consisted entirely of AMs, with no differences in the cell counts in the lavage fluid. At the higher concentrations, 0.75 or 1.5 mg DPM/m 3 , the count of AM increased proportionally with the exposure concentration; the results were identical for AMs at both 6 and 11 or 12 mo of exposure. The increase in AM counts was much larger after exposure to 1.5 mg/m 3 DPM for 6 mo than after exposure to 0.75 mg/m 3 for 1 year, although the total mass (calculated as C × T)5-57 of deposited particulate burden was the same. These data suggested to the authors that the number of lavaged AMs was proportional to the mass influx of particles rather than to the actual DPM burden in the lung. These results further implied that there may be a threshold for the rate of mass influx of DPM into the lungs of rats above which there was an increased recruitment of AMs. Henderson et al. (1988a) reported similar findings of significant increases of AMs in rats and mice exposed to 7.1 mg/m 3 DPM for 18 and 24 mo, respectively, for 7 h/day, 5 days/week, but not at concentrations of 3.5 or 0.35 mg/m 3 for the same exposure durations. Chen et al. (1980), using an exposure regimen of 0.25 and 1.5 mg/m 3 DPM for 2 mo and 20 h/day and 5.5 days/week, found no significant changes in absolute numbers of AMs from guinea pig BALF, nor did Castranova et al. (1985) in rat BALF following exposure to 2 mg/m 3 DPM for 7 h/day, 5 days/week for 2 years. A similar inflammatory response was noted by Henderson et al. (1988a) and Strom (1984), as evidenced by an increased number of PMNs present in BALF from rodents exposed to DE. Henderson et al. (1988) found these changes in rats and mice exposed to 7.1 and 3.5 mg/m 3 DPM for 7 h/day, 5 days/week. Significant increases in BALF PMNs were observed in mice at 6 mo of exposure and thereafter at the 7.1 and 3.5 mg/m 3 exposure levels, but in rats only the 7.1 mg/m 3 exposure level showed an increase in BALF PMNs at 6 mo of exposure and thereafter. Significant increases in BALF PMNs occurred in rats at 12, 18, and 24 mo of exposure to 3.5 mg/m 3 DPM. Although increases in PMNs were usually greater in mice in terms of absolute numbers, the PMN response in terms of increase relative to controls was only about one-third that of rats. Strom (1984) reported that the increased numbers of PMNs in BALF were proportional to the inhaled concentrations and/or duration of exposure. The PMNs also appeared to be affiliated with clusters of aggregated AMs rather than to the diesel particles per se. Proliferation of Type II cells likewise occurred in response to the formed aggregates of AMs (White and Garg, 1981). The integrity of pulmonary defense mechanisms can also be ascertained by assessing if exposure to DE affects colonization and clearance of pathogens and alters the response of the challenged animals to respiratory tract infections. Campbell et al. (1980, 1981) exposed mice to DE followed by infectious challenge with Salmonella typhimurium, Streptococcus pyogenes , or A/PR8-3 influenza virus and measured microbial-induced mortality. Exposures to DE were to 6 mg/m 3 DPM for 8 h/day, 7 days/week for up to 321 days. Exposure to DE resulted in enhanced susceptibility to the lethal effects of S. pyogenes infection at all exposure durations (2 h, 6 h; 8, 15, 16, 307, and 321 days). Tests with S. typhimurium were inconclusive because of high mortality rates in the controls. Mice exposed to DE did not exhibit an enhanced mortality when challenged with the influenza virus. Hatch et al. (1985) found no changes in the susceptibility of mice to Group C Streptococcus sp. infection following intratracheal injection of 100 

### : g of DPM suspended in unbuffered saline.5-58 Hahon et al. (1985) assessed virus-induced mortality, virus multiplication with concomitant IFN levels (lungs and sera), antibody response, and lung histopathology in mice exposed to DE prior to infectious challenge with Ao/PR/8/34 influenza virus. Weanling mice were exposed to DE containing 2 mg/m 3 DPM for 7 h/day, 5 days/week. In mice exposed for 1, 3, and 6 mo, mortality was similar between the exposed and control mice. In mice exposed for 3 and 6 mo, however, there were significant increases in the percentage of mice having lung consolidation, higher virus growth, depressed IFN levels, and a fourfold reduction in hemagglutinin antibody levels; these effects were not seen after the 1-mo exposure. The effects of DE on the pulmonary defense mechanisms appear to be determined by three critical factors related to exposure: the concentrations of the pollutants, the exposure duration, and the exposure pattern. Higher doses of DE as determined by an increase in one or more of these three variables have been reported to increase the numbers of AMs, PMNs, and Type II cells in the lung, whereas lower doses fail to produce such changes. In rats, the single most significant contributor to the impairment of the pulmonary defense mechanisms appears to be an excessive accumulation of DPM, particularly as particle-laden aggregates of AMs. Such an accumulation would result from an increase in deposition and/or a reduction in clearance. The deposition of particles does not appear to change significantly following exposure to equivalent DE doses over time. Because of the significant nonlinearity in particle accumulation between low and high doses of DE exposure, coupled with no evidence of increased particle deposition, an impairment in one or more of the mechanisms of pulmonary defense appears to be responsible for the DPM accumulation and subsequent pathological sequelae. The time of onset of pulmonary clearance impairment was dependent both on the magnitude and on the duration of exposures. For example, for rats exposed for 7 h/day, 5 days/week for 104 weeks, the concentration needed to induce pulmonary clearance impairment appears to lie between 0.35 and 2.0 mg/m 3 DPM.

5.1.3.3.5. Effects on the immune system—inhalation studies

. The effects of DE on the immune system of guinea pigs were investigated by Dziedzic (1981). Exposures were to 1.5 mg/m 3 DPM for 20 h/day, 5.5 days/week for up to 8 weeks. There was no effect of diesel exposure when compared with matched controls for the number of B and T lymphocytes and null 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. The results of this study and others on the effects of exposure to DE on the immune system are
summarized in Table 5-8. Mentnech et al. (1984) examined the effect of DE on the immune system of rats. Exposures were to 2 mg/m 3 DPM for 7 h/day, 5 days/week for up to 2 years. Rats exposed for 12 and 24 mo were tested for immunocompetency by determining antibody-producing cells in5-59 the spleen 4 days after immunization with sheep erythrocytes. The proliferative response of splenic T-lymphocytes to the mitogens concanavalin A and phytohemagglutinin was assessed in rats exposed for 24 mo. There were no significant differences between the exposed and control animals. Results obtained from these two assays indicate that neither humoral immunity (assessed by enumerating antibody-producing cells) nor cellular immunity (assessed by the lymphocyte blast transformation assay) were markedly affected by the exposures. Bice et al. (1985) evaluated whether or not exposure to DE would alter antibody immune responses induced after lung immunization of rats and mice. Exposures were to 0.35, 3.5, or 7.1 mg/m 3 DPM for 7 h/day, 5 days/week for 24 mo. Chamber controls and 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 elevated in rats, but not in mice, exposed to 3.5 or 7.1 mg/m 3 DPM; after 6 mo of exposure, only the 7.1 mg/m 3 level was found to have caused this response in rats. The number of AFCs per 10 6 lymphoid cells in lung-associated lymph nodes and the levels of specific IgM, IgG, or IgA in rat sera were not significantly altered. The investigators concluded that the increased cellularity and the presence of DPM in the lung-associated lymph nodes had only a minimal effect on the immune and antigen filtration function of these tissues. The effects of inhaled DE and DPM have been studied in a murine model of allergic asthma (Takano et al., 1998a,b). ICR mice were exposed for 12 h/day, 7 days/week for 40 weeks to DE (0.3, 1.0, or 3.0 mg/m 3 ). The mice were sensitized with ovalbumin (OA) after 16 weeks exposure and subsequently challenged with aerosol allergen (1% OA in isotonic saline for 6 min) at 3-week intervals during the last 24 weeks of exposure. Exposure to DE enhanced allergen-related eosinophil recruitment to the submucosal layers of the airways and to the
bronchoalveolar space, and increased protein levels of GM-CSF and IL-5 in the lung in a dose-dependent manner. In the DE-exposed mice, increases in eosinophil recruitment and local cytokine expression were accompanied by goblet-cell proliferation in the bronchial epithelium

a nd airway hyperresponsiveness to inhaled acetylcholine. In contrast, mice exposed to clean air or
DE without allergen provocation showed no eosinophil recruitment to the submucosal layers of
the airways or to the bronchoalveolar space, and few goblet-cells in the bronchial epithelium. The5-60Table 5-8. Effects of inhalation of diesel exhaust on the immune system of laboratory animalsSpecies/sexExposureperiodParticles(mg/m3)C × T(mgAh/m3)CO(ppm)NO2(ppm)SO2(ppm)Effects StudyGuinea Pig,Hartley, M20 h/day5.5 days/week4 or 8 weeks1.50.19:mMDD660 or 7,280 7.5 — — No alterations in numbers of B, T, and nulllymphocytes or cell viability among lymphocytesisolated from tracheobronchial lymph nodes, spleen,or bloodDziedzic(1981)Rat, F344, M 7h/day5 days/week52 or 104 weeks2.00.23–0.36:m MDD3,640 or7,28011.5 1.5 0.8 Neither humoral immunity (assessed by enumeratingantibody-producing cells) nor cellular immunity(assessed by the lymphocyte blast transformationassay) were markedly affectedMentnech etal. (1984)Rat, F344;Mouse, CD-17 h/day5 days/week104 weeks0.353.57.10.25:mMDD1,27412,74025,4802.916.529.70.050.340.68———Total number of anti-sheep red blood cell IgM AFC inthe lung-associated lymph nodes was elevated in ratsexposed to 3.5 or 7.0 mg/m3DPM (no such effects inmice); total number of AFC per 106lymphoid cells inlung-associated lymph nodes and level of specificIgM, IgG, or IgA in rat sera were not alteredBice et al.(1985)Mouse,BALB/C, M12 h/day,7 days/week,3 weeksMice administered OAintranasally before,immediately after, and3 weeks after exposure3.06.00.4:m7561,512——2.84.11.72.7Spleen weights in mice exposed to DE (6 mg/m3)increased significantly. Serum anti-OA IgE antibodytiters in mice exposed to 6 mg/m3significantly higherthan control. Antigen-stimulated IL-4 and IL-10production increased while IFN-g productiondecreased significantly in spleen cells from DEexposed (6 mg/m3) mice stimulated with OA in vitro.DE inhalation may affect antigen-specific IgEantibody production through alteration of thecytokine network.Fujimaki et al.(1997)Mouse,C3H/Hen, M12 h /day,for 12 weeks. Beforeexposure mice injected IPwith OA. After 3 weeksand every 3 weeksthereafter, micechallenged with OAaerosol.1.03.01,0083,024— 1.424.020.871.83DE + antigen challenge induced airwayhyperresponsiveness and inflammation with increasedeosinophils, mast cells, and goblet cells.DE alone induced airway hyperresponsiveness, butnot eosinophilic infiltration or increased goblet cells.DE inhalation enhanced airway hyperresponsivenessand airway inflammation caused by OA sensitization.Miyabara etal. (1998a)5-61Table 5-8. Effects of inhalation of diesel exhaust on the immune system of laboratory animals (continued)Species/sexExposureperiodParticles(mg/m3)C × T(mgAh/m3)CO(ppm)NO2(ppm)SO2(ppm)Effects StudyMouse,C3H/HeN,M12 h/day,for 5 weeks. After 7 daysmice injected IP with OA.At end of exposure micechallenged with OAaerosol for 15 minutes.3.0 1,260 — 4.08 1.26 DE alone increased neutrophils and macrophages inBAL fluid; after DE + OA challenge eosinophilsincreased.OA alone increased eosinophils but the increase wasenhanced by DE.DE + OA, but not DE alone, increased goblet cells,respiratory resistance, production of OA-specific IgEand Ig1 in the serum, and overexpression of IL-5 inlung tissue.Miyabara etal. (1998b)Mouse,ICR(murine modelof allergicasthma)12 h/day, 7days/week,40 weeks.After 16 weeks sensitizedto OA and challengedwith OA aerosol for6 min, at 3-week intervalsduring the last 24 weeksof exposure.0.31.03.01,0083,36010,080— — — DE exposure enhanced allergen-related recruitment tothe submucosal layers of the airways and thebronchoalveolar space, and increased GM-CSF andIL-5 in the lung in a dose-dependent manner.Increases in eosinophil recruitment and local cytosineexpression accompanied by goblet cell proliferationin the bronchial epithelium and airwayhyperresponsiveness to inhaled acetylcholine. Miceexposed to clean air or DE without allergenprovocation showed no eosinophil recruitment to thesubmucosal layers of the airways nor to thebronchoalveolar space, and few goblet cells in thebronchial epithelium. Daily inhalation of DE mayenhance allergen-related respiratory diseases such asallergic asthma, and effect may be mediated by theenhanced local expression of IL-5 and GM-CSF.Takano et al.(1998a)DPM = Diesel particulate matter.AFC = Antibody-forming cells.5-62 authors concluded that daily inhalation of DE can enhance allergen-related respiratory diseases such as allergic asthma, and that this effect may be mediated by the enhanced local expression of IL-5 and GM-CSF. The effect of DPM on a second characteristic of allergic asthma, airway hyperresponsiveness, was examined by Takano et al. (1998b). Laboratory mice were administered OA, DPM, or OA and DPM combined by intratracheal instillation for 6 wk. Respiratory resistance (Rrs) after acetylcholine challenge was measured 24 h after the final instillation. Rrs was significantly greater in the mice treated with OA and DPM than in the other treatments. The authors concluded that DPM can enhance airway responsiveness associated with allergen exposure. In a series of inhalation studies following earlier instillation studies, Miyabara and co-workers investigated whether inhalation of DE could enhance allergic reactions in laboratory mice. C3H/HeN mice were exposed to DE (3 mg DPM/m 3 ) by inhalation for 5 weeks (Miyabara et al., 1998b) and, after 7 days of exposure, were sensitized to OA injected intraperitoneally. At the end of the DE exposure, the mice were challenged with an OA aerosol for 15 min. DE caused an increase in the numbers of neutrophils and macrophages in bronchoalveolar lavage fluid independent of OA sensitization, whereas a significant increase in eosinophil numbers occurred only after DE exposure was combined with antigen challenge. Even though OA alone caused an increase in eosinophil numbers in lung tissue, this response was enhanced further by DE. DE exposure combined with OA sensitization enhanced the number of goblet-cells in lung tissue, respiratory resistance, production of OA-specific IgE and IgG 1 in the serum, and overexpression of IL-5 in lung tissue. In a second study, C3H/HeN mice were sensitized with OA injected intraperitoneally and then exposed to DE by inhalation for 12 h/day for 3 mo at either 1 or 3 mg/m 3 (Miyabara et al., 1998a). After 3 weeks of DE exposure, and every 3 weeks thereafter, the mice were challenged with an OA aerosol. Exposure to DE with antigen challenge induced airway hyperresponsiveness and airway inflammation, which was characterized by increased numbers of eosinophils and mast cells in lung tissue. The increase in inflammatory cells was accompanied by an increase in goblet cells in the bronchial epithelium. Airway hyperresponsiveness, but not eosinophilic infiltration or increased goblet cells, was increased by DE exposure alone. These workers concluded that inhalation of DE can enhance airway hyperresponsiveness and airway inflammation caused by OA sensitization in mice. The effects of DE on IgE antibody production were investigated in BALB/c mice sensitized with OA and exposed by inhalation to DE (3.0 and 6.0 mg/m 3 ) for 3 weeks (Fujimaki et al., 1997). The mice were sensitized by intranasal administration of OA alone before, immediately after, and 3 weeks after DE inhalation. While body and thymus weights were unchanged in the DE-exposed and control mice, spleen weights in mice exposed to 6 mg/m 3 DE increased significantly. Anti-OA IgE antibody titers in the sera of mice exposed to 6 mg/m 3 DE were significantly higher than control. Total IgE and anti-OA IgG in sera from DE-exposed and5-63 control mice remained unchanged. Cytokine production was measured in vitro stimulated with OA in spleen cells from mice exposed to DE (6 mg/m 3 ). Antigen-stimulated interleukin-4 (IL-4) and -10 (IL-10) production increased significantly in vitro in spleen cells from DE-exposed mice compared with controls, while IFN- 

### ( production decreased markedly. The authors concluded that DE inhalation in mice may affect antigen-specific IgE antibody production through alteration of the cytokine network.

5.1.3.3.6. Effects on the immune system—noninhalation studies. The immune response of laboratory animals to DPM has been studied in various noninhalation models, and the results of these studies are presented in Table 5-9. Takafuji et al. (1987) evaluated the IgE antibody response of mice inoculated intranasally at intervals of 3 weeks with either 0.5 or 25 

### : g of DPM in ovalbumin per mouse. Antiovalbumin IgE antibody titers, assayed by passive cutaneous anaphylaxis, were enhanced by doses as low as 1 

### : g of particles compared with immunization
with ovalbumin alone. Muranaka et al. (1986) studied the effects of DPM on IgE antibody production in immunized mice. A greater IgE antibody response was noted in mice immunized by ip injection of ovalbumin (OA) mixed with DPM, either 0.02, 0.2, or 2mg per mouse, than in animals immunized with OA alone. This effect of DPM on IgE antibody production in mice was also demonstrated in mice immunized with repeated injections of dinitrophenylated-OA. Moreover, a persistent IgE-antibody response to Japanese cedar pollen (JCPA), a common pollen allergen causing allergic rhinitis in Japan, was observed in mice immunized with JCPA mixed with DPM but not in animals immunized with JCPA alone. The results suggest an association between the adjuvant activity of DPM and allergic rhinitis caused by JCPA. Takano et al. (1997) designed a study to evaluate the effects of DPM on the manifestations of allergic asthma in mice, with emphasis on antigen-induced airway inflammation; the local expression of IL-5, GM-CSF, IL-2, and IFN- 

### ( ; and the production of antigen-specific IgE and IgG. Male ICR mice were intratracheally instilled with ovalbumin(OVA), DPM, and DPM+OVA. DPM was obtained from a 4JB1-type, light-duty 2.74 L, fourcylinder Isuzu diesel engine operated at a steady speed of 1,500 rpm under a load of 10 torque (kg/m). The OVA-group mice were instilled with 1 

### : g OVA at 3 and 6 weeks. The mice receiving DPM alone were instilled with 100 

### : g DPM weekly for 6 weeks. The OVA + DPM group received the combined treatment in the same protocol as the OVA and the DPM groups, respectively. Additional groups were exposed for 9 weeks. DPM aggravated OVA-induced airway inflammation, characterized by infiltration of eosinophils and lymphocytes and an5-64Table 5-9. Effects of diesel particulate matter on the immune response of laboratory animalsModel TreatmentEffectsReferenceMouse,BDFI, FIntranasally delivered doses of DPM as low as 1 mg exerted an adjuvant activity for IgE antibodyproduction.Takafuji et al.(1987)Mouse,ICR, w/w-, MIntratracheal instillation of DPM, once/week for16 weeksInfiltration of inflammatory cells, proliferation of goblet cells, increased mucus secretion, respiratoryresistance, and airway constriction. Increased eosinophils in the submucosa of the proximal bronchi andmedium bronchioles. Eosinophil infiltration suppressed by pretreatment with PEG-SOD. Bound sialicacid, an index of mucus secretion, in bronchial alveolar lavage fluids increased, but was suppressed byPEG-SOD. Increased respiratory resistance suppressed by PEG-SOD. Oxygen radicals produced byinstilled DPM may cause features characteristic of bronchial asthma in mice.Sagai et al. (1996)Mouse,A/J, MMice immunized intranasally with Der f II +pyrene, or Der f II + DPM 7 times at 2-weekintervalsIgE antibody responses to Der f II enhanced in mice immunized with Der f II+ pyrene or Der f II +DPM compared with Der f II alone. Response was dose related. DPM and pyrene contained in DPMhave adjuvant activity on IgE and IgG1 antibody production in mice immunized with house dust miteallergen.Suzuki et al. (1996)Mouse,BDF1, MMice were administered 25 mg of each of 5 fineparticles (Kanto loam dust, fly ash, CB, DPM,and aluminum hydroxide [alum]) intranasallyand exposed to aerosolized Japanese cedarpollen allergens (JCPA) for intervals up to 18wkMeasurements were made of JCPA-specific IgE and IgG antibody titers, the protein-adsorbing capacityof each type of particle, and nasal rubbing movements (a parameter of allergic rhinitis in mice). Theincreases in anti-JPCA IgE and IgG antibody titers were significantly greater in mice treated withparticles and aerosolized JCPA than in mice treated with aerosolized JCPA alone. In a subsequentexperiment, 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 therewere no significant differences in the anti-JCPA IgE and IgG production, nasal rubbing, orhistopathological changes. The workers concluded that the nature of the particle, the ability of theparticle to absorb antigens, and/or particle size is not related to the enhancement of IgE antibodyproduction or symptoms of allergic rhinitis. However, IgE antibody production did appear to occurearlier in mice treated with particles than in mice immunized with allergens alone.Maejima et al.(1997)Mouse,BALB/C,nu/nu, FInoculated OA with DPM or CB into hindfootpad measured response using popliteallymph node assayIncreased response (increased weight, cell numbers, cell proliferation) and longer response observedwith DPM and OA, compared to DPM or OA alone. Response was specific and not an unspecificinflammatory response. CB was slightly less potent than DPM. Nonextractable carbon core contributessubstantially to adjuvant activity of DPM.Løvik et al. (1997)Mouse,BALB/cA, FIntranasal administration of DPM. Miceimmunized with OA or OA combined withDPM or CBIncreased response to antigen in animals receiving DPM or CB. Increased number of respondinganimals and increased serum anti OA IgE antibody. Both DPM and CB have adjuvant activity for IgEproduction. DPM response more pronounced than CB, indicating both organic matter adsorbed to DPMand the nonextractable carbon core responsible for adjuvant activity.Nilsen et al. (1997)Mouse,ICR, MIntratracheal instillation of OA, DPM, or OVAand DPM combined, once/week for 6 wkRespiratory resistance (Rrs) measured 24 h after the final instillation. Rrs after acetylcholine challengewas significantly greater in the mice treated with OVA and DPM than other treatments. DPM canenhance airway responsiveness associated with allergen exposure.Takano et al.(1998b)OA - Ovalbumin.DPM - Diesel particulate matter.CB - Carbon black.PEG-SOD - Polyethyleneglycol-conjugated superoxide dismutase.IL-4 - Interleukin-4.IL-5 - Interleukin-5.IL-10 - Interleukin-10.IFN - Interferon-g.GM-CSF -Granulocyte-colony stimulating factor.IP - Intraperitoneally.5-65 increase in goblet cells in the bronchial epithelium. DPM in combination with antigen markedly increased IL-5 protein levels in lung tissue and bronchoalveolar lavage supernatants compared with either antigen or DPM alone. The combination of DPM and antigen induced significant increases in local expression of IL-4, GM-CSF, and IL-2, whereas expression of IFN- 

### ( was not affected. In addition, DPM exhibited adjuvant activity for the antigen-specific production of IgG and IgE. The potential role of oxygen radicals in injury caused by DPM was investigated by Sagai et al. (1996). These workers reported that repeated intratracheal instillation of DPM (either 0.1 or 0.2 mg per mouse, once/week for 16 weeks) in mice caused marked infiltration of inflammatory cells, proliferation of goblet cells, increased mucus secretion, respiratory resistance, and airway constriction. Eosinophils in the submucosa of the proximal bronchi and medium bronchioles increased eightfold following instillation. Eosinophil infiltration was significantly suppressed by pretreatment with polyethyleneglycol-conjugated superoxide dismutase (PEG-SOD), an inhibitor of oxygen radicals. Bound sialic acid concentrations in bronchial alveolar lavage fluids, an index of mucus secretion, increased with DPM, but were also suppressed by pretreatment with PEG-SOD. Goblet cell hyperplasia, airway narrowing, and airway constriction also were observed with DPM. Respiratory resistance to acetylcholine in the DPM group was 11 times higher than in controls, and the increased resistance was significantly suppressed by PEG-SOD pretreatment. These findings indicate that oxygen radicals caused by intratracheally instilled DPM elicit responses characteristic of bronchial asthma. Potential adjuvant effects of DPM on the response to the model allergen OA were investigated in BALB/c mice using the popliteal lymph node (PLN) assay (Løvik et al., 1997). DPM inoculated together with OA into one hind footpad (0.02 mL of a 5 mg/mL DPM suspension) gave a significantly augmented response (increase in weight, cell numbers, and cell proliferation) in the draining popliteal lymph node as compared to DPM or OA alone. The duration of the local lymph node response was also longer when DPM was given with the allergen. The lymph node response appeared to be of a specific immunologic character and not an unspecific inflammatory reaction. The OA-specific response IgE was increased in mice receiving OA together with DPM as compared with the response in mice receiving OA alone. Further studies using carbon black (CB) as a surrogate for the nonextractable core of DPM found that while CB resembled DPM in its capacity to increase the local lymph node response and serum-specific IgE response to OA, CB appeared to be slightly less potent than DPM. The results indicate that the nonextractable particle core contributes substantially to the adjuvant activity of DPM. Nilsen et al. (1997) investigated which part of the particle was responsible, the carbon core and/or the adsorbed organic substances, for the adjuvant activity of DPM. Female5-66 BALB/cA mice were immunized with OA alone or in combination with DPM or CB particles by intranasal administration a total of four times, once weekly, at 25 

### : g/inoculation. There was an increased response to the antigen in animals receiving OA together with DPM or CB, compared with animals receiving OA alone. The response was seen as both an increased number of responding animals and increased serum anti OA IgE response. The workers concluded that both DPM and CB have an adjuvant activity for specific IgE production, but that the activity of DPM may be more pronounced than that of CB. The results suggest that both the organic matter adsorbed to DPM and the nonextractable carbon are responsible for the observed adjuvant effect
of DPM. The effects of DPM and its components (extracted particles and particle extracts) on the release of proinflammatory cytokines, interleukin-1 (IL-1), and tumor necrosis factor- 

### " (TNF-



### " ) by alveolar macrophages (AMs) were investigated by Yang et al. (1997). Rat AMs were incubated with 0, 5, 10, 20, 50, or 100 

### : g/10 6 AM/mL of DPM, methanol-extracted DPM, or equivalent concentrations of DPM at 37 o C for 24 h. At high concentrations, both DPM and DPM extracts were shown to increase IL-1-like activity secreted by AMs, whereas extracted particles had no effect. Neither particles, particle extracts, or extracted particles stimulated
secretion of TNF- 

### " . DPM inhibited lipid polysaccharide (LPS)-stimulated production of IL-1
and TNF- 

### " . In contrast, interferon (IFN)- 

### ( -stimulated production of TNF- 

### " was not affected by DPM. Results of this study indicate that the organic fraction of exhaust particles is responsible for the effects noted. Stimulation of IL-1 but not TNF- 

### " suggests that IL-1, but not
TNF- 

### " , may play an important role in the development of DPM-induced inflammatory and immune responses. The cellular mechanism involved in inhibiting increased release of IL-1 and
TNF- 

### " by LPS is unknown, but may be a contributing factor to the decreased AM phagocytic activity and increased susceptibility to pulmonary infection after prolonged exposure to DPM. Fujimaki et al. (1994) investigated the relationship between DPM and IgE antibody production, interleukin 4 (IL-4) production in BALB/c mice treated with DPM mixed with antigen OA or JCP antigen by intratracheal instillation. BALB/c mice were injected with DPM (300 

### : g) plus OA or OA alone and, after the last instillation, the proliferative response and lymphokine production by mediastinal lymph node cells (LNC) were examined in vitro. The proliferative response to OA in mediastinal LNC from mice injected with DPM plus OA was enhanced to 4-17 times that of control mice. IL-4 production by OA stimulation was also enhanced in mediastinal LNC from 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 levels of IL-4, estimated by JCP antigen in mediastinal LNC, from mice injected with DPM plus JCP antigen were twofold higher than those from mice injected with JCP antigen alone. These results suggest that intratracheal5-67 instillation of DPM affects antigen-specific IgE antibody responses via local T-cell activation, especially enhanced IL-4 production. Suzuki et al. (1993) investigated the adjuvant activity of pyrene, one of many PAHs contained in DPM, on IgE antibody production in mice. In the first experiment, mice were immunized with 1 mg of OA alone, 1 mg of OA plus 1 mg of pyrene, or 1 mg of OA plus 1 mg of DPM, respectively. The IgE antibody responses to OA in mice immunized with OA plus pyrene or OA plus DPM were enhanced as compared to those in mice immunized with OA alone; the highest responses were observed in mice immunized with OA plus DPM. In the second experiment, mice were immunized with 10 mg of JCPA alone or 10 mg of JCPA plus 5 mg of pyrene. The IgE antibody responses to JCPA in mice immunized with JCPA plus pyrene were higher than those in mice immunized with JCPA alone. The results indicate that pyrene contained in DPM acts as an adjuvant in IgE antibody production in immunized mice. Suzuki et al. (1996) investigated the effect of pyrene on IgE and IgG1 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 mite ( Dermatophagoides farinae ). Allergen mice were grouped and immunized with Der f II (5 

### : g), Der f II (5 

### : g) plus pyrene (200 

### : g), and Der f II (5 

### : g) plus DPM (100 

### : g) intranasally seven times at 2-week intervals. The separate groups of mice were also immunized with Der f II (10 

### : g) plus the same dose of adjuvants in the same way. The IgE antibody responses to Der f II in mice immunized with Der f II plus pyrene or Der f II plus DPM were markedly enhanced compared with those immunized with Der f II alone. The anti-Der f II IgE antibody production increased with increasing the dose of Der f II from 5 

### : g to 10 

### : g in mice immunized with Der f II plus the same dose of adjuvants. The IgG1 antibody responses to Der f II in mice immunized with Der f II (10 

### : g) plus pyrene (200 

### : g) or Der f II (10 

### : g) plus DPM (100 

### : g) were greater than those immunized with 10 

### : g of Der f II alone. In addition, when peritoneal macrophages obtained from normal mice were incubated with pyrene or DPM in vitro, an enhanced IL-1a production by the macrophages was observed. When spleen lymphocytes obtained from the mice immunized with Der f II (10 

### : g) plus DPM (100 

### : g) or Der f II (10 

### : g) plus pyrene (200 

### : g) were stimulated with 10 

### : g of Der f II in vitro, an enhanced IL-4 production of the lymphocytes was also observed compared with those immunized with Der f II alone. This study indicates that DPM and pyrene (one of the many PAHs adsorbed onto DPM) have an adjuvant activity on IgE and IgG1 antibody production in mice immunized intranasally with a house dust mite allergen. Maejima et al. (1997) examined the potential adjuvant activity of several different fine particles. These workers administered 25 

### : g of each of 5 particles (Kanto loam dust, fly ash, CB, DPM, and aluminum hydroxide [alum]) intranasally in mice and exposed them to aerosolized JCPA for intervals up to 18 weeks. Measurements were made of JCPA-specific IgE5-68 and IgG antibody titers, 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 titers were significantly greater in mice treated with particles and plus 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 JCP were dropped onto the tip of the nose of each mouse twice a week for 16 weeks. After 18 weeks 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 particle size are 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. The potential for DPM to modulate cytokine production has been demonstrated in cultured mouse bone marrow-derived mast cells (BMMC). Saneyoshi et al. (1997) examined the production of cytokines in BMMC treated with DPM (0.8, 2 and 4 mg/mL). Production of interleukin-4 (IL-4) and IL-6 was higher in BMMC stimulated with A23187 and treated with low concentrations of DPM than in controls, but no increase was seen in BMMC treated with high DPM. After pretreatment with low DPM for 24 h, IL-4 production in BMMC stimulated with A23187 was lower than in controls. Antigen-induced IL-4 production increased significantly in BMMC treated with 0.4 or 0.8 mg/mL DPM, but did not increase with low DPM. Although the enhancement of IL-4 production of BMMC stimulated with A23187 plus DPM was not completely inhibited by 2-mercaptoethanol, treatment with dexamethasone inhibited further IL-4 production. Thus, DPM may affect the immune response via the modulation of cytokine production in mast cells. Ormstad et al. (1998) investigated the potential for DPM as well as other suspended particulate matter (SPM) to act as a carrier for allergens into the airways. These investigators found both Can f 1 (dog) and Bet v 1 (birch pollen) on the surface of SPM collected in air from different homes. In an extension of the study, they found that DPM adhered to polycarbonate filters had the potential of binding both of these allergens as well as Fel d 1 (cat) and Der p 1 (house mite). The authors conclude that soot particles in indoor air house dust may act as carrier of several allergens in indoor air. Knox et al. (1997) investigated whether free grass pollen allergen molecules, released from pollen grains by osmotic shock (Suphioglu et al., 1992) and dispersed in microdroplets of water in aerosols, can bind to DPM mounted on copper grids in air. Using natural highly purified Lol p 1, immunogold labeling with specific monoclonal antibodies, and a high-voltage transmission electron-microscopic imaging technique, these workers demonstrated binding of the major grass pollen allergen, Lol p 1, to DPM in vitro. These workers conclude that binding of5-69 DPM with Lol p 1 might be a mechanism by which allergens can become concentrated in air and trigger attacks of asthma.Murphy et al. (1999) examined the comparative toxicities to the lung of four differentsized CB particles and DPM, in primary cultures of mouse Clara and rat type II epithelial cells. Particle toxicity was assessed by cell attachment to an extracellular matrix substratum. The CB particles varied in toxicity to Clara and type II cells. DPM stored for 2 weeks was equally toxic to both cell types. DPM became progressively less toxic to type II cells with time of storage. Both primary epithelial cell types internalized the particles in culture. These workers concluded that bioreactivity was related to CB particle size and surface area, with the smaller particles having the larger surface area being the more toxic. Although freshly prepared DPM was equally toxic to type II and Clara cells, DPM became progressively less toxic to the type II cells
with time. Exposure studies in laboratory animals and isolated cell systems derived from animals also indicate that DPM can elicit both inflammatory and immunological changes. Moreover, the effects appear to be due to both the nonextractable carbon core and the adsorbed organic fraction of the diesel particle. Changes in IgE, goblet cell hyperplasia, mast cell influx, and cytokines in various animal models and in vitro model systems are all key markers of asthma. The data further indicate a role for oxygen radicals in DPM injury because the extent of the injury can be reduced by treatment with antioxidants. DPM also has the capacity to bind and transport airborne allergens.

5.1.3.3.7. Effects on the liver . Meiss et al. (1981) examined alterations in the hepatic parenchyma of hamsters by using thin-section and freeze-fracture histological techniques. Exposures to DE were for 7 to 8 h/day, 5 days/week, for 5 mo at about 4 or 11 mg/m 3 DPM. The livers of the hamsters exposed to both concentrations of DE exhibited moderate dilatation of the sinusoids, with activation of the Kupffer cells and slight changes in the cell nuclei. Fatty deposits were observed in the sinusoids, and small fat droplets were occasionally observed in the peripheral hepatocytes. Mitochondria often had a loss of cristae and exhibited a pleomorphic character. Giant microbodies were seen in the hepatocytes, which were moderately enlarged, and gap junctions between hepatocytes exhibited a wide range in structural diversity. The results of this study and others on the effect of exposure of DE on the liver of laboratory animals are
summarized in Table 5-10.5-70Table 5-10. Effects of exposure to diesel exhaust on the liver of laboratory animalsSpecies/sexExposureperiodParticles(mg/m3)C × T(mgAh/m3)CO(ppm)NO2(ppm)SO2(ppm)Effects StudyRat, F344, M,F 7 h/day5 days/week52 weeks2.00.23–0.36:mMDD3,640 12.7 1.6 0.83No changes in absolute liver weight orliver/body weight ratioGreen et al. (1983)Hamster, Syrian 7-8 h/day5 days/week22 weeks4.08.011.03,080-9,680 12.019.025.00.51.01.53.06.07.0Enlarged sinusoids, with activated Kupffer’scells and slight changes of nuclei; fattydeposits; mitochondria, loss of cristae andpleomorphic character; gap junctions betweenhepatocytes had wide range in structuraldiversityMeiss et al. (1981)Cat, inbred, M 8h/day7 days/week124 weeks6.0a12.0b41,66483,32820.233.32.74.42.15.0No change in the absolute liver weight Plopper et al. (1983)a1 to 61 weeks of exposure.b62 to 124 weeks of exposure.5-71 Green et al. (1983) and Plopper et al. (1983) reported no changes in liver weights of rats exposed to 2 mg/m 3 DPM for 7 h/day, 5 days/week for 52 weeks or of cats exposed to 6 to 12 mg/m 3 , 8 h/day, 7 days/week for 124 weeks. The use of light and electron microscopy revealed that long-term inhalation of varying high concentrations of DE caused numerous alterations to the hepatic parenchyma of guinea pigs. A less sensitive index of liver toxicity, increased liver weight, failed to detect an effect of DE on the liver of the rat and cat following long-term exposure to DE. These results are too limited to understand potential impacts on the
liver.

5.1.3.3.8. Blood and cardiovascular systems. Several studies have evaluated the effects of DE exposure on hematological and cardiovascular parameters of laboratory animals. These studies are summarized in Table 5-11. Standard hematological indices of toxicological effects on red and white blood cells failed to detect dramatic and consistent responses. Erythrocyte (RBC) counts were reported as being unaffected in cats (Pepelko and Peirano, 1983), rats and monkeys (Lewis et al., 1989), guinea pigs and rats (Penney et al., 1981), and rats (Karagianes et al., 1981); lowered in rats (Heinrich et al., 1982); and elevated in rats (Ishinishi et al., 1988; Brightwell et al., 1986). Mean corpuscular volume was significantly increased in monkeys, 69 versus 64 (Lewis et al., 1989), and hamsters (Heinrich et al., 1982), and lowered in rats (Ishinishi et al., 1988). The only other parameters of erythrocyte status and related events were lowered mean corpuscular hemoglobin and mean corpuscular hemoglobin concentration in1 rats (Ishinishi et al., 1988), a 3% to 5% increase in carboxyhemoglobin saturation in rats (Karagianes et al., 1981), and a suggestion of an increase in prothrombin time (Brightwell et al., 1986). The biological significance of these findings regarding adverse health effects is deemed to be inconsequential. Three investigators (Pepelko and Peirano, 1983; Lewis et al., 1989; Brightwell et al., 1986) reported an increase in the percentage of banded neutrophils in cats and rats. This effect was not observed in monkeys (Lewis et al., 1989). The health implications of an increase in abnormal maturation of circulating neutrophils are uncertain but indicate a toxic response of leukocytes following exposures to DE. Leukocyte counts were reported to be reduced in hamsters (Heinrich et al., 1982); increased in rats (Brightwell et al., 1986); and unaffected in cats, rats, and monkeys (Pepelko and Peirano, 1983; Ishinishi et al., 1988; Lewis et al., 1989). These inconsistent findings indicate that the leukocyte counts are more indicative of the clinical status of the laboratory animals than any direct effect of exposure to DE. No significant changes in heart mass were found in guinea pigs or rats exposed to DE (Wiester et al., 1980; Penney et al., 1981; Lewis et al., 1989). Rats exposed to DE showed a greater increase in the medial wall thickness of pulmonary arteries of differing diameters and5-72Table 5-11. Effects of exposure to diesel exhaust on the hematological and cardiovascular systems of laboratory animalsSpecies/sexExposureperiodParticles(mg/m3)C × T(mgAh/m3)CO(ppm)NO2(ppm)SO2(ppm)Effects StudyMonkey,Cynomolgus, M7 h/day5 days/week104 weeks20.23–0.36:m MDD7,280 11.5 1.5 0.8Increased MCV Lewis et al. (1989)Rat, F344, M,F 7 h/day5 days/week104 weeks20.23–0.36:m MDD7,280 11.5 1.5 0.8 Increase inbanded neutrophils; no effect on heartor pulmonary arteriesLewis et al. (1989)Vallyathan et al.(1986)Guinea Pig,Hartley, M, F20 h/day7 days/week8 weeks6.3a6.8b7,0567,61617.416.72.32.92.11.9No effect on heart mass or ECG; small decrease inheart rate (IE only)Wiester et al.(1980)Hamster, Syrian,M, F7-8 h/day5 days/week75 weeks3.90.1:m MDD10,238-11,700 18.5 1.2 3.1 At 29weeks, lower erythrocyte count; increasedMCV; reduced leukocyte countHeinrich et al.(1982)Rat, F344;Guinea Pig,Hartley20 h/day5.5 days/week78 weeks0.250.751.50.19:m MDD2,1456,43512,8703.04.86.90.110.270.49———No changes in heart mass or hematology at anyexhaust level or duration of exposure in eitherspeciesPenney et al.(1981)Rat, Wistar, M 6h/day5 days/week78 weeks8.30.71:m MDD19,422 50.0 4-6 —3% increase in COHb Karagianes et al.(1981)Rat, F3444/Jcl,M, F16 h/day6 days/week130 weeks0.11c0.41c1.08c2.31c3.72d0.1:m MDD1,3735,11713,47828,82946,4261.232.123.967.1012.90.080.260.701.413.000.381.062.424.704.57At higher concentrations, RBC, Hb, Hct slightlyelevated; MCV and mean corpuscular hemoglobinand concentration were loweredIshinishi et al.(1988)Rat, F34416 h/day5 days/week104 weeks0.72.26.65,82418,30454,912——32.0——————Increases in RBC, Hb, Hct, and WBC, primarilybanded neutrophils; suggestion of an increase inprothrombin time; increased heart/body weightand right ventricular/heart ratios and decreased leftventricular contractility in 6.6 mg/m3groupBrightwell et al.(1986)Cat, Inbred, M 8h/day7 days/week124 weeks6.0e12.0f41,66483,32820.233.32.74.42.15.0Increases in banded neutrophils; significant at 12mo, but not 24 moPepelko andPeirano (1983)aNonirradiated DE.dHeavy-duty engine.bIrradiated DE.e1 to 61 weeks of exposure.cLight-duty engine.f62 to 124 weeks of exposure.Key: MCV = Mean corpuscular volume.5-73 right ventricular wall thickness; these increases, however, did not achieve statistically significant levels (Vallyathan et al., 1986). Brightwell et al. (1986) reported increased heart/body weight and right ventricular/heart weight ratios and decreased left ventricular contractility in rats exposed to 6.6 mg/m 3 DPM for 16 h/day, 5 days/week for 104 weeks. 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 with suspensions of DPM (10-100 

### : g/mL) markedly attenuated acetylcholine-induced EDR. The mechanism of this effect was studied further in cultured porcine endothelial cells (CPE). A 10-min incubation of CPE with DPM (0.1-100 

### : g/mL) inhibited endothelium-dependent relaxing factor (EDRF) or nitric oxide (NO) release. A 10-min incubation of DPM with NO synthase inhibited formation of NO 2 - , a product of NO metabolism. The authors concluded that DPM, at the concentrations tested, neither induced cell damage nor inhibited EDRF release from CPE, but scavenged and thereby blocked the physiological action of NO.

5.1.3.3.9. Serum chemistry

. A number of investigators have studied the effects of exposure to DE on serum biochemistry, and no consistent effects have been found. Such studies are
summarized in Table 5-12. The biological significance of changes in serum chemistry reported by Lewis et al. (1989) in female but not male rats exposed at 2 mg/m 3 DPM for 7 h/day, 5 days/week for 104 weeks is difficult to interpret. Not only were the effects noted in one sex (females) only, but the serum enzymes, lactate dehydrogenase (LDH), serum glutamic-oxaloacetic transaminase (SGOT), and serum glutamic-pyruvic transaminase (SGPT), were elevated in the control group, a circumstance contrary to denoting organ damage in the exposed female rats. The elevations of liver-related serum enzymes in the control versus the exposed female rats appear to be a random event among these aged subjects. The incidence of age-related disease, such as mononuclear cell leukemia, can markedly affect such enzyme levels, seriously compromising the usefulness of a comparison to historical controls. The serum sodium values of 144 versus 148 mmol/L in control and exposed rats, respectively, although statistically different, would have no biological significance. The increased serum enzyme activities, alkaline phosphatase, SGOT, SGPT, gamma-glutamyl transpeptidase, and decreased cholinesterase activity suggest an impaired liver; however, such an impairment was not established histopathologically (Heinrich et al., 1982; Ishinishi et al., 1988; Brightwell et al., 1986). The increased urea nitrogen, electrolyte levels, and gamma globulin concentration and reduction in total blood proteins are indicative of impaired kidney function. Again, there was no histopathological confirmation of impaired kidneys in these studies.5-74Table 5-12. Effects of chronic exposures to diesel exhaust on serum chemistry of laboratory animalsSpecies/sexExposureperiodParticles(mg/m3)C × T(mgAh/m3)CO(ppm)NO2(ppm)SO2(ppm)Effects StudyRat, F344, M,F 7 h/day5 days/week104 weeks2.00.230.36:m MDD7,280 11.5 1.5 0.8Decreased phosphate, LDH, SGOT, and SGPT;increased sodium in females but not malesLewis et al.(1989)Hamster, Syrian, M,F 7-8 h/day5 days/week75 weeks3.90.1:m MDD10,238-11,700 18.5 1.2 3.1After 29 weeks, increases in SGOT, LDH, alkalinephosphatase, gamma-glutamyl transferase, and BUNHeinrich et al.(1982)Rat, F344/JcL, M,F 16 h/day6 days/week130 weeks0.11a0.41a1.08a2.31a3.72b0.19–0.28:mMDD1,3735,11713,47828,82946,4261.232.123.967.1012.90.080.263.967.103.000.381.062.424.704.57Lower cholinesterase activity in males in both thelight-and heavy-duty series and elevated gammaglobulin and electrolyte levels in males and femalesin both seriesResearchCommittee forHERP Studies(1988)Rat, F344; Hamster,Syrian16 h/day5 days/week104 weeks0.72.26.65,82418,30454,912——32.0——————Rats, 6.6 mg/m3, reduction in blood glucose, bloodproteins, triglycerides, and cholesterol; increase inBUN, alkaline phosphate alamine, and aspartateaminotransferases (SGPT and SGOT); hamsters, 6.6mg/m3, decrease in potassium, LDH, aspartate aminotransferase; increase in albumin and gamma-glutamyltransferaseBrightwell et al.(1986)Cat inbred, M 8h/day7 days/week124 weeks6.0c12.0d41,66483,32820.233.32.74.42.15.0BUN unaltered; SGOT and SGPT unaffected; LHDincrease after 1 year of exposurePepelko andPeirano (1983)aLight-duty engine.bHeavy-duty engine.c1 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.5-75 Clinical chemistry studies suggest impairment of both liver and kidney functions in rats and hamsters chronically exposed to high concentrations of DE. The absence of histopathological confirmation, the appearance of such effects near the end of the lifespan of the laboratory animal, and the failure to find such biochemical changes in cats exposed to a higher dose, however, tend to discredit the probability of hepatic and renal hazards to humans exposed at atmospheric levels of DE.

5.1.3.3.10. Effects on microsomal enzymes

. Several studies have examined the effects of DE exposure on microsomal enzymes associated with the metabolism and possible activation of xenobiotics, especially polynuclear aromatic hydrocarbons (PAH). These studies are summarized in Table 5-13. Lee et al. (1980) measured the activities of aryl hydrocarbon hydroxylase (AHH) and epoxide hydrase (EH) in liver, lung, testis, and prostate gland of adult male rats exposed to 6.32 mg/m 3 DPM 20 h/day for 42 days. Maximal significant AHH activities (pmol/min/mg microsomal protein) occurred at different times during the exposure period, and differences between controls and exposed rats, respectively, were as follows: prostate 0.29 versus 1.31, lung 3.67 versus 5.11, and liver 113.9 versus 164.0. There was no difference in AHH activity in the testis between exposed and control rats. Epoxide hydrase activity was not significantly different from control values for any of the organs tested. Pepelko and Peirano (1983) found no statistically significant differences in livermicrosomal cytochrome P448-450 levels and liver microsomal AHH between control and dieselexposed mice at either 6 or 8 mo of exposure. Small differences were noted in the lung microsomal AHH activities, but these were believed to be artifactual differences, due to increases in nonmicrosomal lung protein present in the microsomal preparations. Exposures to 6 mg/m 3 DPM were for 8 h/day, 7 days/week. Rabovsky et al. (1984) investigated the effect of chronic exposure to DE on microsomal cytochrome P450-associated benzo[ a ]pyrene (B[ a ]P) hydroxylase and 7-ethoxycoumarin deethylase activities in rat lung and liver. Male rats were exposed for 7 h/day, 5 days/week for 104 weeks to 2 mg/m 3 DPM. The exposure had no effect on B[ a ]P hydroxylase or 7-ethoxycoumarin deethylase activities in lung or liver. In related studies, Rabovsky et al. (1986) examined the effects of DE on viral induced enzyme activity and interferon production in female mice. The mice were exposed for 7 h/day, 5 days/week for 1 mo to DE diluted to achieve a concentration of 2 mg/m 3 DPM. After the exposure, the mice were inoculated intranasally with influenza virus. Changes in serum levels of interferon and liver microsomal activities of 7-ethoxycoumarin, ethylmorphine demethylase, and nicotinamide adenine dinucleotide5-76Table 5-13. Effects of chronic exposures to diesel exhaust on microsomal enzymes of laboratory animalsSpecies/sexExposureperiodParticles(mg/m3)C × t(mgAh/m3)CO(ppm)NO2(ppm)SO2(ppm)Effects StudyRat, F344, M — — — — — — Intratracheal administration of DPM extract requireddoses greater than 6 mg/m3before the lung AHH wasbarely doubled; liver AHH activity was unchangedChen (1986)Mouse, CD-1, F 7h/day5 days/week4 weeks2.00.2–0.36:m mdd280 11.5 1.5 0.8 Mice inoculated intranasally with influenza virus hadsmaller increases in ethylmorphine demethylaseactivity on days 2 to 4 postvirus infection and abolitionof day 4 postinfection increase in NADPH-dependentcytochrome c reductaseRabovsky et al. (1986)Rat, SpragueDawley, M20 h/day7 days/week1-7 weeks6.3 882-6,174 17.4 2.3 2.1AHH induction occurred in lung, liver, and prostategland but not in testes; maximum significant activitiesoccurred at different times; liver has greatest overallactivity, percent increase highest in prostate; expoxidehydrase activity was unaffectedLee et al. (1980)Rat, F344, M20 h/day5.5 days/week4, 13, 26, or39 weeks20 h/day5.5 days/week4, 13, 26, or39 weeks0.751.50.19:m mdd0.751.50.19:m mdd330-6,435330-6,4354.87.54.87.5————————Inhalation exposure had no significant effect on liverAHH activity; lung AHH activity was slightly reducedafter 6-mo exposure to 1.5 mg/m3DPM; an ip dose ofdp extract, estimated to be equivalent to inhalationexposure, had no effect on AHH activity in liver andlungs; cyt. P-50 was unchanged in lungs and liverfollowing inhalation or ip administrationChen and Vostal (1981)Rat, F344, F 7h/day5 days/week12, 26, or104 weeks2.00.23-0.36:m mdd840-7,280 11.5 1.5 0.8No effect on B[a]p hydrolase or 7-exthoxycoumarindeethylase activities in the liverRabovsky et al. (1984)Rat, F344, M20 h/day5.5 days/week8-53 weeks0.251.50.19:m mdd220-8,745 2.97.5————After 8 weeks, no induction of cyt. P-450, cyt. P-448,or NADPH-dependent cyt. c reductase; after 1 year ofexposure, liver microsomal oxidation of B[a]p was notincreased; 1 year of exposure to either 0.25 or1.5 mg/m3DPM impaired lung microsomal metabolismof B[a]pNavarro et al. (1981)Mouse, A/J, M 8days/week7 days/week26 or 35 weeks6.0 17.4 17.4 2.3 2.1 No differences in lung and liver AHH activities andliver P-448, P-450 levelsPepelko and Peirano(1983)AHH = aryl hydrocarbon hydroxylase.B[a]p = benzo[a]pyrene.5-77 phosphate (NADPH)-dependent cytochrome c reductase were measured. In the absence of viral inoculation, exposure to DE had no significant effects on the activity levels of the two liver microsomal monooxygenases and NADPH-dependent cytochrome c reductase. Exposure to DE produced smaller increases in ethylmorphine demethylase activity on days 2 to 4 postvirus infection and also abolished the day 4 postinfection increase in NADPH-dependent cytochrome c reductase when compared with nonexposed mice. These data suggested to the authors that the relationship that exists between metabolic detoxification and resistance to infection in unexposed mice was altered during a short-term exposure to DE. Chen and Vostal (1981) measured the activity of AHH and the content of cytochrome P450 in the lungs and livers of rats exposed by inhalation of DE or intraperitoneal (i.p.) injection of a dichloromethane extract of DPM. In the inhalation exposures, the exhaust was diluted to achieve concentrations of 0.75 or 1.5 mg/m 3 DPM, and the exposure regimen was 20 h/day, 5.5 days/week for up to 9 mo. The concentration of total hydrocarbons and particle-phase hydrocarbons was not reported. Parenteral administration involved repeated injections at several dose levels for 4 days. Inhalation exposure had no significant effect on liver microsomal AHH activity; however, lung AHH activity was slightly reduced after 6 mo exposure to 1.5 mg/m 3

. An i.p. dose of DPM extract, estimated to be equivalent to the inhalation exposure, had no effect on AHH activity in liver or lungs. No changes were observed in cytochrome P450 contents in lungs or liver following inhalation exposure or i.p. treatment. Direct intratracheal administration of a dichloromethane DPM extract required doses greater than 6 mg/kg body weight before the activity of induced AHH in the lung was barely doubled; liver AHH activity remained unchanged (Chen, 1986). In related studies, Navarro et al. (1981) evaluated the effect of exposure to DE on rat hepatic and pulmonary microsomal enzyme activities. The same exposure regimen was employed (20 h/day, 5.5 days/week, for up to 1 year), and the exhaust was diluted to achieve concentrations of 0.25 and 1.5 mg/m 3 DPM (a few studies were also conducted at 0.75 mg/m 3 ). After 8 weeks of exposure, there was no evidence for the induction of cytochrome P450, cytochrome P448, or NADPH-dependent cytochrome c reductase in rat liver microsomes. One year of exposure had little, if any, effect on the hepatic metabolism of B[ a ]P. However, 1 year of exposure to 0.25 and 1.5 mg/m 3 significantly impaired the ability of lung microsomes to metabolize B[ a ]P (0.15 and 0.02 nmole/30 min/mg protein, respectively, versus 0.32 nmole/30 min/mg protein for the controls). There are conflicting results regarding the induction of microsomal AHH activities in the lungs and liver of rodents exposed to DE. One study reported induction of AHH activity in the lungs, liver, and prostate of rats exposed to DE containing 6.32 mg/m 3 DPM for 20 h/day for 42 days; however, no induction of AHH was observed in the lungs of rats and mice exposed to 6 mg/m 3 DPM for 8 h/day, 7 days/week for up to 8 mo or to 0.25 to 2 mg/m 3 for periods up to 25-78 years. Exposure to DE has not been shown to produce adverse effects on microsomal cytochrome P450 in the lungs or liver of rats or mice. The weight of evidence suggests that the absence of enzyme induction in the rodent lung exposed to DE is caused either by the unavailability of the adsorbed hydrocarbons or by their presence in quantities insufficient for enzyme induction.

5.1.3.3.11. Effects on behavior and neurophysiology . Studies on the effects of exposure to DE on the behavior and neurophysiology of laboratory animals are summarized in Table 5-14. Laurie et al. (1978) and Laurie et al. (1980) examined behavioral alterations in adult and neonatal rats exposed to DE. Exposure for 20 h/day, 7 days/week, for 6 weeks to exhaust containing 6 mg/m 3 DPM produced a significant reduction in adult spontaneous locomotor activity (SLA) and in neonatal pivoting (Laurie et al., 1978). In a follow-up study, Laurie et al. (1980) found that shorter exposure (8 h/day) to 6 mg/m 3 DPM also resulted in a reduction of SLA in adult rats. Laurie et al. (1980) conducted additional behavioral tests on adult rats exposed during their neonatal period. For two of three exposure situations (20 h/day for 17 days postparturition, or 8 h/day for the first 28 or 42 days postparturition), significantly lower SLA was observed in the majority of the tests conducted on the adults after week 5 of measurement. When compared with control rats, adult 15-month-old rats that had been exposed as neonates (20 h/day for 17 days) also exhibited a significantly slower rate of acquisition of a bar-pressing task to obtain food. The investigators noted that the evidence was insufficient to determine whether the differences were the result of a learning deficit or due to some other cause (e.g., motivational or arousal differences). These data are difficult to interpret in terms of health hazards to humans under ambient environmental conditions because of the high concentration of DE to which the laboratory rats were exposed. Additionally, there are no further concentration-response studies to assess at what exposure levels these observed results persist or abate. A permanent alteration in both learning ability and activity resulting from exposures early in life is a health hazard whose significance to humans should be pursued further. Neurophysiological effects from exposure to DE were investigated in rats by Laurie and Boyes (1980, 1981). Rats were exposed to diluted DE containing 6 mg/m 3 DPM for 8 h/day, 7 days/week from birth up until 28 days of age. Somatosensory evoked potential, as elicited by a 1 mA electrical pulse to the tibial nerve in the left hind limb, and visual evoked potential, as elicited by a flash of light, were the endpoints tested. An increased pulse latency was reported for the rats exposed to DE, and this was thought to be caused by a reduction in the degree ofTable 5-14. Effects of chronic exposures to diesel exhaust on behavior and neurophysiologySpecies/sexExposureperiodParticles(mg/m3)C × T(mgAh/m3)CO(ppm)NO2(ppm)SO2(ppm)Effects StudyRat, SpragueDawley, M8 h/day7 days/week1-4 weeks6 336-1,344 19 2.5 1.8Somatosensory and visual evokedpotentials revealed longer pulselatencies in pups exposed neonatallyLaurie and Boyes(1980, 1981)Rat, SpragueDawley, F20 h/day7 days week6 weeks6 5,040 19 2.51.8 Reduction in adult SLA and in neonatalpivotingLaurie et al. (1978)Rat, SpragueDawley, F8 or 20 h/day7 days/week3, 4, 6, or16 weeks6 1,008-13,440 19 2.51.8 Reduction in SLA in adults; neonatalexposures for 20 or 8 h/day causedreductions in SLA. Neonatalexposures for 20 h/day for 17 daysresulted in a slower rate of abar-pressing task to obtain foodLaurie et al. (1980)SLA = Spontaneous locomotor activity.

5-795-80 nerve myelinization. There was no neuropathological examination, however, to confirm this supposition. Based on the data presented, it is not possible to specify the particular neurological impairment(s) induced by the exposure to DE. Again, these results occurred following exposure to a high level of DE and no additional concentration-response studies were performed.

5.1.3.3.12. Effects on reproduction and development . Studies of the effects of exposure to DE on reproduction and development are summarized in Table 5-15. Twenty rats were exposed 8 h/day on days 6 through 15 of gestation to diluted DE containing 6 mg/m 3 DPM (Werchowski et al., 1980a,b; Pepelko and Peirano, 1983). There were no signs of maternal toxicity or decreased fertility. No skeletal or visceral teratogenic effects were observed in 20-day-old fetuses (Werchowski et al., 1980a). In a second study, 42 rabbits were exposed to 6 mg/m 3 DPM for 8 h/day on gestation days 6 through 18. No adverse effects on body weight gain or fertility were seen in the does exposed to DE. No visceral or skeletal developmental abnormalities were observed in the fetuses (Werchowski et al., 1980b). Pepelko and Peirano (1983) evaluated the potential for DE to affect reproductive performance in mice exposed from 100 days prior to exposure throughout maturity of the F 2 generation. The mice were exposed for 8 h/day, 7 days/week to 12 mg/m 3 DPM. In general, treatment-related effects were minimal. Some differences in organ and body weights were noted, but overall fertility and survival rates were not altered by exposure to DE. The only consistent change, an increase in lung weights, was accompanied by a gross pathological diagnosis of anthracosis. These data denoted that exposure to DE at a concentration of 12 mg/m 3 did not affect reproduction. See Section 5.3, which reports a lack of effects of exposure to DE on rat lung development (Mauderly et al., 1987b). Several studies have evaluated the effect of exposure to DE on sperm. Lewis et al. (1989) found no adverse sperm effects (sperm motility, velocity, densities, morphology, or incidence of abnormal sperm) in monkeys exposed for 7 h/day, 5 days/week for 104 weeks to 2 mg/m 3 DPM. In another study in which A/Strong mice were exposed to DE containing 6 mg/m 3 DPM for 8 h/day for 31 or 38 weeks, no significant differences were observed in sperm morphology between exposed and control mice (Pereira et al., 1981). It was noted, however, that there was a high rate of spontaneous sperm abnormalities in this strain of mice, and this may have masked any small positive effect. Quinto and De Marinis (1984) reported a statistically significant and dose-related increase in sperm abnormalities in mice injected intraperitoneally for 5 days with 50, 100, or 200 mg/kg of DPM suspended in corn oil. A significant decrease in sperm number was seen at the highest dose, but testicular weight was unaffected by the treatment.Table 5-15. Effects of chronic exposures to diesel exhaust on reproduction and development in laboratory animalsSpecies/sexExposureperiodParticles(mg/m3)C × T(mgAh/m3)CO(ppm)NO2(ppm)SO2(ppm)Effects StudyMouse,[C57BL]/6XC3H]F1, M5 days 50, 100, or200 mg/kgin corn oil;i.p.injection— — — — Dose-related increase insperm abnormalities;decrease in sperm number athighest dose; testicularweights unaffectedQuinto and DeMarinis (1984)Rat, Sprague-Dawley, F8 h/day7 days/week1.7 weeks6 571 20 2.7 2.1 No signs of maternal toxicityor decreased fertility; noskeletal or visceralteratogenic effects in 20-day-old fetusesWerchowski et al.(1980a)Pepelko andPeirano (1983)Rabbit, NewZealand Albino,F8 h/day7 days/week1.9 weeks6 638 20 2.7 2.1 No adverse effects onmaternal weight gain orfertility; no skeletal orvisceral teratogenic effects inthe fetusesWerchowski et al.(1980a)Pepelko andPeirano (1983)Monkey,Cynomolgus, M7 h/day5 days/week104 weeks2 7,280 11.5 1.5 0.8 No effects on sperm motility,velocity, density,morphology, or incidence ofabnormalitiesLewis et al. (1989)Mouse,A/Strong, M8 h/day7 days/week31 or38 weeks6 10,416-12,76820 2.7 2.1 No effect on spermmorphology; high rate ofspontaneous spermabnormalities may havemasked small effectsPereira et al. (1981)Mouse, CD-1,M, F8 h/day7 days/week6 to 28weeks12 4,032-18,816 33 4.4 5.0 Overall fertility and survivalrates were unaffected in thethree-generationreproductive study; onlyconsistent change noted, anincrease in lung weights, wasdiagnosed as anthracosisPepelko andPeirano (1983)

5-815-82 Watanabe and Oonuki (1999) investigated the effects of diesel engine exhaust on reproductive endocrine function in growing rats. The rats were exposed to whole diesel engine exhaust (5.63 mg/m 3 DPM, 4.10 ppm NO 2 , and 8.10 ppm NO
x ); a group was exposed to filtered exhaust without DPM, and a group was exposed to clean air. Exposures were for 3 mo beginning at birth (6 h rs /day for 5 days/week).Serum levels of testosterone and estradiol were significantly higher and folliclestimulating hormone significantly lower in animals exposed to whole DE and filtered exhaustcompared to controls. Luteinizing hormone was significantly decreased in the whole-exhaustexposed group as compared to the control and filtered groups. Sperm production and activity of testicular hyaluronidase were significantly reduced in both exhaust-exposed groups as compared to the control group. This study suggests that DE stimulates hormonal secretion of the adrenal cortex, depresses gonadotropin-releasing hormone, and inhibits spermatogenesis in rats. Because these effects were not inhibited by filtration, the gaseous phase of the exhaust appears more responsible than particulate matter for disrupting the endocrine system. The effects of freshly generated DE particles on the reproductive system of male Fischer 344 rats were investigated by Tsukue et al. (2001). Groups (n=25) of 13-mo

. old male rats were exposed to whole DE diluted to 0.33, 0.99 or 3.24 mg/m 3 (MMAD = 0.4 μ m) for 8 months
12 h rs /d ay , 7 d ays /w ee k. Subsequent to this exposure, evaluation of potential reproductive effect

was performed, including measurement of reproductive organ weights, sperm characteristics and number, gonadotrophins, testosterone, and inhibin. Results showed either no effect or effects with an inconsistent dose-response character that typically were not different from controls even at the highest exposure concentration. No teratogenic, embryotoxic, fetotoxic, or female reproductive effects were observed in mice, rats, or rabbits at exposure levels up to 12 mg/m 3 DPM. Effects on sperm morphology and number were reported in hamsters and mice exposed to high doses of DPM; however, no adverse effects were observed in sperm obtained from monkeys exposed at 2 mg/ m 3 for 7 h rs /day, 5 days/week for 104 weeks. Concentrations of 12 mg/m 3 DPM did not affect male rat reproductive fertility in the F 0 and F 1 generation breeders. Thus, exposure to DE would not appear to be a reproductive or developmental hazard.

5.2. MODE OF ACTION OF DIESEL EXHAUST-INDUCED NONCANCER EFFECTS 5.2.1. Comparison of Health Effects of Filtered and Unfiltered Diesel Exhaust There exist a total of four chronic toxicity studies of DE, in which the experimental protocol included exposing test animals to exhaust containing no particles. Comparisons were then made between the effects caused by whole, unfiltered exhaust and those caused by the gaseous components of the exhaust. Concentrations of components of the exposure atmospheres in these four studies are given in Table 5-16.5-83 Heinrich et al. (1982) compared the toxic effects of whole and filtered DE on hamsters and rats. Exposures were at 3.9 mg/m 3 for 7 to 8 h rs /day and 5 days/week. Rats exposed for 24 mo to either whole or filtered exhaust exhibited no significant changes in respiratory frequency, respiratory minute volume, compliance or resistance as measured by a whole-body plethysmography, or heart rate. In the hamsters, histological changes (adenomatous proliferations) were seen in the lungs of animals exposed to either whole or filtered exhaust; however, in all groups exposed to the whole exhaust the number of hamsters exhibiting such lesions was significantly higher than for the corresponding groups exposed to filtered exhaust or clean air. Severity of the lesions was, however, not reported. In a second study, Heinrich et al. (1986a, see also Stöber, 1986) compared the toxic effects of whole and filtered DE on hamsters, rats, and mice. The test animals (96 per test group) were exposed to 4.24 mg DPM/m 3 for 19 h rs /day, 5 days/week for 120 (hamsters and mice) or 140 (rats) weeks. Body weights of hamsters were unaffected by either exposure. Body weights of rats and mice were reduced by the whole exhaust but not by the filtered exhaust. Exposure-related higher mortality rates occurred in mice after 2 years of exposure to whole exhaust. After 1 year of exposure to the whole exhaust, hamsters exhibited increased lung weights, a significant increase in airway resistance, and a nonsignificant reduction in lung compliance. For the same time period, rats exhibited increased lung weights, a significant decrease in dynamic lung compliance, and a significant increase in airway resistance. Test animals exposed to filtered exhaust did not exhibit such effects. Histopathological examination indicated that different levels of response occurred in the three species. In hamsters, filtered exhaust caused no significant histopathological effects in the lung; whole exhaust caused thickened alveolar septa, bronchioloalveolar hyperplasia, and emphysematous lesions. In mice, whole exhaust, but not filtered exhaust, caused multifocal bronchioloalveolar hyperplasia, multifocal alveolar lipoproteinosis, and multifocal interstitial fibrosis. In rats, there were no significant morphological changes in the lungs following exposure to filtered exhaust. In rats exposed to whole exhaust, there were severe inflammatory changes in the lungs, thickened alveolar septa, foci of macrophages, crystals of cholesterol, and hyperplastic and metaplastic lesions. Biochemical studies of lung lavage fluids of hamsters and mice indicated that exposure to filtered exhaust caused fewer changes than did exposure to whole exhaust. The latter produced significant increases in lactate dehydrogenase, alkaline phosphatase, glucose-6- phosphate dehydrogenase (G6PDH), total protein, protease (pH 5.1), and collagen. The filtered exhaust had a slight but nonsignificant effect on G6PDH, total protein, and collagen. Similarly, cytological studies showed that while the filtered exhaust had no effect on differential cell counts, the wholeTable 5-16. Composition of exposure atmospheres in studies comparing unfiltered and filtered diesel exhaustaSpecies/sexExposurebperiodParticles(mg/m3)C × t(mgAh/m3)CO(ppm)NO2(ppm)SO2(ppm)Effects StudyRat, Wistar, F;Hamster, Syrian7 h/day5 days/week104 weeksUfFC3.9——14,196 18.518.0—1.21.0—3.12.8—No effect on pulmonary function or heart rate in rats; increases inpulmonary adenomatous proliferations in hamsters, UFsignificantly higher than F or CHeinrich et al.(1982)Rat, F344, F 8h/day7 days/week104 weeksUfFcC4.9——28,538 7.0——1.8——13.1——Body weight decrease after 6 mo in UF, 18 mo in f; lung/bodyrate weight rate higher in both groups at 24 mo; at 2 years,fibrosis and epithelial hyperplasia in lungs of uf; nominal lungand spleen histologic changesIwai et al. (1986)Rat, F344, M, F;Hamster, Syrian,M, F16 h/day5 days/week104 weeksUfUfUfFdC0.72.26.6——5,82418,30454,912——32.032.01.0——————————Uf: elevated red and white cell counts, hematocrit and hemoglobin; increased heart/body weight and right ventricular/heartweight ratios; lower left ventricular contractility; changes in bloodchemistry; obstructive and restrictive lung disease; F: no effectsBrightwell et al.(1986)Rat, Wistar, F;Hamster, Syrian, F;Mouse NMRI, F19 h/day5 days/week120 to140 weeksUfFdC4.24——48,33656,39212.511.10.161.51.2—3.11.02—Uf: decreased body wt in rats and mice but not hamsters; increased mortality, mice only; decreased lung compliance and increasedairway resistance, rats and hamsters; species differences in lunglavage enzymes and cell counts and lung histopathology andcollagen content, most pronounced in rats; F: no effect onglucose-6-phosphate dehydrogenase, total protein, and lungcollagenHeinrich et al.(1986a)Mouse, NMRI, F,C57BL/6N, F18 h/day5 days/week23 mo(NMRI)24 mo(C57BL/6N)UfFC4.50.010.0140,365 14.214.20.22.32.90.012.82.40.1Uf: increased lung wet weight starting at 3 moF: no noncancer effects reportedHeinrich et al.(1995)aMan values.bUF= unfiltered whole exhaust, F = filtered exhaust, C = control.cReported to have the same component concentrations as the unfiltered, except particles were present in undetectable amounts.dConcentrations reported for high concentration level only.

5-845-85 counts, the whole exhaust resulted in an increase in leukocytes (161 ± 43.3 / μ L versus 55.7 ± 12.8/μL controls), a decrease in AMs (30.0 ± 12.5 versus 51.3 ± 12.5 / μ L in the controls),
and an increase in granulocytes (125 ± 39.7 versus 1.23 ± 1.14 / μ L in the controls). All values presented for this study are the mean with its standard deviation. The differences were significant for each cell type. There was also a small increase in lymphocytes (5.81 ± 4.72 versus 3.01 ± 1.23 μL in the controls). Iwai et al. (1986) exposed rats (24 per group) to whole or filtered DE 8 h/day, 7 days/week for 24 mo. The whole exhaust was diluted to achieve a concentration of 4.9 ± 1.6 mg/m 3 DPM. Body weights in the whole exhaust group began to decrease after 6 mo and in both exposed groups began to decrease after 18 mo when compared with controls. Lung-to-body weight ratios of the rats exposed to the whole exhaust showed a significant increase ( p <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 mo of exposure to whole exhaust, DPM accumulated in AMs, and Type II cell hyperplasia was observed. After 2 years of exposure, the alveolar walls had become fibrotic with mast cell infiltration and epithelial hyperplasia. In rats exposed to filtered exhaust, after 2 years there were only minimal histologic changes in the lungs, with slight hyperplasia and stratification of bronchiolar epithelium and infiltration of atypical lymphocytic cells in the spleen. Brightwell et al. (1986) evaluated the toxic effects of whole and filtered DE on rats and hamsters. Three exhaust dilutions were tested, producing concentrations of 0.7, 2.2, and 6.6 mg/m 3 DPM. The test animals (144 rats and 312 hamsters per exposure group) were exposed for five 16-h periods per week for 2 years. The four exposure types were gasoline, gasoline catalyst, diesel, and filtered diesel. The results presented were limited to statistically significant differences between exhaust-exposed and control animals. The inference from the discussion section of the paper was that there was a minimum of toxicity in the animals exposed to filtered DE: “It is clear from the results presented that statistically significant differences between exhaust-exposed and control animals are almost exclusively limited to animals exposed to either gasoline or unfiltered diesel exhaust.” Additional results are described in Section 5.1.3.3. Heinrich et al. (1995) exposed female NMRI and C57BL/6N mice to a DE dilution that resulted in a DPM concentration of 4.5 mg/m 3 and to the same dilution after filtering to remove the particles. This study is focused on the carcinogenic effects of DPM exposure, and inadequate information was presented to compare noncancer effects in filtered versus unfiltered
exhaust. A comparison of the toxic responses in laboratory animals exposed to whole exhaust or filtered exhaust containing no particles demonstrates across studies that when the exhaust is sufficiently diluted to limit the concentrations of gaseous irritants (NO 2 and SO 2 ), irritant vapors (aldehydes), CO, or other systemic toxicants, the diesel particles are the prime etiologic agents5-86 of noncancer health effects, although additivity or synergism with the gases cannot be ruled out. These toxic responses are both functional and pathological and represent cascading sequelae of lung pathology based on concentration and species. The diesel particles plus gas exposures produced biochemical and cytological changes in the lung that are much more prominent than those evoked by the gas phase alone. Such marked differences between whole and filtered DE are also evident from general toxicological indices, such as decreases in body weight and increases in lung weights, pulmonary function measurements, and pulmonary histopathology (e.g., proliferative changes in Type II cells and respiratory bronchiolar epithelium, fibrosis). Hamsters, under equivalent exposure regimens, have lower levels of retained DPM in their lungs than rats and mice do and, consequently, less pulmonary function impairment and pulmonary pathology. These differences may result from lower DPM inspiration and deposition during exposure, greater DPM clearance, or lung tissue less susceptible to the cytotoxicity of deposited
DPM.

5.2.2. Mode of Action for the Noncarcinogenic Effects of DPM As noted in Chapter 2, diesel emissions are a complex mixture that includes both a vapor phase and a particle phase. The particle phase consists of poorly soluble carbon particles on the surfaces of which are adsorbed a large number of organic and inorganic compounds. Although the effects to be discussed are considered attributable to the particle phase (termed diesel particulate matter or DPM), additive or synergistic effects due to the vapor phase cannot be totally discounted. This may be especially so in the human studies and the animal toxicology studies where exposure is to various dilutions of diesel emissions, or in the in vitro studies in which the test material was captured by filtration. The mechanisms by which DPM is inhaled, deposited, and cleared from the respiratory tract are discussed in Chapter 3. DPM deposited upon airway surfaces may be cleared from the respiratory tract completely, or may be translocated to other sites within the respiratory system. In rats, the pathogenic sequence following the deposition of inhaled DPM begins with the interaction of DPM with airway epithelial cells and phagocytosis by AMs. The airway epithelial cells and activated AMs release chemotactic factors that attract neutrophils and additional AMs. As the lung burden of DPM increases, there is an aggregation of particle-laden AMs in alveoli adjacent to terminal 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 macrophages engulfing the DPM may release cytokines, growth factors, and proteases, which may cause inflammation, cell injury, cell proliferation, hyperplasia, and fibrosis. This is especially true under lung overload conditions occurring in laboratory rats when the rate of deposition exceeds the rate of alveolar clearance. This phenomenon is described in5-87 Chapter 3. The mechanisms leading to the generation of oxygen radicals and subsequent lung injury are described in Chapter 7, Section 7.4.3. DPM is a poorly soluble particle whose rate of clearance by dissolution is likely insignificant compared to its rate of clearance as an intact particle. The organic material adsorbed to the surface is desorbed from the DPM and may enter into metabolic reactions and be
activated and enter into reactions with other macromolecules or be detoxified and excreted (Figure 7-1). The diesel particle may be cleared directly by the clearance mechanisms described in Chapter 3. The organic material desorbed from the particle (described in Chapter 7, Section 7.4.7) appears to be associated with the immunological changes described above. The potential adjuvant effects of DPM have also been studied. The results indicate that the nonextractable particle core and the organic matter adsorbed to the core both contribute to the adjuvant activity of DPM. Further, it is possible that any of the plethora of compounds present in the organic fraction of DPM, including various PAH, may elicit this response. Thus, the available evidence indicates that DPM has the potential to produce pathological and immunological changes in the respiratory tract. Moreover, the magnitude of these responses is determined by the dose delivered to the respiratory tract and is attributable to both the carbon core and the adsorbed organic materials.

5.3. INTERACTIVE EFFECTS OF DIESEL EXHAUST A multitude of factors may influence the susceptibility to exposure to DE as well as the resulting response. Some of these have already been discussed in detail (e.g., the composition of DE and concentration-response data); others will be addressed in this section (e.g., the interaction of DE with factors particular to the exposed individual and the interaction of DE components with other airborne contaminants). In a study discussed already in this chapter, Mauderly et al. (1990a) compared the susceptibility of normal rats and rats with preexisting laboratory-induced pulmonary emphysema exposed for 7 h/day, 5 days/week for 24 mo to DE containing 3.5 mg/m 3 DPM or to clean air (controls). Emphysema was induced in one-half of the rats by intratracheal instillation of elastase 6 weeks before exhaust exposure. Measurements included lung burdens of DPM, respiratory function, bronchoalveolar lavage, clearance of radiolabeled particles, pulmonary immune responses, lung collagen, excised lung weight and volume, histopathology, and mean linear intercept of terminal air spaces. None of the data for the 63 parameters measured suggest that rats with emphysematous lungs were more susceptible than rats with normal lungs to the effects of DE exposure. In fact, each of the 14 emphysema-exhaust interactions detected by statistical analysis of variance indicated that emphysema acted to reduce the effects of DE exposure. DPM accumulated much less rapidly in the lungs of emphysematous rats than in those5-88 of normal rats. The mean lung burdens of DPM in the emphysematous rats were 39%, 36%, and 37% of the lung burdens of normal rats at 12, 18, and 24 mo, respectively. No significant interactions were observed among lung morphometric parameters. Emphysema prevented the exhaust-induced increase for three respiratory indices of expiratory flow rate at low lung volumes, reduced the exhaust-induced increase in nine lavage fluid indicators of lung damage, prevented the expression of an exhaust-induced increase in lung collagen, and reduced the exhaust-induced delay in DPM clearance. Mauderly et al. (1987b) evaluated the relative susceptibility of developing and adult rat lungs to damage by exposure to DE. Rats (48 per test group) were exposed to DE containing 3.5 mg/m 3 DPM and about 0.8 ppm NO 2 . Exposures were for 7 h/day, 5 days/week through gestation to the age of 6 mo, or from the age of 6 to 12 mo. Comparative studies were conducted on respiratory function, immune response, lung clearance, airway fluid enzymes, protein and cytology, lung tissue collagen, and proteinases in both age groups. After the 6-moexposure, adult rats, compared with controls, exhibited (1) more focal aggregates of particlecontaining AMs in the alveolar ducts near the terminal bronchioles, (2) a sixfold increase in the neutrophils (as a percentage of total leukocytes) in the airway fluids, (3) a significantly higher number of total lymphoid cells in the pulmonary lymph nodes, (4) delayed clearance of DPM and radiolabeled particles (t 1/2 = 90 days versus 47 days for controls), and (5) increased lung weights. These effects were not seen in the developing rats. On a weight-for-weight (milligrams of DPM per gram of lung) basis, DPM accumulation in the lungs was similar in developing and adult rats immediately after the exposure. During the 6-mo postexposure period, DPM clearance was much more rapid in the developing rats, approximately 2.5-fold. During postexposure, diesel particle-laden macrophages became aggregated in the developing rats, but these aggregations were located primarily in a subpleural position. The authors concluded that exposure to DE, using pulmonary function, structural (qualitative or quantitative) biochemistry as the indices, did not affect the developing rat lung more severely than the adult rat lung. As a result of the increasing trend of using diesel-powered equipment in coal mining operations and the concern for adverse health effects in coal miners exposed to both coal dust or coal mine dust and DE, Lewis et al. (1989) and Karagianes et al. (1981) investigated the interaction of coal dust and DE. Lewis et al. (1989) exposed rats, mice, and cynomolgus monkeys to (1) filtered ambient air, (2) 2 mg/m 3 DPM, (3) 2 mg/m 3 respirable coal dust, and (4) 1 mg/m 3 of both DPM and respirable coal dust. Gaseous and vapor concentrations were identical in both DE exposures. Exposures were for 7 h/day, 5 days/week for up to 24 mo. Synergistic effects between DE and coal dust were not demonstrated; additive toxic effects were the predominant effects noted. Karagianes et al. (1981) exposed rats (24 per group) to DE containing 8.3 mg/m 3 of DPM alone or in combination with about 6 mg/m 3 of coal dust. No synergistic effects were found5-89 between DE and coal dust; additive effects in terms of visual dust burdens in necropsied lungs were related to dose (i.e., length of exposure and airborne particulate concentrations). The health effects of airborne contaminants from sources other than diesel engines may be altered in the presence of DPM by their adsorption onto the diesel particles. When adsorbed 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 physiological responses may be enhanced. Nitrogen dioxide adsorbed onto carbon particles caused pulmonary parenchymal lesions in mice, whereas NO 2 alone produced edema and inflammation but no lesions (Boren, 1964). Exposure to formaldehyde and acrolein adsorbed onto carbon particles (1 to 4 

### : m) resulted in the recruitment of PMNs to tracheal and intrapulmonary epithelial tissues but not when the aldehydes were tested alone (Kilburn and McKenzie, 1978). Madden et al. (2000) observed that O 3 exposure increased the bioactivity of DPM. DPM, preexposed to O 3 for 48 h or nonozone-exposed DPM (1 to 500 

### : g), was instilled into the lungs of laboratory rats. Lung inflammation and injury were examined 24 h after instillation by lung lavage. DPM pre-exposed to 0.1 PPM O 3 was more potent in increasing neutrophilia, lavage total protein, and LDH compared to unexposed DPM. Treatment of DPM with higher
concentrations of O 3 (1.0 PPM) decreased the bioactivity of the particles. There is no direct evidence that DE, at concentrations found in the ambient environment, interacts with other substances in the exposure environment or the physiological status of the exposed subject other than impaired resistance to respiratory tract infections. Although there is experimental evidence that gases and vapors can be adsorbed onto carbonaceous particles, enhancing the toxicity of these particles when deposited in the lung, there is no evidence for an increased health risk from such interactions with DPM under urban atmospheric conditions. Likewise, there is no experimental evidence in laboratory animals that the youth or preexisting emphysema of an exposed individual enhances the risk of exposure to DE.

5.4. COMPARATIVE RESPONSIVENESS AMONG SPECIES TO THE
HISTOPATHOLOGIC EFFECTS OF DIESEL EXHAUST

There is some evidence indicating that species may differ in pulmonary responses to DE. Mauderly (1994) compared the pulmonary histopathology of rats and mice after 18 mo of exposure to DE. There was less aggregation of macrophages in mice. Diffuse septal thickening was noted in the mice, but there were few inflammatory cells, no focal fibrosis, little epithelial hyperplasia, and no epithelial metaplasia, as was observed in rats. Heinrich et al. (1986a) reported that wet lung weight of hamsters increased only 1.8-fold following chronic exposure to DE, compared with an increase of 3.4-fold in rats. Smaller increases in neutrophils, lactic acid dehydrogenase, collagen, and protein supported the conclusion of a lesser inflammatory response5-90 in Syrian hamsters. The histopathologic changes in the lungs of Chinese hamsters after 6 mo exposure to DE, on the other hand, was similar to that of rats (Pepelko and Peirano, 1983). Guinea pigs respond to chronic DE exposure with a well-defined epithelial proliferation, but it is based on an eosinophilic response in contrast to the neutrophil-based responses in other species. Epithelial hyperplasia and metaplasia were quite striking in the terminal and respiratory bronchioles of cats exposed for 27 mo to DE (Plopper et al., 1983). This study is of particular interest because the terminal airways of cats are more similar to those of humans than rodent species are. It should be noted, however, that exposure concentrations were very high (12 mg/m 3 ) for most of the period. Lewis et al. (1989) exposed rats and cynomolgus monkeys 8 h per day, 5 days per week for 2 years to DE at a particle concentration of 2 mg/m 3

. Unfortunately, this exposure rate was sufficiently low that few effects were noted in either species other than focal accumulations of particles, primarily in the alveolar macrophages, interstitium, and lymphoid tissue. It is apparent that species do vary in their pulmonary responses to DE exposure, despite the difficulty in making direct comparisons because of differences in exposure regimes, lifespans, and pulmonary anatomy. Most species do respond, however, suggesting that humans are likely to be susceptible to induction of pulmonary pathology during chronic exposure to DE at some level.

5.5. DOSE-RATE AND PARTICULATE CAUSATIVE ISSUES The purpose of animal toxicological experimentation is to elucidate mechanisms of action and identify the hazards and dose-response effects posed by a chemical substance or complex mixture and to extrapolate these effects to humans for subsequent health assessments. The cardinal principle in such a process is that the intensity and character of the toxic action are a function of the dose of the toxic agent(s) that reaches the critical site of action. The considerable body of evidence reviewed clearly denotes that major noncancerous health hazards may be presented to the lung following the inhalation of DE. Based on pulmonary function and histopathological and histochemical effects, a determination can be made concerning which dose/exposure rates of DE (expressed in terms of the DPM concentration) result in injury to the lung and which appear to elicit no effect. The inhalation of poorly soluble particles, such as those found in DE, increases the pulmonary particulate burden. When the dosing rate exceeds the ability of the pulmonary defense mechanisms to achieve a steady-state lung burden of particles, there is a slowing of clearance and the progressive retention of particles in the lung that can ultimately approach a complete cessation of lung clearance (Morrow, 1988). This phenomenon, which is reviewed in Chapter 3, has practical significance both for the interpretation of experimental inhalation data and for the prevention of disease in humans exposed to airborne particles.5-91 The data for exposure intensities that cause adverse pulmonary effects demonstrate that they are less than the exposure intensities reported to be necessary to induce lung tumors. Using 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 pulmonary effects was 56 mg



### A h



### A m -3 /week (Brightwell et al., 1986). The lowest-observed-effect level for adverse pulmonary effects (noncancer) in rats was 70 mg



### A h



### A m -3 /week (Lewis et al., 1989), and for pulmonary tumors, 122.5 mg



### A h



### A m -3 /week (Mauderly et al., 1987a). The results clearly show that noncancerous pulmonary effects are produced at lower exposure intensities than are pulmonary tumors. Such data support the position that inflammatory and proliferative changes in the lung may play a key role in the etiology of pulmonary tumors in exposed rats (Mauderly et al., 1990b). The effects of DE on the developing lung and on a model of a preexisting disease state have been studied in rats (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 was faster in the younger lung. Mauderly et al. (1990a) compared the pulmonary response to inhalation of DE in rats with elastase-induced emphysema with normal rats. They found that respiratory tract effects were not more severe in emphysematous rats and that the lung burden of particles was less in the compromised rat. These studies provide limited evidence that some factors that are often considered to result in a wider distribution of sensitivity among members of the population may not have this effect with diesel exposure. However, these studies have no counterpart in human studies and extrapolation to humans remains uncertain. There is also the issue of whether the noncancerous health effects related to exposure to DE are caused by the carbonaceous core of the particle or substances adsorbed onto the core, or
both. Current understanding, derived primarily from studies in rats, suggests that much of the toxicity resulting from the inhalation of DE relates to the carbonaceous core of the particles. Several studies on inhaled aerosols demonstrate that lung reactions characterized by an appearance of particle-laden AMs and their infiltration into the alveolar ducts, adjoining alveoli, and tracheobronchial lymph nodes; hyperplasia of Type II cells; and the impairment of 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; Karagianes et al., 1981), titanium dioxide (Heinrich et al., 1995; Lee et al., 1985), CB (Nikula et al., 1995; Heinrich et al., 1995), titanium tetrachloride hydrolysis products (Lee et al., 1986), quartz (Klosterkötter and Bünemann, 1961), volcanic ash (Wehner et al., 1986), amosite (Bolton et al., 1983), and manmade mineral fibers (Lee et al., 1988) among others. In more recent studies, animals have been exposed to CB that is similar to the carbon core of the DE particle. Nikula et al. (1995) exposed rats for 24 mo to CB or DE at target exposure concentrations of 2.55-92 and 6 mg/m 3 (exposure rates of 200 or 520 mg



### A h



### A m -3 /week). Both concentrations induced AM accumulation, epithelial proliferation, inflammation, and fibrosis. They observed essentially no difference in potency of nonneoplastic or in tumor responses based on a regression analysis. Dungworth et al. (1994) reported moderate to severe inflammation characterized by multifocal bronchoalveolar hyperplasia, alveolar histiocytosis, and focal segmental fibrosis in rats exposed to CB for up to 20 mo at exposure rates of 510 to 540 mg



### A h



### A m -3 /week. The observed lung pathology reflects notable dose-response relationships and usually evolves in a similar manner. With increasing dose, there is an increased accumulation and aggregation of particle-laden AMs, Type II cell hyperplasia, a foamy (degenerative) macrophage response, alveolar proteinosis, alveolar bronchiolization, cholesterol granulomas, and often squamous cell carcinomas and bronchioalveolar adenomas derived from metaplastic squamous cells in the areas
of alveolar bronchiolization. Heinrich et al. (1995) compared effects of diesel exposure in rats and mice with exposure to titanium dioxide or carbon black. Exposures to TiO 2 and carbon black were adjusted during the exposure to result in a similar lung burden for the three types of particles. At similar lung burdens in the rat, DPM, TiO 2 , and CB had nearly identical effects on lung weights and on the incidence of lesions, both noncancer and cancer. Also, a similar effect on clearance of a labeled test aerosol was measured for the different particles. A comparison of the effect of DPM, TiO 2 , and carbon black exposures in mice also showed a similar effect on lung weight, but noncancer effects were not reported and no significant increase in tumors was observed. Murphy et al. (1998) compared the toxicological effects of DPM with three otherparticles chosen for their differing morphology and surface chemistry. One mg each of wellcharacterized crystalline quartz, amorphous silica, CB, and DPM was administered to laboratory rats by a single intratracheal instillation. The laboratory rats were sacrificed at 48 h, and 1, 6, and 12 weeks after instillation. Crystalline quartz produced significant increases in lung permeability, persistent surface inflammation, progressive increases in pulmonary surfactant and activities of epithelial marker enzymes up to 12 wk after primary exposure. Amorphous silica did not cause progressive effects but did produce initial epithelial damage with permeability changes that regressed with time after exposure. By contrast, CB had little if any effect on lung permeability, epithelial markers, or inflammation. Similarly, DPM produced only minimal changes, although 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 that the surface chemistry of the particle is more important than ultrafine size in explaining biological activity. These experiments provide strong support for the idea that DE toxicity results from a mechanism that is analogous to that of other relatively inert particles in the lung. This5-93 qualitative similarity exists along with some apparent quantitative differences in the potency of various particles for producing effects on the lung or on particle clearance. The exact relationship between toxicity and particle size within the ultrafine particle mode, including DPM (BéruBé et al., 1999), remains unresolved. Studies reviewed in the PM CD (U.S. EPA, 1996) suggest a greater inherent potential toxicity of inhaled ultrafine particles. Exposure to ultrafine particles may increase the release of proinflammatory mediators that could be involved in lung disease. For example, Driscoll and Maurer (1991) compared the effects of fine (0.3 

### : m) and ultrafine (0.02 

### : m) TiO 2 particles instilled into the lungs of laboratory rats. Although both size modes caused an increase in the numbers of AMs and PMNs in the lungs, and release of TNF and fibronectin by AMs, the 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 macrophages, some necrosis of macrophages, and slight interstitial inflammation of the alveolar duct region. Moreover, collagen increased only with exposure to ultrafine particles. Oberdörster et al. (1992) compared the effects of fine (0.25 

### : m) and ultrafine (0.02 

### : m)
TiO 2 particles instilled into the lungs of laboratory rats on various indicators of inflammation. Instillation of ultrafine particles increased the number of total cells recovered by lavage, decreased the percentage of AMs, and increased the percentage of PMNs and protein. Instillation with fine particles did not cause statistically significant effects. Thus, the ultrafine particles had greater pulmonary inflammatory potency than did larger sizes of this material. The investigators attributed the enhanced toxicity to greater interaction of the ultrafine particles with their large surface area, with alveolar and interstitial macrophages, which resulted in enhanced release of inflammatory mediators. They suggested that ultrafine particles of low in vitro solubility appear to enter the interstitium more readily than do larger sizes of the same material, which accounted for the increased contact with macrophages in this compartment of the lung. Driscoll and Maurer (1991) noted that the pulmonary retention of ultrafine TiO 2 particles instilled into rat lungs was greater than for the same mass of fine-mode TiO 2 particles. Thus, the available evidence tends to suggest a potentially greater toxicity for inhaled ultrafine particles. Particle size, volume, surface area, and composition may be the critical elements in the overload phenomenon following exposure to particles, which could explain those quantitative differences. The overloaded AMs secrete a variety of cytokines, oxidants, and proteolytic enzymes that are responsible for inducing particle aggregation and damaging adjacent epithelial tissue (Oberdörster, 1994). For a more detailed discussion of mechanism, see Chapter 3. On the basis of currently available laboratory animal data, the principal noncancerous health hazard to humans posed by exposure to DE is a structural or functional injury to the lung.
Such effects are demonstrable at dose rates or cumulative doses of DPM lower than those5-94 reported to be necessary to induce lung tumors in rats. An emerging human health issue concerning short-term exposure to ambient DE/DPM is the potential for allergenic responses in several studies. Heightened allergenic responses including increased cytokine production as well as increased numbers of inflammatory cells have been detected in nasal lavage from humans exposed to inhaled or instilled DE/DPM. In individuals already allergic to ragweed, exposure to DE/DPM with the allergen was observed to result in an enhanced allergenic response, particularly IgE production. Current knowledge indicates that the carbonaceous core of diesel particles is the major causative factor in the injury to the lung and that other factors such as the cytotoxicity of adsorbed substances on the particles also may play a role. The lung injury appears to be mediated through effects on pulmonary AMs. Because noncancerous pulmonary effects occur at lower doses than tumor induction does in the rat, and because these effects may be cofactors in the etiology of DE-induced tumors, noncancerous pulmonary effects must be considered in the total evaluation of DE, notably the particulate component.

5.6. SUMMARY AND DISCUSSION
5.6.1. Effects of Diesel Exhaust on Humans The most readily identified acute noncancer health effect of DE on humans is its ability to elicit subjective complaints of eye, throat, and bronchial irritation and neurophysiological symptoms such as headache, lightheadedness, nausea, vomiting, and numbness and tingling of the extremities. Studies of the perception and offensiveness of the odor of DE and a human volunteer study in an exposure chamber have demonstrated that the time of onset of the human subjective symptoms is inversely related to increasing concentrations of DE and the severity is directly related to increasing concentrations of DE. In one study in which a diesel engine was operated under varying load conditions, a dilution factor of 140 to 475 was needed to reduce the
exhaust level to an odor-detection threshold level. A public health issue is whether short-term exposure to DE might result in an acute decrement in ventilatory function and whether the frequent repetition of such acute 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 in occupationally exposed workers at the beginning and end of a workshift. In studies of underground miners, bus garage workers, dockworkers, and locomotive repairmen, increases in respiratory symptoms (cough, phlegm, and dyspnea) and decreases in lung function (FVC,
FEV 1 , PEFR, and FEF 25-75 ) over the course of a workshift were generally found to be minimal and not statistically significant. In a study of acute respiratory responses in diesel bus garage workers, there was an increased reporting of cough, labored breathing, chest tightness, and wheezing, but no reductions in pulmonary function were associated with exposure to DE. Pulmonary function was affected in stevedores over a workshift exposure to DE but normalized5-95 after a few days without exposure to DE fumes. In a third study, there was a trend toward greater ventilatory function changes during a workshift among coal miners, but the decrements were similar in miners exposed and not exposed to DE. Smokers appeared to demonstrate larger workshift respiratory function decrements and increased incidence of respiratory symptoms. Acute sensory and respiratory symptoms were earlier and more sensitive indicators of potential health risks from diesel exposure than were decrements in pulmonary function. Studies on the acute health effects of exposure to DE 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, equivocal evidence for an exposure-response relationship. The environmental contaminants have frequently been below permissible workplace exposure limits; in those few cases where health effects have been reported, the authors have failed to identify conclusively the individual or collective causative agents in the DE. Chronic effects of DE exposure have been evaluated in epidemiologic studies of occupationally exposed workers (metal and nonmetal miners, railroad yard workers, stevedores, and bus garage mechanics). Most of the epidemiologic data indicate an absence of an excess risk of chronic respiratory disease associated with exposure to DE. In a few studies, a higher prevalence of respiratory symptoms, primarily cough, phlegm, or chronic bronchitis, was observed among the exposed. These increased symptoms, however, were usually not accompanied by significant changes in pulmonary function. Reductions in FEV 1 and FVC and, to a lesser extent, FEF 50 and FEF 75 , also have been reported. Two studies detected statistically significant decrements in baseline pulmonary function consistent with obstructive airway disease. One study of stevedores had a limited sample size of 17 exposed and 11 controls. The second study in coal miners showed that both underground and surface workers at diesel-use mines had somewhat lower pulmonary performance than their matched controls. The proportion of workers in or at diesel-use mines, however, showed equivalent evidence of obstructive airway disease, and for this reason the authors of the second paper felt that factors other than diesel exposure might have been responsible. A doubling of the prevalence of minor restrictive airway disease was also observed in workers in or at diesel-use mines. These two studies, coupled with other reported nonsignificant trends in respiratory flow-volume measurements, suggest that exposure to DE may impair pulmonary function among occupational populations. Epidemiologic studies of the effects of DE on organ systems other than the pulmonary system are scant. Whereas a preliminary study of the association of cardiovascular mortality and exposure to DE found a fourfold higher risk ratio, a more comprehensive epidemiologic study by the same investigators found no significant difference between the observed and expected number of deaths caused by cardiovascular disease.5-96 Caution is warranted in the interpretation of results from the epidemiologic studies that have addressed noncarcinogenic health effects from exposure to DE. These investigations suffer from myriad methodological problems, including (1) incomplete information on the extent of exposure to DE, necessitating in some studies estimations of exposures from job titles and resultant misclassification; (2) the presence of confounding variables such as smoking or occupational exposures to other toxic substances (e.g., mine dusts); and (3) the short duration and low intensity of exposures. These limitations restrict drawing definitive conclusions as to the cause of any noncarcinogenic DE effect, observed or reported. It is also apparent that at some level of exposure DE as measured by DPM appears to have the potential to induce airway inflammation in humans without disease. Also, in one other study peripheral blood changes were noted. An emerging area of concern is the immunological changes that have been documented in response to DE exposure and the potential relationship of these changes to the explosive growth of asthma in human populations.

5.6.2. Effects of Diesel Exhaust on Laboratory Animals Laboratory animal studies of the toxic effects of DE have involved acute, subchronic, and chronic exposure regimens. In acute exposure studies, toxic effects appear to have been associated primarily with high concentrations of carbon monoxide, nitrogen dioxide, and aliphatic aldehydes. In short- and long-term studies, toxic effects have been associated with exposure to the complex exhaust mixture. Effects of DE in various animal species are summarized in Tables 5-2 to 5-15. In short-term studies, health effects related to function, when found, are mild and result from extremely high DPM concentrations of about 6 mg/m 3 and extensive durations of exposure approximating 20 h/day. There is ample evidence, however, that other pathophysiological effects such as accumulation of DPM in pulmonary tissues, evidence of inflammatory response, AM aggregation and accumulation near the terminal bronchioles, Type II cell proliferation, and the thickening of alveolar walls adjacent to AM aggregation do occur under short-term exposures at lower levels of DE. Little evidence exists, however, from short-term studies that exposure to DE impairs lung function. Chronic exposures cause lung pathology that results in altered pulmonary function and increased DPM retention in the lung. Exposures to DE have also been associated with increased susceptibility to respiratory tract infection, neurological or behavioral changes, an increase in banded neutrophils, and morphological alterations in the liver.

5.6.2.1. Effects on Survival and Growth The data presented in Table 5-3 show limited effects on survival in mice and rats and some evidence of reduced body weight in rats following chronic exposures to concentrations of 1.5 mg/m 3 DPM or higher and exposure durations of 16 to 20 h/day, 5 days/week for 104 to5-97 130 weeks. Increased lung weights and lung to body-weight ratios in rats, mice, and hamsters; an increased heart to body weight ratio in rats; and decreased lung and kidney weights in cats have been reported following chronic exposure to DE. No evidence was found of an effect of DE on other body organs (Table 5-4). The lowest-observed-effect level in rats approximated 1 to 2 mg/m 3 DPM for 7 h/day, 5 days/week for 104 weeks.

5.6.2.2. Effects on Pulmonary Function Pulmonary function impairment has been reported in rats, hamsters, cats, and monkeys exposed to DE and included lung mechanical properties (compliance and resistance), diffusing capacity, lung volumes, and ventilatory performance (Table 5-5). The effects generally appeared only after prolonged exposures. The lowest exposure levels (expressed in terms of DPM concentrations) that resulted in impairment of pulmonary function occurred at 2 mg/m 3 in cynomolgus monkeys (the only level tested), 1.5 and 3.5 mg/m 3 in rats, 4.24 and 6 mg/m 3 in hamsters, and 11.7 mg/m 3 in cats. Exposures in monkeys, cats, and rats (3.5 mg/m 3 ) were for 7 to 8 h/day, 5 days/week for 104 to 130 weeks. While this duration is considered to constitute a lifetime study in rodents, it is a small part of the lifetime of a monkey or cat. Exposures in hamsters and rats (1.5 mg/m 3 ) varied in hours per day (8 to 20) and weeks of exposure (26 to 130). In all species but the monkey, the testing results were consistent with restrictive lung disease; alteration in expiratory flow rates indicated that 1.5 mg/m 3 DPM was a LOAEL for a chronic exposure (Gross, 1981). Monkeys demonstrated evidence of obstructive airway disease. The nature of the pulmonary impairment is dependent on the dose of toxicants delivered to and retained in the lung, the site of deposition and effective clearance or repair, and the anatomy and physiology of the affected species; these variables appear to be factors in the disparity of the airway disease in monkey versus the other species tested.

5.6.2.3. Histopathological and Histochemical Effects Histological studies have demonstrated that chronic exposure to DE can result in effects on respiratory tract tissue (Table 5-6). Typical findings include alveolar histiocytosis, AM aggregation, tissue inflammation, increase in PMNs, hyperplasia of bronchiolar and alveolar Type II cells, thickened alveolar septa, edema, fibrosis, and emphysema. Lesions in the trachea and bronchi were observed in some studies. Associated with these histopathological findings were various histochemical changes in the lung, including increases in lung DNA, total protein, alkaline and acid phosphatase, glucose-6-phosphate dehydrogenase; increased synthesis of collagen; and release of inflammatory mediators such as leukotriene LTB and prostaglandin
PGF 2 " . Although the overall laboratory evidence is that prolonged exposure to DPM results in histopathological and histochemical changes in the lungs of exposed animals, some studies have also demonstrated that there may be a threshold of exposure to DPM below which pathologic5-98 changes do not occur. These no-observed-adverse-effect levels for histopathological effects were reported to be 2 mg/m 3 for cynomolgus monkeys (the only concentration tested), 0.11 to 0.35 mg/m 3 for rats, and 0.25 mg/m 3 DPM for guinea pigs exposed for 7 to 20 h/day, 5 to 5.5 days/week for 104 to 130 weeks.

5.6.2.4. Effects on Airway Clearance The pathological effects of DPM appear to be strongly dependent on the relative rates of pulmonary deposition and clearance (Table 5-7). Clearance of particles from the alveolar region of the lungs is a multiphasic process involving phagocytosis by AMs. Chronic exposure to DPM concentrations of about 1 mg/m 3 or above, under varying exposure durations, causes pulmonary clearance to be reduced, with concomitant focal aggregations of particle-laden AMs, particularly in the peribronchiolar and alveolar regions, as well as in the hilar and mediastinal lymph nodes. The exposure concentration at which focal aggregates of particle-laden AMs occur may vary from species to species, depending on rate of uptake and pulmonary deposition, pulmonary clearance rates, the relative size of the AM population per unit of lung tissue, the rate of recruitment of AMs and leukocytes, and the relative efficiencies for removal of particles by the mucociliary and lymphatic transport system. The principal means by which PM clearance is reduced is through a decrease in the function of pulmonary AMs. Impairment of particle clearance seems to be nonspecific and applies primarily to dusts that are persistently retained in the lungs. Lung dust levels of approximately 0.1 to 1 mg/g lung tissue appear to produce this effect in the Fischer 344 rat (Health Effects Institute, 1995). Morrow (1988) suggested that the inability of particle-laden AMs to translocate to the mucociliary escalator is correlated to an average composite particle volume per AM in the lung. When this particle volume exceeds approximately 60 

### : m 3 per AM in the Fischer 344 rat, impairment of clearance appears to be initiated. When the particulate volume exceeds approximately 600 

### : m 3 per cell, evidence suggests that AM-mediated particulate clearance virtually ceases, agglomerated particle-laden macrophages remain in the alveolar region, and increasingly nonphagocytized dust particles translocate to the pulmonary interstitium. Data for other laboratory animal species and humans are, unfortunately, limited.

5.6.2.5. Neurological and Behavioral Effects Behavioral effects have been observed in rats exposed to DE from birth to 28 days of age (Table 5-14). Exposure caused a decreased level of spontaneous locomotor activity and a detrimental effect on learning in adulthood. In agreement with the behavioral changes was physiological evidence for delayed neuronal maturation. Exposures were to 6 mg/m 3 DPM for 8 h/day, 7 days/week from birth to about 7, 14, 21, or 28 days of age.5-99 5.6.2.6. Effects on Immunity and Allergenicity Several laboratory animal studies have indicated that exposure to DPM can reduce an animal’s resistance to respiratory infection. This effect, which can occur even after only 2 or 6 h of exposure to DE containing 5 to 8 mg/m 3 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 exposed mice. As with humans, there are animal data suggesting that DPM is a possible factor in the increasing incidence of allergic hypersensitivity. The effects have been demonstrated primarily in acute human and laboratory animal studies and appear to be associated with both the nonextractable carbon core and the organic fraction of DPM. It also appears that synergies with DPM may increase the potency of known airborne allergens. Both animal and human cell culture studies indicate that DPM also has the potential to act as an adjuvant.

5.6.2.7. Other Noncancer Effects Essentially no effects (based on the weight of evidence of a number of studies) were noted for reproductive and teratogenic effects in mice, rats, rabbits, and monkeys; clinical chemistry and hematology in the rat, cat, hamster, and monkeys; and enzyme induction in the rat and mouse (Tables 5-11 through 5-13 and 5-15).

5.6.3. Comparison of Filtered and Unfiltered Diesel Exhaust The comparison of the toxic responses in laboratory animals exposed to whole DE or filtered exhaust containing no particles demonstrates across laboratories that diesel particles are the principal etiologic agent of noncancerous health effects in laboratory animals exposed to DE (Table 5-16). Whether the particles act additively or synergistically with the gases cannot be determined from the designs of the studies. Under equivalent exposure regimens, hamsters have lower levels of retained DPM in their lungs than rats and mice do and consequently less pulmonary function impairment and pulmonary pathology. These differences may result from a lower intake rate of DPM, lower deposition rate and/or more rapid clearance rate, or lung tissue that is less susceptible to the cytotoxicity of DPM. Observations of a decreased respiration in hamsters when exposed by inhalation favor lower intake and deposition rates.

5.6.4. Interactive Effects of Diesel Exhaust There is no direct evidence that DE interacts with other substances in an exposure environment, other than an impaired resistance to respiratory tract infections. Young animals were not more susceptible. In several ways, animals with laboratory-induced emphysema were more resistant. There is experimental evidence that both inorganic and organic compounds can5-100 be adsorbed onto carbonaceous particles. When such substances become affiliated with particles, these substances can be carried deeper into the lungs where they might have a more direct and potent effect on epithelial cells or on AM ingesting the particles. Few specific studies to test interactive effects of DE with atmospheric contaminants, other than coal dust, have been conducted. Coal dust and DPM had an additive effect only.

5.6.5. Conclusions Conclusions concerning the principal human hazard from exposure to DE are as follows:

• Allergenic inflammatory disorders of the airways to responses typical of asthma have been demonstrated under short-term exposure scenarios to either DE or DPM. The evidence indicates that the immunological changes appear to be due to the DPM component of DE and that the immunological changes are caused by both the nonextractable carbon core and the adsorbed organic fraction of the diesel particle. The toxicological significance of these effects has yet to be
resolved. • Some occupational studies of acute exposure to DE during work shifts suggest that increased acute sensory and respiratory symptoms (cough, phlegm, chest tightness, wheezing) are more sensitive indicators of possible health risks from exposure to DE than pulmonary function decrements (which were consistently found not to be significantly associated with DE exposure) • Noncancer effects in humans from long-term chronic exposure to DPM are not evident. Noncancer effects from long-term exposure to DPM of several laboratory animal species, conducted to assess the pathophysiologic effects of DPM in humans showed pulmonary histopathology (principally fibrosis) and
chronic inflammation.

Although the mode of action of DE is not clearly evident for any of the effects documented in this chapter, the respiratory tract effects observed under acute scenarios are suggestive of an irritant mechanism, while lung effects observed in chronic scenarios indicate an underlying inflammatory response. Current knowledge indicates that the carbonaceous core of the diesel particle is the causative agent of the lung effects, with the extent of the injury being mediated at least in part by a progressive impairment of AMs. It is noted that lung effects occur in response to DE exposure in several species and occur in rats at doses lower than those inducing particle overload and a tumorigenic response (see above); it follows that lung effects such as inflammation and fibrosis are relevant in the development of risk assessments for DE.5-101 REFERENCES FOR CHAPTER 5

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5-1166-1 6. ESTIMATING HUMAN
NONCANCER HEALTH RISKS OF DIESEL EXHAUST

6.1. INTRODUCTION As discussed earlier in this document (C hapter 2, Section 2.2.7, 2.2.8), diesel engine exhaust (DE) consists of a complex mixture of gaseous pollutants and particles. In attempting to estimate potential health risks associated with human exposure to DE, researchers have focused attention mostly on the particul ate matter (PM) components. Th ey have done so, in part, by comparing the relative toxicity of unfiltered versus filtered DE (with gaseous components removed), as discussed in Chapter 5. Diesel particulate matter (DPM) consists mainly of: (a) el emental carbon (E C) particles having relatively large surface areas, (b) sol uble organic carbon, including 5-ring or higher polycyclic aromatic hydrocarbons (PAHs) such as benzo( a )pyrene, and other 3- or 4-ring compounds distributed between gas and particle phases, and (c) meta llic compounds. DPM also typically contains small amounts of sulfate/sulf uric acid and nitrates, trace elements, and water, plus some unidentified components. DPM is made up almost entirel y of fine par ticles (i.e., all
below 1–3 : m) with a significant subset of ultrafine particles (i.e., those with a mass median
diameter below about 0.1 : m). Health concerns have long focused on DPM. Toxicological data described in Chapter 5 (Section 5.2) indicate DPM to be the prime etiolo gic agent of noncancer health effects when DE is sufficiently diluted to limit the concentrations of gaseous irritants (NO 2 and SO 2 ), irritant vapors (aldehydes), CO, or other systemic toxicants. The large surface areas of DPM allow for adsorption of organics from the diesel combus tion process and for adsorption of additional compounds during transport in ambient air. The sm all size of DPM, combined with their large surface area, likely enhance the potential for subcellular interactions with important cellular components of respiratory tissues once the par ticles are inhaled by humans or other species (Johnston et al., 2000; Ob erdörster et al., 2000).
The content of DPM as described above a nd in Chapter 2 is of clear toxicological significance. The experimental evidence described in Chapter 5 concerning DPM’s association with and etiology of noncancer effects is extensive and compelling. These points, along with the fact that DPM is easily and most frequently m easured and reported in toxicological studies of diesel emissions, make DPM a reasonable choice as a measure of diesel emissions. As a surrogate, DPM is as valid as any other compone nt of DE to show what is currently known—and probably what is not yet known—about diesel e missions. Therefore, DPM is the quantitative focus of this chapter.1 A new PM CD is now being prepared to reflect the latest scientific studies on ambient PM available since the last document was completed.

6-2 The usual agency approach to evaluating non cancer risks from inhaled exposures to toxic air pollutants such as ambient DE has been doc umented by EPA in the methods for derivation of
an inhalation reference concentr ation (RfC) (U.S. EPA, 1994). For DPM exposures, this means combining key elements derived from evaluations of specific DPM noncancer effects in animals and humans (described in Chapter 5) with the use of quantitative dosimetry models (described in Chapter 3). The goal is to estimate DPM concentrations to which humans might be exposed throughout their lives (i.e., chronically) without experiencing any untowar d or adverse effects. Such an effort can be accomplished through anal ysis of dose-response relationships where the adverse response is considered as a function of a corresponding measure of dose. Chapter 5 is replete with dose-response information on adve rse (but nonlethal) noncancer health effects observed in long-term (chronic/lifetime) exposure studies to DE in general and to DPM in particular, albeit mostly in anima ls. Chapter 3 analyzes availa ble methods to convert external exposure concentrations of DPM in animal studies to estimates of a human-equivalent concentration (HEC). The following sections of this chapter (Secti ons 6.2, 6.3, and 6.5) assess and integrate this information to derive a chronic RfC, using the above-cited methodology in developing dose-response asse ssments of the noncancer effects of toxic air pollutants. Yet another approach to consider in deriving quantitative estimates of potential human health risks associated with ambient (nonoccupa tional) DPM exposures is the extent to which
DPM could contribute to the advers e health effects that have been associated with exposure to ambient fine PM, PM 2.5 . Such associations with adverse health effects are based primarily on epidemiologic studies evaluated in EPA’s Air Quality Criteria Document for Particulate Matter (PM CD) (U.S. EPA, 1996a). 1 This PM CD served as the scie ntific basis for the last periodic review of the national ambient air quality st andards (NAAQS) for PM, which resulted in the
establishment of revised PM standa rds in 1997, including standards for PM 2.5 . DPM is a component of ambient fine PM (see Chapter 2) and should be considered as a toxicologically important component of ambient fine PM. Any guidelines established for DPM, then, should be
concordant with information on fi ne PM in general, as presente d in the PM CD. To more fully consider the implications of the relations hip between ambient DPM and fine PM, the epidemiological evidence on fine PM and the basis for the PM 2.5 standards are summarized, and the relationship between ambient DPM and fine PM is discussed later in this chapter (Section 6.4). This relationship is of inte rest with respect to the noncancer assessment of DE. As is noted here , however, and reflected in Sections 6.2–6.4 be low, the definitions, procedures, and statutory mandates that apply to criteria pollutants such as PM (regulated through the establishment of6-3 NAAQS under sections 108 and 109 of the Clean Ai r Act) are fundament ally different from those that apply to toxic air pollutants such as DE and to the derivation of RfCs for such pollutants. Thus, the ambient PM 2.5 concentrations that are specified as the levels of the PM 2.5 NAAQS should not be compared directly with an y RfC that may be derived for DPM. It is reasonable to observe, however, that the annual PM 2.5 standard would be expected to provide a measure of protection from DP M, reflecting DPM’s current approximate pr oportion to PM 2.5 . Estimates of DE levels associ ated with effects occurring under less than lifetime exposure scenarios (such as acute exposure) are not addresse d in this chapter. Studies of acute exposure to DE are discussed in Chapter 5, but are accomp anied by scant dose-response information, with single-exposure studies for vari ous specialized endpoints (e.g., a llergenicity/adjuva ncy) and other multiple-exposure-level studies reporting data on mortality only. Based on currently available methodologies, these studies do not yet appear to provide a sufficient basis from which to derive a dose-response assessment for an acute DE exposure scenario.

6.2. THE INHALATION REFERE NCE CONCENTRAT ION APPROACH Historically, approaches such as the Accep table Daily Intake (ADI) were developed whereby effect levels, such as no-observed-adverse-effect leve ls (NOAELs) or lowest-observedadverse-effect levels (LOAELs) from human or animal data, were combined with certain “safety
factors” to accommodate areas of uncertainty to make quantitative es timates of a safe dose, i.e., a
level at which no adverse effect would be likely to occur. In response to the National Academy of Sciences (NAS) report entitled “Risk Assessment in the Federal Government: Managing the Process” (National Research Council, 1983), EPA developed two approaches similar to the ADI, i.e., the oral reference dose (RfD) (Barnes and Dourson, 1988) and th e parallel inhalation reference concentration, the RfC, with its form al methodology (U.S. EPA, 1994). Similar to the ADI in intent, the RfD/C approach is used for dose-response assessment of noncancer effects, using an explicitly delineated, rigorous methodology th at adheres to the principles set forth in the 1983 NRC report. The RfC methodology incl udes comprehensive guidance on a number of complex issues, including consistent application to effect levels of un certainty factors (UFs) rather than the ADI safety factors for consider ation of uncertainty. Ba sically, these approaches attempt to estimate a likely subthreshold concen tration in the human population. Use of the RfD/C approach is one of the principal curre nt agency methods for deriving dose-response

assessments. A chronic RfC is currently defined as:2 BMCL
x is defined as the lower 95% confidence limit of the dose that will result in a level of “x” response (e.g., BMCL 10 is the lower 95% confidence limit of a dose for a 10% increase in a particular response). See Appendix B for further specifics.

6-4 

### RfC = NOAEL UF HEC (6-1) An estimate (with uncertainty spanning perh aps an order of magnitude) of a continuous inhalation exposure to the human population (inc luding sensitive subgr oups) that is likely to be without an appreciable risk of de leterious noncancer eff ects during a lifetime.

The RfC approach involves the following general steps:

• Identification of a critical eff ect relevant to humans, i.e., an adverse effect that occurs at the lowest exposure/dose in human or animal studies and whose prevention avoids the occurrence of all other adverse effects; • Selection of appropriate dose -response data to derive a poi nt of departure (POD) for extrapolation of a key study (or studi es) that provides a NOAEL, LOAEL, or benchmark concentration (BMCL

x ) 2 ; • Estimation of HECs when animal exposur e-response data are used (via use of PBPK/dosimetry models); • Application of UFs to the point of departure (e.g., NOAEL, LOAEL, BMCL

x ) to address extrapolation uncertainties (e.g., interindividual variability, interspecies differences, adequacy of database); and
• Characterization of the “confidence” in the dose-response asse ssment and resultant
RfC.

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 as an HEC, and UFs. The units of an RfC are typically mg/m 3 or : g/m 3

. Alternatively, the numerator in E quation 6-1 may be a LOAEL or BMCL

x . The

benchmark concentration (BMC) a pproach and its application in this assessment are documented in Appendix B and described furthe r below. Also, a modifying fact or (MF) may be used in the denominator of this equation to account for scientific uncertaintie s, usually relating to the study chosen as the basis for the RfC. Further specifi cs of RfC derivation pro cedures are discussed as6-5 they are used in the following sections. All su ch procedures are descri bed in detail in the RfC Methodology (U.S. EPA, 1994).

6.3. CHRONIC REFERENCE CONCE NTRATION FOR DIESEL EXHAUST As concluded in Chapter 5, chronic respirat ory effects are the principal noncancer hazard to humans from long-term environmental exposure to DE. Other effects (e.g., neurological, liver-related) are observed in animal studies at higher exposures than those producing the respiratory effects. The human and animal da ta for the immunological effects of DE are currently considered inadequate for dose-response evaluation. T hus, the respiratory effects are
considered the “critical effect” for the derivation of a chronic RfC for DE. The evidence for chronic respiratory effects is based mainly on animal studies showing consistent findings of inflamma tory, histopathological (includi ng fibrosis), and functional changes in the pulmonary and tracheobronchial re gions of laboratory animals, including the rat, mouse, hamster, guinea pig, and monkey. Occupational studies of DE provide some corroborative evidence of possible respiratory e ffects (e.g., respiratory symptoms and possible lung function changes), although those studies ar e generally deficient in exposure information.
Mode-of-action information about respiratory effects from DE exposure indicates that, at least in rats, the pathogenic sequence followi ng the inhalation of DPM begins with the phagocytosis of diesel particle s by alveolar macrophages (AMs). These activated AMs release chemotactic factors that attract neutrophils a nd additional AMs. As the lung burden of DPM increases, there are aggregations of particle-laden AMs in alveoli adjacent to terminal 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 mediat ors of inflammation and oxygen radicals, and particle-laden macrophages are f unctionally altered, re sulting in decreased viability and impaired phagocytosis and clearance of particles. This series of events may result in pulmonary inflammation, fibrosis, and eventu ally lesions like those describe d in the studies reviewed in Chapter 5. Although information describing the possible pathogenesis of respiratory effects in humans is not available, the effects reported in studies of humans exposed to DE are not inconsistent with the findings in controlled laboratory animal studies. Several reasons explain why the dose-response data from rats are considered especially appropriate for use in characterizing noncancer h ealth effects in humans and deriving a chronic RfC for DE. First, similar noncancer respirator y effects are seen in other species (mouse, hamster, guinea pig, and monkey). Second, ra ts and humans exhibit similar noncancer responses (macrophage response and interstitial fibrosis) to other particles su ch as coal mine dust, silica, and beryllium (Haschek and Witschi, 1991; Ober dörster, 1994). Third, relative to other species6-6 there exists a plethora of long-term, specialize d, and mechanistic studies in rats. Fourth, an expert panel convened by the In ternational Life Sciences In stitute (ILSI) recommended that response data on persistent, inflammatory processes may be used to assess nonneoplastic responses of poorly soluble particles (PSP) such as DPM (ILSI, 2000).

6.3.1. Principal Studies for Dose-Response An alysis: Chronic, Multiple-Dose Level
Rat Studies The experimental protocols a nd results from the long-term, repeated-exposure chronic studies demonstrating and characterizing the critical effects of pulmonary fibrotic changes and
inflammation are discussed in Ch apter 5. Salient points of thes e studies, including species/sex of the test species, the exposure regime and concentrations re ported in mg DPM/m 3 , and effect levels, are abstracted in Table 6- 1 for further consideration. The effect levels are designated as
N for no-observed-adverse-e ffect level, A for advers e-effect level, and BMCL 10 . The purpose of many of the chronic studies lis ted in this table was not the elucidation of the concentration-response charact er of DPM. The studies of Heinrich et al. (1982, 1986) in hamsters, mice, and rats; of Iwai et al. (1986) in rats; of Lewi s et al. (1989) in monkeys; and of Pepelko (1982a) in rats ar e all single-dose-level analyses that have as their genesis mechanistic or species-comparative purposes. As discussed in Chapter 5, many of these studies do provide valuable supporting information for designa tion of the critical effect of pulmonary histopathology. The lack of any clear dose-response data, how ever, precludes consideration of
these studies as a basis for RfC derivation. Likewise, studies of chronic, multiple-level exposure involving speci es other than rats, i.e., hamsters (Pepelko, 1982b), cat s (Plopper et al., 1983), and guin ea pigs (Barnhart et al., 1981, 1982), provide cross-species corr oboration of the critical effect s of pulmonary histopathology and inflammatory alteration. The remaining studies showing exposure-res ponse relationships in rats for the critical
effects include those of Ishini shi et al. (1986, 1988), Mauderly et al. (1987a), Heinrich et al. (1995), and Nikula et al. (1995). As described in Chapter 5, all of these studies were conducted and reported in a thorough, exhaustive manner on the critical effects and little, if any, basis exists for choosing one over another for purposes of Rf C derivation. One way of taking advantage of this high degree of methodological and scientific mer it would be to array data from all these
studies and their effect le vels (NOAEL, LOAEL, BMCL

x ) subsequent to normalization of the exposure conditions, i.e., conversion of the exposur e regimes to yield an HEC. This exercise would result in an interstudy concentrati on-response continuum normalized to a continuous human exposure to DPM that would facilitate the choice of a concentration to use as a point of departure in deriving an RfC.6-7 Table 6-1. Histopathological effects of diesel exhaust in the lungs of laboratory animals

Study Species/sex Exposure period Particles (mg/m 3 ) Effect level a Effects b
Lewis et al. (1989) Monkey, Cynomolgus, M 7 h/day 5 days/wk 104 wks 2.0 N AM aggregation; no fibrosis, inflammation, or emphysema

Bhatnagar et al. (1980) Pepelko (1982a) Rat, F344, M, F 7 h/day 5 days/wk 104 wks 2.0 Multifocal histiocytosis; inflammatory changes; Type II cell proliferation; fibrosis
Pepelko (1982b) Hamster, Chinese, M 8 h/day 5 days/wk 26 wks 6.0 12.0 A Inflammatory changes; AM accumulation; thickened alveolar lining; Type II cell hyperplasia; edema; increase in collagen

Heinrich et al. (1982) Hamster, Syrian, M, F 7-8 h/day 5 days/wk 120 wks 3.9 A Inflammatory changes, 60% adenomatous cell proliferation

Iwai et al. (1986) Rat, F344, F 8 h/day 7 days/wk 104 wks 4.9 A Type II cell proliferation; inflammatory changes; bronchial hyperplasia; fibrosis
Mauderly et al. (1987a) Henderson et al. (1988) Rat, F344, M, F; Mouse, CD-1, M, F 7 h/day 5 days/wk 130 wks 0.35 3.5 7.1 N A A Alveolar and bronchiolar epithelial metaplasia in rats at 3.5 and 7.0 mg/m 3 ; fibrosis at 7.0 mg/m 3 in rats and mice; inflammatory changes; few quantitative data given

Heinrich et al. (1995) Rat, Wistar, F; Mouse, NMRI, F (7 mg/m 3 only) 18 h/day 5 days/wk 24 mo 0.8 2.5 7.0 A A A Bronchioalveolar hyperplasia, interstitial fibrosis in all groups; severity and incidence increase with exposure concentration; text given only

Mouse, NMRI, F; C57BL/6N, F 18 h/day 5 days/wk 13.5 mo (NMRI) 24 mo (C57BL/ N) 7.0 A No increase in tumors; noncancer effects not discussedTable 6-1. Histopathological effects of diesel exhaust in the lungs of laboratory animals (continued)

Study Species/sex Exposure period Particles (mg/m 3 ) Effect level a Effects b

6-8 Ishnishi et al. (1986, 1988) Rat, M, F, F344, /Jcl. 16 h/day 6 days/wk 130 wks 0.11 c 0.41 c 1.08 c 2.32 c

0.46 d 0.96 d 1.84 d N N A A

N A A Inflammatory changes; Type II cell hyperplasia and lung tumors seen at >0.4 mg/m 3 ; shortening and loss of cilia in trachea and bronchi; data given in text only

3.72 d A

Heinrich et al. (1986) Hamster, Syrian, M, F; Mouse, NMRI, F; Rat, Wistar, F 19 h/day 5 days/wk 120 wks 4.24 A Inflammatory changes; thickened alveolar septa; bronchioloalveolarhyperplasia; alveolar lipoproteinosis; emphysema (diagnostic methodology not described); hyperplasia; lung
tumors

Barnhart et al. (1981, 1982); Vostal et al. (1981) Guinea pig, Hartley, M 20 h/day 5.5 days/wk 104 wks 0.25 0.75 1.5 6.0 N A A A Minimal response at 0.25 and ultrastructural changes at 0.75 mg/m 3 ; thickened alveolar membranes; cell proliferation; fibrosis at 6.0 mg/m 3 ; increase in PMN at 0.75 mg/m 3 and 1.5 mg/m 3

Plopper et al. (1983) Hyde et al. (1985) Cat, inbred, M 8 h/day 7 days/wk 124 wks 6.0 c 12.0 d A A Inflammatory changes; AM aggregation; bronchiolar epithelial metaplasia; Type II cell hyperplasia; peribronchiolar fibrosis
Nikula et al. (1995) Rat, F344, M 16 h/day 5 days/wk 23 mo 2.44 6.33 A, A BMCL 10 AM hyperplasia, epithelial hyperplasia, inflammation, septal fibrosis, bronchoalveolar metaplasia

a N= no-observed-adverse-effect level; A = adverse-effect level; BMCL 10 = benchmark concentration, lower limit, at a 10% response level (for incidence); see Appendix A for further specifics.

b AM = Alveolar macrophage; PMN = Polymorphonuclear leukocyte

c Light-duty engine.

d Heavy-duty engine.

c 1 to 61 weeks exposure.

d 62 to 124 weeks of exposure.

e See Appendix A.6-9 6.3.2. Derivation of Human Continuous Equivalent Con centrations, HECs Pharmacokinetic, or PK, models can be used to estimate across species the external
concentrations of a toxicant that would result in equivalent internal doses. When used for these purposes, PK models may be term ed comparative dosimetric models. Chapter 3 reviewed and
evaluated a number of dosimetric models applicable to DPM. This analysis indicated that outputs from the human component of the model de veloped by Yu et al. (1991) specifically for DPM, such as deposition and estimated lung burden, were not substantially different from other available models. The analysis also demonstr ated that the Yu model accounted for several diesel-specific phenomena, incl uding particle overload lung clearance rates and interspecies kinetics of desorption of organics from the carbonaceous core of DPM, both slow- and fastcleared. Of importance, the Yu model was pa rameterized for deposition and clearance in both animals and humans. Also, the animal compone nt of the model was based on data from rats actually exposed to DPM, whereas other models analyzed used data based only on generic particles in the size range of DP M. It was concluded from this an alysis that the Yu model could be used to estimate disposition of DPM both in animals and in humans and would therefore be an acceptable choice in performing animal-to-human ex trapolation in deriving a continuous humanequivalent concentration. Note, however, that use of this or any other available PK model would address species differences in dose (i.e., pharmacokinetics, PK), and not necessarily pharmacodynamics (PD), the other co mponent of uncertainty in an imal-to-human or interspecies extrapolation (U.S. EPA, 1994). Guidance on choosing measures of exposure fo r poorly soluble particles such as DPM (ILSI, 2000) states that some measures of exte rnal dose (e.g., the aerosol exposure parameters of MMAD, F g , particle surface area, and density) s hould be characterized. Likewise, some indication of internal dose resulting from the ex ternal exposure (e.g., lung burden) should be measured so that differences in dose metrics may be considered as new mechanistic insights are developed. The whole particle, as characterized in this assessment and used in the model of Yu et al. (1991), meets this recommended guidance, and DPM, in : g/m 3 , is used as the measure of external exposure. Internal measures of expos ure or dose were also considered in Chapter 3 (Section 3.3.1.1) with the conclusion that the dose metric of lung burde n of DPM in terms of surface area (mg/cm 2 ) at the termination of the exposure pe riod appears to be the most defensible and appropriate measure of internal dose, especially where clearance is involved. More detailed specifics are available in Chapter 3 and in Appendix A. The logical and operational sequence of deri ving a HEC using the Yu model and these metrics, i.e. external air concentration (in : g/m 3 ) and lung burden (in mg/cm 2 ), is demonstrated in Figure 6-1. First, the experimental animal exposures, including external concentration and6-10 Figure 6-1. Flow diagram of pr ocedure for calculating HECs.

daily and weekly duration, are en tered into the animal component of the Yu model to estimate the animal lung burden, in mg DPM/cm 2 , for the specific exposure scenario. The human component of the Yu model is then used by setting desired exposure conditions (continuous for 70 years) and running the model to find an ex ternal exposure DPM con centration that would result in this same lung burden. The human ex ternal DPM concentration matching this lung burden is the human-equivalent c oncentration. The step-by-step specifics and results of this procedure as applied to the various studies in Table 6-1 are shown in Table A-4 and fully explained in Appendix A. The foregoing discussion does not address the variability in outcomes that may be estimated from the Yu et al. (1991) model fro m deposition of DPM. The model comparison exercises in Chapter 3 showed relatively minor differences among the va rious human models for one measure, deposition, and indicated that human lung burdens estimated by the human component of the Yu and ICRP66 models were n early identical at low- exposure concentrations. Variability in output of their model (lung burde n) was also examined by Yu and Yoon (1990), who studied dependency on tidal vol ume, respiration rate, and cl earance (in terms of the overall particle transpor t rate from the alveolar region, 8 A ). Analysis indicated that the model output is sensitive, but not overly so, for these determinative parameters. A ± 20% change in values for
8 A , for example, was estimated to result in a 16%–26% change in soot burden at a 0.1 mg/m 3 continuous diesel exposure for 10 years. For a ± 10% change in tidal volume, the model projected changes in soot burden ranging from 14% to 22% for this same exposure scenario. The fact that the changes in the model outcome were comparable to changes in the input parameters, such as tidal volume, indicates that the variability of the model when applied to the human population would reflect the variability of these physiological parameters across that population. In sum, at low concentrations of DPM (< 0.5 mg/m 3 ), relatively minor differences exist among the models currently available, and the input pa rameters in the human population may be a major source of variability. As discussed below, variability within the human population often is addressed by applying sa fety or uncertainty factors, usually in the range of 10 (Renwick and Lazarus, 1998; U.S. EPA, 1994).6-11 6.3.3. Dose-Response Analysis—C hoice of an Effect Level
HECs were obtained for the dose levels and ex posure scenarios presented in the studies of Mauderly et al. (1987b), Ishinishi et al. (1986, 1988), Nikula et al. (1995), and Heinrich et al. (1995), the specifics of which are presented in Appendix A, specifically Table A-4. The HECs, along with the corresponding speci fic lung burdens in terms of : g/cm 2 , were transcribed from Table A-4 and, along with the accompanying effect level (NOAEL, LOAEL or BMCL 10 ), are arrayed ordinally in Table 6-2. It is acknowledged that Table 6-2 is by no means a full portrayal of the dose-response relationship that may exist for DPM and health effects. As indicated by the BMCL 10 values listed for the Nikula et al. (1995) study in Table 6-2, the BMC analysis was carried out on the DPM data base and is documented in Appendix B. The chronic rat studies identified in this chapter were analyzed for information suitable for BMC analysis. Results yielded only a few datasets of pulmonary toxic ity data from a single study, that of Nikula et al. (1995), that could be used for BMC analysis. These pulmonary data (histopathology incidence data) were extracted, HEC concentrations were calculated using the model of Yu, and the BMCs we re generated. The results yi elded a complex array of BMCL 10 s from three different effects in two sexes (both se parate and combined) with nine different models
that were evaluated based on th e nature of the dataset, on th e goodness-of-fit pa rameters, and on visual inspection of the graphical outputs. Fr om among all the benchm ark data generated, the
BMCL 10 of 0.37 mg/m 3 calculated from combined male and female rat pulmona ry histopathology was judged as the most defensible choice. However, further charact erization of this same
benchmark value indicates that it is not a suitabl e candidate for use as a point of departure for development of a dose-response a ssessment such as the RfC. Limitations included the excessive extent of extrapolation from the observed exper imental range (see Figure B-1 in Appendix B) and the paucity of data points (there were only two exposure groups) overall. Another serious limitation is that the high experimental concentr ations used (and their C × t product) are well in the range where the problematic phenomenon of pulmonary overload in rats occurs (Section 5.1.3.3.4). Inspection of Table 6-2 shows that calculatin g and ordering the HECs created a partial concentration-response continuum reflected in the estimated intern al lung burden also given in
this table. The continuum extends from HECs w ith no observed adverse e ffects at concentrations as low as 0.032 mg/m 3 to as high as 0.144 mg/m 3 to HECs with an adverse effect level that first appears definitively in the continuum probably at 0.33 mg/m 3 and extends out to 1.95 mg/m 3

. It should be noted that the relationship betw een HEC and lung burden is not consistently proportional. For example, at the lowest HEC listed, 0.032 mg/m 3 , a lifetime (70 years) of continuous exposure to this concentration is estima ted to result in a specific burden to the lung of
0.0587 : g/cm 2 . At the other end of this spectrum, a lifetime of continuous exposure to 4.46-12 Table 6-2. Human equivalent continuous concentrat ions: 70-year HECs calculated with the model of Yu et al. (1991) from long-term studies of rats repeatedly exposed to DPM a

Study Exposure concentration (mg/m 3 ) Effect level a Lung burden (modeled) ( : g DPM /cm 2 ) b HEC (mg/m 3 )
Ishinishi et al. (1988) (LD c ) 0.11 NOAEL 0.0587 0.032
Mauderly et al. (1987a) 0.35 NOAEL 0.0685 0.038
Ishinishi et al. (1988) (LD c ) 0.41 NOAEL 0.245 0.128
Ishinishi et al. (1988) (HD c ) 0.46 NOAEL 0.281 0.144
Heinrich et al. (1995) 0.84 LOAEL 0.94 0.33
Nikula et al. (1995) 2.44 & 6.3 d BMCL 10 -inflam 1.34 0.37
Ishinishi et al. (1988) (HD c ) 0.96 LOAEL 3.16 0.883
Ishinishi et al. (1988) (LD c ) 1.18 LOAEL 4.50 1.25
Nikula et al. (1995) 2.44 & 6.3 d BMCL 10 - fibrosis 4.70 1.3
Mauderly et al. (1987a) 3.47 LOAEL 4.95 1.375
Nikula et al. (1995) 2.44 LOAEL 7.00 1.95
Ishinishi et al. (1988) (HD c ) 1.84 AEL 7.63 2.15
Heinrich et al. (1995) 2.5 AEL 8.40 2.35
Ishinishi et al. (1988) (LD c ) 2.32 AEL 9.75 2.75
Mauderly et al. (1987a) 7.08 AEL 10.9 3.05
Ishinishi et al. (1988) (HD c ) 3.72 AEL 15.8 4.4

a Effect levels are based on the critical effects of pulmonary hist opathology and inflammation as reported in the individual studies. NOAEL: no-observed-adverse-effect level; LOAEL: lowest-observed-adverse-effect level; AEL: adverse-effect level; BMCL 10 : lower 95% confidence estimate of the con centration of DPM associated with a 10% incidence of chronic pulmonary inflammation (inflam) or fibrosis (see Appendices A and B for more specifics).

b Lung burdens were derived from data generated from the animal portion of the Yu model using the concentration and duration scenario of each study. The human portion of the Yu model was then used to estimate the continuous, 70-year exposures that would result in this same lung burden, i.e., the HEC. See Table A-4 in Appendix A and accompanying text for further specifics on derivation.

c LD/HD = light-duty/heavy-duty diesel engine.

d These values are the actual exposure levels used in the Nikula study. These values were converted into HEC and entered into BMC equations to obtain the estimate of the BMCL 10 listed. The lung burdens for the two BMCL 10 s listed here were derived by interpolation.6-13 mg/m 3 is estimated to result in a specific lung burden of 15.8 : g/cm 2 . This latter lung burden is disproportionally elevated compared with the burden estimated to result from exposure to the lowest concentration. Applying the absolute ratio of lung burden/HEC at the lowest HEC exposure (i.e., 0.0587/ 0.032 = 1.8) to the highest concentration would result in a lower lung burden, 4.4 × 1.8 = 7.9 : g/cm 2 , which is much lower than the 15.8 : g/cm 2 indicated. This disproportionate increase in lung burden as a f unction of DPM concentra tion would be predicted from the assumption in the Yu model that th e overload phenomena occurs in humans, as is demonstrated in Figure 3-9 in Chapter 3. Insp ection of Table 6-2 shows that this disproportion between lung burden and HEC begi ns to be noticeable around 0.33 mg/m 3 , at the HEC derived from the Heinrich et al. (1995) study. HECs belo w this value are not appreciably influenced by the overload/disproportiona te lung burden phenomenon. Inspection of the combined interstudy dose-re sponse continuum in Table 6-2 to elucidate a point of departure for an RfC entails some in terpretation. Exposures at the lower end of this
table show that elevated chr onic exposures to DPM consistently result in AELs. Conversely, entries in the upper portion of this table show that low-level chronic exposures to DPM have minimal, if any, effects within the capability of these studies to detect them. Intermediate chronic exposures, from 0.128 mg/m 3 to 0.9 mg/m 3 , are, however, less clear and effect levels and exposures either have no or fe w observable effects, or effects that are minimally adverse. In choosing from among levels (e.g., NOAELs, LOAELs, BMCL

x s) as a POD for derivation of an RfC, the methodology (U.S. EPA, 1994) provi des guidance for choice of a highest no-effect level below an effect level; the interim guidance for the BMC suggests that for use as a point of departure, a benchmark (e.g., BMCL 10 ) should be within the range of the observable response
data so as to avoid excessive extrapolation, and take the shap e of the dose-response curve into consideration (Barnes et al., 1995; U.S. EPA, 1995). The highest no-effect HECs (NOAEL HEC ) in this table are 0.128 mg/m 3 and 0.144 mg/m 3 from the Ishinishi et al. (1988) study, nearly
fivefold above other no-eff ect levels of 0.032 and 0.038 mg/m 3 . The lower BMCL 10 (0.37 mg/m 3 ) is at nearly the same concentra tion as the lowest LOAEL of 0.33 mg/m 3 and thus may be too high an estimate for use as a POD based on these data. As disc ussed above, the limitations
on this BMCL 10 , including excessive extrapolation out of the observable range (see Appendix B for more specifics), make it a less than optimal candidate for consideration as a POD in the development of dose-response asse ssments and therefore was not used for this purpose in this assessment. However, this BMCL 10 (i.e., at a response rate of 0.1 or 10%) was generated directly from a modeled dose-response curv e for chronic inflammation and lends credence to the other NOAELs in Table 6-2 as being associated with their respective dose-res ponse curve at incidences of considerably less than 10%. More over, the HECs of less than 0.33 mg/m 3 are not appreciably influenced by the overload phenomenon (see above). Based on this analysis, the value of 0.1446-14 mg/m 3 is chosen as the POD for development of the RfC, because it is the highest NOAEL HEC among those available.

6.3.4. Uncertainty Factors (UF) for the RfC—A Composite Factor of 30 Areas of uncertainty designated in the RfC that are relevant to the DPM assessment are interindividual variability and anima l-to-human extrapolation. Each shall be addr essed in this
section. Considerable qualitative but little, if any, quantitative information exists regarding subgroups that could be sensitive to any respirator y tract effects of DPM. It is acknowledged that exposure to DPM could be additive to many other daily or lifetime exposures to airborne organic compounds and nondiesel ambient PM. It is also likely that individuals who predispose their lungs to increased particle rete ntion through smoking or other hi gh particulate burdens, who have existing respiratory tract inflammation or infections, or who have chronic bronchitis, asthma, or fibrosis could be more susceptible to a dverse impacts from DPM ex posure (U.S. EPA, 1996a, Chapter 5 of this document). Also, infants and children could have a greater susceptibility to the acute/chronic toxicity of DPM because of th eir greater breathing frequency and consequent potential for greater particle deposition in the respiratory tract, whic h has not reached full development. Increased respiratory symptoms and decreased lung function in children versus ambient PM levels, of which DPM is a part, ha ve been observed (U.S. EPA, 1996a). Thus, even though the limited evidence currently available (s ee Chapter 5) produces no clear evidence that children are especially sensitive to effects from breathing DPM, the possibility that they actually may be more susceptible because of their i nherent physiology and anatomy should remain a consideration. Likewise, a num ber of factors may modify nor mal lung clearance, including, aging, gender, and disease. It should be noted th at the results of Mauderl y et al. (1989) discussed in Chapter 5 indicated that rats with diseased lungs (emphysematous) were no more susceptible than rats with normal lungs to the effects of DE exposure. Although the exact role of these factors is not resolved, all would influence the particle dose to the lung tissue from inhalation exposure. Activity patterns related to occupation and habitation in the proximity of major roadways are certain to be contributory for so me subgroups in receiving higher DPM exposures (Chapter 2). In the absence of DE-specific data, this assessmen t relies on a default UF value of 10 to account for possible interindividual human variability (U.S. EPA, 1994; Renwick and Lazarus, 1998). Application of an animal-to-hum an extrapolation or interspeci es uncertainty factor to an assessment may be modified via a number of circumstances. Wh en the assessment is based on human data, no such UF is necessary. When th e assessment is based on animal data, as is the case with DPM, a default UF of 10 typically is applied to the animal effect level. This latter6-15 action implies that the effect observed in the animal study would occur in humans at a 10-fold lower concentration, ostensibly from so me combination of pharmacokinetic and pharmacodynamic factors that woul d reflect greater dose (PK consideration) to the human target or greater sensitivity (PK consideration) of the human tissue.
The circumstances with DPM warrant modifica tion away from application of the default UF for animal-to-human extrapolation. The firs t circumstance is the extensive effort in this assessment to address the pharmac okinetic component of the UF. The point of employing stateof-the-art lung dosimetry models with speci fic parameterization for DPM in conversion of animal exposures to human-equivalent exposures is to derive an estimate of interspecies pharmacokinetics; to know this asp ect of interspecies difference with some degree of certainty. Having made this informed e ffort addresses a major portion of the PK component. It is acknowledged, however, that uncer tainties about the model employ ed here (or any other model) persist. Although the model comparison s hown in Chapter 3 indicates relatively minor variability in output among the various human models examined (see Table 3-3 and Figure 3-9) other sources of uncertainty and variability remain. These include, but are not limited to, matters
such as the estimates if the m odel were applied to the genera l population or variability from the animal portion of the model(s). A second ci rcumstance involves the pharmacodynamic or PD component of the interspecies UF, especially th e aspect as to whether the experimental animal species used in the assessment is more or less sensitive than humans. In the consensus report of ILSI (2000) a specific recommendation is made c oncerning the PD aspect of the interspecies uncertainty factor for poorly so luble particles such as DPM. Because the pulmonary responses from DPM in the principal experimental species , the rat, are present under exposure conditions that do not appear to elicit any response in human s, the experimental species is considered more sensitive than humans. Accordingly, the repor t suggested that no accommodation be made for uncertainty concerning the pha rmacodynamic component of the interspecies UF for DPM and presumably for any other PSP, as the rat appeared to be a sensitive species, more so even than the human. However, other information currently available on DPM suggests that, at least with regard to inflammatory effects, humans may indeed be as sensitive or even more so than rats.
Section 5.1.1.1.3 discusses se veral studies where humans were exposed to airborne DPM and either precursors (Salvi et al ., 2000; Nordenhall et al., 2000) or markers (Nightingale et al., 2000; Salvi et al.,1999) of inflammati on were detected. These indi cators of inflammation were in response to DPM levels of only 200–300 : g/m 3 of 1–2h duration. Note that in Table 6-2,
NOAEL concentrations to which rats actually were exposed were only 100–400 : g/m 3 , clearly within the range of the aforementioned human exposure levels. Thus, adverse effects (inflammation) have been shown to occur in human s at equivalent or po ssibly even lower levels6-16 of DPM than observed in rats, indicating that humans may indeed be at least as sensitive if not
more so than rats.
The sum of these considerations on the an imal-to-human UF is that, although major portions of uncertainty have been addressed, degrees of un certainty persist in both the pharmacodynamic and pharm acokinetic components of the fact or. In considering both this residual uncertainty and the in formation discussed above, it w ould be prudent to acknowledge partial degrees of uncertainty in both these ar eas with a partial uncertainty factor, i.e., 10 0.5 vice
10 1 , such that a factor of 3 would be applied for interspecies extrapolation. In summary, the application of UFs for the two areas discussed a bove, interhuman and animal to human, would result in a composite uncer tainty factor of 30, 10 for interhuman × 3 for animal to human. Use of other UFs, as discu ssed in the RfC methodology (U.S. EPA, 1994) for
deficiencies in database or fo r duration extrapolation, is not c onsidered necessary. It should be noted that, given the emerging research on DE-induced immunological effects, it may be necessary at a later date to reconsider the basi s for selection of the critical effect and UFs and
thus the entire derivation of the DE RfC.

6.3.5. Derivation of the RfC for Diesel Exhaust On the basis of the above analysis, the value of 0.144 mg/m 3 DPM was selected as the point of departure for the RfC evaluation. This value was derived from concentrations in rat
chronic studies that were mode led to obtain HECs. The pulmona ry effects, histopathology and inflammation, were dete rmined to be the critical noncan cer effects. Response data on inflammation also were suggested by a specifi c scientific working group as a satisfactory surrogate for fibrogenic responses in assessi ng the pulmonary responses of poorly soluble particles such as DPM (ILSI, 2000). Suffici ent documentation from other studies showed no effect in the portal-of-entry tissues, the extrat horacic (nasopharyngeal) region of the respiratory system, or in other organs at the lowest le vels 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/m 3 16 hr/day, 6 days/wk, a NOAEL, yielded a NOAEL HEC of 0.144 mg/m 3 . Application of the composite UF yields the RfC:



### NOAEL HEC 

### ÷

÷= = UF = RfC 0.144 mg/m mg/m g/m 3 33 

### 30 0 0048 5

. . μ6-17 6.4. EPIDEMIOLOGICAL EVIDENCE AND NAAQS FOR FINE PM Historically, EPA has estab lished primary NAAQS to protect sensitive human population groups against adverse health eff ects associated with ambient expos ures to certain widespread air pollutants, including PM, ozone (O 3 ), carbon monoxide (CO), sulfur dioxide (SO 2 ), nitrogen dioxide (NO 2 ), and lead (Pb). The U.S. Clean Air Ac t (the Act) requires that EPA periodically review and revise as appropriate the criteria (sci entific bases) and standards for each pollutant or class of pollutants (e.g., PM) for which NAAQS ha ve been established. The primary, healthbased NAAQS must be based on the la test scientific info rmation useful in indicating the kind and extent of all effects on public health expected from the presence of the pollutant in the ambient air, which is evaluated in a “C riteria Document” (CD). The NAAQS are then set at levels that, in the judgment of the EPA Administrator, protect pub lic health (as contrasted with the health of any individual) with an adequate margin of safety. In deter mining the degree of protection that will satisfy this mandate, EPA considers the nature and severity of the effects, the types of health evidence available, the kind and degree of scientific uncertainty that effects would in fact occur at any particular level of pollution, and the size and nature of sensitive populations at risk of experiencing exposures of concer n. The EPA develops a staff pa per to bridge the gap between the scientific criteria and the public health policy considerations the Administrator must take into account in reaching a final judgment. The EPA al so must consider the recommendations of the Clean Air Scientific Advisory Committee (CASAC), an independent committee established by the Act specifically to advise the Administrator on air quality criteria and NAAQS. In contrast to an RfC, the NAAQS are not intended to identify a concentration that is protective against a hypothetical continuous lifetime exposure to a give n level, but rather ta ke into account expected actual exposure conditions of U.S. populations. The original PM NAAQS were set in 1971 in te rms of total suspende d particulate matter (TSP) and included both inhala ble and noninhalable pa rticles, ranging in size up to 25–50 : m. A later periodic review of the PM criteri a and NAAQS led to the setting in 1987 of PM 10 NAAQS (150 : g/m 3 , 24-h average; 50 : g/m 3 , annual average) aimed at protecting against health effects associated with those inhalable particles capable of penetrating to lower (thoracic) regions of the human respiratory tract and depositing in tracheobronchial and alveolar tissue of the lung ( # 10.0 : m) (52 FR 24634, July 1, 1987). The most recently completed PM NAAQS review was
based on an assessment of the latest available scientific information characterized in the EPA PM CD (U.S. EPA, 1996a) and additiona l staff assessments contained in an associated PM Staff Paper (U.S. EPA, 1996b). In 1997, on the basis of this information and taking into account CASAC recommendations and extensive pub lic comments, EPA established new PM 2.5 NAAQS (15 : g/m 3 , annual average; 65 : g/m 3 , 24-h average) to protect against adverse health effects associated with exposures to fine PM. At the same time, EPA retained, in modified form, the3 At present, the 1997 PM 2.5 standards are the subject of ongoing litigation, although they legally remain in effect, as do the 1987 PM 10 standards.

6-18 PM 10 NAAQS originally set in 1987 to protect agains t effects associated w ith coarse fraction PM (62 FR 38652, July 18, 1997). 3 The 1997 PM NAAQS decisions were based, in part, on important distinctions already highlighted by information present in the PM CD between the fine and coarse fractions of PM 10 with regard to size, chemical composition, sources , and transport. Also of key importance were the assessment and interpretation of new epidemiological findings on health effects associated with ambient PM. The epidemiological evid ence and basis for the NAAQS for fine PM are summarized below, followed by a discussion of the relevance of this information for noncancer
assessment of DE.

6.4.1. Epidemiological Evidence for Fine PM The PM CD (U.S. EPA, 1996a) and Staff Pape r (U.S. EPA, 1996b) highlighted more than 80 newly published community epidemiologic studi es, of which more than 60 found significant associations between increased mortality and/or morbidity risks and various ambient PM indicators. The main findings of concern we re community epidemiology results showing ambient PM exposures to be statistically associated with increased mortality (especially among people over 65 years of age and those with preexisting cardiopulmonary conditions) and morbidity (indexed by increased hospital admissions, respiratory symptom rates, and decrements in lung function). Time-series mortality studies reviewed in the 1996 PM CD (U.S. EPA, 1996a) provide strong evidence that ambient PM air pollution is associated with increases in daily human mortality and morbidity (e.g., increased hospital admissions and respiratory symptoms). These studies provided evidence that such effects occur at routine ambient PM levels, extending to 24-h
concentrations below the 150 : g/m 3 level of the PM 10 NAAQS set in 1987. Overall, as shown in Table 6-3, the PM 10 effects estimates deri ved from the recent PM 10 total mortality studies suggest
that an increase of 50 : g/m 3 in 24-h average PM 10 is significantly associated with an increase in total mortality, with an RR on the order of 1.025 to 1.05 in the general population. Table 6-3 also shows higher relative risks for increased hospital admissions for the elderly and for those with preexisting respiratory conditions, both of wh ich represent subpopulations at special risk for mortality implications of acute exposures to air pollution, including PM; higher relative risks are also shown for increased respiratory symptoms and decreased lung function in children. Results are very similar over a range of statistical models used in the analyses, and are not artifacts of the methods by which the data were analyzed. Further, these st udies suggest a possible linear,6-19 Table 6-3. Effect estimates per 50 : g/m 3 increase in 24-h PM 10 concentrations from U.S. and Canadian studies

Study location RR (± CI) only PM in model RR (± CI) other pollutants in model Reported PM 10 levels mean (min/max) †

Increased total ac ute mortality

Six Cities a

— Portage, WI 1.04 (0.98, 1.09) — 18 (±11.7) Boston, MA 1.06 (1.04, 1.09) — 24 (±12.8) Topeka, KS 0.98 (0.90, 1.05) — 27 (±16.1) St. Louis, MO 1.03 (1.00, 1.05) — 31 (±16.2) Kingston/Knoxville, TN 1.05 (1.00, 1.09) — 32 (±14.5) Steubenville, OH 1.05 (1.00, 1.08) — 46 (±32.3)
St. Louis, MO c 1.08 (1.01, 1.12) 1.06 (0.98, 1.15) 28 (1/97) Kingston, TN c 1.09 (0.94, 1.25) 1.09 (0.94, 1.26 30 (4/67) Chicago, IL h 1.04 (1.00, 1.08) — 37 (4/365) Chicago, IL g 1.03 (1.02, 1.04) 1.02 (1.01, 1.04) 38 (NR/128) Utah Valley, UT b 1.08 (1.05, 1.11) 1.19 (0.96, 1.47) 47 (11/297) Birmingham, AL d 1.05 (1.01, 1.10) — 48 (21, 80) Los Angeles, CA f 1.03 (1.00, 1.055) 1.02 (0.99, 1.036) 58( 15/177)
Increased hospital admission s (for elderly > 65 yrs.) Respiratory Disease
Toronto, CAN i 1.23 (1.02, 1.43) ‡ 1.12 (0.88, 1.36) ‡ 30-39 *
Tacoma, WA j 1.10 (1.03, 1.17) 1.11 (1.02, 1.20) 37 (14, 67)
New Haven, CT j 1.06 (1.00, 1.13) 1.07 (1.01, 1.14) 41 (19, 67)
Cleveland, OH k 1.06 (1.00, 1.11) — 43 (19, 72) Spokane, WA l 1.08 (1.04, 1.14) — 46 (16, 83)

COPD Minneapolis, MN n 1.25 (1.10, 1.44) — 36 (18, 58) Birmingham, AL m 1.13 (1.04, 1.22) — 45 (19, 77) Spokane, WA l 1.17 (1.08, 1.27) — 46 (16, 83)
Detroit, MI o 1.10 (1.02, 1.17) — 48 (22, 82)

Pneumonia Minneapolis, MN n 1.08 (1.01, 1.15) — 36 (18,58) Birmingham, AL m 1.09 (1.03, 1.15) — 45 (19, 77)Table 6-3. Effect estimates per 50 : g/m 3 increase in 24-h PM 10 concentrations from U.S. and Canadian studies (continued)

Study location RR (± CI) only PM in model RR (± CI) other pollutants in model Reported PM 10 levels mean (min/max) †

6-20 Spokane, WA l 1.06 (0.98, 1.13) — 46 (16, 83)
Detroit, MI o — 1.06 (1.02, 1.10) 48 (22, 82)

Ischemic HD
Detroit, MI p 1.02 (1.01, 1.03) 1.02 (1.00, 1.03) 48 (22, 82)
Increased respiratory symptoms Lower Respiratory

Six Cities q 2.03 (1.36, 3.04) Similar RR 30 (13,53) Utah Valley, UT r 1.28 (1.06, 1.56) J — 46 (11/195) 1.01 (0.81, 1.27) B Utah Valley, UT s 1.27 (1.08, 1.49) — 76 (7/251) Cough
Denver, CO x 1.09 (0.57, 2.10) — 22 (0.5/73)

Six Cities q 1.51 (1.12, 2.05) Similar RR 30 (13, 53) Utah Valley, UT s 1.29 (1.12, 1.48) — 76 (7/251) Decrease in Lung Function Utah Valley, UT r 55 (24, 86) ** — 46 (11/195) Utah Valley, UT s 30 (10, 50) ** — 76 (7/251) Utah Valley, UT w 29 (7,51) *** — 55 (1,181)

References:

a Schwartz et al. (1996a). l Schwartz (1996). x Ostro et al. (1991)

b Pope et al. (1992, 1994)/O 3 . m Schwartz (1994e). † Min/Max 24-h PM 10 in parentheses unless noted

c Dockery et al. (1992)/O 3 . n Schwartz (1994f). otherwise as standard deviation (± S.D), 10 and

d Schwartz (1993). o Schwartz (1994d). 90 percentile (10, 90). NR = not reported.

f Kinney et al. (1995)/O 3 , CO. p Schwartz and Morris (1995)/O 3 , CO, SO 2 . J Children.

g Ito and Thurston (1996)/O 3 . q Schwartz et al. (1994). B Asthmatic children and adults.

h Styer et al. (1995). r Pope et al. (1991). * Means of several cities.

i Thurston et al. (1994)/O 3 . s Pope and Dockery (1992). ** PEFR decrease in ml/sec. j Schwartz (1995)/SO 2 . t Schwartz (1994g) *** FEV 1 decrease.

k Schwartz et al. (1996b). w Pope and Kanner (1993). ‡ RR refers to total population, not just>65 years.

Source: Adapted from U.S. EPA, 1996b, Tables V-3, V-6, and V-7. See U.S. EPA (1996a,b) for all reference
citations.6-21 non-threshold PM/mortality relationship, but the data do not rule out the existence of an underlying nonlinear, threshold relationshi p (U.S. EPA, 1996a, 12-310-311; 1996b, VI-16). Figure 6-2 illustrates the consistency and coherence of the PM 10 epidemiology findings for increased total and cause-specific mortality and morbidity risks in adults and children. In addition, Table 6-4 summarizes results from a wide array of U.S. and Canadian studies that showed increased risks of mortality and morbidity to be related to changes in short-term (24-h) fine PM (indexed by PM 2.5 and other fine pa rticle indicators). As summarized below, long-term exposure st udies reviewed in the 1996 PM CD (U.S. EPA, 1996a) also provide evidence of associati ons between indicators of PM, including fine particle indicators, and chronic mortality and morbidity. Table 6-5 shows the direct comparisons of two key prospective studies of long-term PM mortality: the Harvard Six Cities Study (Dockery et al., 1993) and the American Cancer Society (ACS) Study (Pope et al., 1995). These two studies agree in their findings of strong a ssociations between fine particles and increased mortality. The RR estimates for total mortality are large and highly significant in the Six Cities study. With their 95% confidence intervals, the RR estimate for a 50 : g/m 3 increase in PM 15/10 is 1.42 (1.16, 2.01), the RR estimate for a 25 : g/m 3 increase in PM 2.5 is 1.31 (1.11, 1.68), and the
RR estimate for a 15 : g/m 3 increase in SO 4 is 1.46 (1.16, 2.16). The ACS study estimates for total mortality are smaller, but also more precise: RR = 1.17 (1.09, 1.26) for a 25 : g/m 3 increase
in PM 2.5 , and RR = 1.10 (1.06, 1.16) for a 15 : g/m 3 increase in SO 4 . Both studies used Cox regression models and were adju sted for similar sets of indivi dual covariates. In each case, however, caution must be applied in use of the stated quantitative risk estimates, given that the lifelong cumulative exposures of the study cohorts (especially in the di rtiest cities) included distinctly higher past PM exposures than thos e indexed by the more current PM measurements used to estimate long-term PM exposures in the study. Thus, somewhat lower relative risk estimates than the published ones may well apply. A third st udy by Abbey et al. (1991, 1995) reported no association between long-term PM exposure (indexed by TSP and other estimated PM indices) after 10 years, although the PM CD (U.S. EPA, 1996a) noted TSP may have been an inadequate index for exposure to inhalable particles and that additional follow-up might still
reveal chronic effects. An additional line of evidence concerning long-term effects may be seen in comparing cause-specific deaths in the Six Cities and ACS studies. The relative risks for the most versus the least polluted cities in the two studies ar e very similar for mortality from cardiopulmonary causes (U.S. EPA, 1996b, V-17). These two long-t erm exposure studies, taken together, suggest that there may be increases in mortality for specific disease categories that are consistent with long-term exposure to ambient fine particles. Moreover, at least some fraction of these deaths is6-22 Figure 6-2. Relative risk (RR) estimates for increased mortality and morbidity endpoints associated with 50 μg/m 3 increments in PM 10 concentrations as derived from studies cited by numbers listed above each given type of health endpoint.

Note: Notice 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 (see U.S. EPA, 1996b for full refe rence citations for each study identified in figure.)6-23 Table 6-4. Effect estimates per variable increments in 24-h concentrations of fine particle indicators (PM 2.5 , SO 4 = , H + ) from U.S. and Canadian studies

Acute mortality Indicator RR (± CI) per 25 : g/m 3 PM increase Reported PM levels mean (min/max) †

Six Cities a Portage, WI PM 2.5 1.030 (0.993, 1.071) 11.2 (±7.8) Topeka, KS PM 2.5 1.020 (0.951, 1.092) 12.2 (±7.4)
Boston, MA PM 2.5 1.056 (1.038, 1.0711) 15.7 (±9.2)
St. Louis, MO PM 2.5 1.028 (1.010, 1.043) 18.7 (±10.5) Kingston/Knoxville, TN PM 2.5 1.035 (1.005, 1.066) 20.8 (±9.6)

Steubenville, OH PM 2.5 1.025 (0.998, 1.053) 29.6 (±21.9)
Increased hospitalization
Ontario, CAN b SO 4 = 1.03 (1.02, 1.04) R = 3.1-8.2
Ontario, CAN c SO 4 = O 3 1.03 (1.02, 1.04) 1.03 (1.02, 1.05) R = 2.0-7.7

NYC/Buffalo, NY d SO 4 = 1.05 (1.01, 1.10) NR

Toronto d H + (Nmol/m 3 ) SO 4 = PM 2.5 1.16 (1.03, 1.30) * 1.12 (1.00, 1.24) 1.15 (1.02, 1.78) 28.8 (NR/391) 7.6 (NR, 48.7) 18.6 (NR, 66.0)
Increased respiratory symptoms

Southern California e SO 4 = 1.48 (1.14, 1.91) R = 2-37

Six Cities f (Cough) PM 2.5 PM 2.5 Sulfur H + 1.19 (1.01, 1.42) ** 1.23 (0.95, 1.59) ** 1.06 (0.87, 1.29) ** 18.0 (7.2, 37) *** 2.5 (3.1, 61) *** 18.1 (0.8, 5.9) ***

Six Cities f (Lower Resp. Symp.) PM 2.5 PM 2.5 Sulfur H + 1.44 (1.15-1.82) ** 1.82 (1.28-2.59) ** 1.05 (0.25-1.30) ** 18.0 (7.2, 37) *** 2.5 (0.8, 5.9) *** 18.1 (3.1, 61) ***

Decreased lung function

Uniontown, PA g PM 2.5 PEFR 23.1 (-0.3, 36.9) (per 25 : g/m 3 ) 25/88 (NR/88)
References:

a Schwartz et al. (1996a) † Min/Max 24-h PM indicator level shown in parentheses unless

b Burnett et al. (1994) otherwise noted as (± S.D.), 10 and 90 percentile (10,90)

c Burnett et al. (1995) O 3 or R = range of values from min-max, no mean value reported.

d Thurston et al. (1992, 1994) * Change per 100 nmoles/m 3

.
e Ostro et al (1993) ** Change per 20 : g/m 3 for PM 2.5 ; per 5 : g/m 3 for

f Schwartz et al. (1994) PM 2.5 sulfur; per 25 nmoles/m 3 for H +

. g Neas et al. (1995) *** 50th percentile value (10,90 percentile).

Source: Adapted from U.S. EPA, 1996b, Table V-12. See U.S. EPA (1996a,b) for all reference citations.6-24 Table 6-5. Effect estimates per increments a in annual average levels of fine particle indicators from U.S. and Canadian studies

Type of health effect and loca tion Indicator Change in health indicator per increment in PM a Range of city PM levels mean ( : g/m 3 )

Increased total chronic mo rtality in adults Relative risk (95% CI) Six City b PM 15/10 1.42 (1.16-2.01) 18-47

PM 2.5 1.31 (1.11-1.68) 11-30

SO 4 = 1.46 (1.16-2.16) 5-13

ACS Study c (151 U.S. SMSA) PM 2.5 1.17 (1.09-1.26) 9-34 *

SO 4 = 1.10 (1.06-1.16) 4-24

Increased bronchitis in ch ildren Odds ratio (95% CI) Six City d PM 15/10 3.26 (1.13, 10.28) 20-59 Six City e TSP 2.80 (1.17, 7.03) 39-114 24 City f H + 2.65 (1.22, 5.74) 6.2-41.0 24 City f SO 4 = 3.02 (1.28, 7.03) 18.1-67.3 24 City f PM 2.1 1.97 (0.85, 4.51) 9.1-17.3 24 City f PM 10 3.29 (0.81, 13.62) 22.0-28.6

Southern California g SO 4 = 1.39 (0.99, 1.92) —
Decreased lung function in children Six City d,h PM 15/10 NS Changes 20-59 Six City e TSP NS Changes 39-114 24 City i,j H + (52 nmoles/m 3 )

! 3.45% (-4.87, -2.01) FVC — 24 City i PM 2.1 (15 : g/m 3 )

! 3.21% (-4.98, -1.41) FVC — 24 City i SO 4 = (7 : g/m 3 )

! 3.06% (-4.50, -1.60) FVC — 24 City i PM 10 (17 : g/m 3 )

! 2.42% (-4.30, -.0.51) FVC —

a Estimates calculated annual-average PM increments assume: a 100 : g/m 3 increase for TSP; a 50 : g/m 3 increase for PM 10 and PM 15 ; a 25 : g/m 3 increase for PM 2.5 ; and a 15 : g/m 3 increase for SO 4 = , except where noted otherwise; a 100 nmole/m 3 increase for H +

. b Dockery et al. (1993). g Abbey et al. (1995a,b,c).

c Pope et al. (1995). h NS Changes = No significant changes.

d Dockery et al. (1989). i Raizenne et al. (1996).

e Ware et al. (1986). j Pollutant data same as for Dockery et al. (1996).

f Dockery et al. (1996). *Range of annual median values for subset of 50 cities.

Source: Adapted from U.S. EPA, 1996a, Table 12-6 and U.S. EPA, 1996b, Table V-8. See U.S. EPA (1996a,b) for all reference citations.4 As an initial matter, EPA concluded that the existing PM 10 standards were not adequate to protect public health, that fine and coarse fraction particles should be considered separately, that PM 2.5 was the appropriate indicator to use for fine particles, and that an annual PM 2.5 standard could provide the requisite reduction in risk associated with both annual and 24-h averaging times in most areas of the United States. This annual standard, together with a 24-h standard, could provide supplemental protection against extreme peak fine particle levels that

6-25 likely to be a consequence of cumulative, long-term exposure effects. These effects extend beyond the additive impacts of short-term expos ure episodes, in terms of producing marked increases above the expected number of daily d eaths among especially susceptible groups, such as the elderly and those with pulmonary disease. The PM CD (U.S. EPA, 1996a) also highlig hted a growing body of evidence directly comparing fine and coarse fraction PM effects that suggests that fine par ticles are more stronglyrelated than coarse fraction particles to increased mortality and morbidity in both short- and longterm exposure studies. Such evidence notably includes the results of analyses of the type illustrated in Figure 6-3 through 6-5. More specifically, Figure 6-3 shows a stronger relationship between changes in short-term (24-h) concen trations of fine pa rticles (indexed by PM 2.5 ) and increased mortality risks than for changes in short-term concentrations of coarse fraction particles (indexed by PM 15-2.5 ). Similarly, a stronger relationship is seen between chronic mortality and long-term exposure to fine par ticles (including both the sulfat e and nonsulfate components) than exposure to coarse fraction partic les (Figure 6-4), and a much st ronger relationship between lung function decrements and long-term exposure to fine particles than to coarse fraction particles (Figure 6-5).

6.4.2. NAAQS for Fine PM
The health effects evidence discussed above is relevant to this current HAD, as both this document (Chapter 2) and the PM CD present information that clearly shows DPM to be a constituent of ambient fine particles. Therefore, it is reasonable to conclude that DPM is associated, but to an undetermin ed degree, with the health e ffects described above. Whereas broader public health factors ar e taken into account in setting NAAQS than are relevant for this
noncancer assessment of lifetime exposure for DPM, the annual PM 2.5 NAAQS based primarily on this evidence is of interest in considering the extent to which the RfC for DE (as derived above in Section 6.3) is concordant with the info rmation on fine particles. As presented in the Federal Register fi nal rule notice (62 FR 38652, July 18, 1997), EPA drew upon the quantitative epidemiology informa tion concisely summarize d above to derive a
rationale for selection of an annual-average PM 2.5 standard. 4 First, to appropriately reflect themight occur in some localized situations or in areas with di stinct variations in seasonal fine particle levels (62 FR 38652).

6-26 Topeka

Portage

Steubenville

St. Louis

Harriman

Boston —

—

—

—

—

0.9 1 1.1 Relative Risk for 50 g/m PM in Six City Acute Study 3 15 μ

Relative Risk

Topeka

Portage

Steubenville

St. Louis

Harriman

Boston —

—

—

—

—

0.9 1 1.1 Relative Risk for 25 g/m Fine Particles (PM ) in Six City Acute Study 3

2.5 μ

Relative Risk Topeka

Portage

Steubenville

St. Louis

Harriman

Boston —

—

—

—

—

0.9 1 1.1 Relative Risk for 25 g /m Coarse Particles

(Pm – PM ) in Six City Acute Study 3

15 2.5 μ

Relative Risk

Figure 6-3. Relative risks of acute mortality in Harvard Six Cities Study, for inhalable thoracic particles (PM 15 /PM 10 ), fine particles (PM 2.5 ), and coarse fraction particles (PM 15 -PM 2.5 ).

Note: The coarse fraction effects are smaller and statisti cally nonsignificant (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 (R 2 = 0.69).

Source: PM CD (U.S. EPA, 1996a) graphical depiction of results from Schwartz et al. (1996).6-27 Figure 6-4. Adjusted rel ative risks for mortality are plotted against each of seven longterm average particle indices in the Harvard Six Cities Study, from largest range (total suspended particles, upper right) through sulfate and nonsulfate fine particle concentrations (lower left).

Note: A relatively strong linear relationship is seen for fine particles, and for sulfate and nonsulfate 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).6-28 Figure 6-5. Percent of ch ildren with <85% normal FVC versus annualaverage fine (PM 2.1 ) particle concentr ations and coarse fraction (PM 10-2.1 ) levels for 22 North American cities.

Note: A much stronger connection appears between fine particles and lung function decrements (top panel) than for coarse fraction particles (bottom panel). Source: PM Staff Paper (1996b) graphical depiction of results from Razienne et al. (1996).6-29 weight of evidence as a whole, EPA concl uded that it was appropr iate to limit annual PM 2.5 concentrations to somewhat below those wher e the body of epidemiological evidence is most consistent and coherent, recognizing both the strengths and limitations of the full range of information on the health effects of PM, as well as associated uncertainties. In accordance with EPA staff and CASAC views on the relative strengths of the epidemiologic studies, major reliance was placed on several short-term ( 24-h) exposure studies showing significantly increased risks of daily mortality (Schwartz et al., 1996) and morbidity indexed by hospital admissions (Thurston et al., 1994) and respirat ory symptoms/lung function decrements in children (Schwartz et al., 1994; Neas et al., 1995) in relationship to increased fine particle (PM 2.5 ) concentrations. Whereas it was recognized that health effects may occur over the full range of concentrations observed in these studies, it was concluded that the strongest evidence for shortterm PM 2.5 effects occurs at concentr ations near the long-term (e.g., annual) average. More specifically, the strength of the evidence of effects increases for concentrations of PM 2.5 that are at or above the long-term mean le vels reported for these studies. Given the serious nature of the potential effects, EPA judged that it was both pr udent and appropr iate to select a level for an
annual standard at or below such concentrations. More specifi cally, statistically significant increases in relative risks for daily mortality or morbidity were most clearly observed in these studies to be associated with 24-h fine particle concentrations in cities with long-term mean fine particle concentra tions ranging from about 16 to about 21 : g/m 3 , leading to the judgment that an
annual standard level of 15 : g/m 3 would be appropriate. Before reaching a final conclusion, the epide miologic studies of long-term exposures to fine particles were al so considered, which may reflect th e accumulation of daily effects over time as well as potential effects uniquely associated with long-term exposures. Even subject to additional uncertainties, these studi es were judged to provide impor tant insights with respect to the overall protection afforded by an annual standard. In particular, the annual mean PM 2.5 concentrations for the multiple cities included in the two key long-term exposure mortality studies (Dockery et al., 1993; Pope et al., 1995) were 18 : g/m 3 and about 21–22 : g/m 3

, respectively, with most of the 50 cities in the Pope, et al. (1995) having mean PM 2.5 concentrations above 15 : g/m 3 . Taken together with other long-term exposure studies and considering other factors di scussed in the final rule (62 FR 38676, July 18, 1997), EPA
concluded that the concordance of evidence for PM effects and associated levels provides clear support for an annual standard set at 15 : g/m 3

.6-30 6.4.3. DPM as a Component of Fine PM Chapter 2 of this document, as well as the PM CD (U.S. EPA, 1996a), report the extent to which DPM may contribute to ambient PM 2.5 concentrations. In so me urban situations, the annual average fraction of PM 2.5 attributable to DPM (according to mass concentrations) is about 35% on the high end, although the proportion appears to be more typically in the range of about 10% (see Chapter 2, Table 2-23 and Section 2.4.2.1). An approach to considering the relati onship of toxicity between DPM and PM 2.5 would be simply to assume that, as DPM is c ontributory to the content of ambient PM 2.5 , so too would it be contributory to toxicity of PM 2.5 . This approach is qualitative only because no firm basis currently exists for apportioning toxic ity among the various components of PM 2.5 . Nevertheless, some qualitative information from laboratory animal studies does exist, showing that DPM is no more potent at eliciting pulmonary pathology than other poorly soluble pa rticles such as talc, titanium dioxide, or carbon black in rats, or talc or titanium dioxide in mice. No data suggest that DPM is any more potent in eliciting pulm onary pathology than any other poorly soluble particle that typically may be present in ambient PM 2.5 . It may be reasonable to suggest, then, that DPM is no more likely to be toxicologically potent than any other fine particle constituents that typically make up ambient PM 2.5 . Based on the foregoing aspects of such an a pproach, a conclusion could be drawn that as long as DPM constituted its curre nt approximate proportion to PM 2.5 , the annual PM 2.5 standard would also be expected to provide a measure of protection for DPM. Even if a basis did exist to apportion toxicity among the va rious components of ambient PM 2.5 , such as DPM, use of such information in an approach to derive a safe ai r level for DPM would result in only a generalized, nonspecific estimate limited by a variety of factor s including the accuracy of the apportionment
of DPM from PM 2.5 . The RfC derived in Section 6.3 wa s based on an approach that utilized toxicological information from actual DPM exposures , a more direct approach that would result in a more specific estimate not limited by any appor tionment scheme.

6.5. CHARACTERIZATION OF THE NONCANCER ASSESSMENT FOR DIESEL
EXHAUST

Adverse health effects from short-term acute (high-level) exposures to DE such as occupational reports of decreases in lung function, wheezing, chest tightness, increases in airway resistance, and reports in laborat ory animals of inflammatory ai rway changes and lung function changes are acknowledged but are not assessed qua ntitatively. The focus of this dose-response
assessment is on the adverse noncancer health c onsequences of a lifetime , low-level, continuous air exposure by humans to DE.6-31 This assessment uses the whole particle, term ed DPM, as the key index or measure of DE dose. DPM includes any and all adsorbed orga nics, among which are a large number of PAHs, heterocyclic compounds, and their derivatives (Chapter 2), as we ll as the carbon core. It is not possible to separate the carbon core of DPM from the adsorbed organics to compare the toxicity in exposures other than with limited in-vitro-type scenarios. The dosimetric model used in the derivation of the RfC (Yu et al., 1991) is consistent with this de signation, as it considers DPM as well as the adsorbed organics as two types, slow-c leared and fast-cleared. Studies with diesel do occasionally report levels of accompanyi ng gaseous components of DE (e.g., NO

x , CO), but nearly all report particle concentration and characteristics. Adverse responses occurring in the rat lung have been used in this assessment as the basis for characterizing nonneoplastic human lung responses , yet use of these data in hazard evaluation
for cancer is not considered relevant to human s. The basis for this use of these noncancer pulmonary effects in rats for de rivation of an RfC includes the f act that humans and rats exhibit similar responses to other poorly soluble particles and also that similar noncancer effects are seen in other species (ILSI, 2000; Freedman and R obinson, 1988). Thus, when viewed across species (including humans), the nonneoplastic pulmonary e ffects of inflammation and fibrosis used in
this assessment are dissociable from the cancer response and are of likely relevance to humans. As a part of the RfC methodology (U.S. EPA, 1994), dose-response assessments are assigned levels of confidence that are intended to reflect the strengths and limitations of an
assessment as well as to indicate the likelihood of the assessment changing with any additional
information. Confidence levels of either low, medium, or high are assigned both to the study (or studies) used in the assessment to characterize the critical effects and to the overall toxicological
database of the substance. An overall confidence level also is assigned to the entire assessment. Usually, it is the same, or in any case no higher than the level assigned to the database. Compared with the databases of most other toxicants, the basic toxicological database for DE is substantial. The critical effects are characterized using not one but multiple long-term chronic studies conducted independe ntly of one another (Tables 6-1 and 6-2). The exhaustive
manner in which these studies were conducted and reported al so imparts a high degree of confidence. Both developmenta l and reproductive areas are addr essed. Also, ancillary studies that address mechanistic aspects of DE toxicity, either as the whole particle with adsorbed organics, or segregated as a poorly soluble partic le and extracted organics, are available and used in this assessment. Although only limited human da ta are available, exte nsive consideration has been given to the relevancy of the animal studies to the human condition. On the other hand,
data from related toxicants such as general ambient PM indi cate effects in endpoints (e.g., cardiovascular measures) that ha ve not been addressed in the DPM database. A major point to6-32 consider in assigning confidence in this assessment, and a reason that the value of the RfC may change in the future, is the emerging issue of allergenicity caused or exacerbated by DE. Although information to evaluate allergenicity in pa rallel to the present effects (pulmonary inflammation and histopathology) is currently lackin g, future efforts to elucidate and characterize this effect may well be a driver to make a reevaluation of the noncancer RfC derivation for DE appropriate. With respect to the current RfC for DE, the confiden ce level is medium, both for the
database and overall. The level reflects the relevan ce of (and information lacking on) allergenicity effects associated with DE in humans, and the possibility that the current RfC could change as a consequence of this information becoming available from the scientific community. In the introductory portion of this chapte r, DPM is acknowledged as a constituent of ambient PM (U.S. EPA, 1996a,b). A discussion of the quantitative epid emiology, particularly regarding fine PM, indicated that public health effects, including premature mortality, increased hospital admissions, respiratory symptoms, a nd decreased lung func tion, were observed in populations living in areas with long-term mean PM 2.5 levels generally ranging above 15 : g/m 3

. Application of the RfC method, which involved critical consider ation of the entirety of the disparate DE database with many chronic studies from several different species, evaluation of a myriad of possible DE-specifi c toxicological endpoints, and use of extrapolation models, produced a value of 5 : g/m 3 . As the accuracy of the RfC is stated in the definition ( “...within an order of magnitude ...” ), this dose-response estimate could be considered to be not different from
the level of 15 : g/m 3 , the lower end of the range identified for PM 2.5 . It is acknowledged here again that the levels of the PM 2.5 NAAQS should not be considered as indicative of the same degree of health protection for DE as intended by the RfC. Nevertheless, the congruence of these estimates tends to enhance the overall confidence that this range of levels is near or inclusive of those that would be expected to be protective of the human popula tion against the health effects
of DE.

6.6. SUMMARY Table 6-6 summarizes the key da ta and factors used in the dose-response analysis leading
to the derivation of the RfC for DE. The DE RfC of 5 : g DPM/m 3 is a chronic exposure likely to be without an appreciable risk of adverse human health effects.
The link between ambient fine PM and DPM with respect to origin, content, and possible health effects has been presen ted and discussed in this chap ter, and the general congruence
between the DE RfC and the level of the a nnual NAAQS for fine par ticles has been noted. Although these values should not be compared dir ectly, it is reasonable to observe that the annual
PM 2.5 standard would be expected to provide a m easure of protection for DPM, reflecting DPM’s6-33 Table 6-6. Decision summary for the quantitative noncancer RfC assessment for continuous exposure to diesel particulate matter (DPM)
Quantitative assessment for noncancer effects from lifetime exposure to DPM 5 : g/m 3
Critical effect Pulmonary inflammation and histopathology in rats
Principal study Array of four chronic rat studies
Designated basis for quantitation (e xposures in rats) 0.46 mg DPM /m 3 , 16 hr/day, 6 d/wk, 130 wks; a NOAEL

NOAEL HEC (HEC) 0.144 mg DPM / m 3
Adjustments for uncertainty factors (interspecies variability and intraspecies extrapolation) 30

NOAEL HEC /UF = RfC 0.144 mg/m 3 / 30 = 5 : g/m 3

current approximate proportion to PM 2.5 . The estimated air concentration of 5 : g/m 3 (the RfC, a lifetime exposure to DE measured as DPM) is above the ambient ai r levels reported in most rural areas but could be below those levels reported under short- term conditions in some urban scenar ios, such as at busy intersections or bus stops (see Chapter 2, Table 2-23). The RfC is intended to address lifetime chronic exposures and aspects of time-averaging for less than lifetime scenarios, such as, for example, acute exposures at busy intersections or bus st ops, which are not addre ssed in this particular

assessment.

REFERENCES FOR CHAPTER 6

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Abbey, DE; Hwang, BL; Burchette, RJ; et al. (1995) Estimated long-term ambient concentrations of PM 10 and development of respiratory symptoms in a nonsmoking population. Arch Environ Health 50:139-152.

Barnes, DG; Dourson, M. (1988) Reference dose (RfD): description and use in health risk assessments. Regul Toxicol Pharmacol 8:471-486.

Barnes, DG; Daston, GP; Evans, JS; et al. (1995) Benchmark dose workshop: criteria for use of a benchmark dose to estimate a reference dose. Regul Toxicol Pharmacol 21:296-306.6-34 Barnhart, MI; Chen, S-T; Salley, SO; et al. (1981) Ultrastructure and morphometry of the alveolar l ung of guinea pigs chronically exposed to diesel engine exhaust: six months’ experience. J Appl Toxicol 1:88-103.

Barnhart, MI; Salley, SO; Chen, S-T; et al. (1982) Morphometric ultrastructural analysis of alveolar l ungs of guinea pigs chronically exposed by inhalation to DE (DE). In: Toxicological effects of emissions from diesel engines: proceedings of the Environmental Protection Agency dies el emissions symposium; Lewtas, J., ed. October, 1981; Raleigh, NC. New York: Elsevier Biomedical; pp. 183-200. (Developments in toxicology and environmental science: v. 10).

Bhatnagar, RS; Hussain, MZ; Sorensen, KR; et al. (1980) Biochemical alterations in l ung connective tissue in rats and mice exposed to diesel emissions. In: Health effects of diesel engine emissions: proceedings of an international symposium. Pepelko, WE; Danner, RM; Clarke, NA., eds., v. 1; December 1979; Cincinnati, OH. Cincinnati, OH: U.S. Environmental Protection Agency, Health Effects Research Laboratory; pp. 557-570; EPA report no. EPA-600/9-80-057a. Available from: NTIS, Springfield, VA; PB81-173809.

Dockery, DW; Pope, CA, III; Xu, X; et al. (1993) An a ssociation between air pollution and mortality in six U.S. cities. N Engl J Med 329:1753-1759.

Freedman, AP; Robinson, SE. (1988) Noninvasive magnetopneumographic studies of l ung dust retention and clearance in coal miners. In: Respirable dust in the minera l industries: health effects, characterization, and control: proceedings of the international symposium on respirable dust in the mineral industries; October 1986; University Park, PA. Frantz, RL; Ramani, RV., eds. University Park, PA: Pennsylvania State University Press; pp. 181-186.

Heinrich, U; Peters, L; Funcke, W; et al. (1982) Investigation of toxic and carcinogenic effects of DE in long-term 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).

Heinrich, U; Muhle, H; Takenaka, S; et al. (1986) Chronic effects on the respiratory tract of hamsters, mice, and rats after long-term inhalation of high concentrations of filtered and unfiltered diesel engine emissions. J Appl Toxicol 6:383-395.

Heinrich, U; Fuhst, R; Rittinghausen, S; et al. ( 1995) Chronic inhalation exposure of Wistar rats and two different strains of mice to diesel engine exhaust, carbon black, and titantium dioxide. Inhal Toxicol 7:533-556.

Henderson, RF; Pickrell, JA; Jones, RK; et al. (1988) Response of rodents to inhaled diluted DE: biochemical and cytological changes in bronchoalveolar lavage fluid and in lung tissue. Fundam Appl Toxicol 11: 546-567.

Hyde, DM; Plopper, CG; Weir, AJ; et al. (1985) Peribronchiolar fibrosis in l ungs of cats chronically exposed to DE. Lab Invest 52:195-206.

International Life Sciences Institute Risk Science Institute Workshop Participants (ILSI). (2000) The relevance of the rat lung response to particle overload for human risk assessment: a workshop consensus report. In: ILSI Risk Science Institute Workshop: the relevance of the rat lung re sponse to particle overload for human risk assessment; March 1998. Gardner, DE., ed. Inhal Toxicol 12:1-17.

Ishinishi, N; Kuwabara, N; Nagase, S; et al. (1986) Long-term inhalation studies on effects of exhaust from heavy and light duty diesel engines on F344 rats. In: Carcinogenic and mutagenic effects of diesel engine exhaust: proceedings of the international satellite symposium on toxicological effects of emissions from diesel engines; July; 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).6-35 Ishinishi, N; Kuwabara, N; Takaki, Y; et al. (1988) Long-term inhalation experiments on DE. In: DE and health risks: results of the HERP studies. Tsukuba, Ibaraki, Japan: Japan Automobile Research Institute, Inc., Research Committee for HERP Studies; pp. 11-84.

Iwai, K; Udagawa, T; Yamagishi, M; et al. (1986) Long-term inhalation studies of DE on F344 SPF rats. Incidence of lung cancer and lymphoma. In: Carcinogenic and mutageni c effects of diesel engine exhaust: proceedings of the international satellite symposium on toxicological effects of emissions from diesel engines; July; Tsukuba Science City, Japan. Ishinishi, N; Koizumi, A; McClellan, RO; et al., eds. Amsterdam, Holland: Elsevier Science Publishers B. V; pp. 349-360. (Developments in toxicology and environmental science: v. 13).

Johnston, CJ; Finkelstein, JN; Mercer, P, et al. (2000) Pulmonary effects induced by ultrafine PTFE particles. Toxicol Appl Pharmacol 168:208-215.

Lewis, TR; Green, FHY; Moorman, WJ; et al. (1989) A chronic inhalation toxicity study of diesel engine emissions and coal dust, alone and combined. J Am. Coll Toxicol 8:345-375.

Mauderly, JL; Jones, RK; Griffith, WC; et al. (1987a) DE is a pulmonary carcinogen in rats exposed chronically by inhalation. Fundam. Appl Toxicol 9:208-221.

Mauderly, JL; Bice, DE; Carpenter, RL; et al. (1987b) Effects of inhaled nitrogen dioxide and DE on developing lung. Cambridge, MA: Health Effects Institute; research report no. 8.

Mauderly, JL; Bice, DE; Cheng, YS; et al. (1989) Influence of experimental pulmonary emphysema on toxicologicaleffects from inhaled nitrogen dioxide and diesel exhaust. Cambridge, MA: Health Effects Institute; report no. HEIRR-89/30. Available from: NTIS, Springfield, VA; PB90-247347.

National Research Council (NRC). (1983) Risk assessment in the federal government: managing the process. Washington, DC: National Academy Press.

Neas, LM; Dockery, DW; Koutrakis, P; et al. (1995) The association of ambient air pollution with twice daily peak expiratory flow rate measurements in children. Am. J Epidemiol 141:111-122.

Nightingale, JA; Maggs, S; Cullinan, P; et al. (2000) Airway inflammation after controlled exposure to diesel exhaust particulates. Am.J Respir Crit Care Med 162(1):161-166.

Nikula, KJ; Snipes, MB; Barr, EB; et al. (1995) Comp arative pulmonary toxicities and carcinogenicities of chronically inhaled DE and carbon black in F344 rats. Fundam Appl Toxicol 25:80-94.

Nordenhall, C; Pourazar, J; Blomberg, A; (2000) Airway inflammation following exposure to diesel exhaust: a study of time kinetics using induced sputum. Eur Respir J 15(6):1046-1051.

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Plopper, CG; Hyde, DM; Weir, AJ. (1983) Centriacinar alterations in l ungs of cats chronically exposed to DE. Lab Invest 49:391-399.6-36 Pope, CA, III; T hun, MZ; Nam boodiri, MM; et al. (1995) Par ticulate air pollution as a predictor of mortality in a prospective study of U.S. adults. Am J Respir Crit Care Med 151:669-674.

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Salvi, S; Blomberg,A; Rudell, B; et al. (1999) Acute inflammatory responses in the airways and peripheral blood after short-term exposure to diesel exhaust in healthy human volunteers. Am J Respir Crit Care Med 159 (3):702- 709.

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Schwartz, J; Dockery, DW; Neas, LM; Wypij, D; Ware, JH ; Spengler, JD; Koutrakis, P; Speizer, FE; Ferris, BG, Jr. (1994) Acute effects of summer air pollution on respiratory symptom reporting in children. Am. J. Respir. Crit. Care Med. 150: 1234-1242.

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Yu, CP; Yoon, KJ. (1990) Retention modeling of DE particles in rats and humans. Amherst, NY: State University of New York at Buffalo (Health Effects Institute research report no. 40).

Yu, CP; Yoon, KJ; Chen, YK. (1991) Retention modeling of DE particles in rats and humans. J Aerosol Med 4:79-115.1 As noted in Chapter 6, a new PM CD is now being prepared to reflect the latest scientific studies on ambient PM available since the last document was completed.

7-1 7. CARCINOGENICITY OF DIESEL EXHAUST

7.1. INTRODUCTION Initial health hazard concerns regarding the potential carcinogenicity of diesel engine exhaust (DE) were based on the reported induction of skin papillomas by diesel particle extracts (Kotin et al., 1955), evidence for mutagenicity of extracts (Huisingh et al., 1978), evidence that components of diesel extract act as weak tumor promoters (Zamora et al., 1983), and the knowledge that diesel particles and their associated organics are respirable. During the 1980s, both human epidemiologic studies and long-term animal cancer bioassays were initiated. In 1981, Waller published the first epidemiologic investigation, a retrospective mortality study of London transport workers. Since then a large number of retrospective cohort and case-control studies have been carried out with railroad workers, dockworkers, truck drivers, construction workers, miners, and bus garage employees. During 1986 and 1987, several chronic animal cancer bioassays were published. These studies and numerous laboratory investigations carried out since then have been directed toward assessing the carcinogenic potential of whole exhaust, evaluating the importance of various exhaust components in the induction of cancer, and understanding the mode of action and implications of deposition, retention, and clearance of DE particles.

7.1.1. Overview This chapter evaluates the carcinogenic potential of DE in both humans (Section 7.2) andanimals (Section 7.3), discusses mode(s) of action (Section 7.4), and provides an overall weightof-evidence evaluation (Section 7.5) for carcinogenicity in humans. This chapter also summarizes evaluations of DE conducted by other organizations (Section 7.6) and the final conclusions (Section 7.7) identify major uncertainties for which additional research is needed. This assessment focuses on DE, although it should be noted that diesel particles make up a portion of ambient particulate matter (PM) (Chapter 2, Section 2.2.3; Chapter 6, Section 6.4.3), and thus, the ambient PM data may have some relevance.

7.1.2. Ambient PM-Lung Cancer Relationships A brief overview of the data regarding exposure to ambient PM and lung cancer is provided as background information and is based on analyses contained in the 1996 Air Quality Criteria for PM (PM CD) (U.S. EPA, 1996a). 1 With DE being part of ambient PM, the question7-2 of what is seen in the ambient PM data is of interest, as epidemiologic evidence for an effect of ambient PM on lung cancer mortality or incidence could possibly contribute to evaluation of DE-specific epidemiologic data. Chapters 2 and 5 noted that DPM, consisting mostly of fine particles (<1.0 mm diameter), represents a toxicologically important component of typical ambient fine particle mixes. As discussed in Chapter 6, several large-scale prospective studies (Harvard Six Cities Study; American Cancer Society (ACS) Study; Adventist Health Study of Smog (AHSMOG) provide important evidence regarding associations between chronic exposures to ambient fineparticles and increased risks of noncancer mortality/morbidity effects (e.g., cardiorespiratoryrelated deaths or hospital admissions) (U.S. EPA, 1996a). As summarized below, these same studies also evaluated relationships between chronic PM exposures and lung cancer mortality
and/or incidence. As an initial matter, both the Harvard Six Cities Study (Dockery et al., 1993), of approximately 8,000 adults in six cities comprising a transect across the northcentral and northeastern United States, and the ACS Study (Pope et al., 1995), of 550,000 adults in 151 cities across all U.S. geographic regions, found markedly increased relative risks (RR) of lung cancer mortality associated with smoking. More specifically, the Six City Study reported increased risks of smoking for current (RR = 8.00, 95% CI = 2.97-21.6) and former (RR = 2.54, CI = 0.90-7.18) smokers, with the ACS Study reporting striking similar increased risks for current smokers (RR = 9.73, 95% CI = 5.96-15.9). After controlling for smoking and other risk factors, both the Six Cities Study and the ACS Study (using a subset of 50 of the 151 cities) evaluated relationships between long-term exposure to fine PM (indexed by PM 2.5 ), from the least to the most polluted of the cities in each study, and lung cancer mortality. In both studies, lung cancer mortality risks were not statistically significantly associated with ambient PM 2.5 concentrations in combined analyses of data for both males and females (RR = 1.37, 95% CI = 0.81-2.31, in the Six Cities Study; RR = 1.03, 95% CI = 0.80-1.33, in the ACS Study). Also, lung cancer mortality risks were not statistically significantly associated with ambient PM 2.5 concentrations in the ACS Study for smaller sample size subgroups broken out by sex and smoking status. In addition, analyses of data from the AHSMOG series of studies, of 6,338 nonsmoking long-term California adult residents, found no statistically significant associations between PM 2.5 (estimated from visibility data) and lung cancer mortality or total mortality (Abbey et al., 1995); further, no such associations were reported for PM 10 (estimated from total suspended particulate matter [TSP] data) in the same study. Earlier AHSMOG analyses (Abbey et al., 1991) reported no statistically significant associations between TSP (which includes not only fine PM but also larger coarse-7-3 mode particles ranging up to 25-50 

### : m) and respiratory cancer for either sex (only respiratory symptoms and any-site female cancers were reported to be associated with TSP in this study). The ACS Study and the later AHSMOG analyses (Abbey et al., 1995) also evaluated relationships between long-term exposures to sulfates (SO 4 ) (which are predominantly but not exclusively found in fine-mode particles, and can be considered an index for ambient fine particles) and lung cancer mortality. The ACS Study reported somewhat elevated and statistically significant lung cancer risk (RR = 1.36, 95% CI = 1.11-1.66) across 151 cities in combined analyses of data for both males and females. However, in further analyses of subgroups broken out by sex and smoking status (and thus having smaller sample sizes in each than for the above overall combined analyses), only the lung cancer mortality risks for male “ever-smokers” (RR = 1.44, 95% CI = 1.14-1.83) were statistically significant; no statistically significant relationships were reported for male “never-smokers” (RR = 1.36, 95% CI = 0.40 -4.66), for female “ever-smokers” (RR = 1.10, 95% CI = 0.72-1.68), or for female “neversmokers” (RR = 1.61; 95% CI = 0.66 -3.92). In the later AHSMOG analyses, Abbey et al. (1995) found no statistically significant associations between sulfates and lung cancer or total mortality. In summary, the three key prospective cohort studies summarized above, and discussed in more detail in the 1996 PM CD (U.S. EPA, 1996a), provide an equivocal array of results with regard to possible associations between chronic exposures to ambient PM and lung cancer mortality and/or incidence. None of the analyses of fine particles (as indexed by PM 2.5 ) in these three studies reported statistically significant relationships between long-term PM 2.5 concentrations and lung cancer mortality. Only the ACS Study found a statistically significant association of increased risk of lung cancer with one indicator of ambient fine particles (sulfates). Overall, then, these studies support a conclusion that there continues to be little epidemiologic evidence for an effect of ambient PM on lung cancer mortality or incidence. It is recognized, however, that subsequent AHSMOG analyses and other studies, published since completion of the 1996 PM CD, have further analyzed relationships between ambient PM and lung cancer. Results from these more recent studies are now being evaluated as part of the integrated assessment of ambient PM that will be part of the new PM CD targeted for completion in 2002.

7.2. EPIDEMIOLOGIC STUDIES OF THE CARCINOGENICITY OF
EXPOSURE TO DIESEL EXHAUST An increased risk from malignancies of the lung, bladder, and lymphatic tissue has been reported in populations potentially exposed to higher levels of DE than typically seen in the environment. A few authors have reported other malignancies, including testicular cancer7-4 (Garland et al., 1988), gastrointestinal cancer (Balarajan and McDowall, 1988; Guberan et al., 1992), and prostate cancer (Aronsen et al., 1996). A detailed review of 22 lung cancer studies is presented in this section; a few more studies exist, but these 22 are judged to be the key ones. A detailed review of other health effect studies is not presented because findings are equivocal. Excess risk of bladder cancer has been reported in several studies (Howe et al., 1980; Wynder et al., 1985; Hoar and Hoover et al., 1985; Silverman et al., 1983; Vineis and Magnani 1985; Silverman et al., 1986; Jensen et al., 1987; Steenland et al., 1987; Isocovich et al., 1987; Risch et al., 1988; Iyer et al., 1990; Steineck et al., 1990; Cordier et al., 1993; Notani et al., 1993). Very few studies found significant excesses after adjustment for cigarette smoking. Most studies failed to show any association between exposure to DE and occurrence of bladder cancer. Some authors have reported excess mortality from lymphohematopoietic system cancers in people potentially exposed to diesel fumes. Rushton and Alderson (1983) and Howe and Lindsay (1983) found increased mortality from lymphatic neoplasms. Balarajan and McDowall (1983) found raised mortality for malignant lymphomas. Flodin et al. (1987) observed increased risk for multiple myeloma, and Bender et al. (1989) reported excess mortality from leukemia. Because evidence for bladder cancer and lymphohematopoietic cancer was found to be equivocal, detailed reviews of these studies are not presented here. The potential for elevated DE exposure in the occupational setting generally includes miners, railroad workers, truckers, bus and taxi drivers, heavy equipment operators, farm tractor drivers, and those involved with heavy duty marine engines. Regarding the mining industry some assert that excess lung cancer should be observed in the miners if exposure to DE is causally associated with the occurrence of lung cancer since DE is allegedly present in the mines. Our review of the mining industry data dose not support this assertion for the following reasons. In the United States, the introduction of diesel engines into metal mines dates from the early to the mid 1960s. Currently, there are approximately 265 underground metal/nonmetal mines in the United States. Virtually all of these mines use diesel powered equipment for various tasks, such as haulage, roof bolting etc. (Department of Labor, Mine Safety and Health Administration, 2001). Introduction of diesel equipment into coal mines was even later. Of 910 existing underground coal mines in the United States, only 145 currently use diesel-powered equipment. Of these 145 mines, 32 mines are currently using diesel equipment for face coal haulage. The remaining mines use diesel equipment for transportation, materials handling, and other support operations (Department of Labor, Mine Safety and Health Administration, 2001). It should be noted that there is a paucity of epidemiologic studies in miners where exposure to DE and health effects are explored. Furthermore, the majority of epidemiologic studies in miners do not mention exposure to diesel equipment use. Thus, it is impossible to know how many miners were exposed to DE and for how long and at what concentrations in a given study,7-5 if any. Hence the studies in miners (coal, metal, and nonmetal with the exception of potash miners) are not reviewed in this chapter, because the available studies are uninformative relative
to DE. In this section, various mortality and morbidity studies of lung cancer from potential exposure to diesel engine emissions are reviewed. Although an attempt was made to cover all the relevant studies, a number of studies are not included for several reasons. In the United States the change from 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% complete by 1959 (Garshick et al., 1988). Therefore, exposure to DE was less common, and the follow-up period for studies conducted prior to 1960 (Raffle, 1957; Commins et al., 1957; Kaplan, 1959) was not long enough to cover the long latency period of lung cancer. The usefulness of these studies in evaluating the carcinogenicity of DE is greatly reduced; thus, they are not considered here. On the other hand, the trucking industry changed to diesel trucks by the 1960s. In the 1960s sales of diesel-powered Class 8 trucks (long-haul trucks) were 48% of the market, and by the 1970s sales had risen to 85%. Thus, studies conducted among truck drivers prior to the 1970s may reflect exposures to gasoline exhaust as well as DE. Hence, studies with ambiguous exposures or studies that examined several occupational risk factors were excluded because they would have contributed little to the evaluation of the carcinogenicity of DE (Waxweiler et al., 1973; Williams et al., 1977; Ahlberg et al., 1981; Stern et al., 1981; Buiatti et al., 1985; Gustafsson et al., 1986; Siemiatycki et al., 1988). A study by Coggon et al. (1984) was excluded because occupational information abstracted from death certificates had not 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 meta-analysis. In the cohort studies, cohorts of heavy construction equipment operators, railroad and locomotive workers, bus garage employees, and miners were studied retrospectively to determine increased mortality and morbidity resulting from exposures to varying levels of diesel emissions in the workplace. The evaluation of each study presents the study population, methodology used for the study, i.e., data collection and verification, analysis, results, and a critique of the study. There are some methodologic limitations that are common to studies with similar design. The total evidence, including limitations, is discussed at the end of the chapter in the summary and discussion section.7-6 7.2.1. Cohort Studies
7.2.1.1. Waller (1981): Trends in Lung Cancer in London in Relation to Exposure
to Diesel Fumes A retrospective mortality study of a cohort of London transport workers was conducted to determine if there was an excess of deaths from lung cancer that could be attributed to DE exposure. From nearly 20,000 male employees in the early years, those aged 45 to 64 were followed for the 25-year period between 1950 and 1974 (the actual number of employees is not given in the paper), constituting a total of 420,700 man-years at risk. These workers were distributed among five job categories: drivers, garage engineers, conductors, motormen or guards, and engineers (works). Lung cancer were ascertained from death certificates of individuals who died while still employed, or if retired, following diagnosis. Expected death rates were calculated by applying greater London death rates to the population at risk within each job category. Data were calculated in 5-year periods and 5-year age ranges, and the results were combined to obtain the total expected deaths in the required age range for the calendar period. A total of 667 cases of lung cancer was reported, compared with 849 expected, to give a cancer mortality ratio of 79%. In each of the five job categories, the observed numbers were below those expected. Engineers in garages had the highest mortality ratio, 90%, motormen and guards had a mortality ratio of 87%, and both the bus drivers and conductors had mortality ratios of 75%. The engineers in the central works had a mortality ratio of 66%. These mortality ratios did not differ significantly from each other. Environmental sampling was done at one garage, on one day in 1979, for benzo[ a ]pyrene (B[ a ]P) concentrations and was compared with corresponding values recorded in 1957. Concentrations of B[ a ]P recorded in 1957 were at least 10 times greater than those measured in 1979. This study failed to find any association between DE and occurrence of lung cancer, which may be due to several methodologic limitations. The lung cancer deaths were ascertained while the workers were employed (the worker either died of lung cancer or retired after lung cancer was diagnosed). Although man-years at risk were based on the entire cohort, no attempt was made to trace or evaluate the individuals who had resigned from the London transport company for any other reason. Hence, information on resignees who may have had significant exposure to DE, and on lung cancer deaths among them, was not available for analysis. This may have led to a dilution effect, resulting in underascertainment of observed lung cancer deaths and underestimation of mortality ratios. Eligibility criteria for inclusion in the cohort, such as starting date and length of service with the company, were not specified. Therefore, there may not have been sufficient latency for the development of lung cancer. Use of greater London population death rates to obtain expected number of deaths may have resulted in a deficit in mortality ratios reflecting the “healthy worker effect.” Investigators did not categorize the five7-7 job categories either by qualitative or quantitative levels of DE exposure; neither did they use an internal comparison group to derive risk estimates. The age range considered for this study was limited (45 to 64 years of age) for the period between 1950 and 1974. It is not clear whether this age range was applied to calendar year 1950 or 1974, or at the midpoint of the 25-year follow-up period. No analyses were presented either by latency or by duration of employment (surrogate for exposure). The environmental survey based on B[ a ]P 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 when the reduction in B[ a ]P concentration occurred, because there are no environmental readings available between 1957 and 1979. It is also important to note that the concentrations of B[ a ]P inside the garage in 1957 were not very different from those outside the garage, thus indicating that exposure for garage workers was not much different from that of the general population. Thus, this study fails to provide either positive or negative association between the DE exposure and the occurrence of lung cancer.

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 This is a retrospective cohort study of the mortality experience of 43,826 male pensioners of the Canadian National Railroad (CNR) between 1965 and 1977. Members of this cohort consisted of male CNR pensioners who had retired before 1965 and who were known to be alive at the start of that year, as well as those who retired between 1965 and 1977. The records were obtained from a computer file that is regularly updated and used by the company for payment of 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 Canada since 1950. Of the 17,838 deaths among members of the cohort between 1965 and 1977, 16,812 (94.4%) were successfully linked to a record in the mortality file. A random sample manual check on unlinked data revealed that failure to link was due mainly to some missing information on the death records. Occupation at time of retirement was used by the Department of Industrial Relations to classify workers into three diesel fume and coal dust exposure categories: (1) nonexposed, (2) possibly exposed, and (3) probably exposed. Person-years of observation were calculated and classified by age at observation in 5-year age groups (35 to 39, 40 to 44, . . . , 80 to 84, and 

### $ 85 years). The observed deaths were classified by age at death for different cancers, for all cancers combined, and for all causes of death combined. Standard mortality ratios (SMRs) were then calculated using rates of the Canadian population for the period between 1965 and 1977. The7-8 relative risks were calculated using the three exposure categories: nonexposed, possibly exposed, and probably exposed. Both total mortality (SMR = 95, p <0.001) and all cancer deaths (SMR = 99, p >0.05) were close to that expected for the entire cohort. Analysis by exposure to diesel fume levels in the three categories (nonexposed, possibly exposed, and probably exposed) revealed an increased relative risk for lung cancer among workers with increasing exposure to diesel fumes. The relative risk for nonexposed workers was presumed to be 1.0; for those possibly exposed, the relative risk was significantly elevated to 1.2 ( p =0.013); and for those probably exposed, it was significantly elevated to 1.35 ( p =0.001). The corresponding rates for exposure to varying levels of coal dust were very similar at 1.00, 1.21 ( p =0.012), and 1.35 ( p =0.001), respectively. The trend tests were highly significant for both exposures ( p <0.001). Analysis performed after the exclusion of individuals who worked in the maintenance of steam engines, and hence were exposed to high levels of asbestos, yielded a risk of lung cancer of 1.00, 1.21, and 1.33 for those nonexposed, possibly exposed, and probably exposed to DE, respectively, with a highly significant trend ( p <0.001). An analysis done on individuals who retired prior to 1950 showed the relative risk of lung cancer among nonexposed, possibly exposed, and probably exposed to be 1.00, 0.70, and 0.44, respectively, based on fewer than 15 deaths in each category. A similar analysis of individuals who retired after 1950 found the results in the same categories to be 1.00, 1.23, and 1.40, respectively. Although retirement prior to 1950 indicated exposure to coal combustion fumes alone, retirement after 1950 shows the results of mixed exposure to coal combustion fumes and diesel fumes. As there was considerable overlap between occupations involving probable exposure to diesel fumes and probable exposure to coal, and as most members of the cohort were employed during the years in which the transition from coal to diesel occurred, it was difficult to distinguish whether lung cancer was associated with exposure to coal
combustion fumes or diesel fumes or a mixture of both. Although this study showed a highly significant dose-response relationship between diesel fumes and lung cancer, it has some methodological limitations. There were concurrent exposures to both diesel fumes and coal combustion fumes during the transition period; therefore, misclassification of exposure may have occurred, because only occupation at retirement was available for analysis. It is possible that the elevated response observed for lung cancer was due to the combined effects of exposure to both coal dust/coal combustion products and diesel fumes and not just one or the other. However, deaths due to lung cancer were not elevated among workers who retired prior to the 1950s and thus would have been primarily exposed to coal dust/coal combustion products. Furthermore, it should be noted that so far coal dust has not been demonstrated to be a pulmonary carcinogen in studies of coal miners. This7-9 study was restricted to deaths among retired workers; therefore, it is unclear if a worker who developed lung cancer when actively employed and filed for a disability claim instead of retirement claim would be included in the study or not. Thus, it is possible that workers with heavy exposure might have been excluded from the study. Neither information on duration of 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 risk. Although information on potential confounders such as smoking is lacking, the use of an internal comparison group to compute the relative risks minimizes the potential for confounding by smoking, as there is no reason to assume different smoking patterns among individuals exposed to DE versus those not exposed. Despite these limitations, this study provides suggestive evidence toward a causal association between exposure to DE and excess lung cancer.

7.2.1.3. Rushton et al. (1983): Epidemiological Survey of Maintenance Workers in the London Transport Executive Bus Garages and Chiswick Works This is a retrospective mortality cohort study of male maintenance workers employed for at least 1 continuous year between January 1, 1967, and December 31, 1975, at 71 London transport bus (also known as rolling stock) garages and at Chiswick Works. The following information was obtained from computer listings: surname with initials, date of birth, date of joining company, last or present job, and location of work. For those individuals who left their 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 an alphabetical card index in the personnel department. Additional tracing of individuals who had left was carried out through social security records. The area of residence was assumed to be close to their work; therefore place of work was coded as residence. One hundred different job titles were coded into 20 broader groups. These 20 groups were not ranked for DE exposure, however. The reason for leaving was coded as died in service, retired, or other. The underlying cause of death was coded using the eighth revision of the International Classification of Diseases (ICD). Person-years were calculated from date of birth and dates of entry to and exit from the study using the man-years computer language program. The workers were then subdivided into 5-year age and calendar period groups. The expected number of deaths was calculated by applying the 5-year age and calendar period death rates of the comparison population with the person-years of corresponding groups. The mortality experience of the male population in England and Wales was used as the comparison population. Significance values were calculated for the difference between the observed and expected deaths, assuming a Poisson distribution.7-10 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. Observed deaths from all causes were significantly lower than expected (O = 495,

p <0.001). Observed deaths from all neoplasms and cancer of the lung were approximately the same as those expected. The only significant excess observed, for cancer of the liver and gall bladder at Chiswick Works, was based on four deaths ( p <0.05). A few job groups showed a significant excess of risks for various cancers. All the excess deaths observed for the various job groups, except for the general hand category, were based on very small numbers (usually fewer than five) and merited cautious interpretation. Only a notable excess in the general hand category for lung cancer was based on as many as 48 cases (SMR = 133, p <0.03). This mortality study did not demonstrate any cancer excess. Details of work history were not obtained to permit any analysis by DE exposure. The study’s limitations, including small sample size, short duration of follow-up (average of only 6 years), and lack of sufficient latency period, make it inadequate to draw any conclusions.

7.2.1.4. Wong et al. (1985): Mortality Among Members of a Heavy Construction Equipment Operators Union With Potential Exposure to DE Emissions This retrospective mortality study was conducted on a cohort of 34,156 male members of a heavy construction equipment operators union with potential exposure to DE emissions. Study cohort members were identified from records maintained at Operating Engineers’ Local Union No. 3-3A in San Francisco, CA. This union has maintained both work and death records on all its members since 1964. Individuals with at least 1 year of membership in this union between January 1, 1964, and December 31, 1978, were included in the study. Work histories of the cohort were obtained from job dispatch computer tapes. The study follow-up period was January 1964 to December 1978. Death information was obtained from a trust fund, which provided information on retirement dates, vital status, and date of death for those who were entitled to retirement and death benefits. Approximately 50% of the cohort had been union members for less than 15 years, whereas the other 50% had been union members for 15 years or more. The average duration of membership was 15 years. As of December 31, 1978, 29,046 (85%) cohort members were alive, 3,345 (9.8%) were dead, and 1,765 (5.2%) remained untraced. Vital status of 10,505 members who had left the union as of December 31, 1978, was ascertained from the Social Security Administration. Death certificates were obtained from appropriate State health departments. Altogether, 3,243 deaths (for whom death certificates were available) in the cohort were coded using the seventh revision of the ICD. For 102 individuals, death certificates could not be obtained, only the date of death; these individuals
were included in the calculation of the SMR for all causes of death but were deleted from the7-11 cause-specific SMR analyses. Expected deaths and SMRs were calculated using the U.S. national age-sex-race 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 15-year study period. A total of 3,345 deaths was observed, compared with 4,109 expected. The corresponding SMR for all causes was 81 ( p =0.01), which is consistent with the “healthy worker effect.” A total of 817 deaths was attributed to malignant neoplasms, slightly fewer than the 878 expected based on U.S. white male cancer mortality rates (SMR = 93, p =0.05). Mostly there were SMR deficits for cause-specific cancers, including lung cancer for the entire cohort (SMR = 99, O = 309). The only significant excess SMR was observed for cancer of the liver (SMR = 167, O = 23, p <0.05). Analysis by length of union membership as a surrogate of duration for potential exposure showed statistically significant increases in SMRs of cancer of the liver (SMR = 424, p <0.01) in the 10- to 14-year membership group and of the stomach (SMR = 248, p <0.05) in the 5- to 9- year membership group. No cancer excesses were observed in the 15- to 19-year and 20+-year membership groups. Although the SMR for cancer of the lung had a statistically significant deficit in the less-than-5-year duration group, it showed a positive trend with increasing length of membership, which leveled off after 10 to 14 years. Cause-specific mortality analysis by latency period showed a positive trend for SMRs of all causes of death, although all of them were statistically significant deficits, reflecting the diminishing “healthy worker effect.” This analysis also demonstrated a statistically significant SMR excess for cancer of the liver (10- to 19-year group, SMR = 258). The SMR for cancer of the lung showed a statistically significant deficit for a <10-year latency but showed a definite positive trend with increasing latency. In addition to these analyses of the entire cohort, similar analyses were carried out in various subcohorts. Analyses of retirees, 6,678 individuals contributing to 32,670 person-years, showed statistically significant increases ( p <0.01) in SMRs for all cancers; all causes of death; cancers of the digestive system, large intestine, respiratory system, and lung; emphysema; and cirrhosis of the liver. The other two significant excesses ( p <0.01) were for lymphosarcoma and reticulosarcoma and nonmalignant respiratory diseases. Further analysis of the 4,075 retirees (18,678 person-years) who retired at age 65 or who retired earlier but had reached the age of 65 revealed statistically significant SMR increases ( p <0.05) for all cancers, cancer of the lung, and lymphosarcoma and reticulosarcoma. To analyze cause-specific mortality by job held (potential exposure to DE emissions), 20 functional job titles were used, which were further grouped into three potential categories: high exposure, low exposure, and unknown exposure. A person was classified in a job title if he ever7-12worked on that job. Based on this classification system, if a person had ever worked in a highexposure job title he was included in that group, even though he may have worked for a longer time in a low-exposure group or in an unknown exposure group. Information on length of work in any particular job, hence indirect information on potential length of exposure, was not
available either. For the high-exposure group a statistically significant excess was observed for cancer of the lung among bulldozer operators who had 15 to 19 years of membership and 20+ years of follow-up (SMR = 343, p <0.05). This excess was based on 5 out of 495 deaths observed in this group of 6,712 individuals, who contributed 80,328 person-years of observation. The cause-specific mortality analysis in the low-exposure group revealed statistically significant SMR excesses in individuals who had ever worked as engineers. These excesses were for cancer of the large intestine (SMR = 807, O = 3, p <0.05) among those with 15 to 19 years of membership and length of follow-up of at least 20 years, and cancer of the liver (SMR = 872, O = 3, p <0.05) among those with 10 to 14 years of membership and length of follow-up of 10 to 19 years. There were 7,032 individuals who contributed to 78,403 person-years of observation in the low-exposure group. For the unknown exposure group, a statistically significant SMR was observed for motor vehicle accidents only (SMR = 174, O = 21, p <0.05). There were 3,656 individuals who contributed to 33,388 person-years of observation in this category. No work histories were available for those who started their jobs before 1967 and for those who held the same job prior to and after 1967. This group comprised 9,707 individuals (28% of the cohort) contributing to 104,448 person-years. Statistically significant SMR excesses were observed for all cancers (SMR = 112, O = 339, p <0.05) and cancer of the lung (SMR = 119, O = 141, p <0.01). A significant SMR elevation was also observed for cancer of the stomach (SMR = 199, O = 30, p <0.01). This study demonstrates a statistically significant excess for cancer of the liver but also shows statistically significant deficits in cancers of the large intestine and rectum. It may be, as the authors suggested, that the liver cancer cases actually resulted from metastases from the large intestine and/or rectum, as tumors of these sites will frequently metastasize to the liver. The excess in liver cancer mortality and the deficits in mortality that are due to cancer of the large intestine and rectum could also, as the authors indicate, be due to misclassification. Both possibilities have been considered by the investigators in their discussion. Cancer of the lung showed a positive trend with length of membership as well as with latency, although none of the SMRs were statistically significant except for workers without any work histories. The individuals without any work histories may have been the ones who were in their jobs for the longest period of time, because workers without job histories included those7-13 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 DE emissions was very common for this prolonged time, then the increase in lung cancer, as well as stomach cancer, might be linked to DE. Further information on those without work histories should be obtained if possible, because such information may be quite informative with regard to the evaluation of the carcinogenicity of DE. The study design is adequate, covers about a 15-year observation period, has a large enough population, and is appropriately analyzed; however, it has too many limitations to permit any conclusions. First, no exposure histories are available; one has to make do with job histories, which provide limited information on exposure level. Any person who ever worked at the job, or any person working at the same job over any period of time, is included in the same category; this would have a dilution effect, because extremely variable exposures were considered in the study. Second, the length of time worked in any particular job is not available.Third, work histories were not available for 9,707 individuals, who contributed 104,448 personyears, a large proportion of the study cohort (28%). These individuals happen to show the most evidence of a carcinogenic effect. Confounding by alcohol consumption for cancer of the liverand smoking for emphysema and cancer of the lung was not ruled out. Fourth, 15 years’ followup 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 still considered in the denominator of all the analyses. The investigators fail to mention how the person-year calculation for these individuals was handled. Also, some of the person-years might have been overestimated, as people may have paid the dues for a particular year and then left work. These two causes of overestimation of the denominator may have resulted in some or all the SMRs being underestimated.

7.2.1.5. Edling et al. (1987): Mortality Among Personnel Exposed to DE This retrospective cohort mortality study of bus company employees investigated a possible increased mortality of cardiovascular diseases and cancers from DE exposure. The cohort comprised all males employed at five different bus companies in southeastern Sweden between 1950 and 1959. Based on information from personnel registers, individuals were classified into one or more categories and could have contributed person-years at risk in more than one exposure category. The study period was from 1951 to 1983; information was collected from the National Death Registry, and copies of death certificates were obtained from 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 investigators). The cause-, sex-, and age-specific national death rates in Sweden were applied to7-14 the 5-year age categories of person-years of observation to determine expected deaths for all causes, malignant diseases, and cardiovascular diseases. A Poisson distribution was used to
calculate p -values and confidence limits for the ratio of observed to expected deaths. The total cohort of 694 men (after loss of 5 men to follow-up) was divided into three exposure categories: (1) clerks with lowest exposure, (2) bus drivers with moderate exposure, and (3) bus garage workers with highest exposure. The 694 men provided 20,304 person-years of observation, with 195 deaths compared with 237 expected. A deficit in cancer deaths largely accounted for this lower-than-expected mortality in the total cohort. Among subcohorts, no difference between observed and expected deaths for total mortality, total cancers, or cardiovascular causes was observed for clerks (lowest diesel exposure), bus drivers (moderate diesel exposure), and garage workers (high diesel exposure). The risk ratios for all three categories were less than 1 except for cardiovascular diseases among bus drivers, which was 1.1. When the analysis was restricted to members who had at least a 10-year latency period and either any exposure or an exposure exceeding 10 years, similar results were obtained, with fewer neoplasms than expected, whereas cardiovascular diseases showed risk around or slightly above unity. Five lung cancer deaths were observed among bus drivers who had moderate DE exposure, whereas seven were expected. The only other lung cancer death was observed among bus garage workers who had the highest DE exposure. This study’s major limitations, including small size and poor data on DE exposure, make it inadequate to draw any conclusions.

7.2.1.6. Boffetta and Stellman (1988): DE Exposure and Mortality Among Males in the American Cancer Society Prospective Study Boffetta and Stellman conducted a mortality analysis of 461,981 males with known smoking history and vital status at the end of the first 2 years of follow-up. The analysis was restricted to males aged 40 to 79 years in 1982 who enrolled in the American Cancer Society’s prospective mortality study of cancer. Mortality was analyzed in relation to exposure to DE and to employment in selected occupations related to DE exposure. In 1982, more than 77,000 American Cancer Society volunteers enrolled more than 1.2 million men and women from all 50 States, the District of Columbia, and Puerto Rico in a long-term cohort study, the Cancer Prevention Study II (CPS-II). Enrollees were usually friends, neighbors, or relatives of the volunteers; enrollment was by family groups, with at least one person in the household 45 years of age or older. Subjects were asked to fill out a four-page confidential questionnaire and 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; and7-15 information on diet, drinking, smoking, and other habits. The questionnaire also included three questions on occupation: (1) current occupation, (2) last occupation, if retired, and (3) job held for the longest period of time, if different from the other two. Occupations were coded to an ad hoc two-digit classification in 70 categories. Exposures at work or in daily life to any of the 12 groups of substances were also ascertained. These included diesel engine exhausts, asbestos, chemicals/acids/solvents, dyes, formaldehyde, coal or stone dusts, and gasoline exhausts. Volunteers checked whether their enrollees were alive or dead and recorded the date and place of all deaths every other year during the study. Death certificates were then obtained from State health departments and coded by a trained nosologist according to a system based on the ninth
revision of the ICD. The data were analyzed to determine the mortality for all causes and lung cancer in relation to DE exposure, mortality for all causes and lung cancer in relation to employment in selected occupations with high DE exposure, and mortality from other causes in relation to DE exposure. The incidence-density ratio was used as a measure of association, and test-basedconfidence limits were calculated by the Miettinen method. For stratified analysis, the MantelHaenszel method was used for testing linear trends. Although data on 476,648 subjects comprising 939,817 person-years of risk were available for analysis, 3% of the subjects (14,667) had not given any smoking history, and 20% (98,026) did not give information on DE exposure and were therefore excluded from the main DE analysis. Among individuals who had provided DE exposure history, 62,800 were exposed and 307,143 were not exposed. Comparison of the population with known information on DE exposure with the excluded population with no information on DE exposure showed that the mean ages were 54.7 and 57.7 years, the nonsmokers were 72.4% and 73.2%, and the total mortality rates per 1,000 per year were 23.0% and 28.8%, respectively. All-cause mortality was elevated among railroad workers (relative risk [RR] = 1.43, 95% confidence interval [CI] = 1.2, 1.72), heavy equipment operators (RR = 1.7, 95% CI = 1.19, 2.44), miners (RR = 1.34, 95% CI = 1.06, 1.68), and truck drivers (RR = 1.19, 95% CI = 1.07, 1.31). The age-adjusted lung cancer relative risk was elevated significantly (RR = 1.41, 95% CI = 1.19,1.66), which was slightly decreased to 1.31 (95% CI = 1.10, 1.54). For lung cancer 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 were also elevated, but not significantly, for railroad workers (RR = 1.59, 95% CI = 0.94, 2.69) and truck drivers (RR = 1.24, 95% CI = 0.93, 1.66). These risks were calculated with the Mantel-Haenszel method, controlling for age and smoking. Although the relative risk was nonsignificant for truck drivers, a small dose-response effect was observed when duration of DE exposure was examined. For drivers who worked for 1 to 15 years, the relative risk was 0.87,7-16 whereas for drivers who worked for more than 16 years, the relative risk was 1.33 (95% CI = 0.64, 2.75). Relative risks for lung cancer were not presented for other occupations. Mortality analysis for other causes and DE exposure showed a significant excess of deaths ( p <0.05) in the following categories: cerebrovascular disease, arteriosclerosis, pneumonia, influenza, cirrhosis of the liver, and accidents. The main strength of this study is detailed information on smoking. The two main methodologic concerns are the representativeness of the study population and the quality of information on exposure. The sample, though very large, was composed of volunteers. Thus, the cohort was healthier and less frequently exposed to important risk factors such as smoking and alcohol. Self-administered questionnaires were used to obtain data on occupation and DE exposure. None of this information was validated. Nearly 20% of the individuals had an unknown exposure status to DE, and they experienced a higher mortality for all causes and lung cancer than both the DE exposed and unexposed groups. This could have introduced a substantial bias in the estimate of the association. Given that all DE exposure occupations, such as heavy equipment operators, truck drivers, and railroad workers, showed elevated lung cancer risk, this study is suggestive of a causal association. It should be noted that after adjusting for smoking, the RR reduced slightly from 1.41 to 1.31 and remained significant, indicating that observed excess of lung cancer was associated mainly with DE exposure.

7.2.1.7. Garshick et al. (1988): A Retrospective Cohort Study of Lung Cancer and DE Exposure in Railroad Workers An earlier case-control study of lung cancer and DE exposure in U.S. railroad workers by these investigators had demonstrated a relative odds of 1.41 (95% CI = 1.06, 1.88) for lung cancer with 20 years of work in jobs with DE exposure. To confirm these results, a large retrospective cohort mortality study was conducted by the same investigators. Data sources for the study were the work records of the U.S. Railroad Retirement Board (RRB). The cohort was selected based on job titles in 1959, which was the year by which 95% of the locomotives in the United States were diesel powered. DE exposure was considered to be a dichotomous variable depending on yearly job codes between 1959 and death or retirement through 1980. Industrial hygiene evaluations and descriptions of job activities were used to classify jobs as exposed or unexposed to diesel emissions. A questionnaire survey of 534 workers at one of the railroads where workers were asked to indicate the amount of time spent in railroad locations, either near or away from sources of DE, was used to validate this classification. Workers selected for this survey were actively employed at the time of the survey, 40 to 64 years of age, started work between 1939 and 1949 in the job codes sampled in 1959, and eligible for railroad benefits. To qualify for benefits, a worker must have had 10 years or more of service with the railroad and7-17 should not have worked for more than 2 years in a nonrailroad job after leaving railroad work. Workers with recognized asbestos exposure, such as repair of asbestos-insulated steam locomotive boilers, passenger cars, and steam pipes, or railroad building construction and repairs, were excluded from the job categories selected for study. However, a few jobs with some potential for asbestos exposure were included in the cohort, and the analysis was done both ways, with and without them. The death certificates for all subjects identified in 1959 and reported by the RRB to have died through 1980 were searched. Twenty-five percent of them were obtained from the RRB and 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 assigned as a secondary cause of death, so that the ascertainment of all cases was complete. Workers not reported by the RRB to have died by December 31, 1980, were considered to be alive. Deceased workers for whom death certificates had not been obtained or, if obtained, did not indicate cause of death, were assumed to have died of unknown causes. Proportional hazard models were fitted that provided estimates of relative risk for death caused by lung cancer using the partial likelihood method described by Cox, using the time dimension being the time since first entry into the cohort. The model also controlled for the birth year and the calendar time. The 95% confidence intervals were constructed using the asymptotic normality of the estimated regression coefficients of the proportional hazards model. Exposure was analyzed by DE-exposed jobs in 1959 and by cumulative number of years of DE exposure through 1980. Directly standardized rate ratios for deaths from lung cancer were calculated for DE exposed compared with unexposed for each 5-year age group in 1959. The standardized rates were based on the overall 5-year person-year time distribution of individuals in each age group starting in 1959. The only exception to this was between 1979 and 1980,when a 2-year person-year distribution was used. The Mantel-Haenszel analogue for personyear data was used to calculate 95% confidence intervals for the standardized rate ratios. The cohort consisted of 55,407 workers, 19,396 of whom had died by the end of 1980. Death certificates were not available for 11.7% of all deaths. Of the 17,120 deaths for whom death certificates were obtained, 48.4% were attributable to diseases of the circulatory system, whereas 21% were attributable to all neoplasms. Of all neoplasms, 8.7% (1,694 deaths) were due to lung cancer. A higher proportion of workers in the younger age groups, mainly brakemen and conductors, were exposed to DE, while a higher proportion of workers in the older age groups were potentially exposed to asbestos. In a proportional hazards model, analyses by age in 1959 found a relative risk of 1.45 (95% CI = 1.11, 1.89) among the age group 40 to 44 years and a relative risk of 1.33 (95% CI = 1.03, 1.73) for the age group 45 to 49 years. Risk estimates in7-18 the older age groups 50 to 54, 55 to 59, and 60 to 64 years were 1.2, 1.18, and 0.99, respectively, and were not statistically significant. The two youngest age groups in 1959 had workers with the highest prevalence and longest duration of DE exposure and lowest exposure to asbestos. When potential asbestos exposure was considered as a confounding variable in a proportional hazards model, the estimates of relative risk for asbestos exposure were all near null value and not significant. Analysis of workers exposed to DE in 1959 (n = 42,535), excluding workers with potential past exposure to asbestos, yielded relative risks of 1.57 (95% CI = 1.19, 2.06) and 1.34 (95% CI = 1.02, 1.76) in the 1959 age groups 40 to 44 years and 45 to 49 years. Directly standardized rate ratios were also calculated for each 1959 age group based on DE exposure in 1959. The results confirmed those obtained by using the proportional hazards model. Relative risk estimates were then obtained using duration of DE exposure as a surrogate for dose. In a model that used years of exposure up to and including exposure in the year of death, no exposure duration-response relationship was obtained. When analysis was done by disregarding exposure in the year of death and 4 years prior to death, the risk of dying from lung cancer increased with the number of years worked in a diesel-exhaust-exposed job. In this analysis, exposure to DE was analyzed by exposure duration groups and in a model entering age in 1959 as a continuous variable. The workers with greater than 15 years of exposure had a relative risk of lung cancer of 1.72 (95% CI = 1.27, 2.33). The risk for 1 to 4 years of cumulative exposure was 1.20 (95% CI = 1.01, 1.44); for 5 to 9 years of cumulative exposure, it was 1.24 (95% CI = 1.06, 1.44); and for 10 to 14 years of cumulative exposure, it was 1.32 (95% CI = 1.13, 1.56). The results of this study, demonstrating a positive association between DE exposure and increased lung cancer, are consistent with the results of the case-control study 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 epidemiologic studies. The large sample size (55,400) allowed sufficient power to detect small risks and also permitted the exclusion of workers with potential past exposure to asbestos. The stability of job career paths in the cohort ensured that of the workers 40 to 44 years of age in 1959 classified as DE-exposed, 94% of the cases were still in DE-exposed jobs 20 years later. The main limitation of the study is the lack of quantitative data on exposure to DE in either individual workers or overall job categories. This is one of the few studies in which industrial hygiene measurements of DE were done. These measurements were correlated with job titles to divide the cohort in dichotomous exposure groups of exposed and nonexposed. This may have led to an underestimation of the risk of lung cancer because exposed groups included individuals with low to high exposure. The number of years exposed to DE was used as a surrogate for dose. The dose, based on duration of employment, was inaccurate because7-19 individuals were working on steam and diesel locomotives during the transition period. It should be noted that the investigators only included exposures after 1959; the duration of exposure prior to 1959 was not known. If the categories of exposure to DE 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 exposure prior to 1959 and latency. No adjustment for smoking was made in this study. However, an earlier case-control study done in the same cohort (Garshick et al., 1987) showed no significant difference in the risk estimate after adjusting for smoking. Despite these limitations, the results of this study indicate that occupational exposure to DE is associated with a modest risk (1.5) of lung cancer. The data of this study were used by Crump et al. (1991) to explore the development of dose-response-based quantitative estimates of lung cancer associated with DE exposure by using diesel exposure estimate data from the industrial hygiene (IH) studies conducted by Hammond (1998) and Woskie et al. (1988a,b). These studies were conducted in conjunction with the Garshick et al. (1988) study. The Woskie et al. (1988a,b) IH studies were conducted in four small northern railroads where the workers were exposed to DE in the early 1980s, prior to the Garshick et al. (1988) epidemiologic study. A total of 39 job titles were identified by Woskie et al. (1988a,b), which were subsequently combined into 13 job groups and finally merged into 5 career exposure job codes as follows: brakers, conductors, and hostlers; clerks; engineers and firers; signal maintainers; and shop workers. The average exposure estimates were assigned to the cohort members by Crump et al. (1991) based on the job codes in 1959. Cumulative exposures were calculated using these average exposures for each job code. The exposures in the IH study by Hammond (1998) were defined as the concentrations of respirable-sized particles (RSP), the adjusted respirable particles (ARP) concentrations, and the adjusted extractable mass (AEM). The concentrations of ARP were estimated in the IH study by removing the particle contribution of environmental tobacco smoke (ETS). Crump et al. (1991) also used another index called total extractable material (TEX), which was the extractable RSP including the particle contribution of ETS. Using these four exposure indices and the regional climates for the United States, Crump et al. (1991) constructed various exposure metrics. They conducted more than 50 analyses based on calendar year, age in 1959, attained age, and five job codes identified in 1959: brakers, conductors, and hostlers; clerks; engineers and firers; signal maintainers; and shop workers; using the exposure metrics. Crump et al. (1991) used the U.S. general population age- and year-specific death rates for comparison and found that the relative risk can be positively or negatively related to the duration of exposure depending on how age was controlled in a model. Their use of the U.S. general population rates instead of the internal unexposed group of railroad workers that was used by Garshick et al. (1988) identified that the7-20 death ascertainment between 1977 and 1980 as incomplete. The Crump et al. (1991) analysis, limited to 1959 through 1976, found an excess lung cancer risk similar to the subsequent Garshick analysis ( letter from Garshick, Harvard Medical School, to Chao Chen, U.S. EPA, dated August 15, 1991)

. Garshick conducted some additional analyses after confirming the underascertainment of deaths by RRB identified by Crump et al. (1991). He reported that the relationship between years of exposure, when adjusted for attained age and calendar year, was flat to negative depending upon which model was used. He also found that in the years 1977-1980 the death ascertainment was incomplete; approximately 20% to 70% of deaths were missing depending upon the calendar year. Garshick’s analysis, based on job titles in 1959 and limited to deaths occurring through 1976, showed that even though the relative risk for all exposure groups was elevated, the youngest workers still had the highest risk of dying of lung cancer. Crump (1999), on the other hand, reported that the negative dose-response continued to be upheld in his latest analysis when age was controlled more carefully and years of exposure quantified more accurately. Crump (1999) asserted that the negative dose-response trends for lung cancer observed either with the cumulative exposure or with duration of exposure may be due to underascertainment of deaths in the last 4 years of follow-up of the Garshick et al. (1988) study as well as incomplete follow-up in earlier years. California EPA’s (Cal EPA, 1998) Office of Environmental Health Hazard Assessment (OEHHA) used the same railroad worker data for its quantitative risk assessment. The five job categories defined by Woskie et al. (1988a,b) and used by Crump et al. (1991) were combined into three exposure categories: exposed (engineers and firers; brakers, conductors, and hostlers; collectively known as “train workers”), unexposed (clerks and signalmen), and uncertain exposure (shop workers). In its analysis, OEHHA found a positive dose-response and a steadily increasing risk of lung cancer with increasing duration of exposure by using age in 1959 butallowing for an interaction term of age and calendar year in the model. This positive doseresponse finding was contradictory to the negative to flat dose-response findings of both Crump et al. (1991) and Garshick ( letter from Garshick, Harvard Medical School, to Chao Chen, U.S. EPA, dated August 15, 1991 ). The Health Effects Institute (HEI, 1999) convened an expert panel specifically to evaluate strengths and limitations of two epidemiologic studies that had some exposure data, for quantitative risk estimation and to resolve the discrepancies in the dose-response results reported by Garshick et al. (1988), Crump et al. (1991), and OEHHA (Cal EPA, 1998). In their evaluation of the epidemiologic study of railroad worker data for quantitative risk assessment, the panel conducted their own analysis of the Garshick et al. (1988) data. They excluded the last 4 years of follow-up (1977-1980) because of underascertainment of deaths during these years.7-21 The panel categorized the duration of exposure in 12 categories that were basically the duration of employment. The exposure was assumed to be linearly increasing for 15 years prior to 1959. Lags of 5 and 10 years were also considered in the analysis. The job categories based on job held in 1959 were classified as clerks, signalmen, engineers and firers, conductors and brakers, hostlers, and shop workers. For final analysis these were collapsed into three groups: clerks and signalmen, train workers (engineers and firers, conductors and brakers, and hostlers), and shop workers. Seven different models were used. The panel’s analysis revealed consistently elevated lung cancer risk for train workers compared with clerks for each duration of employment (1-4, 5-9, 10-14, 15-17, 18+) in years and that shop workers had an intermediate risk of lung cancer. Their analysis also revealed decreasing risk of lung cancer with increasing duration of employment in all three job categories. These findings were similar to those of Garshick ( letter from Garshick, Harvard Medical School, to Chao Chen, U.S. EPA, dated August 15, 1991 ) and Crump et al. (1991). In addition to differences in adjusting the age (age in 1959 versus attained age) in their respective analyses, these three investigators made different assumptions in estimating exposure patterns in these railroad workers. Garshick et al. (1988) assumed that there was no exposure to DE prior to 1959 and that the exposure to DE was constant throughout the period of follow-up, i. e., 1959 to 1980 (block exposure pattern). Crump et al. (1991) assumed that the exposure to DE increased steadily from 1945 to 1959 to the same level as assumed in the block exposure pattern by Garshick et al. (1988) and then remained constant from 1959 through 1980 (ramp exposure pattern). OEHHA assumed that the exposure increased steeply from 1945 to 1959. The peak exposure attained in 1959 according to OEHHA was twice as high as assumed in the block and ramp exposure patterns by Garshick et al. (1988) and Crump et al. (1991), respectively. The exposures then declined steeply from 1959 to reach the levels assumed in the block and ramp exposure patterns in 1980 (roof exposure pattern). The roof exposure pattern was constructed on the assumption that diesel engines were “smokier” in the past. A detailed discussion of divergent results observed by Crump and Cal EPA can be found in Chapter 8. The panel discussed various possibilities for the negative dose-response found among train workers and to a lesser extent among shop workers. They asserted that several types of biases could affect the data, alone or in combination, and mask a true positive association. The biases enumerated by the panel were: unmeasured confounding by smoking, exposure to other sources of pollution, previous occupational exposures, exposure misclassification, use of “duration of employment” as a surrogate measure for exposure, healthy worker survivor effect, and differential or incomplete ascertainment of lung cancer deaths (for detailed discussion of how an individual bias affects the results, please see HEI, 1999). The panel concluded, “However, despite the reason or reasons why the relative risks in these data decrease with7-22 duration of employment, the lack of a positive exposure-response association in the railroad worker cohort substantially weakens that study’s potential to provide a reliable quantitative estimate of risk of exposure to diesel engine emissions.” Thus, the panel recommended against using the current railroad worker data as the basis for quantitative risk assessment in ambient settings. The panel also reported that the Garshick et al. study (1987, 1988) had several strengths, such as a large number of study subjects (55,407 subjects, including 1,694 lung cancer deaths in the cohort study and 1,256 lung cancer cases for the case-control study). The workers were employed in an industry where many of them were exposed to DE. Confounding by asbestos was handled by either excluding certain job categories from the analyses or controlling for it in the analyses. Confounding by smoking was controlled in the analyses of case-control study. The panel concluded that the overall results of the Garshick studies were generally consistent with findings of a weak association between exposure to DE and occurrence of lung cancer. Thus, it should be noted that although the railroad worker data are unsuitable for quantitative risk assessment, they provide qualitative support for a positive association between exposure to DE and occurrence of lung cancer.

7.2.1.8. Gustavsson et al. (1990): Lung Cancer and Exposure to DE Among Bus Garage Workers A retrospective mortality study (from 1952 to 1986), cancer incidence study (from 1958 to 1984), and nested case-control study were conducted among a cohort of 708 male workers from five bus garages in Stockholm, Sweden, who had worked for at least 6 months between 1945 and 1970. Thirteen individuals were lost to follow-up, reducing the cohort to 695. Information was available on location of workplace, job type, and beginning and ending of work periods. Workers were traced through a computerized register of the living population, death and burial books, and data from the Stockholm city archives. For the cohort mortality analyses, death rates of the general population of greater Stockholm were used. Death rates of occupationally active individuals, a subset of the general population of greater Stockholm, were used as a second comparison group to reduce the bias from “healthy worker effect.” Mortality analysis was conducted using the “occupational mortality analysis program” (OCMAP-PC). For cancer incidence analysis, the “epidemiology in Linköping” (EPILIN) program was used, with the incidence rates obtained from the cancer registry. For the nested case-control study, both dead and incident primary lung cancers identified in the register of cause of deaths and the cancer register were selected. Six controls matched on age ± 2 years, selected from the noncases at the time of the diagnosis of cases, were drawn at7-23 random without replacements. Matched analyses were done to calculate odds ratios using conditional logistic regression. The EGRET and Epilog programs were used for these analyses. DE and asbestos exposure assessments were performed by industrial hygienists based on the intensity of exposure to DE and asbestos, specific for workplace, work task, and calendar time period. A DE exposure assessment was based on (1) amount of emission (number of buses, engine size, running time, and type of fuel), (2) ventilatory equipment and air volume of the garages, and (3) job types and work practices. Based on detailed historical data and very few actual measurements, relative exposures were estimated (these were not absolute exposure levels). The scale was set to 0 for unexposed and 1 for lowest exposure, with each additional unit increase corresponding to a 50% increase in successive intensity (i.e., 1.5, 2.25, 3.38, and 5.06). Based on personal sampling of asbestos during 1987, exposures were estimated and timeweighted annual mean exposures were classified on a scale of three degrees (0, 1, and 2). Cumulative exposures for both DE and asbestos were calculated by multiplying the level of exposure by the duration of every work period. An exposure index was calculated by adding for every individual contribution from all work periods for both DE and asbestos. Four DE index classes were created: 0 to 10, 10 to 20, 20 to 30, and >30. The four asbestos index classes were 0 to 20, 20 to 40, 40 to 60, and >60. The cumulative exposure indices were used for the nested case-control study. Excesses were observed for all cancers and some other site-specific cancers using both comparison populations for the cohort mortality study, but none of them was statistically significant. Based on 17 cases, SMRs for lung cancer were 122 and 115 using Stockholm occupationally active and general population, respectively. No dose-response was observed with increasing cumulative exposure in the mortality study. The cancer incidence study reportedly confirmed the mortality results (results not given). The nested case-control study, on the other hand, showed increasing risk of lung cancer with increasing exposure. Using 0 to 10 DE exposure index as the comparison group 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 = 1.32 to 4.47) for the DE 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 asbestos exposure did not change the lung cancer risk for DE. The main strength of this study is the detailed exposure matrices constructed for both DE and asbestos exposure, although they were based primarily on job tasks and very few actual measurements. There are a few methodological limitations to this study. The cohort is small and there were only 17 lung cancer deaths; thus the power is low. Exposure or outcome may be misclassified, although any resulting bias in the relative risk estimates is likely to be toward7-24 unity, because exposure classification was done independently of the outcome. Although the analysis by dose indices was done, no latency analysis was performed. Although data on smoking were missing, it is unlikely to confound the results because this is a nested case-control study; therefore, smoking is not likely to be different among the individuals irrespective of their exposure status to DE. Overall, this study provides some support to the excess lung cancer results found earlier among populations exposed to DE.

7.2.1.9. Hansen (1993): A Follow-up 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 group included all truck drivers employed in the road delivery or long-haul business (14,225). The unexposed group included all laborers in certain selected occupational groups considered tobe unexposed to fossil fuel combustion products and to resemble truck drivers in terms of workrelated physical demands and various personal background characteristics (43,024). Through automatic record linkage between the 1970 census register (the Central Population Register 1970 to 1980) and the Death Certificate Register (1970 to 1980), the population was followed for cause-specific mortality or emigration up to November 9, 1980. Expected number of deaths among truck drivers was calculated by using the 5-year age group and 5-year time period death rates of the unexposed group and applying them to the person-years accumulated by truck drivers. ICD Revision 8 was used to code the underlying cause of death. Test-based CIs were calculated using Miettinen’s method. A Poisson distribution was assumedfor the smaller numbers, and CI was calculated based on exact Poisson distribution (CibaGeigy). Total person-years accrued by truck drivers were 138,302, whereas for the unexposed population, they were 407,780. There were 627 deaths among truck drivers and 3,811 deaths in the unexposed group. Statistically significant excesses were observed for all cancer mortality (SMR = 121, 95% CI = 104 to 140); cancer of respiratory organs (SMR = 160, 95% CI = 128 to 198), which was due mainly to cancer of bronchus and lung (SMR = 160, 95% CI = 126 to 200); and multiple myeloma (SMR = 439, 95% CI = 142 to 1,024). When lung cancer mortality was further explored by age groups, excesses were observed in most age groups (30 to 39, 45 to 49, 50 to 54, 55 to 59, 60 to 64, and 65 to 74), but there were small numbers of deaths in each group when stratified by age, and the excesses were statistically significant for the 55 to 59 (SMR = 229, O = 19, 95% CI = 138 to 358) and 60 to 64 (SMR = 227, O = 22, 95% CI = 142 to 344) age groups only. As acknowledged by the author, the study has quite a few methodologic limitations. The exposure to DE is assumed in truck drivers based on use of diesel-powered trucks, but no validation of qualitative or quantitative exposure is attempted. It is also not known whether any7-25 of these truck drivers or any other laborers had changed jobs after the census of November 9, 1970, thus creating potential misclassification bias in exposure to DE. The truck drivers and the unexposed laborers were from the same socioeconomic class and may have the same smoking habits. Still, the lack of information on smoking data and a 36% rural population (usually consuming less tobacco) in the unexposed group may potentially confound the lung cancer results. However, a population survey carried out in 1988 showed very little difference in smoking habits of residents of rural areas and the total Danish male population. The investigator reports that diesel trucks were introduced in Denmark after World War II, and since the late 1940s the majority of the Danish fleet has been composed of diesel trucks. Consequently, even though the follow-up period is relatively short, the truck drivers may have had exposure to DE 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.

7.2.1.10. Saverin et al. (1999): DE and Lung Cancer Mortality in Potash Mining This is a cohort mortality study conducted in male potash miners in Germany. The mines began using mobile diesel-powered vehicles in 1969 and 1970. Miners who had worked underground for at least 1 year after 1969 to 1991, when the mines were closed, were followed from 1970 to 1994. A total of 5,981 individuals were identified from the medical records by a team of medical personnel familiar with the mining technology. A total of 5,536 were eligible for follow-up after 5.5% were excluded due to implausible or incomplete work history and 1.9% were lost to follow-up. A subcohort of 3,258 miners who had worked for at least 10 years underground (80% had held a single job) was also identified. The miners’ biannual medical examination records were used to extract the information about personal data, smoking data, and pre-mining occupation, and to reconstruct a chronology of workplaces occupied by the worker since hire for each person. 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 all workplaces was obtained. Most were personal dust samples; some were area dust samples. Cumulative exposure was calculated for each miner, for each year of observation, using the work chronology and the work category. For the workshop category years of employment were considered as exposure time; for production and maintenance years of employment was weighted by a factor of 5/8, since these workers for an 8-hour shift worked for only 5 hours underground. As neither the mining technology nor the type of machinery used had changed substantially from 1970 to 1992, the exposure measurements were considered to represent the exposures throughout the study period. Accrued person-years were classified into cumulative7-26 exposures and were expressed in intervals of 0.5 ymg/m 3 . Both the exposure data and the smoking data obtained from the medical files were validated by personal interviews with 1,702
cohort members. Death certificates were obtained from local health centers for 94.4% of deceased members. Autopsy data were available for 13% of the deceased. Internal comparison was done between production and workshop categories. Using East German general male population rates, SMRs were computed for the total cohort as well as the subcohort. Analyses were done using Poisson and Cox regression models. The concentrations of total carbon for production, maintenance, and workshop categories were 0.39 mg/m 3 , 0.23 mg/m 3 , and 0.12 mg/m 3 , respectively. The cumulative exposure ranged from 0.25 ymg/m 3 to 6.25 ymg/m 3 . The regression analysis showed that the cohort’s smoking habits were homogenous and that smoking had an even distribution over cumulative exposure. A total of 424 deaths were observed for the entire cohort (SMR = 54). The all-cancer deaths were 133, of which 38 were from lung cancer (SMR = 78). Analysis for the subcohort using the internal comparison group of low exposure (workshop category, mean cumulative exposure = 2.12 ymg/m 3 ) RR of 2.17 (95% CI = 0.79, 5.99) was found for the production category (mean cumulative exposure = 4.38 ymg/m 3 ). The relative risks for lung cancer for 20 years of exposure in the production category (highest exposure = cumulative exposure of 4.9 ymg/m 3 ) were calculated using Poisson and Cox regression methods. RRs of 1.16 and 1.68 were observed for the total cohort, while RRs of 1.89 and 2.7 were observed for the subcohort by Poisson and Cox regression methods respectively. The main strengths of the study are the information available on DE exposure and smoking. Although these potash miners were exposed to salt dust and nitric gases, exposures to other confounders such as heavy metals and radon were absent. Smoking does not seem to be a confounder in this study but cannot be completely ruled out. Unfortunately, the age distribution of the cohort is not available. Since there were only 424 deaths in 25 years of follow-up in this cohort of 5,536, it appears that the cohort is young. Although lung cancer risk was elevated by twofold in the production category of the subcohort of miners who had worked for at least 10 years underground at the same job for 80% of their time and did not have more than 3 jobs, it was not statistically significant. The follow-up period for this study was 25 years, but the cohort members could have entered the cohort any time between 1970 and 1990, as long as they worked underground for a year, i.e., they could have worked in the mines for 1 year to 21 years. Thus, the authors may not have had enough follow-up or latency to observe the lung cancer excess. Despite these limitations, the results of this study provide suggestive evidence for the causal association between DE and excess lung cancer.
Table 7-1 summarizes the above cohort studies.Table 7-1. Epidemiologic studies of the health effects of exposure to DE: cohort mortality studiesAuthors Population studied DEexposure assessment Results LimitationsWaller(1981)Approximately 20,000 maleLondon transportation workersAged 45 to 64 years25 years follow-up (1950-1974)Five job categories used todefine exposureEnvironmental B[a]Pconcentrations measured in1957 and 1979SMR = 79 for lung cancer for thetotal cohortSMRs for all five job categorieswere less than 100 for lungcancerExposure measurement of B[a]Pshowed very little difference betweeninside and outside the garageIncomplete information on cohortmembersNo adjustment for confounding suchas other exposures, cigarettesmoking, etc.No latency analysisHowe et al.(1983)43,826 male pensioners of theCanadian National RailwayCompanyMortality between 1965 and1977 among these pensionerswas compared with mortalityof general Canadian populationExposure groupsclassified by a groupof experts based onoccupation at the timeof retirementThree exposure groups:NonexposedPossibly exposedProbably exposedRR = 1.2 (p=0.013) andRR = 1.3 (p=0.001) for lungcancer for possible and probableexposure, respectivelyA highly significantdose-response relationshipdemonstrated by trendtest (p<0.001)Incomplete exposure assessment dueto lack of lifetime occupationalhistoryMixed exposures to coaldust/combustion products and DENo validation of method was used tocategorize exposureLack of data on smoking but use ofinternal comparison group tocompute RRs minimizes the potentialconfounding by smokingNo latency analysis

7-27Table 7-1. Epidemiologic studies of the health effects of exposure to DE: cohort mortality studies(continued)Authors Population studied DEexposure assessment Results LimitationsRushtonet al. (1983)8,490 male London transportmaintenance workersMortality of workers employedfor 1 continuous year betweenJanuary 1, 1967, and December31, 1975, was compared withmortality of general populationof England and Wales100 different job titles weregrouped in20 broad categoriesThe categories were notranked for DE exposureSMR = 133 (p<0.03) for lungcancer in the general hand jobgroupSeveral other jobcategories showed SS increasedSMRs for several other sitesbased on fewer than five casesIll-defined DE exposure without anyrankingAverage 6-year follow-up i.e., notenough time for lung cancer latencyNo adjustment for confoundersWong et al.(1985)34,156 male heavy constructionequipment operatorsMembers of the local union forat least 1 year betweenJanuary 1, 1964, and December 1,197820 functional job titlesgrouped into three jobcategories for potentialexposureExposure groups (high,low, and unknown) basedon job description andproximity to source of DEemissionsSMR = 166 (p<0.05) for livercancer for total cohortSMR = 343 (observed = 5,p<0.05) for lung cancer for highexposure bulldozer operatorswith 15-19 years of membership,20+ years of follow-upSMR = 119 (observed = 141,p<0.01) for workers with nowork historiesNo validation of exposure categories,which were based on surrogateinformationIncomplete employment recordsEmployment history other than fromthe union not available15 year follow-up may not providesufficient time for lung cancerlatencyNo data on confounders such as otherexposures, alcohol, smoking, etc.Edling et al.(1987)694 male bus garage employeesFollow-up from 1951 through1983Mortality of these men wascompared with mortality ofgeneral population of SwedenThree exposure groupsbased on job titles:High exposure, busgarage workersIntermediate exposure,bus driversLow exposure, clerksNo SS differences were observedbetween observed and expectedfor any cancers by differentexposure groupsSmall sample sizeNo validation of exposureNo data on confounders such as otherexposures, smoking, etc.

7-28Table 7-1. Epidemiologic studies of the health effects of exposure to DE: cohort mortality studies(continued)Authors Population studied DEexposure assessment Results LimitationsBoffetta andStellman(1988)46,981 male volunteers enrolledin the American Cancer Society’sProspective Mortality Study ofCancer in 1982Aged 40 to 79 years at enrollmentFirst 2-year follow-upSelf-reported occupationswere coded into 70 jobcategoriesEmployment in high DEexposure jobs werecompared with nonexposedjobsTotal mortality (SS) elevated forrailroad workers (RR=1.43),heavy equipment operators(RR=1.7), miners (RR=1.34),and truck drivers (RR=1.19)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 workersand truck driversTruck drivers also showed adose-responseExposure information based on selfreported occupation for which novalidation was doneVolunteer population, probablyhealthy population

7-29Table 7-1. Epidemiologic studies of the health effects of exposure to DE: cohort mortality studies(continued)Authors Population studied DEexposure assessment Results LimitationsGarshicket al. (1988)Garshick (ltrto ChaoChen, EPA,dtd 8/15/9155,407 white male railroadworkersAged 40 to 64 years in 1959Started work 10-20 years earlierthan 1959Industrial hygiene datacorrelated with job titles todichotomize the jobs as“exposed” or “not exposed”RR = 1.45 (40-44 year agegroup)RR = 1.33 (45-49 year agegroup)Both SSAfter exclusion of workersexposed to asbestosRR = 1.57 (40-44 year agegroup)RR = 1.34 (45-49 year agegroup)Both SSDose response indicated byincreasing lung cancer risk withincreasing cumulative exposureFurther analysis using attainedage, limited through 1976showed youngest workers stillhad the highest riskYears of exposure used as surrogatefor doseNot possible to separate the effect oftime since first exposure and durationof exposureLack of smoking data but casecontrol study showed very littledifference between those exposed toDE versus those who were notCrump et al.(1991)Crump(1999)Reanalysis of Garshick et al.,1988 dataDose response found to bepositive or negative dependingupon how the age was controlledin the modelNegative dose-response upheldin the latest analysisCaliforniaEPA (1998)Reanalysis of Garshick et al.,1988Positive dose response using ageat 1959 and interaction term ofage & calendar year

7-30Table 7-1. Epidemiologic studies of the health effects of exposure to DE: cohort mortality studies(continued)Authors Population studied DEexposure assessment Results Limitations

7-31Gustavssonet al. (1990)695 male workers from 5 busgarages in Stockholm, Sweden,who had worked for 6 monthsbetween 1945 and 197034 years follow-up (1952-1986)Nested case-control study17 cases, six controls for eachcase matched on age ± 2 yearsFour DE indices werecreated:0 to 10, 10 to 20, 20-30,and >30 based on job tasksand duration of workSNS SMRs of 122 and 115 (OAand GP), respectivelyCase-control study resultsshowed dose response:RR = 1.34 (10 to 20)RR = 1.81 (20 to 30)RR = 2.43 (>30)All SS with 0-10 as comparisongroupExposure matrix based on job tasks(not on actual measurements)Small cohort, hence low powerLack of smoking data is unlikely toconfound the results since it is anested case-control studyHansen(1993)Cohort of 57,249 unskilledlaborers, ages 15 to 74, inDenmark (nationwide census file)November 9, 1970Follow-up through November 9,1980DE exposure assumedbased on diesel-poweredtrucksSS SMRs for lung cancer :SMR = 160 for total populationSMR = 229 for age 55-59 yearsSMR = 227 for age 60-64 yearsNo actual exposure data availableLack of smoking data but populationsurvey showed very little differencebetween rural and urban smokinghabitsJob changes may have occurred fromlaborer to driverShort follow-up periodSaverin etal. (1999)Cohort of 5,536 potash minerswho had worked underground forat least 1 year after 1969Subcohort of 3,258 who hadworked for at least 10 yearsundergroundFollow-up from 1970 to 1994DE exposure categoriesdefined as:production (high)maintenance (medium)workshop (low)225 air samples obtained:for total carbon, organics,& fine dust in 1992SNS increased RRs adjusted forsmoking: 1.68 and 2.7 for totalcohort & subcohort, respectivelySmall, young cohortFew deathsNo latency analysisAbbreviations: RR = relative risk; SMR = standardized mortality ratio; SNS = statistically nonsignificant; SS = statisticallysignificant;O = occupationally active; GP = general population.7-32 7.2.2. Case-Control Studies of Lung Cancer
7.2.2.1. Hall and Wynder (1984): A Case-Control Study of DE Exposure and Lung Cancer Hall and Wynder (1984) conducted a case-control study of 502 male lung cancer cases
and 502 controls without tobacco-related diseases that examined an association between occupational DE exposure and lung cancer. Histologically confirmed primary lung 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 without tobacco-related diseases, were matched to cases by age (± 5 years), race, hospital, and hospital room status. The number of male lung cancer cases interviewed totaled 502, which was 64% of those who met the study criteria for eligibility. Of the remaining 36%, 8% refused, 21% were too ill or had died, and 7% were unreliable. Seventy-five percent of eligible controls completed interviews. Of these interviewed controls, 49.9% were from the all-cancers category, whereas 50.1% were from the all-noncancers category. All interviews were obtained in hospitals to gather detailed information on smoking history, coffee consumption, artificial sweetener use, residential history, and abbreviated medical history as well as standard demographic variables. Occupational information was elicited by a question on the usual lifetime occupation and was coded by the abbreviated list of the U.S. Bureau of Census Codes. The odds ratios were calculated to evaluate the association between DE exposure and risk of lung cancer incidence. Summary odds ratios were computed by the Mantel-Haenszel method after adjusting for potential confounding by age, smoking, and socioeconomic class. Two-sided, 95% confidence intervals were computed by Woolf’s method. Occupational exposure to DE 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 workers, and heavy equipment operators and repairmen. The second method used the National Institute for Occupational Safety and Health (NIOSH) criteria to analyze occupations by diesel exposure. In this method, the estimated proportion of exposed workers was computed for each occupational category by using the NIOSH estimates of the exposed population as the numerator and the estimates of individuals employed in each occupational category from the 1970 census as the denominator. Occupations estimated to have at least 20% of their employees exposed to DE were defined as “high exposure,” those with 10% to 19% of their employees exposed were defined as “moderate exposure,” and those with less than 10% of their employees exposed were defined as “low exposure.” Cases and controls were compared with respect to exposure. The relative risk was 2.0 (95% CI = 1.2, 3.2) for those workers who were exposed to DE versus those who were not. The risk, however, decreased to a nonsignificant 1.4 when the data were adjusted for smoking.7-33 Analysis by NIOSH criteria found a nonsignificant relative risk of 1.7 in the high- 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 DE exposure and smoking, the lung cancer risks were calculated for different smoking categories. The relative risks were 1.46 among nonsmokers and ex-smokers, 0.82 among current smokers of <20 cigarettes/day, and 1.3 among current smokers of 20+ cigarettes/day, indicating a lack of synergistic effects. The major strength of this study is the availability of a detailed smoking history for all the study subjects. However, this is offset by lack of DE exposure measurements, use of a poor surrogate for exposure, and lack of consideration of latency period. Information was collected on only one major lifetime occupation, and it is likely that those workers who had more than one major job may not have reported the occupation with the heaviest DE exposures. Furthermore, the exposure categories based on job titles were broad, and thus would have made a true effect of
DE difficult to detect.

7.2.2.2. Damber and Larsson (1987): Occupation and Male Lung Cancer, a Case-Control Study in Northern Sweden A case-control study of lung cancer was conducted in northern Sweden to determine the occupational risk factors that could explain the large geographic variations of lung cancer incidence in that country. The study region comprised the three northernmost counties of Sweden, with a total male population of about 390,000. The rural municipalities, with 15% to 20% of the total population, have forestry and agriculture as dominating industries, and the urban areas have a variety of industrial activities (mines, smelters, steel factories, paper mills, and mechanical workshops). All male cases of lung cancer reported to the Swedish Cancer Registry during the 6-year period between 1972 and 1977 who had died before the start of the study were selected. Of 604 eligible cases, 5 did not have microscopic confirmation, and in another 5 the diagnosis was doubtful, but these cases were included nevertheless. Cases were classified as small-cell carcinomas, squamous cell carcinomas, adenocarcinomas, and other types. For each case a dead control was drawn from the National Death Registry matched by sex, year of death, age, and municipality. Deaths in controls classified as lung cancer and suicides were excluded. A living control matched to the case by sex, year of birth, and municipality was also drawn from the National Population Registry. Postal questionnaires were sent to close relatives of cases and dead controls, and to living controls themselves to collect data on occupation, employment, and smoking habits. Replies were received from 589 cases (98%), 582 surrogates of dead controls (96%), and 453 living controls (97%).7-34 Occupational data were collected on occupations or employment held for at least 1 year and included type of industry, company name, task, and duration of employment. Supplementary telephone interviews were performed if occupational data were lacking for any period between age 20 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 analyses were performed with dead and living controls, and on the whole there was good agreement between the two control groups. A person who had been active for at least 1 year in a specific occupation was in the analysis assigned to that occupation. Using dead controls, the odds ratios adjusted for smoking were 1.0 (95% CI = 0.7, 1.5) and 2.7 (95% CI = 1.0, 8.1) for professional drivers ( 

### $ 1 year of employment) and underground miners ( 

### $ 1 year of employment), respectively. For 20 or more years of employment in those occupations, the odds ratios adjusted for smoking were 1.2 (95% CI = 0.9, 2.6) and 9.8 (95% CI = 1.5, 414). These were the only two occupations listed with potential DE exposure. An excess significant risk was detected for copper smelter workers, plumbers, electricians, and asbestos workers, as well as concrete and asphalt workers. All the odds ratios were calculated by adjusting for age, smoking, and municipality. A comparison with the live controls resulted in the odds ratios being lower than those observed with dead controls, and none were statistically significant in this comparison. This study did not detect any excess risk of lung cancer for professional drivers, who, among all the occupations listed, had the most potential for exposure to motor vehicle exhaust. However, it is not known whether these drivers were exposed exclusively to gasoline exhaust, DE, or varying degrees of both. An excess risk was detected for underground miners, but it is not known if this was due to diesel emissions from engines or from radon daughters in poorly ventilated mines. Although a high response rate (98%) was obtained by the postal questionnaires, the use of surrogate respondents is known to lead to misclassification errors that
can bias the results in either direction.

7.2.2.3. Lerchen et al. (1987): Lung Cancer and Occupation in New Mexico This is a population-based case-control study conducted in New Mexico that examined the association between occupation and occurrence of lung cancer in Hispanic and non-Hispanic whites. Cases involved residents of New Mexico, 25 through 84 years of age, and diagnosed between January 1, 1980, and December 31, 1982, with primary lung cancer, excluding bronchioalveolar carcinoma. Cases were ascertained through the New Mexico Tumor Registry, which is a member of the Surveillance Epidemiology and End Results (SEER) Program of the National Cancer Institute. Controls were chosen by randomly selecting residential telephone7-35 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 females) and 771 controls (499 males and 272 females) were interviewed, with a nonresponse rate of 11% for cases. Next of kin provided interviews for 50% and 43% of male and female cases, respectively. Among controls, only 2% of the interviews were provided by next of kin for each sex. Data were collected by personal interviews conducted by bilingual interviewers in the participants’ homes. A lifetime occupational history and a self-reported history of exposure to specific agents were obtained for each job held for at least 6 months since age 12. Questions were asked about the title of the position, duties performed, location and nature of industry, and time at each job title. A detailed smoking history was also obtained. The variables on occupational exposures were coded according to the Standard Industrial Classification scheme by a single person and reviewed by another. To test the hypothesis about high-risk jobs for lung cancer, the principal investigator created an a priori listing of suspected occupations and industries by a two-step process involving a literature review for implicated industries and occupations. The principal investigator also determined the appropriate Standard Industrial Classification and Standard Occupational Codes associated with job titles. For four agents—asbestos, wood dust, DE, and formaldehyde—the industries and occupations determined to have exposure were identified, and linking of specific industries and occupations was based on literature review and consultation with local industrial hygienists. The relative odds were calculated for suspect occupations and industries, classifying individuals as ever employed for at least 1 year in an industry or occupation and defining the reference group as those subjects never employed in that particular industry or occupation. Multiple logistic regression models were used to control simultaneously for age, ethnicity, and smoking status. For occupations with potential DE exposure, the analysis showed no excess risks for diesel engine mechanics and auto mechanics. Similarly, when analyzed by exposure to specific agents, the odds ratio (OR) adjusted for age, smoking, and ethnicity was not elevated for DE fumes (OR = 0.6, 95% CI = 0.2, 1.6). Significantly elevated ORs were found for uranium miners (OR = 2.8), underground miners (OR = 2.4), construction workers, and welders (OR = 4.3). No excess risks were detected for the following industries: shipbuilding, petroleum refining, printing, blast furnace, and steel mills. No excess risks were detected for the following occupations: construction workers, painters, plumbers, paving equipment operators, roofers, engineers and firemen, woodworkers, and shipyard workers. Females were excluded from detailed analysis because none of the Hispanic female controls had been employed in high-risk jobs; among the non-Hispanic white controls, employment in a high-risk job was recorded for at7-36 least five controls for only two industries, construction and painting, for which the OR were not significantly elevated. Therefore, the analyses were presented for males only. 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, southwestern Hispanic and non-Hispanic white males. The major limitations pertain to the occupational exposure data. Job titles obtained from occupational histories were used as proxy for exposure status, but these were not validated. Further, for nearly half the cases, next of kin provided occupational histories. The authors acknowledge the above sources of bias but state without substantiation that these biases would not strongly affect their results. They also did not use a job exposure matrix to link occupations to exposures and did not provide details on the method they used to classify individuals as DE exposed based on reported occupations. The observed absence of an association for exposure to asbestos, a well-established lung carcinogen, may be explained by the misclassification errors in exposure status or by sample size constraints (not enough power). Likewise, the association for DE reported by only 7 cases and 17 controls also may have gone undetected because of low power. In conclusion, there is insufficient evidence from this study to confirm or refute an association between lung cancer and DE

exposure.

7.2.2.4. Garshick et al. (1987): A Case-Control Study of Lung Cancer and DE Exposure in
Railroad Workers An earlier pilot study of the mortality of railroad workers by the same investigators (Schenker et al., 1984) found a moderately high risk of lung cancer among workers exposed to DE compared with those who were not. Based on these findings the investigators conducted acase-control study of lung cancer in the same population. The population base for this casecontrol study was approximately 650,000 active and retired male U.S. railroad workers with 10 years or more of railroad service who were born in 1900 or later. The U.S. Railroad Retirement Board (RRB), which operates the retirement system, is separate from the Social Security System, and to qualify for the retirement or survivor benefits the workers had to acquire 10 years or more of service. Information on deaths that occurred between March 1, 1981, and February 28, 1982, was obtained from the RRB. For 75% of the deceased population, death certificates were obtained from the RRB, and, for the remaining 25%, they were obtained from the appropriate State departments of health. Cause of death was coded according to the eighth revision of the ICD. The cases were selected from deaths with primary lung cancer, which was the underlying
cause of death in most cases. Each case was matched to two deceased controls whose dates of birth were within 2.5 years of the date of birth of the case and whose dates of death were within 31 days of the date of death noted in the case. Controls were selected randomly from workers7-37 who did not have cancer noted anywhere on their death certificates and who did not die of
suicide or of accidental or unknown causes. Each subject’s work history was determined from a yearly job report filed by his employer with the RRB from 1959 until death or retirement. The year 1959 was chosen as the effective start of DE exposure for this study since by this time 95% of the locomotives in the United States were diesel powered. Investigators acknowledge that because the transition to diesel-powered engines took place in the early 1950s, some workers had additional exposure prior to 1959; however, if a worker had died or retired prior to 1959, he was considered unexposed. Exposure to DE was considered to be dichotomous for this study, which was assigned based on an industrial hygiene evaluation of jobs and work areas. Selected jobs with and without regular DE exposure were identified by a review of job title and duties. Personal exposure was assessed in 39 job categories representative of workers with and without DE exposure. Those jobs for which no personal sampling was done were considered exposed or unexposed based on similarities in job activities and work locations and by degree of contact with diesel equipment. Asbestos exposure was categorized based on jobs held in 1959, or on the last job held if the subject retired before 1959. Asbestos exposure in railroads occurred primarily during the steam engine era and was related mostly to the repair of locomotive steam boilers that were insulated with asbestos. Smoking history information was obtained from the
next of kin. Death certificates were obtained for approximately 87% of the 15,059 deaths reported by the RRB, from which 1,374 cases of lung cancer were identified. Fifty-five cases of lung cancer were excluded from the study for either incomplete data (20) or refusal by two States to use information on death certificates to contact the next of kin. Successful matching to at least one control with work histories was achieved for 335 (96%) cases 

### # 64 years of age at death and 921 (95%) cases 

### $ 65 years of age at death. In both age groups, 90% of the cases were matched with two controls. There were 2,385 controls in the study; 98% were matched within ± 31 days of the date of death, whereas the remaining 2% were matched within 100 days. Deaths from diseases of the circulatory system predominated among controls. Among the younger workers, approximately 60% had exposure to DE, whereas among older workers, only 47% were exposed
to DE. Analysis by a regression model, in which years of DE exposure were the sum total of the number of years in diesel-exposed jobs, used as a continuous exposure variable, yielded an odds ratio of lung cancer of 1.39 (95% CI = 1.05, 1.83) for >20 years of DE exposure in the 

### # 64 years of age group. After adjustment for asbestos exposure and lifetime smoking (pack-years), the odds ratio was 1.41 (95% CI = 1.06, 1.88). Both crude odds ratio and asbestos exposure as well as lifetime smoking-adjusted odds ratio for the 

### $ 65 years of age group were not significant.7-38 Increasing years of DE exposure, categorized as 

### $ 20 diesel years and 5 to 19 diesel years, with 0 to 4 years as the referent group, showed significantly increased risk in the 

### # 64 years of age group after adjusting for asbestos exposure and pack-year category of smoking. For individuals
who had 

### $ 20 years of DE exposure, the odds ratio was 1.64 (95% CI = 1.18, 2.29), whereas among individuals who had 5 to 19 years of DE exposure, the odds ratio was 1.02 (95% CI = 0.72, 1.45). In the 

### $ 65 years of age group, only 3% of the workers were exposed to DE for more than 20 years. Relative odds for 5 to 19 years and 

### $ 20 years of diesel exposure were less than 1 ( p >0.01) after adjusting for smoking and asbestos exposure. Alternative models to explain past asbestos exposure were tested. These were variables for regular and intermittent exposure groups and an estimate of years of exposure based on estimated years worked prior to 1959. No differences in results were seen. The interactions between DE exposure and the three pack-year categories (<50, >50, and missing pack-years) were explored. The cross-product terms were not significant. A model was also tested that excluded recent DE exposure occurring within the 5 years before death and gave 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. The many strengths of the study are consideration of confounding factors such as asbestos exposure and smoking; classification of DE exposures by job titles and industrial hygiene sampling; exploration of interactions between smoking, asbestos exposure, and DE exposure; and good ascertainment (87%) of death certificates from the 15,059 deaths reported by
the RRB. The investigators also recognized and reported the following limitations: overestimation of cigarette consumption by surrogate respondents, which may have exaggerated the contribution of smoking to lung cancer risk, and use of the Interstate Commerce Commission (ICC) job classification as a surrogate for exposure, which may have led to misclassification of DE exposure jobs with low intensity and intermittent exposure, such as railroad police and bus drivers, as unexposed. These two limitations would result in underestimation of the lung cancer risk. This source of error could have been avoided if DE exposures were categorized by a specific dose range associated with a job title that could have been classified as heavy, medium, low, and zero exposure instead of a dichotomous variable. The use of death certificates to identify cases and controls may have resulted in misclassification. Controls may have had undiagnosed primary lung cancer, and lung cancer cases might have been secondary lesions misdiagnosed as primary lung cancer. However, the investigators quote a third National Cancer Survey report in which the death certificates for lung cancer were coded appropriately in 95% of the cases. Last, as in all previous studies, there is a lack of data on the contribution of unknown7-39 occupational or environmental exposures and passive smoking. In conclusion, this study provides strong evidence that occupational DE emission exposure increases the risk of lung

cancer.

7.2.2.5. Benhamou et al. (1988): Occupational Risk Factors of Lung Cancer in a French Case-Control Study This is a case-control study of 1,625 histologically confirmed cases of lung cancer and 3,091 matched controls, conducted in France between 1976 and 1980. This study was part of an international study to investigate the role of smoking and lung cancer. Each case was matched with one or two controls, whose diseases were not related, to tobacco use, sex, age at diagnosis (± 5 years), hospital of admission, and interviewer. Information was obtained from both cases and controls on place of residence since birth, educational level, smoking, and drinking habits. A complete lifetime occupational history was obtained by asking participants to give their occupations from the most recent to the first. Women were excluded because most of them had listed no occupation. Men who smoked cigars and pipes were excluded because there were very few in this category. Thus, the study was restricted to nonsmokers and cigarette smokers. Cigarette smoking exposure was defined by age at the first cigarette (nonsmokers, 

### # 20 years, or >20 years), daily consumption of cigarettes (nonsmokers, <20 cigarettes a day, and 

### $ 20 cigarettes a day), and duration of cigarette smoking (nonsmokers, <35 years, and 

### $ 35 years). The data on occupations were coded by a panel of experts according to their own chemical or physical exposure determinations. Occupations were recorded blindly using the International Standard Classification of Occupations. Data on 1,260 cases and 2,084 controls were available for analysis. The remaining 365 cases and 1,007 controls were excluded because they did not satisfy the required smoking status criteria. A matched logistic regression analysis was performed to estimate the effect of each occupational exposure after adjusting for cigarette status. Matched relative risk ratios were calculated for each occupation with the baseline category, which consisted of patients who had never been engaged in that particular occupation. The matched RR ratios, adjusted for cigarette smoking for the major groups of occupations, showed that the risks were significantly higher for production and related workers, transport equipment operators, and laborers (RR = 1.24, 95% CI = 1.04, 1.47). On further analysis of this group, for occupations with potential diesel emission 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 with smoking status was found in any of the occupations. The only other significant excess was observed for miners and quarrymen (RR = 2.14, 95% CI = 1.07, 4.31). None of the significant associations showed a dose-response relationship with duration of exposure.7-40 This study was designed primarily to investigate the relationship between smoking (not occupations or environmental exposures) and lung cancer. Although an attempt was made to obtain complete occupational histories, the authors did not clarify whether, in the logistic regression analysis, they used the subjects’ first occupation, predominant occupation, last occupation, or ever worked in that occupation as the risk factor of interest. The most important limitation of this study is that the occupations were not coded into exposures for different chemical and physical agents, thus precluding the calculation of relative risks for diesel exposure. Using occupations as surrogate measures of diesel exposure, an excess significant risk was obtained for motor vehicle drivers and transport equipment operators, but not for motor mechanics. However, it is not known if subjects in these occupations worked with diesel engines or nondiesel engines.

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 case-control studies that examine in detail the association between employment in motor exhaust-related (MER) occupations and lung cancer risk adjusted for confounding by smoking and other risk factors. The three studies were carried out by the National Cancer Institute in Florida (1976 to 1979), New Jersey (1980 to 1981), and Louisiana (1979 to 1983). These three studies were selected because the combined group would provide a sufficient sample to detect a risk of lung cancer in excess of 50% among workers in MER occupations. The analyses were restricted to males who had given occupational history. The Florida study was hospital based, with cases ascertained through death certificates. Controls were randomly selected from hospital records and death certificates, excluding psychiatric diseases, matched by age and county. The New Jersey study was population based, with cases ascertained through hospital records, cancer registry, and death certificates. Controls were selected from among the pool of New Jersey licensed drivers and death certificates. The Louisiana study was hospital based (it is not specified how the cases were ascertained), and controls were randomly selected from hospital 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,including the industry, occupation, and number of years employed. The proportion of next-ofkin interviews varied by site from 50% in Louisiana to 85% in Florida. The coding schemes were reviewed to identify MER occupations, which included truck drivers and heavy equipment operators (cranes, bulldozers, and graders); bus drivers, taxi drivers, chauffeurs, and other motor vehicle drivers; and automobile and truck mechanics. Truck drivers were classified as routemen and delivery men and other truck drivers. All jobs were also classified with respect to potential7-41 exposure to known and suspected lung carcinogens. ORs were calculated by the maximum likelihood method, adjusting for age by birth year, usual amount smoked, and study area. Logistic regression models were used to examine the interrelationship of multiple variables. A statistically significant excess risk was detected for employment of 10 years or more for all MER occupations (except truck drivers) adjusted for birth cohort, usual daily cigarette use, and study area. The odds ratio for lung cancer using data gathered by direct interviews was 1.4 (95% CI = 1.1, 2.0), allowing for multiple MER employment, and 2.0 (95% CI = 1.3, 3.0), excluding individuals with multiple MER employment. ORs for all MER employment, except truck drivers who were employed for less than 10 years, were 1.3 (95% CI = 1.0, 1.7) and 1.3 (95% CI = 0.9, 1.8) including and excluding multiple MER employment, respectively. ORs were then derived for specific MER occupations and, to avoid the confounding effects of multiple MER job classifications, analyses were also done excluding subjects with multiple MER job exposures. Truck drivers employed for more than 10 years had an odds ratio of 1.5 (95% CI = 1.1, 1.9). A similar figure was obtained excluding subjects with multiple MER employment. An excess risk was not detected for truck drivers employed less than 10 years. The only other job category that showed a statistically significant excess for lung cancer included taxi drivers and chauffeurs who worked multiple MER jobs for less than 10 years (OR = 2.5, 95% CI = 1.4, 4.8). For the same category, the risk for individuals working in that job for more than 10 years was 1.2 (95% CI = 0.5, 2.6). A statistically significant positive trend ( p <0.05) with increasing employment of <2 years, 2 to 9 years, 10 to 19 years, and 20+ years was observed for truck drivers but not for other MER occupations. A statistically nonsignificant excess risk was also observed for heavy equipment operators, bus drivers, taxi drivers and chauffeurs, and mechanics employed for 10 years or more. All of the above-mentioned ORs were derived, adjusted for birth cohort, usual daily cigarette use, and State of residence. Exposure to other occupational suspect lung carcinogens did not account for the excess risks
detected. Results of this large study provide evidence that workers in MER jobs are at an excess risk of lung cancer that is not explained by their smoking habits or exposures to other lung carcinogens. Because no information on type of engine had been collected, it was not possible to determine if the excess risk was due to exposure to DE or gasoline exhaust or a 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 would make the effect of DE harder to detect due to broad categorization of jobs and the problems in classifying individuals into uniform occupational groups based on the pooled data in the three studies that used different occupational classification schemes.7-42 7.2.2.7. Steenland et al. (1990): A Case-Control Study of Lung Cancer and Truck Driving in
the Teamsters Union Steenland et al. conducted a case-control study of lung cancer deaths in the Teamsters Union to determine the risk of lung cancer among different occupations. Death certificates were obtained from the Teamsters Union files in the central States for 10,485 (98%) male decedents who had filed claims for pension benefits and who had died in 1982 and 1983. Individuals were required to have 20 years’ tenure in the union to be eligible to claim benefits. Cases comprised all deaths (n = 1,288) from lung cancer, coded as ICD 162 or 163 for underlying or contributory cause on the death certificate. The 1,452 controls comprised every sixth death from the entire file, excluding deaths from lung cancer, bladder cancer, and motor vehicle accidents. Detailed 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 80% for controls. Using these interview data and the 1980 census occupation and industry codes, subjects were classified either as nonexposed or as having held other jobs with potential DE exposure. Data on job categories were missing for 12% of the study subjects. A second work history file was also created based on the Teamsters Union pension application that lists occupation, employer, and dates of employment. A three-digit U.S. census code for occupation and industry was assigned to each job for each individual. This Teamsters Union work history file did 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 dockworkers. Subjects were assigned the job category in which they had worked the longest. The case-control analysis was done using unconditional logistic regression. Separate analyses were conducted for work histories from the Teamsters Union pension file and from next-of-kin interviews. Covariate data were obtained from next-of-kin interviews. Analyses were also performed for two time periods: employment after 1959 and employment after 1964. 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 obtained for 994 cases and 1,085 controls using Teamsters Union work history and for 872 cases and 957 controls using next-of-kin work history. When exposure was considered as a dichotomous variable, for both Teamsters Union and next-of-kin work history, no single job category had an elevated risk. From the next-of-kin data, diesel truck drivers had an odds ratio of 1.42 (95% CI = 0.74, 2.47) and diesel truck mechanics had an odds ratio of 1.35 (95% CI = 0.74, 2.47). ORs by duration of employment as a categorical variable were then estimated. For the Teamsters Union work history data, when only employment after 1959 was considered, both long-haul ( p <0.04) and short-haul drivers (not significant) showed an increase in risk with7-43 increased years of exposure. The length-of-employment categories for which the trends were analyzed were 1 to 11 years, 12 to 17 years, and 18 years or more. Using 1964 as the cutoff date, long-haul drivers continued to show a significant positive trend ( p =0.04), with an odds ratio of 1.64 (95% CI = 1.05, 2.57) for those who worked for 13+ years, the highest category. Short-haul drivers, however, did not show a positive trend when 1964 was used as the cutoff date. Similar trend analysis was done for most next-of-kin data. A marginal increase in risk with increasing duration of employment as a truck driver ( p =0.12) was observed. For truck drivers who primarily drove diesel trucks for 35 years or longer, the odds ratio for lung cancer was 1.89 (95% CI = 1.04, 3.42). Similarly, the corresponding odds ratio was 1.34 (95% CI = 0.81, 2.22) for both gasoline truck drivers and drivers who drove both types of trucks, and 1.09 (95% CI = 0.44, 2.66) for truck mechanics. No significant interactions between age and DE exposure or smoking and DE exposure were observed. All the ORs were adjusted for age, smoking, and asbestos in addition to various exposure categories. This is a well-designed and analyzed study. The main strengths of the study are the availability of detailed records from the Teamsters Union, a relatively large sample size, availability of smoking data , and measurements of exposures. The authors acknowledge some limitations of this study, which include possible misclassifications of exposure and smoking habits, as information was provided by next of kin; lack of sufficient latency to observe lung cancer excess; and a small nonexposed group (n = 120). Also, they could not evaluate the concordance between Teamsters Union and next-of-kin job categories easily because job categories were defined differently in each data set. No data were available on levels of diesel exposure for the different job categories. Despite these limitations, the positive findings of this study, which are probably underestimated, provide a positive evidence toward causal association between DE exposure and excess lung cancer.

7.2.2.8. Steenland et al. (1998): DE and Lung Cancer in the Trucking Industry: Exposure-Response Analyses and Risk Assessment Steenland et al. (1998) conducted an exposure-response analysis by supplementing the data from their earlier case-control study of lung cancer and truck drivers in the Teamsters Union (Steenland et al., 1990) with exposure estimates based on a 1990 industrial hygiene survey of elemental carbon exposure, a surrogate for DE in the trucking industry. Study subjects were long-term Teamsters enrolled in the pension system who died during the period 1982-1983. Using death certificate information, the researchers identified 994 cases of lung cancer for the study period, and 1,085 non-lung-cancer deaths served as controls. Subjects were divided into job categories based on the job each held the longest. Most had held7-44 only one type of job. The job categories were short-haul driver, long-haul driver, mechanic, dockworker, other jobs with potential diesel exposure, and jobs outside the trucking industry without occupational diesel exposure. Smoking histories were obtained from next of kin. ORs were calculated for work in an exposed job category at any time and after 1959 (an estimated date when the majority of heavy-duty trucks had converted to diesel) compared with work in nonexposed jobs. ORs were adjusted for age, smoking, and potential asbestos exposure. Trends 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 carbon measurements were generally consistent with the epidemiologic results, in that mechanics were found to have the highest exposures and relative risk, followed by long-haul and then short-haul drivers, although dockworkers had the highest exposures and the lowest relative risks. Past exposures were estimated assuming that they were a function of (1) the number of heavy-duty trucks on the road, (2) the particulate emissions (grams/mile) of diesel engines over time, and (3) leaks from truck exhaust systems for long-haul drivers. Estimates of past exposure to elemental carbon, as a marker for DE exposure, for subjects in the case-control study were made by assuming that average 1990 levels for a job category could be assigned to all subjects in that category, and that levels prior to 1990 were directly proportional to vehicle miles traveled by heavy-duty trucks and the estimated emission levels of diesel engines. A 1975 exposure level of elemental carbon in terms of micrograms per cubic meter was estimated by the following equation: 1975 level = 1990 level*(vehicle miles 1975/vehicle miles 1990) (emissions 1975/emissions 1990). Once estimates of exposure for each year of work history were derived for each subject, analyses were conducted by cumulative level of estimated carbon exposure. 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). Logistic regression was used to estimate ORs adjusted for five categories of age, race, smoking (never, former-quitting before 1963, former-quitting in 1963 or later, current-with <1 pack per day, and current-with 1 or more packs per day), diet, and reported asbestos exposure. A variety of models for cumulative exposure were considered, including a log-linear model with cumulative exposure, a model adding a quadratic term for cumulative exposure, a log transform of cumulative exposure, dummy variables for quartile of cumulative exposure, and smoothing splines of cumulative exposure. The estimates of rate ratios from logistic regression for specific levels of exposure to elemental carbon were then used to derive excess risk estimates for lung cancer after lifetime exposure to elemental carbon. The survey found that mechanics had the highest current levels of DE exposures and dockworkers who mainly used propane-powered forklifts had the lowest exposure. ORs of 1.697-45 and 0.93 were observed for the mechanics and dockworkers, respectively. The finding of the highest lung cancer risk for mechanics and lowest for dockworkers is indicative of causal association between the DE exposure and development of lung cancer. The log of cumulative exposure was found to be the best-fitting model and was a significant predictor ( p = 0.01). However, the risk among mechanics did not increase with increasing duration of employment. OR for quartile of cumulative exposure show a pattern of significantly increasing trends in risk with increasing exposure, ranging between 1.08 and 1.72, depending on the exposure level and lag structure used. The lifetime excess risk of lung cancer death (through age 75) for a male truck driver was estimated to be in the range of 1.4%-2.3% (95% confidence limits ranged from 0.3% to 4.6%) above the background risk, depending on the emissions scenarios assumed. The authors found that current exposures indicated that truck drivers are exposed to DE at levels about the same as ambient levels on the highways, which are about double the background levels in urban air. They conclude that the data suggest a positive and significant increase in lung cancer risk with increasing estimated cumulative exposure to DE among workers in the trucking industry. They assert that these estimates suggest that the lifetime excess risk for lung cancer is 10 times higher than the OSHA standards, but caution that the results should be viewed as exploratory. The authors acknowledge that the increasing trend in risk with increasing estimates of cumulative exposure is partly due to the fact that a component of cumulative dose is simple duration of exposure, and that analyses by simple duration also exhibit a positive trend with duration. This analysis essentially weights the duration by contrived estimates of exposure intensity, and the authors acknowledge that this weighting depends on very broad assumptions. This is not an analysis of new data that provides independent estimates of relative risk for DE and lung cancer incidence. Instead, it is an attempt to convert the data from Steenland's earlier study of lung cancer for the purpose of estimating a different risk metric, “lifetime excess risk of lung cancer,” by augmenting these data with limited industrial hygiene data and rationalizations about plausible models for cumulative exposure. The Health Effects Institute (HEI, 1999) and others have raised some concerns about the exposure estimations, selection of controls, and control for confounding variables, and hence, this study’s usefulness for quantitative risk assessment. EPA and NIOSH will address these concerns in the year 2001. The HEI (1999) panel noted that some of the strengths of this study include the relevance of exposure levels to the general population and the use of an exposure marker for diesel engine emissions that was an improvement over the concentration of respirable-size particles (RSP). The number of study subjects (996 lung cancer cases) is large. Histories of exposures to asbestos and smoking were obtained, and confounding by these two variables was controlled in the analysis. Thus, it should be noted that these concerns are about7-46 the use of these data for quantitative risk assessment, due to limitations of the exposure data. As far as qualitative risk assessment is concerned, this study is still considered to be positive and strong.

7.2.2.9. Boffetta et al. (1990): Case-Control Study on Occupational Exposure to DE and Lung Cancer Risk This is an ongoing (since 1969) case-control study of tobacco-related diseases in 18 hospitals (six U.S. cities). Cases comprise 2,584 males with histologically confirmed primary lung cancers. Sixty-nine cases were matched to 1 control, whereas 2,515 were matched to 2 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 questionnaire was modified to include the usual occupation and up to five other jobs as well as duration (in years) worked in those jobs. After 1985, information was also obtained on exposure to 45 groups of chemicals, including DE at the workplace or during hobby activities. A priori aggregation of occupations was categorized into low probability of DE exposure (reference group), possible exposure (19 occupations), and probable exposure (13 occupations). Analysis was conducted based on “usual occupation” on all study subjects, and any occupation with sufficient cases was eligible for further analysis. In addition, cases enrolled after 1985 for which there were self-reported DE exposure and detailed work histories were also analyzed separately. Both matched and unmatched analyses were done by calculating the adjusted (for smoking and education) relative odds using the Mantel-Haenzael method and calculating the test-based 95% confidence interval using the Miettinen method. Unconditional logistic regression was used to adjust for potential confounders (the PROC LOGIST of SAS). Linear trends for risk 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 population. Although slight excesses were observed for the self-reported DE exposure group and the subset of post-1985 enrollees for highest duration of exposure (for self-reported exposure, occupations with probable exposure, and truck drivers), none was statistically significant. Trend tests for the risk of lung cancer among self-reported DE exposure, probable exposure, and truck drivers with increasing exposure (duration of exposure used as surrogate for increasing dose) were nonsignificant too. Statistically significant lung cancer excesses were observed for cigarette smoking only. The major strength of this study is availability of detailed smoking history. Even though detailed information was obtained for the usual and five other occupations (1985), because it7-47 was difficult to estimate or verify the actual exposure to DE, duration of employment was used as a surrogate for dose instead. The numbers of cases and controls were large; however, the number of individuals exposed to DE was relatively few, thus reducing the power of the study. This study did not attempt latency analysis either. Due to these limitations, the findings of this study are unable to provide either positive or negative evidence for a causal association between DE and occurrence of lung cancer.

7.2.2.10. Emmelin et al. (1993): DE Exposure and Smoking: A Case-Referent Study of Lung Cancer Among Swedish Dock Workers This case-control study of lung cancer was drawn from a cohort defined as all male workers who had been employed as dockworkers for at least 6 months between 1950 and 1974. In the population of 6,573 from 20 ports, there were 90 lung cancer deaths (cases), identified through Swedish death and cancer registers, during the period 1960 to 1982. Of these 90 deaths, the 54 who were workers at the 15 ports for which exposure surrogate information was available were chosen for the case-control study. Four controls, matched on port and age, were chosen for each case from the remaining cohort who had survived to the time of diagnosis of the case. Both live and deceased controls were included. The final analyses were done on 50 cases and 154 controls who had complete information on employment dates and smoking data. The smoking strata were created by classifying ex-smokers as nonsmokers if they had not smoked for at least 5 years prior to the date of diagnosis of the case; otherwise they were classified as smokers. Relative odds and regression coefficients were calculated using conditional logistic regression models. Comparisons were made both with and without smoking included as a variable, and the possible interaction between smoking and DE was tested. Both the weighted linear regressions of the adjusted relative odds and the regression coefficients were used to test mortality trends with all three exposure variables. Exposure to DE was assessed indirectly by initially measuring: (1) exposure intensity based on exhaust emission, (2) characteristics of the environment in terms of ventilation, and (3) measures of proportion of time in higher exposed jobs. For exhaust emissions, annual diesel fuel consumption at a port was used as the surrogate. For ventilation, the annual proportion of ships with closed or semiclosed holds was used as the surrogate. The proportion of time spent below decks was used as the surrogate for more exposed jobs. Although data were collected for all three measures, only the annual fuel consumption was used for analysis. Because every man was likely to rotate through the various jobs, the authors thought using annual consumption of diesel fuel was the appropriate measure of exposure. Consequently, in a second analysis, the annual fuel consumption was divided by the number of employees in the same port that year to come up with the fuel-per-person measure, which was further used to create a second measure,7-48“exposed time.” The “annual fuel” and exposed-time data were entered in a calendar timeexposure matrix for each port, from which individual exposure measures were created. A third measure, “machine time” (years of employment from first exposure), was also used to compare the results with other studies. All exposure measures were accumulated from the first year of employment or first year of diesel machine use, 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 three exposure variables: machine time, fuel, and exposed time. Conditional logistic regression models, adjusting for smoking status and using low exposures and/or nonsmokers as a comparison group, yielded positive trends for all exposure measures, but no trend test results were reported, and only the relative odds for the exposed-time exposure measure in the high-exposure group (OR = 6.8, 90% CI = 1.3 to 34.9) was reported as statistically significant. For smokers, adjusting for DE exposure level, the relative odds were statistically significant and about equal for all three exposure variables: machine time, 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 = 6.2 (90% CI = 2.6 to 14.6). Interaction between DE and smoking was tested by conditional logistic regression in the exposed-time variable. Although there were positive trends for both smokers and nonsmokers, the trend for smokers was much steeper: low, OR = 3.7 (90% CI = 0.9 to 14.6); medium, OR = 10.7 (90% CI = 1.5 to 78.4); and high, OR = 28.9 (90% CI = 3.5 to 240), indicating more than additive interaction between these two variables. In the weighted linear regression model with the exposed-time variable, the results were 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 DE and smoking. However, with just six nonsmokers and no further categorization of smoking amount or duration, these results are of limited value. The DE exposure matrices created using three different variables are intricate. Analyses by any of these variables yield essentially the same positive results and positive trends, providing consistent support for a real effect of DE exposure, at least in smokers. However, methodological limitations to this study prevent a more definitive conclusion. The numbers of cases and controls are small. There are very few nonsmokers; thus, testing the effects of DE exposure in them is futile. Lack of information on asbestos exposure, to which dockworkers are usually exposed, may also confound the results. Also, no latency analyses are presented. Overall, despite these limitations, this study supports the earlier findings of excess lung cancer mortality among individuals exposed to DE.7-49 7.2.2.11. Swanson et al. (1993): Diversity in the Association Between Occupation and Lung Cancer Among Black and White Men This population-based case-control study of lung cancer was conducted in metropolitan Detroit. The cases and controls for this study were identified from the Occupational Cancer Incidence Surveillance Study (OCISS). A total of 3,792 incident lung cancer cases and 1,966 colon and rectal cancer cases used as controls, diagnosed between 1984 and 1987 among white and black males aged 40 to 84 years, were selected for the study. Information was obtained by telephone interview either with the individual or a surrogate about lifetime work history and smoking history, as well as medical, demographic, and residential history. Occupation and industry data were coded using the 1980 U.S. Census Bureau classification codes. The investigators selected certain occupations and industries as having little or no exposure to carcinogens and defined them as an unexposed group. Analysis was done using logistic regression method and adjusting for age at diagnosis, pack-years of cigarette smoking, and race. The results were presented by various occupations and industries; those with potential exposures to DE were drivers of heavy trucks and light trucks, farmers, and railroad workers, respectively. Among white males, increasing lung cancer risks were observed with increasing duration of employment for drivers of heavy trucks, drivers of light trucks, and farmers. Although none of the individual ORs were statistically significant, trend tests were significant for all three occupations ( p 

### # 0.05). On the other hand, among black males increasing lung cancer risks with increasing duration of employment were observed for farmers only, with an OR of 10.4 (95% CI = 1.4, 77.1) reaching significance for employment of 20+ years. As for the railroad industry, increasing lung cancer risks with increasing duration of employment were 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. The main strengths of the study are large sample size, availability of lifetime work history and smoking history, and the population-based study format, precluding selection bias. The major limitation, as in other studies, is lack of direct information on specific exposures. The interesting result of this study is lung cancer excesses observed in farmers, mainly among crop farmers, who have potential exposure to DE from their tractors in addition to pesticides, herbicides, and other PM 10. The authors point out that this is the first study to find excess lung cancer in this occupation.

7.2.2.12. Hansen et al. (1998): Increased Risk of Lung Cancer Among Different Types of Professional Drivers in Denmark This is a population-based case-control study of lung cancer, conducted in professional drivers in Denmark. The cases first diagnosed as primary lung cancer between 1970 and 19897-50 among males born between 1897 and 1966 were identified from the Danish Cancer Registry. The registry provided the information on diagnosis from ICD-7, name, sex, and unique personal 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 even after the employee has retired or died. Information about current employment was obtained from the Danish Central Population Registry (CPR) by linkage with the PIDN. Of 37,597 cases identified from the Registry, 8,853 did not have any employment records. Controls (1:1) for 28,744 lung cancer cases with employment histories were selected randomly from CPR, matched with the case by year of birth and sex. Furthermore, these controls had to be alive, cancer free, and employed prior to the diagnosis of lung cancer in the corresponding case. Employment histories were obtained for the controls in the same fashion as cases from the pension fund. The employment record search resulted in a total of 1,640 lorry/bus drivers and 426 taxi drivers. They were further divided into subgroups by their duration of employment. Information about smoking in drivers was acquired from two national surveys conducted in 1970-72 and 1983. No direct information on smoking was available in either cases or controls. A separate case-control study of mesothelioma indirectly looked at asbestos exposure among professional drivers. OR, adjusting for socioeconomic status and 95% CI, were computed using conditional logistic regression (PECAN procedure in the statistical package EPICURE). Significant ORs for lung cancer were found for lorry/bus drivers (OR = 1.31, 95% CI = 1.17, 1.46), taxi drivers (OR = 1.64, 95% CI = 1.22, 2.19), and unspecified drivers (OR = 1.39, 95% CI = 1.30, 1.51). Significant ORs were found for both lorry/bus drivers and taxi drivers by duration of employment in 1-5 years and >5 years categories, with no lag time and with a 10- year lag time. The ORs remained the same for lorry/bus drivers in these employment categories for no lag time and 10-year lag time. Among taxi drivers, on the other hand, the OR of 2.2 in >5 year employment in no-lag-time analysis increased to 3.0 in the 10-year lag time analysis. The authors asserted that the higher risk seen in the taxi drivers may be due to higher exposure attributable due to longer time spent in traffic congestion. The trend tests for increasing risk with increasing duration of employment (surrogate for exposure) were statistically significant ( p <0.001) for both lorry/bus drivers and taxi drivers in no-lag-time and 10-year lag time analysis. All the ORs were adjusted for socioeconomic status. The main strengths of the study are the large sample size, availability of information on socioeconomic status, and detailed employment records. The main limitation, however, is lack of information on what type of fuel these vehicles used. It is probably safe to assume that the lorry/buses were diesel powered, whereas the taxis could be either diesel or gasoline powered. A7-51 personal communication with Dr. Johnni Hansen confirmed that dieselization in Denmark was completed in the late 1940s and lorries, buses, and taxis have been using diesel fuel since then. Although direct adjustments were not done for smoking and exposure to asbestos, indirect information on both these confounders indicates that they are unlikely to explain the observed excesses and the increasing risk with increasing duration of employment. Thus, the results of this study are strongly supportive of DE being associated with increased lung cancer.

7.2.2.13. Brüske-Hohlfeld et al. (1999): Lung Cancer Risk in Male Workers Occupationally Exposed to Diesel Motor Emissions in Germany This paper presents a pooled analysis of two case-control studies of lung cancer. The first study, by Jöckel et al. (1995, 1998), was conducted between 1988 and 1993 and had 1,004 cases and 1,004 controls matched for sex, age, and region of residence, selected randomly from the compulsory municipal registries. The inclusion criteria for cases were: they should have been born in or after 1913, should have been of German nationality, and should have been diagnosed with lung cancer within 3 months prior to the interview. The second study, by Wichmann et al. (1998), was ongoing when it was included in this study. The study span covered the years 1990 to 1996. By 1994 a total of 3,180 cases and 3,249 controls, randomly selected from the compulsory population registries, were frequency matched on sex, age, and region. The cases were less than 76 years old, were residents of the region and living in Germany for more than 25 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 interview was conducted with each study participant. Data were collected on basic demographic information, detailed smoking history, and lifelong occupational history about jobs held and industries worked in. The job titles and industries were classified into 33 and 21 categories, respectively, using the German Statistical Office codes. Based on job codes with potential exposure to diesel motor emission (DME), four exposure groups were constituted. Group A comprised professional drivers of trucks, buses, taxis, etc. Group B comprised other traffic-related jobs such as switchmen, diesel locomotive drivers, and diesel forklift truck drivers. Group C comprised bulldozer operators, graders, and excavators. Group D comprised full-time farm tractor drivers. Validation of the jobs was done by written evaluation of the job task descriptions, which also avoided misclassification. The following information was acquired for the construction of job task descriptions: (1) What were your usual tasks at work and how often (in % of daily working hours) were they performed? (2) What did you produce, manufacture, or transport? (3) Which material was used? (4) What kind of machine did you operate? Some individuals had more than one job task7-52 with DME exposure. 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 beginning and end of exposure. The first year of exposure was defined as < 1945, 1946-1955,
and > 1956 while the last year of exposure was defined as < 1965, 1966-1975, and > 1976. For professional drivers, hours driven per day were accumulated and were classified as “driving
hours.” A smoker was defined as any individual who had smoked regularly for at least 6 months. Smoking information was acquired in series with the starting time, type of tobacco, amount smoked, duration in years, and calender year of quitting. Asbestos exposure was estimated by certain job-specific supplementary questions. The cases and controls were post-hoc stratified into 6 age and 17 region categories. ORs adjusted for smoking and asbestos exposure were calculated by conditional logistic regression, using “never exposed” workers as the reference group. The adjustment for cigarette smoking was done by using pack-years as a continuous variable; adjustment for other tobacco products was done by considering them as a binary variable. A total of 716 cases and 430 controls were 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 smoking and asbestos exposure (West Germany, >10-20 years and >20-30 years of exposure, first year of exposure in 1946-1955 and 1956+, end of exposure in 1966-1975 and 1976+, and for the job categories of Group A, B, and C). The risk increased with increasing years of exposure, and for both the first year of exposure (< 1945, 1946-1955, and > 1956) and end year of exposure (< 1965, 1966-1975, and > 1976). Separate analyses by four job categories (all the ORs were adjusted for smoking and asbestos exposure) showed that for professional drivers (Group A) the overall OR was 1.25 (95% CI = 1.05, 1.47). Significant ORs were found for various factors in West Germany only. The factors were: >0-3 years and >10-20 years of exposure (OR = 1.69, 95% CI = 1.13, 2.53, and OR = 2.02, 95% CI = 1.32, 3.08, respectively), beginning of exposure in 1956+ and end of exposure in 1976+ (OR = 1.56, 95% CI = 1.21, 2.03, and OR = 1.5, 95% CI = 1.14, 1.98, respectively), and 1,000-49,999 driving hours (OR = 1.54, 95% CI = 1.15, 2.07). None of the ORs were significant in East Germany in this group. For other traffic-related jobs (Group B) the overall OR was 1.53 (95% CI = 1.04, 2.24). The ORs for beginning of exposure in 1956+ and end of exposure in 1976+ were OR = 1.71, 95% CI = 1.05, 2.78, and OR = 2.68, 95% CI = 1.47, 4.90, respectively. The risk increased with7-53 increasing duration of exposure and was statistically significant for >10-20 years (OR = 2.49) and more than 20 years (OR = 2.88). No separate analyses for West Germany and East Germany were presented in this category. For heavy equipment operators (Group C) the overall OR of 2.31 (95% CI = 1.44, 3.7) was highest among all the job categories. Significant ORs were observed for beginning exposure in 1946-1955 (OR = 2.83, 95% CI = 1.10, 7.23) and end exposure in 1966-1975 (OR = 3.74, 95% CI = 1.20, 11.64). The risk increased with increasing duration of exposure and was statistically significant for more than 20 years of exposure (OR = 4.3). Although no separate analyses for West Germany and East Germany were presented, investigators mentioned that for this job group hardly any difference was seen between West Germany and East Germany. For drivers of the farming tractors (Group D) the overall OR of 1.29 was not significant. Risk increased with increasing duration of exposure and was significant for exposure of more than 30 years (OR = 6.81, 95% CI = 1.17, 39.51). No separate analyses for West Germany and East Germany were presented in this category. The professional drivers and the other traffic-related job categories probably have mixed exposures to gasoline exhaust in general traffic. On the other hand, it should be noted that exposure to DME among heavy equipment and farm tractor drivers is much higher and not as mixed as in professional drivers. The heavy equipment drivers usually drive repeatedly through their own equipment’s exhaust. Therefore, the observed highest risk for lung cancer in this job category establishes a direct link with the DME. The only other study that found significantly higher risk for heavy equipment operators (RR = 2.6) was conducted by Boffeta et al. (1988). 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 job category with increasing exposure. The investigators stated that the working conditions and the DME of tractors remained fairly constant over the years. This increase may be due mainly to exposure to DME and, in addition, PM 10. This is a well-designed, well-conducted, and well-analyzed study. Its main strengths are large sample size, resulting in good statistical power; inclusion of incident cases that were diagnosed not more than 3 months prior to the interview; use of only personal interviews, reducing recall bias; diagnosis ascertained by cytology or histology; and availability of lifelong detailed occupational and smoking history. Exposure estimation for each individual was based on job codes and industry codes, which were validated by written job descriptions to avoid misclassification. The main limitation of the study is lack of data on actual exposure to DME. The cumulative quantitative exposures were calculated based on time spent in each job with potential exposure to DME and the type of equipment used. Thus, this study provides strong evidence for a causal association between exposure to DE and occurrence of lung cancer. Table 7-2 summarizes the above lung cancer case-control studies.Table 7-2. Epidemiologic studies of the health effects of exposure to DE: case-control studies of lung cancerAuthors Population studied DEexposure assessment Results LimitationsHall andWynder(1984)502 histologically confirmedlung cancersCases diagnosed 12 mo prior tointerviews502 matched hospital controlswithout tobacco-related diseases,matched for age, sex, race, andgeographical areaPopulation from 18 hospitals incontrolsBased on previousIndustrial HygieneStandards for aparticular occupation,usual lifetime occupationcoded as “probably highexposure” and “noexposure”NIOSH standards usedto classify exposures:HighModerateLowSNS excess risk after adjustmentfor smoking for lung cancer:RR = 1.4 (1st criteria)andRR = 1.7 (NIOSH criteria)Complete lifetime employmenthistory not availableSelf-reported occupation history notvalidatedNo analysis by dose, latency, orduration of exposureNo information on nonoccupationaldiesel exposureDamber andLarsson(1987)589 lung cancer cases who haddied prior to 1979 reported toSwedish registry between 1972and 1977582 matched dead controls (sex,age, year of death, municipality)drawn from National Registryof Cause of Death453 matched living controls(sex, year of birth, municipality)drawn from NationalPopulation RegistryOccupations held for atleast 1 year or moreA 5-digit code was used toclassify the occupationsaccording to NordicClassification ofOccupationsFor underground miners: SS OR =2.7 ($1 year of employment)SS OR = 9.8 ($20 years ofemployment)For professional drivers: SNS OR= 1.2 ($20 years of employment)with dead controlsAll ORs adjusted for smokingUncertain DE exposureNo validation of exposure doneUnderground miners data notadjusted for other confounders suchas radon, etc.

7-54Table 7-2. Epidemiologic studies of the health effects of exposure to DE: case-control studies of lung cancer (continued)Authors Population studied DEexposure assessment Results LimitationsLerchenet al. (1987)506 lung cancer cases fromNew Mexico tumor registry(333 males and 173 females)Aged 25-84 yearsDiagnosed between January 1,1980, and December 31, 1982771 (499 males and 272 females)frequency matched with cases,selected from telephone directoryLifetime occupationalhistory and self-reportedexposure history wereobtainedCoded according toStandard IndustrialClassification SchemeNo excess of relative odds wereobserved for DE exposureExposure based on occupationalhistory and self-report, which was notvalidated50% occupational history providedby next of kinAbsence of lung cancer associationwith asbestos suggestsmisclassification of exposureGarshicket al. (1987)1,319 lung cancer cases who diedbetween March 1, 1981,and February 28, 19822,385 matched controls (two each,age and date of death)Both cases and controls drawnfrom railroad worker cohortwho had worked for 10 ormore yearsPersonal exposureassessed for 39 jobcategoriesThis was corrected withjob titles to dichotomizethe exposure into:ExposedNot exposedIndustrial hygienesampling doneSS OR = 1.41 (#64 year age group)SS OR = 1.64 (#64 year age group)for$20 years DE exposure groupwhen compared to 0- to 4-yearexposure groupAll ORs adjusted for lifetimesmoking and asbestos exposureProbable misclassification of DEexposure jobsYears of exposure used as surrogatefor dose13% of death certificates notascertainedOverestimation of smoking historyBenhamouet al. (1988)1,260 histologically confirmedlung cancer cases2,084 non-tobacco-relateddisease matched controls(sex, age at diagnosis,hospital admission, andinterviewer)Occurring between 1976 and1980 in FranceBased on exposuresdetermined by panel ofexpertsThe occupations wererecorded blindly usingInternational StandardClassification ofOccupations as chemicalor physical exposuresSignificant excess risks were foundin motor vehicle drivers(RR = 1.42) andtransport equipment operators(RR = 1.35) (smoking adjusted)Exposure based on occupationalhistories not validatedExposures classified as chemical andphysical exposures, not specific toDE

7-55Table 7-2. Epidemiologic studies of the health effects of exposure to DE: case-control studies of lung cancer (continued)Authors Population studied DEexposure assessment Results LimitationsHayes et al.(1989)Pooled data from three differentstudies consisting of 2,291 malelung cancer cases2,570 controlsOccupational informationfrom next of kin for alljobs heldJobs classified withrespect to potentialexposure to known andsuspected pulmonarycarcinogensSS OR = 1.5 for truck drivers (>10years of employment)SS positive trend with increasingemployment as truck driverAdjusted for age, smoking, & studyareaExposure data based on jobdescription given by next of kin,which was not validatedCould have been mixed exposure toboth diesel and gasoline exhaustsJob description could have led tomisclassificationSteenlandet al. (1990)1,058 male lung cancer deathsbetween 1982 and 19831,160, every sixth death fromentire mortality file, sorted bySocial Security number(excluding lung cancer,bladder cancer, and motorvehicle accidents)Cases and controls were fromCentral State Teamsters whohad filed claims (requiring20-year tenure)Longest job held: dieseltruck driver, gasolinetruck driver, both typesof trucks, truckmechanic, anddockworkersAs 1964 cut-off point:SS OR = 1.64 for long-haul driverswith 13+ years of employmentPositive trend test for long-hauldrivers (p=0.04)SS OR = 1.89 for diesel truckdrivers of 35+ years ofemploymentAdjusted for age, smoking, &asbestosExposure based on job titles notvalidatedPossible misclassification of exposureand smoking, based on next-of-kininformationLack of sufficient latencySteenland etal. (1998)Exposure-response analyses oftheir 1990 case-control studyIndustrial hygiene data ofelemental carbon intrucking industry collectedby Zaebst et al. (1991)used to estimate individualexposuresCumulative exposurescalculated based onestimated lifetimeexposuresFor mechanics: OR = 1.69 (hadthe highest DE exposure)Lowest DE exposure and lowestOR = 0.93 observed fordockworkersIncreasing risk of lung cancer withincreasing exposureAdjusted for age & smoking

7-56Table 7-2. Epidemiologic studies of the health effects of exposure to DE: case-control studies of lung cancer (continued)Authors Population studied DEexposure assessment Results LimitationsBoffetta etal. (1990)From 18 hospitals (since 1969),2,584 male lung cancer casesmatched to either one control (69)or two controls (2,515) weredrawn. Matched on age, hospital,and year of interviewA priori aggregation ofoccupations categorizedinto low probability,possible exposure (19occupations), andprobable exposure (13occupations) to DEOR slightly below unity SNSAdjusted for smokingNo verification of exposureDuration of employment used assurrogate for doseNumber of individuals exposed to DEwas smallEmmelin etal. (1993)50 male lung cancer cases from15 ports (worked for at least6 months between 1950 and1974), 154 controls matched onage and portIndirect DE exposureassessment done based on(1) exposure intensity, (2)characteristics ofventilation, (3) measure ofproportion of time inhigher exposure jobsSS OR for high-exposure group =6.8Positive trend for DE observed(trend much steeper for smokersthan nonsmokers)Adjusted for smokingNumbers of cases and controls aresmallVery few nonsmokersLack of exposure information onasbestosNo latency analysisSwanson etal. (1993)Population based case-controlstudy in metropolitan Detroit3,792 lung cancer cases and 1,966colon cancer (cases) controls,diagnosed between 1984 and 1987in white and black males (agedbetween 40-84)Telephone interviews withthe individual or surrogateabout lifetime workhistoryOccupation and industrydata coded per 1980 U.S.Census Bureauclassification codesCertain occupations andindustries were selected asunexposed to carcinogensSS excess ORs observed for- black farmers OR= 10.4 for 20+years employment- white railroad industry workersOR= 2.4 for 10+ years employmentAmong white trend tests were SSfor-drivers of heavy duty trucks- drivers of light duty trucks- farmers- railroad workersAmong blacks trend test was SS forfarmers onlyAll the ORs were adjusted for ageat diagnosis, pack-years ofcigarette smoking and raceLack of direct information onspecific exposuresNo latency analysis

7-57Table 7-2. Epidemiologic studies of the health effects of exposure to DE: case-control studies of lung cancer (continued)Authors Population studied DEexposure assessment Results LimitationsHansen etal. (1998)Population-based case-controlstudy of professional drivers inDenmarkMale lung cancer cases diagnosedbetween 1970-1989, controlsmatched by year of birth and sexInformation about pastemployment obtained bylinkage with nationwidepension fundEmployment as lorry/busdrivers (n=1,640) and taxidrivers (n=426) was usedas surrogate for exposureto DEFor lorry/bus drivers: SS OR = 1.31For taxi drivers: SS OR = 1.64,which increased to 2.2 in > 5-yearemployment with no lag time &3.0 in > 5 year employment with10- year lag timeSS trend test for increasing riskwith increasing employment forboth lorry/bus drivers & taxidrivers (p<0.001)All ORs adjusted forsocioeconomic statusLack of information on the type offuel (personal communication withthe principal investigator confirmedthat diesel fuel is used for thelorry/buses and taxis since early1960s)Even though direct adjustment wasnot done for smoking/asbestos,indirect methods indicate that theresults are not likely to beconfounded by these factors

7-58Table 7-2. Epidemiologic studies of the health effects of exposure to DE: case-control studies of lung cancer (continued)Authors Population studied DEexposure assessment Results LimitationsBrüskeHohlfeld etal. (1999)Pooled analysis of two casecontrol studies (3,498 cases &3,541 controls)Controls frequency matched onsex, age, & region, randomlyselected from the compulsorypopulation registryInclusion criteria: (1) born in orafter 1913/less than 75 years old,(2) German nationality/resident ofthe region - lived in Germany formore than 25 years, & (3) lungcancer diagnosis should be 3months prior to the studyInformation obtained by personalinterview on:Lifetime detailedoccupational & smokinghistories obtained fromeach individual in apersonal interviewBased on job codes (33job titles & 21 industries)potential DE exposureclassified in 4 categories:A- professional drivers oftrucks, buses, & taxis; Bother traffic related i.e.,switchman, locomotive, &forklift drivers; Cbulldozer operators,graders,& excavators; Dfarm tractor driversCumulative DE exposuresand pack-years (smoking)calculated for eachindividualSS higher risk adjusted for smokingobserved for all 4 categories:A- ORs ranged from 1.25 to 2.53B- ORs ranged from 1.53 to 2.88C- ORs ranged from 2.31 to 4.3D- 6.81 (exposure < 30 years)Risk increased with increasingexposureLack of data on actual exposure todiesel exaustAbbreviations: OR = odds ratio; RR = relative risk; SNS = statistically nonsignificant; SS = statistically significant.

7-597-60 7.2.3. Summaries of Studies and Meta-Analyses of Lung Cancer
7.2.3.1. Cohen and Higgins (1995): Health Effects of DE: Epidemiology The Health Effects Institute (HEI) reviewed all published epidemiologic studies on the health effects of exposure to DE available through June 1993, identified by a MEDLINE search and by reviewing the reference sections of published research and earlier reviews. HEI identified 35 reports of epidemiologic studies (16 cohort and 19 case-control) of the relation of occupational exposure to diesel emissions and lung cancer published between 1957 and 1993. HEI reviewed the 35 reports for epidemiologic evidence of health effects of exposure to DE for lung cancer, other cancers, and nonmalignant respiratory disease. They found that the data were strongest for lung cancer. The evidence suggested that occupational exposure to DE from diverse sources increases the rate of lung cancer by 20% to 40% in exposed workers generally, and to a greater extent among workers with prolonged exposure. They also found that the results are not explicable by confounding caused by cigarette smoking 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 one relative risk greater than one indicating positive association. Twelve studies indicating a relative risk greater than 1 had 95% confidence intervals, which excluded unity. The authors conclude that epidemiologic data consistently show weak associations between exposure to DE and lung cancer. They find that the evidence suggests that long-term exposure to DE in a variety of occupational circumstances is associated with a 1.2- to 1.5-fold increase in the relative risk of lung cancer compared with workers classified as unexposed. Most of the studies that controlled for smoking found that the association between increased risk of lung cancer and exposure to DE persisted after such controls were applied, although in some cases the excess risk was lower. None of the studies measured exposure to diesel emissions or characterized the actual emissions from the source of exposure for the time period most relevant to the development of lung cancer. Most investigators classified exposure based on work histories reported by subjects or their next of kin, or by retirement records. Although these data provide relative rankings of exposure, the absence of concurrent exposure information is the key factor that limits interpretation of the epidemiologic findings and subsequently their utility in making quantitative estimates of cancer risks. This is a comprehensive and thorough narrative review of studies of the health effects of DE. It does not undertake formal estimation of summary measures of effect or evaluation of heterogeneity in the results. The conclusion drawn about the consistency of the results is based on the author's assessment of the failure of potential biases and alternative explanations for the increase in risk to account for the observed consistency. In many if not most studies, the quality of the data used to control confounding was relatively crude. Although the studies do include qualitative assessment of whether control for smoking is taken into account, careful scrutiny of2 Of 35 studies identified in the literature search, 6 pairs of studies represented analyses of the same study population, reducing the number of studies to 29.

7-61 the quality of the control or adjustment for smoking among the studies is absent. This leaves open the possibility that prevalent residual confounding by inadequate control for smoking in many studies may account for the consistent associations seen.

7.2.3.2. Bhatia et al. (1998): DE Exposure and Lung Cancer Bhatia et al. (1998) report a meta-analysis of 29 published 2 cohort and case-control studies of the relation between occupational exposure to DE and lung cancer. A search of the epidemiologic literature was conducted for all studies concerning lung cancer and DE exposure. Occupational studies involving mining were excluded because of concern about the possible influence of radon and silica exposures. Studies in which the 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 inferred, were excluded. When studies presented risk estimates for more than one specific occupational category of DE-exposed workers, the subgroup risk estimates were used in the meta-analysis. Smoking-adjusted effect measures were used when present. Of 29 studies 23 met the criteria for inclusion in the meta-analysis. The observed relative risk estimates were greater than 1 in 21 of these studies; this result is unlikely to be due to chance. The pooled relative risk weighted by study precision was 1.33 (95% CI = 1.24, 1.44), indicating increased relative risk for lung cancer from occupational exposure to DE. Subanalyses by study design (case-control and cohort studies) and by control for smoking produced results that did not differ from those of the overall pooled analysis. Cohort studies using internal comparisons showed higher relative risks than those using external comparisons (see Figure 7-1). Bhatia and colleagues conclude that the analysis shows a small but consistent increase in the risk for lung cancer among workers with exposure to DE. The authors evaluate the dependence of the relative risk estimate on the presence of control for smoking among studies, 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 conclusions for the following reasons: (1) the pooled relative risks for studies adjusted for smoking were the same as those for studies not adjusting for smoking; (2) in those studies giving risk estimates adjusted for smoking and risk estimates not adjusted for smoking, there was only a small reduction in the pooled relative risk from DE exposure; and (3) in studies with internal7-62 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.

comparison populations, in which confounding is less likely, the pooled relative risk estimate
was 1.43. The validity of this assessment depends on the adequacy of control for smoking in the individual studies. If inadequate adjustment for smoking is employed and residual confounding by cigarette smoking pertains in the result of the individual studies, then the comparisons and contrasts of the pooled estimates the authors cite as reasons for dismissing the effect of residual confounding by smoking will remain contaminated by residual confounding in the individual studies. In fact, Bhatia et al. erroneously identify the treatment of the smoking data in the main7-63 analysis for the 1987 report by Garshick et al. as a continuous variable representing pack-years of smoking, whereas the analysis actually dichotomized the pack-years data into two crude dose 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 cigarette consumption was missing for more than 20% of the lung cancer cases. In this instance, the consistency between the adjusted and unadjusted estimates of the relative risk for DE exposure may be attributable to failure of adjustment rather than lack of confounding by cigarette smoking. A similar problem exists for the Bhatia et al. representation of the control for confounding in the study by Boffetta and Stellman (1988). An evaluation of the potential for publication bias is presented that provides reassurance that the magnitude of published effects is not a function of the precision or study power; however, this assessment cannot rule out the possibility of publication bias.

7.2.3.3. Lipsett and Campleman (1999): Occupational Exposure to DE and Lung Cancer: A Meta-Analysis Lipsett and Campleman (1999) conducted electronic searches to identify epidemiologic studies published between 1975 and 1995 of the relationship of occupational exposure to DE and lung cancer. Studies were selected based on the following criteria: (1) Estimates of relative risks and their standard errors must be reported or derivable from the information presented. (2) Studies must have allowed for a latency period of 10 or more years for development of lung cancer after onset of exposure. (3) No obvious bias resulted from incomplete case ascertainment in follow-up studies. (4) Studies must be independent: that is, a single representative study selected from any set of multiple analyses of data from the same population. Studies focusing on occupations involving mining were excluded because of potential confounding by radon, arsenic, and silica, as well as possible interactions between cigarette smoking and exposure to these substances in lung cancer induction. Thirty of the 47 studies initially identified as relevant met the specified inclusion criteria. Several risk estimates were extracted from six studies reporting results from multiple mutually exclusive diesel-related occupational subgroups. If a study reported effects associated with several levels or durations of exposure, the effect reported for the highest level or longest duration of exposure was used. If estimates for several occupational subsets were reported, the most diesel-specific occupation or exposure was selected. Adjusted risk estimates were used
when available. Thirty-nine independent estimates of relative risk and standard errors were extracted. Pooled estimates of relative risk were calculated using a random-effects model. Among study7-64 populations most likely to have had substantial exposure to DE, the pooled smoking- adjusted relative risk was 1.47 (95% CI = 1.29, 1.67) (see Figure 7-2).

Figure 7-2. Pooled estimates of relative risk of lung cancer in epidemiologic studies involving occupational exposure to DE (random-effects models).

Source: Lipsett and Campleman, 1999.

The between-study variance of the relative risks indicated the presence of significant heterogeneity in the individual estimates. The authors evaluated the potential sources of heterogeneity by subset analysis and linear meta-regressions. Major sources of heterogeneity included control for confounding by smoking, selection bias (a healthy worker effect), and exposure patterns characteristic of different occupational categories. A modestly higher, pooled relative risk was derived for the subset of case-control studies, which, unlike the cohort studies, showed little evidence of heterogeneity. This meta-analysis also evaluated the potential for publication bias, which provides reassurance that the magnitude of published effects is not a function of the precision or study7-65 power. Again, as stated in the Bhatia et al. (1998) review, this assessment cannot rule out the possibility of publication bias. Although a relatively technical approach was used in deriving summary estimates ofrelative risk and the evaluation of possible sources of variation in the relative risks in this metaanalysis, this approach should not be confused with rigorous evaluation of the potential weaknesses among the studies included in the analysis. The heterogeneity attributable to statistical adjustment for smoking was evaluated based on 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.

7.2.4. Summary and Discussion Certain extracts of DE have been demonstrated as both mutagenic and carcinogenic in animals and in humans. Animal data suggest that DE is a pulmonary carcinogen among rodents exposed by inhalation to high doses over long periods of time. While rat lung cancer response to DE is not suitable for dose-response extrapolation to humans, the positive lung cancer response doses imply a hazard for humans. Because large working populations are currently exposed to DE and because nonoccupational ambient exposures currently are of concern as well, the possibility that exposure to this complex mixture may be carcinogenic to humans has become an important public health issue. Because diesel emissions become diluted in the ambient air, it is difficult to study the health effects in the general population. Nonoccupational exposure to DE is 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 exposed to DE. Furthermore, the “exposed” group in these studies is based on job titles, which in most instances are not verified or correlated with environmental hygiene measurement. The issue of health effect measurement is further complicated by the fact that occupational cohorts tend to be healthy and have below-average mortality, usually referred to as the “healthy worker effect.” Hence, the usual standard mortality ratios observed in cohort mortality studies are likely to be
underestimations of true risk. A major difficulty with the occupational studies considered here was measurement of actual DE exposure. Because all the cohort mortality studies were retrospective, assessment of health effects from exposure to DE was naturally indirect. In these occupational settings, no systematic quantitative records of ambient air were available. Most studies compared men in job categories with presumably some exposure to DE with either standard populations (presumably no exposure to DE) or men in other job categories from industries with little or no potential for DE exposure. A few studies have included measurements of diesel fumes, but there is no7-66 standard method for the measurement. No attempt is made to correlate these exposures with the cancers observed in any of these studies, nor is it clear exactly which extract should have been measured to assess the occupational exposure to DE. All studies have relied on the job categories or self-report of exposure to DE. Gustavsson et al. (1990), Emmelin et al. (1993), and Brüske-Hohlfeld et al. (1999) estimated exposure levels by getting detailed histories of job tasks/categories and computing cumulative exposures, which unfortunately were not verifiable due to of the lack of industrial hygiene data. In the studies by Garshick et al. (1987, 1988), the diesel-exhaust-exposed job categories were verified based on an industrial hygiene survey done by Woskie et al. (1988a,b). The investigators found that in most cases the job titles were good surrogates for DE exposure. Also, in the railroad industry, where only persons who had at least 10 years of work experience were included in the study, the workers tended not to change job categories over the years. Thus, a job known only at one point in time was a reasonable marker of past DE exposure. Unfortunately, the exposure was only qualitatively verified. Quantitative use of this information would have been much more meaningful. Zaebst et al. (1991) conducted an industrial hygiene survey of elemental carbon exposure in the trucking industry by jobcategories. Using these exposure measurements, Steenland et al. (1998) conducted an exposureresponse analysis of their earlier lung cancer case-control study (Steenland et al., 1990). These exposure data are currently being verified and will be used for quantitative risk assessment in the
near future. Occupations involving potential exposure to DE are miners, truck drivers, transportation workers, railroad workers, and heavy equipment operators. No known studies in metal miners have assessed whether DE is associated with lung cancer. Currently, there are about 265 underground metal/nonmetal mines in the United States (Department of Labor, Mine Safety and Health Administration, 2001). Approximately 20,000 miners are employed, but not all of them are currently working in the mines. Diesel engines were introduced in metal mines in the United States in the early to mid-1960s. Although all these mines use diesel equipment, it is difficult to estimate how many of these miners were actually exposed to diesel fumes. Diesel engines were introduced in coal mines at an even later date in the United States, and their use is still quite limited. There are 910 underground coal mines in the United States, of which only 145 currently use diesel powered equipment (Department of Labor, Mine Safety and Health Administration, 2001). Even if it were possible to estimate how many miners (metal and coal) were exposed to DE, it would be very difficult to separate out the confounding effects of other potential pulmonary carcinogens, such as radon decay products or heavy metals (e.g., arsenic, chromium). Furthermore, the relatively short latency period limits the usefulness of
these cohorts of miners.7-67 Both metal and coal mines in Europe and Australia, on the other hand, have been using diesel equipment for more than 50 years. The epidemiologic studies of coal miners conducted in these countries discuss only exposures to coal dust. In most of the coal miner studies, DE exposures are not even mentioned by the investigators as confounding exposures. Therefore, it is not known how many miners, if any, were exposed to DE, for how long, and at what concentrations. Although studies of coal miners reviewed by IARC (1997) generally found lower than expected lung cancer mortality (with some exceptions where some excess of lung cancer was observed), without knowing the concentrations, duration of exposure, and number of miners exposed to DE, it is inappropriate to conclude that the reported lung cancer mortality deficit in these studies provides a proof positive of absence of causal association between DE exposure and occurrence of lung cancer.

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 demonstrate any statistically significant excess risk of lung cancer, but these studies have certain methodological problems, such as small sample sizes, short follow-up periods (just 6 years in the Rushton et al. study), lack of information on confounding variables, and lack of analysis by duration of exposure, duration of employment, or latency that preclude their use in determining the carcinogenicity of DE. Although the Waller (1981) study had a 25-year follow-up period, the cohort was restricted to employees (ages 45 to 64) currently in service. Employees who left the job earlier, as well as those who were still employed after age 64 and who may have died from cancer, were excluded. Wong et al. (1985) conducted a mortality study of heavy equipment operators that demonstrated a nonsignificant positive trend for cancer of the lung with length of membership and latency. Analysis of deceased retirees showed a significant excess of lung cancer. Individuals without work histories who started work prior to 1967, when records were not kept, may have been in the same jobs for the longest period of time. Workers without job histories included those who had the same job before and after 1967 and thus may have worked about 12 to 14 years longer; these workers exhibited significant excess risks of lung cancer and stomach cancer. If this assumption about duration of jobs is correct, then these site-specific causes can be linked to DE exposure. One of the methodologic limitations of this study is that most of these men worked outdoors; thus, this cohort might have had relatively low exposure to DE. The authors did not present any environmental measurement data either. Because of the absence of detailed work histories for 30% of the cohort and the availability of only partial work histories for the remaining 70%, jobs were classified and ranked according to presumed diesel exposure.7-68 Information is lacking regarding duration of employment in the job categories (used for surrogate of exposure) and other confounding factors (alcohol consumption, cigarette smoking, etc.). Thus, this study cannot be used to support or refute a causal association between exposure to DE and lung cancer. A 2-year mortality analysis by Boffetta and Stellman (1988) of the American Cancer Society’s prospective study, after controlling for age and smoking, demonstrated an excess risk of lung cancer in certain occupations with potential exposure to DE. These excesses were statistically significant among miners (RR = 2.67, 95% CI = 1.63, 4.37) and heavy equipment operators (RR = 2.6, 95% CI = 1.12, 6.06). Recently Brüske-Hohlfeld et al. (1999) also have observed significantly higher risk for lung cancer, in the range of 2.31 to 4.3, for heavy equipment operators. The elevated risks were nonsignificant in railroad workers (RR = 1.59) and truck drivers (RR = 1.24). A dose response was also observed for truck drivers. With the exception of miners, exposure to DE occurred in the three other occupations showing an increase in the risk of lung cancer. Despite methodologic limitations, such as the lack of representiveness of the study population (composed of volunteers only, who were probably healthier than the general population), leading to an underestimation of the risk, and the questionable reliability of exposure data based on self-administered questionnaires that were not validated, this study is suggestive of a causal association between exposure to DE and excess risk of lung cancer. Two mortality studies were conducted by Gustavsson et al. (1990) and Hansen (1993) among bus garage workers (Stockholm, Sweden) and truck drivers, respectively. An SMR of 122 was found among bus garage workers, based on 17 cases. A nested case-control study was also conducted in this cohort. Detailed exposure matrices based on job tasks were assembled for both DE and asbestos exposures. Statistically significant increasing lung cancer relative risks of 1.34, 1.81, and 2.43 were observed for DE indices of 10 to 20, 20 to 30, and >30, respectively, using 0 to 10 as a comparison group. Adjustment for asbestos exposure did not change the results. The main strength of this study is the detailed exposure matrices; some of the limitations are low power (small cohort) and lack of smoking histories. But smoking is not likely to be different among study individuals irrespective of their exposure status to DE. Hansen (1993), on the other hand, found statistically significant SMR of 160 from cancer of 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 quite a few methodologic limitations to this study. Exposure to DE was assumed in truck drivers for diesel-powered trucks, but no validation of exposure was attempted. Follow-up period was short, no latency analysis was done, and smoking data were lacking. However, a population survey carried out in 1988 showed very little difference in smoking habits of residents of rural area and the total Danish male population, thus, smoking is unlikely to confound the finding of7-69 excess lung cancer. The findings of both these studies are consistent with the findings of other truck driver studies and are supportive of causal association. Two mortality studies of railroad workers were conducted by Howe et al. (1983) and Garshick et al. (1988). The Howe et al. study, which was conducted in Canada, found relative risks of 1.2 ( p <0.01) and 1.35 ( p <0.001) among “possibly” and “probably” exposed groups, respectively. The trend test showed a highly significant dose-response relationship with exposure to DE and the risk of lung cancer. The main limitation of the study was the inability to separate overlapping exposures of coal dust/combustion fumes and DE fumes. Information on jobs was available at retirement only. There also was insufficient detail on the classification of jobs by DE exposure. The exposures could have been nonconcurrent or concurrent, but because the data are lacking, it is possible that the observed excess could be due to the effect of both coal dust/combustion fumes and DE fumes and not just one or the other. It should be noted that, so far, coal dust has not been demonstrated to be a pulmonary carcinogen in studies of coal miners. However, lack of data on confounders such as asbestos and smoking (though use of the internal comparison group to compute relative risks minimizes confounding by smoking) makes interpretation of this study difficult. When three DE exposure categories were examined for smoking-related diseases such as emphysema, laryngeal cancer, esophageal cancer, and buccal cancer, positive trends were observed, raising a possibility that the dose response demonstrated for diesel exposure may have been due to smoking. The findings of this study are at best suggestive of DE being a lung carcinogen. The strong evidence for linking DE exposure to lung cancer comes from the Garshick et al. (1988) railroad worker study conducted in the United States. Relative risks of 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 to 49, respectively, after the exclusion of workers exposed to asbestos. The investigators reported that the risk of lung cancer increased with increasing duration of employment. As this 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 factors such as asbestos, the observed association between increased lung cancer and exposure to DE is more meaningful. Even though the reanalysis of these data by Crump et al. (1991) found that the relative risk could be positively or negatively related to duration of exposure depending on how age was controlled, additional analysis by Garshick et al. ( letter from Garshick, Harvard Medical School, to Chao Chen, U.S. EPA, dated August 15, 1991 ) found that the relationship between years exposed when adjusted for the attained age and calendar years was flat to negative, depending on the choice of the model. They also found that deaths were underreported by approximately 20% to 70% between 1977 and 1980, and their analysis based on job titles, limited to 1959-1976, showed that the youngest workers still had the highest risk of dying of7-70 lung cancer. On the other hand, an analysis of the same data by California EPA (CalEPA, 1998) yielded a positive dose response set using age at 1959 and adding an interaction term of age and calendar year in the model. However, Crump (1999) reported that the negative dose-response continued to be upheld in his latest analysis when age was controlled more carefully and years of exposure quantified more accurately. Crump (1999) asserted that the negative dose-response trends for lung cancer observed with either the cumulative exposure or duration of exposure may be due to underascertainment of deaths in the last 4 years of follow-up of the Garshick et al. (1988) study, as well as incomplete follow-up in earlier years. The HEI (1999) special panel conducted its own analyses using Garshick et al. (1988) data to evaluate their usefulness for quantitative risk assessment and found results similar to those of Crump et al. (1991) and Garshick ( letter from Garshick, Harvard Medical School, to Chao Chen, U.S. EPA, dated August 15, 1991 ). The HEI panel reported consistently elevated risk of lung cancer for train workers compared with clerks for each duration of employment, and that shop workers had an intermediate risk of lung cancer. But they found decreasing risk of lung cancer with increasing duration of employment. The panel discussed various possibilities (different types of biases) for the negative dose-response and advised against using the Garshick et al. (1988) data for quantitative risk assessment. The panel also reported the strengths of the Garshick et al. (1988) study such as large population, control for asbestos, and smoking, and concluded that the study was generally consistent with findings of weak association between exposure to DE and occurrence of lung cancer. Hence, the divergent results of these recent analyses do not negatethe positive evidence this study provides for the qualitative evaluation. The observance of doseresponse would have strengthened the causal association, but an absence of a dose-response does not negate it. Suggestive evidence is provided by a recent study of potash miners in Germany. The information on the exposure (including elemental carbon and organics), work chronology, and work category was used by the investigators to calculate cumulative exposures for each worker. Furthermore, information on smoking habits indicated homogeneity in the cohort. A statistically nonsignificant twofold increase in lung cancer was observed in the production workers as compared to workshop workers. The lack of significance for this finding could be due to short follow-up, not enough latency, and relatively young age of the cohort.

7.2.4.2. Summary of the Case-Control Studies of Lung Cancer Among the 11 lung cancer case-control studies reviewed in this chapter, only 2 studies did not find any increased risk of lung cancer. Lerchen et al. (1987) did not find any excess risk of lung cancer, after adjusting for age and smoking, for diesel fume exposure. The major limitation of this study was a lack of adequate exposure data derived from the job titles obtained7-71 from occupational histories. Next of kin provided the occupational histories for 50% of the cases that were not validated. The power of the study was small (analysis done on males only, 333 cases). Similarly, Boffeta et al. (1990) did not find any excess of lung cancer after adjusting for smoking and education. This study had a few methodological limitations. The lung cancer cases and controls were drawn from the ongoing study of tobacco-related diseases. It is interesting to note that the leading risk factor for lung cancer is cigarette smoking. The exposure was not measured. Instead, occupations were used as surrogates for exposure. Furthermore, there were very few individuals in the study who were exposed to DE. On the other hand, statistically nonsignificant excess risks were observed for DE exposure by Hall and Wynder (1984) in workers who were exposed to DE versus those who were not (OR = 1.4 and 1.7 with two different criteria) and by Damber and Larsson (1987) in professional drivers (OR = 1.2). These rates were adjusted for age and smoking. Hall and Wynder (1984) had a high nonparticipation rate of 36%. Therefore, the positive results found in this study are underestimated at best. In addition, the self-reported exposures used in the study by Hall and Wynder (1984) were not validated. This study also had low power to detect excess risk of lung cancer for specific occupations. The study by Benhamou et al. (1988), after adjusting for smoking, found significantly increased risks of lung cancer among French motor vehicle drivers (RR = 1.42) and transport equipment operators (RR = 1.35). The main limitation of the study was the inability to separate exposures to DE from those to gasoline exhaust because both motor vehicle drivers and transport equipment operators probably were exposed to the exhausts of both types of vehicles. Hayes et al. (1989) combined data from three studies (conducted in three different states) to increase the power to detect an association between lung cancer and occupations with a high potential for exposure to DE. 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 study also found a significant trend of increasing risk of lung cancer with increasing duration of employment among truck drivers. The relative odds were computed by adjusting for birth cohort, smoking, and State of residence. The main limitation of this study is again the mixed exposures to diesel and gasoline exhausts, because information on type of engine was lacking. Also, potential bias may have been introduced because the way in which the cause of death was ascertained for the selection of cases varied in the three studies. Furthermore, the methods used in these studies to classify occupational categories were different, probably leading to incompatibility of occupational categories. Emmelin et al. (1993), in their Swedish dockworkers from 15 ports, found increased relative odds of 6.8 (90% CI = 1.3 to 34.9). A strong interaction between smoking and DE was7-72 observed in this study. Of 50 cases and 154 controls, only 6 individuals were nonsmokers. Although intricate exposure matrices were created using three different variables, no direct exposure measurement was done. Despite the limitations of small number of cases and controls; lack of data on asbestos exposure, which is fairly common in dockworkers; and very few nonsmokers; this study provides consistent support for a real effect of DE exposure and occurrence of lung cancer, at least in smokers. The most convincing evidence comes from the case-control studies among railroad workers by Garshick et al. (1987); among truck drivers of the Teamsters Union by Steenland et al. (1990, 1998); among truck drivers, railroad workers, and farmers in a population-based study by Swanson et al. (1993); among different professional drivers in Denmark by Hansen et al. (1998); and among male workers occupationally exposed to diesel motor emissions in Germany by Brüske-Hohlfeld et al. (1999). Garshick et al. (1987) found that after adjustment for asbestos and smoking, the relative odds for continuous exposure were 1.39 (95% CI = 1.05, 1.83). Among the younger workers with longer DE exposure, the risk of lung cancer increased with duration of exposure after adjusting for asbestos and smoking. Even after the exclusion of recent DE exposure (5 years before death), the relative odds increased to 1.43 (95% CI = 1.06, 1.94). This appears to be a well-conducted and well-analyzed study with reasonably good power. Potential confounders were controlled adequately, and interactions between DE and other lung cancer risk factors were tested. Some of the limitations of this study are misclassification of exposure because ICC job classification was used as surrogate for exposure
and use of death certificates for identification of cases and controls. Steenland et al. (1990), on the other hand, created two separate work history files, one from Teamsters Union pension files and the other from next-of-kin interviews. Using duration of employment as a categorical variable and considering employment after 1959 (when presumed dieselization occurred) for long-haul drivers, the risk of lung cancer increased withincreasing years of exposure. Using 1964 as the cutoff, a similar trend was observed for longhaul drivers. For short-haul drivers, the trend was positive with a 1959 cutoff, but not when 1964 was used as the cutoff. For truck drivers who primarily drove diesel trucks and worked for 35 years, the relative odds were 1.89. The main strengths of the study are availability of detailed records from the Teamsters Union, a relatively large sample size, availability of smoking data, and measurements of exposure. The limitations of this study include possible misclassifications of exposure and smoking, lack of levels of diesel exposure, a smaller nonexposed group, and an 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. (1991). The estimates were made for long-haul drivers, short-haul drivers, dockworkers, mechanics, and those outside the trucking industry. The survey found that mechanics had the7-73 highest current levels of DE exposures and dockworkers who mainly used propane- powered forklifts had the lowest exposure. The finding of the highest lung cancer risk for mechanics and lowest for dock workers is indicative of a causal association between the DE exposure and development of lung cancer. However, the risk among mechanics did not increase with increasing duration of employment. The ORs for quartile cumulative exposures, computed by using logistic regression adjusted for age, race, smoking, diet, and asbestos exposure, showed a pattern of increasing trends in risk with increasing exposure, between 1.08 and 1.72 depending upon exposure level and lag structure used. In a population-based lung cancer case-control study Swanson et al. (1993) found statistically significant excess risks adjusted for age at diagnosis, smoking, and race, among white male drivers of heavy trucks employed for 

### $ 20 years and railroad workers employed for



### $ 10 years (OR = 2.5, 95% CI = 1.1, 4.4, and OR = 2.4, 95 % CI = 1.1, 5.1, respectively), and among black farmers employed for 

### $ 20 years (OR = 10.4, 95% CI = 1.4, 77.1). Although individual ORs were not significant for various occupations with potential exposure to DE, statistically significant trends were observed for drivers of heavy trucks, light trucks, farmers, and railroad industry workers among whites, and among black farmers ( p 

### # 0.05). The main strengths of the study are availability of data on lifetime work history and smoking history; the main limitation is absence of actual specific exposure data. This is the first study that found increased lung cancer risk for farmers, who are exposed to DE of their farm tractors. Hansen et al. (1998), in their study of professional drivers in Denmark, found statistically significant ORs (adjusted for socioeconomic status) of 1.31, 1.64, and 1.39 for lorry/bus drivers, taxi drivers, and unspecified drivers, respectively. The lag time analyses for duration of employment were unchanged for lorry/bus drivers but increased to OR = 3 from 2.2 in taxi 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 because of longer time spent in traffic congestion. Furthermore, the trend tests for increasing risk of lung cancer with increasing duration of employment were statistically significant for both lorry/bus drivers and taxi drivers in both 10-year lag time and no lag time. The main strengths of the study are the large sample size, availability of detailed employment records, and information on socioeconomic status. The main limitations are absence of individual data on smoking habits and asbestos exposure, and information about the type of fuel used for the vehicles driven by these professional drivers. A personal communication with the main investigator revealed that the lorries/buses and taxis have been using diesel fuel since the late 1940s. Moreover, indirect information about smoking and asbestos exposure indicated that these two confounders are unlikely to explain the observed excesses or the trends, resulting in strong support of earlier positive studies.7-74 Brüske-Hohlfeld et al. (1999) recently conducted a pooled analysis of two case-control studies among male workers occupationally exposed to DME in Germany. The investigators collected data on demographic information, detailed smoking, and occupational history. Job titles and industries were classified in 33 and 21 categories respectively. Job descriptions were written and verified to avoid misclassification of estimated exposure to diesel emissions. Individual cumulative DME exposures and smoking pack-years were calculated. Asbestos exposures were estimated by certain job-specific supplementary questions. Analysis of 3,498 lung cancer cases and 3,541 controls yielded statistically significant ORs ranging from 1.25 to 2.31 adjusted for smoking and asbestos exposure. The risk increased with increasing years of exposure for both the first year of exposure and the end year of exposure. These investigators presented analyses by various job categories, by years of exposure, first and end years of exposure and, when possible, separately for West and East Germany. Significantly higher risks were found among all four job categories. For professional drivers (of trucks, buses, and taxis) ORs ranged from 1.25 to 2.53. For other traffic-related jobs (switchmen, diesel locomotive drivers, diesel forklift truck drivers), ORs ranged from 1.53 to 2.88. For heavy equipment operators (bulldozers, graders, and excavators), ORs ranged from 2.31 to 4.3, and for drivers of farming equipment the only significant excess (OR = 6.81) was for exposure for <30 years. This study shows increased risk for all the DME-exposed job categories. The professional drivers and the other traffic-related jobs also have some mixed exposures to gasoline exhaust in general traffic. On the other hand, it should be noted that exposure to DME among heavy equipment and farm tractor drivers is much higher and not as mixed as in professional drivers. The heavy equipment drivers usually drive repeatedly through their own equipment’s exhaust. Therefore, the observed highest risk for lung cancer in this job category establishes a strong link with the DME. The only other study that found significantly higher risk for heavy equipment operators (RR = 2.6) was conducted by Boffeta et al. (1988). Although the only significant excess in the group was observed for farming tractor operators with more than 30 years of exposure, a steady increase in risk was observed for this job category with increasing exposure. The investigators stated that the working conditions and the DME of tractors remained fairly constant over the years. This increase may be due mainly to exposure to DME
and PM 10. The main strengths of the study are large sample size, resulting in good statistical power; inclusion of incident cases diagnosed not more than 3 months prior to the interview; use of only personal interviews, reducing recall bias; diagnoses ascertained by cytology or histology; and availability of lifelong detailed occupational and smoking history. Exposure estimation done for each individual was based on job codes and industry codes, which were validated by written job descriptions to avoid misclassification.7-75 The main limitation of the study is lack of data on actual exposure to DME. The cumulative quantitative exposures were calculated based on time spent in each job with potential exposure to DME and the type of equipment used. Thus, this study provides strong evidence for causal association between exposure to DE and occurrence of lung cancer.

7.2.4.3. Summary of the Reviews and Meta-Analyses of Lung Cancer Three summaries of studies concerned with the relationship of DE exposure and lung cancer risk are reviewed. The HEI report is a narrative study of 35 epidemiologic studies (16 cohort and 19 case-control) of occupational exposure to diesel emissions published between 1957 and 1993. Control for smoking was identified in 15 studies. Six of the studies (17%) reported relative risk estimates less than 1, whereas 29 (83%) reported at least 1 excess relative risk, indicating a positive association. Twelve studies indicating a relative risk greater than 1 had 95% confidence intervals that excluded unity. These studies found that the evidence suggests that occupational exposure to DE from diverse sources increases the rate of lung cancer by 20% to 40% in exposed workers generally, and to a greater extent among workers with prolonged exposure. They also found that the results are not explicable by confounding due to cigarette smoking or other known sources of bias.Bhatia et al. (1998) identified 23 studies that met criteria for inclusion in the metaanalysis. The observed relative risk estimates were greater than 1 in 21 of these studies. The pooled relative risk weighted by study precision was 1.33 (95% CI= 1.24, 1.44), which indicated increased relative risk for lung cancer from occupational exposure to DE. Subanalyses by study design (case-control and cohort studies) and by control for smoking produced results that did not differ from those of the overall pooled analysis. Cohort studies using internal comparisons showed higher relative risks than those using external comparisons. Lipsett and Campleman (1999) identify 39 independent estimates of relative risk among 30 eligible studies of DE and lung cancer published between 1975 and 1995. Pooled relative risks for all studies and for study subsets were estimated using a random effect model. Interstudy heterogeneity was also modeled and evaluated. A pooled smoking-adjusted relative risk was 1.47 (95% CI = 1.29, 1.67). Substantial heterogeneity was found in the pooled-risk estimates. Adjustment for confounding by smoking, having a lower likelihood of selection bias, and increased study power were all found to contribute to lower heterogeneity and increased pooled estimates of relative risk. There is some variability in the conclusions of these summaries of the association of DE and lung cancer. The three analyses find that smoking is unlikely to account for the observed effects, and all conclude that the data support a causal association between lung cancer and DE exposure. On the other hand, Stöber and Abel (1996), Muscat and Wynder (1995), and Cox7-76 (1997) call into question the assertions by Cohen and Higgins (1995), Bhatia et al. (1998), and Lipsett and Campleman (1999) that the associations seen for DE 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 confounding by cigarette smoking, and thus, the observed association between exposure to DE and excess risk of lung cancer is more likely to be due to bias. The conclusions of the two meta-analyses are based on magnitude of pooled relative risk estimates and evaluation of potential sources of heterogeneity in the estimates. Despite the statistical sophistication of the meta-analyses, the statistical models used cannot compensate for deficiencies in the original studies and will remain biased to the extent that bias exists in the original studies.

7.2.4.4. Discussion of Relevant Methodologic Issues A persistent association of risk for lung cancer and DE exposure has been observed in more than 30 epidemiologic studies published in the literature over the past 40 years. Evaluation of whether this association can be attributed to a causal relation between DE exposure and lung cancer requires careful consideration of whether chance, bias, or confounding might be likely alternative explanations. A total of 10 cohort and 12 case-control studies are reviewed in this chapter. An increased lung cancer risk was observed in 8 cohort and 10 case-control studies, even though the results were not always statistically significant. There is a consistent tendency for point estimates of relative risk to be greater than one in studies that adjusted (either directly or indirectly) for smoking, had a long enough follow-up, and sufficient statistical power among truck drivers, railroad workers, dock workers, and heavy equipment workers. If this elevated risk was due to chance one would expect almost equal distribution of these point estimates to be above and below one. Many of the studies provide confidence intervals for their estimates of excess risk or statistical tests, which indicate that it is unlikely that the individual study findings were due to random variation. The persistence of this association between DE and lung cancer risk in so many studies indicates that the possibility is remote that the observed association in aggregate is due to chance. It is unlikely that chance alone accounts for the observed relation between DE and lung cancer. The excess risk is observed in both cohort and case-control designs, which contradicts the concern that a methodologic bias specifically characteristic of either design (e.g., recall bias) might account for the observed effect. Selection bias is certainly present in some of the occupational cohort studies that use external population data in estimating relative risks, but this form of selection bias (a healthy worker effect) would only obscure, rather than spuriously produce, an association between DE and lung cancer. Several occupational epidemiologic7-77 studies that use more appropriate data for their estimates are available. Selection biases may be operating in some case-control studies, but it is not obvious how such a bias could be sufficiently uniform in effect, prevalent, and strong enough to lead to the consistent association seen in the aggregate data. Given the variety of designs used in studying the DE 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 association seen. Exposure information bias is certainly a problem for almost all of the studies concerned. Detailed and reliable individual-level data on DE exposure for the period of time relevant to the induction of lung cancer are not available and are difficult to obtain. Generally, the only information from which diesel exposure can be inferred is occupational data, which is a poor surrogate for the true underlying exposure distribution. The variability in actual lifetime exposure to DE in an occupational cohort may not be reflected in differences in job title, and there might be considerable variability in actual exposure despite similar job titles. Study endpoints are frequently mortality data taken from death certificate information, which is frequently inaccurate and often does not fully characterize the lung cancer incidence experience of the population in question. Using inaccurate surrogates for lung cancer incidence and for diesel exposure can lead to substantial bias, and these shortcomings are endemic in the field. In most cases these shortcomings will lead to misclassification of exposure and of outcome, which is nondifferential. Nondifferential misclassification of exposure and/or outcome can bias estimates of a DE–lung cancer association, if one exists, toward the null; but it is unlikely that such misclassification would produce a spurious estimate in any one study. It is even more unlikely that it would bias a sufficient number of studies in a uniform direction to account for the consistent 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 confounding variables, use of death certificates to identify the lung cancer cases, and lack oflatency analysis. The studies with small sample sizes (i.e., not enough power) and short followup periods (i.e., not enough latent period) have been difficult to interpret due to these limitations. The most important confounding variable is smoking which is a strong risk factor for lung cancer. All the studies considered for this report are either cohort retrospective mortality or case-control studies where history of exposures in the past is elicited. Smoking history is usually difficult to obtain in such instances. The smoking histories obtained from surrogates (next of kin, either spouse or offspring) were found to be accurate by Lerchen and Samet (1986) and McLaughlin et al. (1987). Lerchen and Samet did not detect any consistent bias in the report of cigarette consumption. In contrast, overreporting of cigarette smoking by surrogates was observed by Rogot and Reid (1975), Kolonel et al. (1977), and Humble et al. (1984). Kolonel et7-78 al. found that the age at which an individual started smoking was reported within 4 years of actual age 84% of the time. These studies indicate that surrogates were able to provide fairly credible information on the smoking habits of the study subjects. If the surrogates of the cases were more likely to overreport cigarette smoking compared with the controls, then it might be harder to find an effect of DE because most of the increase in lung cancer would be attributed to smoking rather than to exposure to DE. Some studies do not adjust for tobacco smoke exposure. Even though smoking is a strong risk for lung cancer, it is only a confounder if there are differential smoking habits among individuals exposed to DE versus individuals who are not exposed. Most of the occupational cohorts include workers from the same socioeconomic background or used an internal comparison group; hence, it is unlikely that confounding by cigarette smoking is substantial in these studies. Some studies have adjusted for socioeconomic status and some studies have compared the cigarette smoking habits by conducting rural and urban general population surveys. Besides, in studies with long enough latency, adjustment for cigarette smoking did not alter substantially the observed higher risk. Another methodologic concern in these studies is use of death certificates to determine cause of death. Death certificates were used by all of the cohort mortality studies and some of the case-control studies of lung cancer to determine cause of death. Use of death certificates could lead to misclassification bias because of overdiagnosis. Studies of autopsies done between 1960 and 1971 demonstrated that lung cancer was overdiagnosed when compared with hospital discharge, with no incidental cases found at autopsy (Rosenblatt et al., 1971). Schottenfeld et al. (1982) also found an overdiagnosis of lung cancer among autopsies conducted in 1977 and 1978. On the other hand, Percy et al. (1981) noted 95% concordance when comparing 10,000 lung cancer deaths observed in the Third National Cancer Survey from 1969 to 1971 (more than 90% were confirmed histologically) to death-certificate-coded cause of death. These more recent findings suggest that the diagnosis of lung cancer on death certificates is better than anticipated. In reality, lung cancer is one cause of death that has been found to be generally reliably reported on the death certificate. Thus, the misclassification bias probably is minimal in the studies described in this chapter. Finally, several investigators have not conducted latency analysis in their studies. The latent period for lung cancer development is from 20 to 30 years or more. Considering the fact that dieselization was not complete till almost 1959 for locomotives and the 1970s for the trucking industry in the United States, most of the cohort studies conducted in the U.S. population do not have a long enough follow-up period to allow for latency of 20 to 30+ years. In addition, the study inclusion criteria for most of the studies are individuals who worked in the industry for at least 6 months /1 year from the beginning of the follow-up period to the end of3 Hill in his address to the Royal Society of Medicine in 1965 on “The environment and disease: association or causation” explored several aspects of association between exposure and occurrence of an event before deciding that the most likely interpretation of it is causation. He provided nine different aspects of association that he characterized as his viewpoints before interpreting the association being causal. The epidemiologic community universally adopted these (aspects/viewpoints) later as criteria for causality/causal association.

7-79 the follow-up period. Hence, the later the individual enters the cohort, the shorter the follow-up period; thus, the latent period is insufficient for the occurrence of lung cancer in these late entrants. Therefore, the observed slight to moderate increase in risk of lung cancer could be due to insufficient latency. On the other hand, in certain case-control studies the elapsed period between the identification of the lung cancer cases and exposure to DE is long enough to allow for the 30+ years latency needed for the development of lung cancer (Hansen et al., 1998; Brüske-Hohlfeld et al., 1999). These investigators identified lung cancer cases in the early to mid-1990s and found significant excess risks for lung cancer among the individuals exposed to DE. It should be noted that the use of diesel fuel for trucks, buses, and taxis had started in their countries (Denmark and Germany, respectively) in the late 1940s.

7.2.4.5. Evaluation of Causal Association In most situations, epidemiologic data are used to delineate the causality of certain health effects. Several cancers have been causally associated with exposure to agents for which there is no direct biological evidence. Insufficient knowledge about the biological basis for diseases in humans makes it difficult to identify exposure to an agent as causal, particularly for malignant diseases when the exposure was in the distant past. Consequently, epidemiologists and biologists have used the original or modified version of a set of criteria provided by Hill (1965) 3 that define a causal relationship between exposure and the health outcome. A causal interpretation is enhanced for studies that meet these criteria. None of these criteria actually proves causality; actual proof is rarely attainable when dealing with environmental carcinogens. None of these criteria should be considered either necessary (except 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. Thus, applying the Hill criteria (1965) of causal inference, as modified by Rothman (1986), to the studies reviewed here resulted in the following:

• Strength of association. This phrase refers to the magnitude of the ratio of incidence or mortality (RRs or ORs). Several studies found statistically significant RRs and ORs that ranged from 1.2 to 2.6 (Howe et al., 1983; Rushton7-80 et al., 1983; Wong et al., 1985; Gustavsson et al., 1990; Emmlin et al., 1993; Hansen, 1993; Hansen et al., 1998) and, after adjustment for smoking and/or asbestos, RRs and ORs remained statistically significant and in the same range in certain studies (Dambar and Larson 1987; Garshick et al., 1987, 1988; Benhamou et al., 1988; Boffetta and Stellman, 1988; Hays et al., 1989; Steenland et al.,1990; Swanson et al., 1993; Brüsk-Hohlfeld et al., 1999). In addition, two metaanalyses demonstrated that not only did excess in lung cancer remain the same 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 to weak association between DE and occurrence of lung cancer. Even though strong associations are more likely to be causal than modest-to-weak associations, the fact that association is relatively modest or weak does not rule
out the causal link.

• Consistency. Increased lung cancer risk has been observed in several cohort and case-control studies, conducted in several industries and occupations in which workers were potentially exposed to DE. However, not all the excesses were statistically significant. Statistically significant lung cancer excesses adjusted for smoking were observed in truck drivers (Hayes et al., 1989; Hansen, 1993; Swanson et al., 1993; Brüske-Hohlfeld et al., 1999), professional drivers (Benhamou et al., 1988; Brüske-Hohlfeld et al., 1999), railroad workers (Garshick et al., 1987; Swanson et al., 1993), heavy equipment drivers (Boffetta and Stellman, 1988; Brüske-Hohlfeld et al., 1999), and farm tractor drivers (Swanson et al., 1993; Brüske-Hohlfeld et al., 1999). Furthermore, the two recent meta-analyses by Bhatia et al. (1998) and Lipsett and Campleman (1999) found that even though a substantial heterogeneity existed in their initial pooled estimates, stratification on several factors demonstrated a relationship between exposure to DE and excess lung cancer that remained positive throughout various analyses.

• Specificity. This criterion requires that a single cause lead to a single effect. With respect to exposure to DE, excess for lung cancer is the only effect that is found to be consistently elevated and statistically significant in several studies. Quite a few studies have examined DE for other effects such as bladder cancer, leukemia,7-81 gastrointestinal cancers, prostate cancer etc. The evidence for these effects is inadequate.

• Temporality. The only necessary, but not sufficient, criterion described by Hill for causality inference is that exposure to a causal agent precedes the effect in time. This criterion is clearly satisfied in the studies reviewed here. Temporality can be explored further in addressing the latency issue. A certain period is necessary for development of an effect after exposure to a causal agent has occurred. For instance, in cancer-causing agents a latent period can vary from 5 years (childhood leukemia) to 

### $ 30 years (mesothelioma). Most of the studies 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, 1981; Howe et al., 1983; Rushton et al., 1983; Wong et al., 1985, Hansen, 1993) while several studies clearly allowed for an adequate latency period (Garshick et al., 1987; Gustavsson et al., 1990; Steenland et al., 1990; Swanson et al., 1993; Brüske-Hohlfeld et al., 1999). Both type of studies showed mixed results.

• Biological gradient. This criterion refers to the dose-response curve. Due to the lack of quantitative data on DE exposure in most studies reviewed here, analyzing the dose-response curve directly was not possible. In very few studies, exposure to DE was addressed specifically. Most investigators have used job titles/categories and duration of employment as surrogates for exposure and thushave presented response in relation to duration of employment. Significant doseresponse (using duration of employment as a surrogate) was observed in various studies for railroad workers (Howe et al., 1983; Garshick et al., 1987; Garshick et al., 1988; Swanson et al., 1993; Cal EPA, 1998), truck drivers (Boffetta and Stellman, 1988; Hayes et al., 1989; Steenland et al., 1990; Swanson et al., 1993; Hansen et al., 1998; Brüske-Hohlfeld et al., 1999), transportation/heavyequipment operators (Wong et al., 1985; Gustavsson et al., 1990; BrüskeHohlfeld et al., 1999), farmers/farm tractor users (Swanson et al., 1993; BrüskeHohlfeld et al., 1999), and dockworkers (Emmelin et al., 1993).

• Biological plausibility. This criterion refers to the biologic plausibility of the hypothesis, an important concern that may be difficult to judge. The hypothesis considered for this review is that occupational exposure to DE is causally associated with the occurrence of lung cancer and is supported by the following:7-82 First, DE has been shown to cause lung and other cancers in animals (Heinrich et al., 1986b; Iwai et al., 1986b; Mauderly et al., 1987; Pott et al., 1990; Mauderly, 1994). Second, it contains highly mutagenic substances such as polycyclic aromatic hydrocarbons as well as nitroaromatic compounds (Claxton, 1983; Ball et al., 1990; Gallagher et al., 1993; Sera et al., 1994; Nielsen et al., 1996a) that are recognized human pulmonary carcinogens (IARC, 1989). Third, DE consists of carbon core particles with surface layers of 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 mixture of organic constituents, both in particles and vapor phases, have the capacity to interact with DNA and give riseto mutations, chromosomal aberrations, and cell transformations, all wellestablished steps in the process of carcinogenesis. Further, increased levels of peripheral blood cell DNA adducts associated with occupational exposure to DE have been observed in humans (Nielsen et al., 1996a,b). Thus, the above evidence makes a convincing case that occupational exposures to DE are causally associated with the occurrence of lung cancer is highly plausible biologically.

In conclusion, the epidemiologic studies of exposure to DE and occurrence of lung
cancer furnish evidence that is consistent with a causal association. This association observed in several studies is unlikely to be due to chance or bias. Although many studies did not have information on smoking, significant confounding by smoking is unlikely in these studies because the comparison population was from the same socioeconomic class. The strength of association(i.e., RRs/ORs between 1.2 and 2.6) was weak to modest by epidemiologic standards, with doseresponse relationships observed in several studies. Last, but not least, there is highly plausible biological evidence that exposure to DE could result in excess risk of lung cancer in humans.

7.3. CARCINOGENICITY OF DIESEL EXHAUST IN LABORATORY ANIMALS This chapter summarizes studies that assess the carcinogenic potential of DE in laboratory animals. The first portion of this chapter summarizes results of inhalation studies. Experimental protocols for the inhalation studies typically consisted of exposure (usually chronic) to diluted exhaust in whole-body exposure chambers using rats, mice, and hamsters as model species. Some of these studies used both filtered (free of particulate matter) DE and unfiltered (whole) DE to differentiate gaseous-phase effects from effects induced by diesel PM (DPM) and its adsorbed components. Other studies were designed to evaluate the relative importance of the carbon core of the diesel particle versus that of particle-adsorbed compounds.7-83 Finally, a number of exposures were carried out to determine the combined effect of inhaled DE and tumor initiators, tumor promoters, or cocarcinogens. Particulate matter concentrations in the DE used in these studies ranged from 0.1 to 12 mg DPM /m 3 . In this chapter, any mention of statistical significance implies that p 

### # 0.05 was reported in the reviewed publications. A summary of the animal inhalation carcinogenicity studies and their results is presented in Table 7-3. Results of lung implantation and intratracheal instillation studies of whole diesel particles, extracted diesel particles, and particle extracts are reported in Section 7.3.3 and in Tables 7-4 and 7-5. Studies destined to assess the carcinogenic effects of DPM as well as solvent 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 present in the gaseous phase or adsorbed to the particle surface were not included in this review because assessments of those of likely concern (i.e., formaldehyde, acetaldehyde, benzene, polycyclic aromatic hydrocarbons [PAHs]) have been published elsewhere (U.S. EPA, 1993).

7.3.1. Inhalation Studies (Whole Diesel Exhaust)
7.3.1.1. Rat Studies The potential carcinogenicity of inhaled DE was first evaluated by Karagianes et al. (1981). Male Wistar rats (40 per group) were exposed to room air or diesel engine exhaust diluted to a DPM concentration of 8.3 (± 2.0) mg/m 3 , 6 hr/day, 5 days/week for up to 20 months. The animals were exposed in 3,000 L plexiglass chambers. Airflow was equal to 50 liters per minute. Chamber temperatures were maintained between 25 o C and 26.5 o C. Relative humidity ranged from 45% to 80%. Exposures were carried out during the daytime. The connected to an electric generator and operated at varying loads and speeds to simulate operating conditions in an 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. The only tumor detected was a bronchiolar adenoma in the group exposed over 16 months to DE. No lung tumors were reported in controls. The equivocal response may have been caused by the relatively short exposure durations (20 months) and small numbers of animals examined. In more recent studies, for example, Mauderly et al. (1987), most of the tumors were detected in rats exposed for more than 24 months.7-84Table 7-3. Summary of animal inhalation carcinogenicity studiesStudySpecies/strainSex/totalnumberExposureatmosphereParticleconcentration(mg/m3)OthertreatmentExposureprotocolPostexposureobservation Tumor typeand incidence (%)aCommentsAdenomasKaragianeset al. (1981)Rat/Wistar M, 40M, 40Clean airWholeexhaust8.3 NoneNone6 hr/day,5 days/week,for up to20 moNA 0/6 (0)1/6 (16.6)Bronchoalveolar carcinoma0/30 (0)1/30 (3.3)3/30 (10.0)1/30 (3.3)Kaplan et al.(1983)White et al.(1983)Rat/F344 M, 30M, 30M, 30M, 30Clean airWholeexhaustWholeexhaustWholeexhaust00.250.751.5NoneNoneNoneNone20 hr/day,7 days/week,for up to15 mo8 mo8 mo8 mo8 moAdenomasCarcinomasSquamouscell tumorsAll tumorsHeinrich etal. (1986a,b)Mohr et al.(1986)Rat/WistarF, 96F, 92F, 95Clean airFilteredexhaustWholeexhaust4 NoneNoneNone19 hr/day,5 days/weekfor up to35 moNA 0/96 (0)0/92 (0)8/95 (8.4)0/96 (0)0/92 (0)0/95 (0)0/96 (0)0/92 (0)9/95 (9.4)0/96 (0)0/92 (0)17/95(17.8)cAdenomasAdenocarcinomaandadenosquamouscarcinomaLarge cellandsquamouscellcarcinomasAll tumorsIwai et al.(1986a,b)Rat/F344 F, 24F, 24F, 24Clean airFilteredexhaustWholeexhaust4.9 NoneNoneNone8 hr/day,7 days/week,for 24 moNA 1/22 (4.5)0/16 (0)3/19 (0)0/22 (0)0/16 (0)3/19 (15.8)0/22 (0)0/16 (0)2/19 (10.5)1/22 (4.5)f0/16 (0)8/19(42.1)c,gTable 7-3. Summary of animal inhalation carcinogenicity studies (continued)StudySpecies/strainSex/totalnumberExposureatmosphereParticleconcentration(mg/m3)OthertreatmentExposureprotocolPostexposureobservation Tumor typeand incidence (%)aCommentsAdenomaCarcinomaTakemoto etal. (1986)Rat/F344 F, 12F, 21F, 15F, 18Clean airClean airWholeexhaustWholeexhaust002-42-4NoneDIPNhNoneDIPNh4 hr/day,4 days/week,18-24 moNA 0/12 (0)10/21 (47.6)0/15 (0)12/18 (66.7)0/12 (0)4/21 (19)0/15 (0)7/18 (38.9)AdenomasAdenocarcinoma+ squamous cellcarcinomaSquamouscystsAlltumorsMauderly etal. (1987)Rat/F344 M + F,230bM + F, 223M + F, 221M + F, 227Clean airWholeexhaustWholeexhaustWholeexhaust00.353.57.1NoneNoneNoneNone7 hr/day,5 days/week up to30 moNA (0)(0)(2.3)(0.4)(0.9)(1.3)(0.5)(7.5)(0)(0)(0.9)(4.9)(0.9)(1.3)(3.6)c(12.8)cIshinishi etal. (1988a)Heavy-dutyengineRat/F344 M + F, 123M + F, 123M + F, 125M + F, 123M + F, 124Clean airWholeexhaustWholeexhaustWholeexhaustWholeexhaust00.51.01.83.7NoneNoneNoneNoneNone16 hr/day,6 days/week,for up to30 moNAAdenomas0/123 (0)0/123 (0)0/125 (0)0/123 (0)0/124 (0)Adenosquamouscarcinomas1/123 (0.8)0/123 (0)0/125 (0)4/123 (3.3)6/124 (4.8)Squamouscellcarcinomas0/123 (0)1/123 (0.8)0/125 (0)0/123 (0)2/124 (1.6)All tumors1/123 (0.8)1/123 (0.8)0/125 (0)4/123 (3.3)8/124(6.5)c

7-85Table 7-3. Summary of animal inhalation carcinogenicity studies (continued)StudySpecies/strainSex/totalnumberExposureatmosphereParticleconcentration(mg/m3)OthertreatmentExposureprotocolPostexposureobservation Tumor typeand incidence (%)aCommentsAdenomasCarcinomasAll tumorsIshinishi etal.(1988a)Light dutyRat/F344 NS, 5NS, 8NS, 11NS, 5NS, 9NS, 11WholeexhaustWholeexhaustWholeexhaustWholeexhaustWholeexhaustWholeexhaust0.10.10.11.11.11.1NoneNoneNoneNoneNoneNone16 hr/day,6 days/week,for 12 mo6 mo12 mo18 mo6 mo12 mo18 mo0/5 (0)0/8 (0)0/11 (0)0/5 (0)0/9 (0)0/11 (0)0/5 (0)0/8 (0)0/11 (0)0/5 (0)0/9 (0)0/11 (0)0/5 (0)0/8 (0)0/11 (0)0/5 (0)0/9 (0)0/11 (0)Heavy duty NS, 5NS, 9NS 11NS, 5NS, 6NS, 13WholeexhaustWholeexhaustWholeexhaustWholeexhaustWholeexhaustWholeexhaust0.50.50.51.81.81.8NoneNoneNoneNoneNoneNone16 hr/day,6 days/week,for 12 mo6 mo12 mo18 mo6 mo12 mo18 mo0/5 (0)0/9 (0)0/11 (0)0/5 (0)0/6 (0)0/13 (0)0/5 (0)0/9 (0)0/11 (0)0/5 (0)0/6 (0)1/13 (0)0/5 (0)0/9 (0)0/11 (0)0/11 (0)0/6 (0)1/13 (0)

7-86Table 7-3. Summary of animal inhalation carcinogenicity studies (continued)StudySpecies/strainSex/totalnumberExposureatmosphereParticleconcentration(mg/m3)OthertreatmentExposureprotocolPostexposureobservation Tumor typeand incidence (%)aCommentsPrimary lung tumorsBrightwellet al. (1989)Rat/344 M + F, 260M + F, 144M + F, 143M + F, 143M + F, 144M + F, 143Clean airFilteredexhaust(mediumexposure)Filteredexhaust(highexposure)WholeexhaustWholeexhaustWholeexhaust0000.72.26.6NoneNoneNoneNoneNoneNone16 hr/day,5 days/week,for 24 moNA 3/260 (1.2)0/144 (0)0/143 (0)1/143 (0.7)14/144 (9.7)c55/143 (38.5)cTumorincidence forall rats dyingor sacrificed&24/25(96%) after 24mo%12/27(44%)after 24 moHenrich etal. (1989a)Rat/Wistar F, NSF, NSF, NSF, NSF, NSF, NSClean airWholeexhaustFilteredexhaustClean airWholeexhaustFilteredexhaust04.2004.20DPNdDPNdDPNdDPNeDPNeDPNe19 hr/day,5 days/weekfor 24 to30 moNASquamouscellcarcinoma(4.4)(46.8)c(4.4)(16.7)(31.3)c(14.6)All lungtumors(84.8)(83.0)(67.4)(93.8)(89.6)(89.6)Lewis et al.(1989)Rat/F344 M + F,288nClean airWholeexhaust2 NoneNone7 hr/day,5 days/week,24 moNA No tumors 0/192 (0)0/192 (0)

7-87Table 7-3. Summary of animal inhalation carcinogenicity studies (continued)StudySpecies/strainSex/totalnumberExposureatmosphereParticleconcentration(mg/m3)OthertreatmentExposureprotocolPostexposureobservation Tumor typeand incidence (%)aCommentsAdenosquamouscarcinomasSquamous cellcarcinomasAll tumorsTakaki et al.(1989)Light-dutyengineRat/F344 M + F, 123M + F, 123M + F, 125M + F, 123M + F, 124Clean airWholeexhaustWholeexhaustWholeexhaustWholeexhaust00.10.41.12.3NoneNoneNoneNoneNone16 hr/day,6 days/week, forup to30 moNA 1/23 (0.8)1/23 (0.8)1/25 (0.8)0/23 (0)1/24 (8.1)2/123 (1.6)1/23 (0.8)0/125 (0)5/123 (4.1)2/124 (1.6)1/23 (0.8)1/23 (0.8)0/125 (0)0/123 (0)0/124 (0)4/123 (3.3)3/123 (2.4)1/125 (0.8)5/123 (4.1)3/124 (2.4)AdenomasAdenocarcinomasSquamouscellcarcinomasBenignsquamouscelltumorsHeinrich etal.(1995)Rat/Wistar F, 220F, 200F, 200F, 100F, 100F, 100Clean airWholeexhaustWholeexhaustWholeexhaustCarbon blackTiO200.82.57.011.610.0NoneNoneNoneNoneNoneNone18 hr/day,5 days/week,for up to24 mo6 mo 0/217 (0)0/198 (0)2/200 (1)4/100 (4)13/100(13)4/100 (4)1/217 (<1)0/198 (0)1/200 (<1)4/100 (4)13/100 (13)13/100 (13)0/217 (0)0/198 (0)0/200 (0)2/100 (2)4/100 (4)3/100 (3)0/217 (0)0/198 (0)7/200 (3.5)14/100 (14)20/100 (20)20/100 (20)Tumorincidencesafter 30 moAdenomasAdenocarcinomasSquamouscellcarcinomaAdenosquamouscarcinomaOtherneoplasmsNikula et al.(1995)Rat/F344 M + F,214bM + F, 210M + F, 212M + F, 213M + F, 211Clean airWholeexhaustWholeexhaustCarbon blackCarbon black02.56.52.56.5NoneNoneNoneNoneNone16 hr/day,5 days/week forup to24 mo6 weeks 1/214 (<1)7/210 (3)23/212(11)3/213 (1)13/211 (6)1/214 (<1)4/210 (2)22/212 (10)7/213 (3)21/211 (10)1/214 (<1)3/210 (1)3/212 (1)0/213 (0)3/211 (1)0/214 (0)0/210 (0)1/212 (<1)0/213 (0)2/211 (<1)0/214 (0)0/210 (0)0/212 (0)1/213 (<1)0/211 (0)

7-88Table 7-3. Summary of animal inhalation carcinogenicity studies (continued)StudySpecies/strainSex/totalnumberExposureatmosphereParticleconcentration(mg/m3)OthertreatmentExposureprotocolPostexposureobservation Tumor typeand incidence (%)aCommentsIwai et al.(1997)F/344 121, F108, F153, FClean airFiltered airWholeexhaust003.2-9.4NoneNoneNoneNA48-56hr/day48-56hr/dayNA6 mo6 mo5/121(4%) type not stated2/108(4%) type not stated53/153(35%) 61.3% adenoma, 25.8% adenocarcinoma,2.2% benign squamous cell tumor, 7.5% squamous cellcarcinoma, 3.2% adenosquamous carcinomaCumulativeexposure doseranged from154-274mg/m3Orthoefer etal. (1981)(Pepelkoand Peirano,1983)Mouse/Strong AM, 25 Clean airWholeexhaustWholeexhaust06.46.4NoneNoneUVirradiated20 hr/day,7 days/week,for 7weeks26 weeks26 weeks3/22 (13.6)7/19 (36.8)6/22 (27.3)0.13 tumors/mouse0.63 tumors/mouse0.27 tumors/mouse

7-89Table 7-3. Summary of animal inhalation carcinogenicity studies (continued)StudySpecies/strainSex/totalnumberExposureatmosphereParticleconcentration(mg/m3)OthertreatmentExposureprotocolPostexposureobservation Tumor typeand incidence (%)aCommentsLung tumorsMouse/Jackson AMouse/Jackson AM + F, 40M + F, 40F, 60F, 60F, 60F, 60M, 429M, 430Clean airWholeexhaustClean airClean airWholeexhaustWholeexhaustClean airWholeexhaust06.4006.46.406.4NoneNoneNoneUrethanlNoneUrethankNoneNone20 hr/day,7 days/week,for 8weeks20 hr/day,7 days/week,for approx.7 mo.8 weeks8 weeks16/36 (44.4)11/34 (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)0.5 tumors/mouse0.4 tumors/mouse0.09 tumors/mouse0.25 tumors/mouse0.32 tumors/mouse0.39 tumors/mouse0.23 tumors/mouse0.20 tumors/mouseKaplan et al.(1982)MouseA/JM, 458M, 18M, 485Clean airClean airWholeexhaust1.5 NoneUrethankNone20 hr/day,7 days/week,for 3 mo6 mo Pulmonary adenomas144/458 (31.4)18/18 (100)165/485 (34.2)Pulmonary adenomaKaplan et al.(1983)White et al.(1983)Mouse/A/JM, 388M, 388M, 399M, 396Clean airWholeexhaustWholeexhaustWholeexhaust00.250.751.5NoneNoneNoneNone20 hr/day,7 days/week,for up to8 moNA 130/388 (33.5)131/388 (33.8)109/399 (27.3)99/396 (25.0)

7-90Table 7-3. Summary of animal inhalation carcinogenicity studies (continued)StudySpecies/strainSex/totalnumberExposureatmosphereParticleconcentration(mg/m3)OthertreatmentExposureprotocolPostexposureobservation Tumor typeand incidence (%)aCommentsPepelko andPeirano(1983)Mouse/SencarM + F, 260 Clean airClean airClean airWholeexhaustWholeexhaustWholeexhaust121212 NoneBHTlUrethankNoneBHTlUrethanlContinuous for 15moNA Adenomas(5.1)(12.2)(8.1)(10.2)c(5.4)(8.7)Carcinomas(0.5)(1.7)(0.9)(1.0)(2.7)(2.6)All tumors(5.6)(2.8)(9.0)(11.2)c(8.1)(11.2)

7-91Table 7-3. Summary of animal inhalation carcinogenicity studies (continued)StudySpecies/strainSex/totalnumberExposureatmosphereParticleconcentration(mg/m3)OthertreatmentExposureprotocolPostexposureobservation Tumor typeand incidence (%)aComments

7-92All tumorsPepelko andPeirano(1983)Mouse/Strain AM + F, 90 Clean airClean airWholeexhaustWholeexhaustClean airWholeexhaust1212012 NoneExposure(darkness)Exposure(darkness)UrethanmUrethanmNA 21/87 (24)59/237 (24.9)10/80 (12.5)22/250 (0.10)66/75 (88)42/75 (0.95)0.29 tumors/mouse0.27 tumors/mouse0.140.102.800.95AdenomasAdenocarcinomaSquamouscelltumorsAll tumorsHeinrich etal. (1986a,b)Mouse/NMRIM + F, 84M + F, 93M + F, 76Clean airFilteredexhaustWholeexhaust4 NoneNoneNone19 hr/day,5 days/weekfor up to30 moNA 9/84 (11)11/93 (12)11/76 (15)2/84 (2)18/93 (19)c13/76 (17)c———11/84 (13)29/93 (31)c24/76 (32)cTakemoto etal.(1986)Mouse/IRCM + F, 45M + F, 69Clean airWholeexhaust02-4NoneNone4 hr/day,4 days/week, for19-28 moNAAdenomaAdenocarcinomaMouse/C57BLM + F, 12M + F, 38Clean airWholeexhaust02-4NoneNone4 hr/day,4 days/week for19-28 moNA 3/45 (6.7)6/69 (8.7)1/45 (2.2)3/69 (4.3)Table 7-3. Summary of animal inhalation carcinogenicity studies (continued)StudySpecies/strainSex/totalnumberExposureatmosphereParticleconcentration(mg/m3)OthertreatmentExposureprotocolPostexposureobservation Tumor typeand incidence (%)aComments

7-93Heinrich etal.(1995)Mouse/C57BL/6NF, 120F, 120F, 120Clean airWholeexhaustParticle-freeexhaust4.5 NoneNoneNone18 hr/day,5 days/week,for up to21 mo6 mo 1/12 (8.3)8/38 (21.1)0/12 (0)3/38 (7.9)5.1% tumorrate8.5% tumorrate3.5% tumorrateMouse/NMRIF, 120F, 120Clean airWholeexhaustCarbon blackTiO204.511.610NoneNoneNoneNone18 hr/day,5 days/week forup to13.5 mo9.5 moAdenomasAdenocarcinomas(25)(21.8)(11.3)(11.3)(15.4)(15.4)(10)(2.5)Mouse/NMRIF, 120F, 120F, 120Clean airWholeexhaustParticle-freeexhaust4.5 NoneNoneNone18 hr/day,5 days/week,23 moNone (25)(18.3)(31.7)(8.8)(5.0)(15)MultipleadenomasMultiplecarcinomasAdenomas/carcinomaAlveolar/bronchiolaradenomaAlveolar/bronchiolarcarcinomaMauderly etal. (1996)Mouse/CD-1M + F,157bM + F, 171M + F, 155M + F, 186Clean airWholeexhaustWholeexhaustWholeexhaust00.353.57.1NoneNoneNoneNone7 hr/day, 5days/week,for up to24 moNone 1/157 (0.6)2/171 (1.2)0/155 (0)0/186 (0)2/157 (1.3)1/171 (0.6)1/155 (0.6)0/186 (0)1/157 (0.6)1/171 (0.6)0/155 (0)0/186 (0)10/157(6.4)16/171(9.4)8/155 (5.2)10/186(5.4)7/157 (4.5)5/171 (2.9)6/155 (3.9)4/186 (2.2)AdenomasAdenocarcinomaSquamouscelltumorsAll tumorsHeinrich etal.(1986a,b)Hamster/SyrianM + F, 96M + F, 96M + F, 96Clean airFilteredexhaustWholeexhaust 4NoneNoneNone19 hr/day5days/weekfor up to30 mo NA0/96(0)0/96(0)0/96(0)0/96(0)0/96(0)0/96(0)0/960/960/960/96(0)0/96(0)0/96(0)Table 7-3. Summary of animal inhalation carcinogenicity studies (continued)StudySpecies/strainSex/totalnumberExposureatmosphereParticleconcentration(mg/m3)OthertreatmentExposureprotocolPostexposureobservation Tumor typeand incidence (%)aCommentsPrimary lungtumorsBrightwellet al.(1989)Hamster/SyrianGoldenM + F,M + F, 202M + F, 104M + F, 104M + F, 101M + F, 102M + F, 101M + F, 204M + F, 203Clean airClean airFilteredexhaust(mediumdose)Filteredexhaust(high dose)WholeexhaustWholeexhaustWholeexhaustFilteredexhaust(high dose)Wholeexhaust00000.72.26.606.6NoneDENjDENjDENjDENjDENjDENjNoneNone16 hr/day,5 days/week, for24 moNA 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)Respiratorytract tumorsnot related toexhaustexposure forany of thegroupsaTable values indicate number with tumors/number examined (% animals with tumors).bNumber of animals examined for tumors.cSignificantly different from clean air controls.dDipentylnitrosamine; 6.25 mg/kg/week s.c. during first 25 weeks of exposure.eDipentylnitrasamine; 12.5 mg/kg/week s.c. during first 25 weeks of exposure.fSplenic lymphomas also detected in controls (8.3%), filtered exhaust group (37.5%) and wholeexhaust group (25%).g5.3% incidence of large cell carcinomas.h1 g/kg, i.p. 1/week for 3 weeks starting 1 mo into exposure.iIncludes adenomas, squamous cell carcinomas, adenocarcinomas, adenosquamous cell carcinoma,and mesotheliomas.j4.5 mg/diethylnitrosamine (DEN)/kg, s.c., 3 days prior to start of inhalation exposure.kSingle i.p. dose 1 mg/kg at start of exposure.lButylated hydroxytoluene 300 mg/kg, i.p. for week 1, 83 mg/kg for week 2, and 150 mg/kg forweeks 3 to 52.m12 mg/m3from 12 weeks of age to termination of exposure. Prior exposure (in utero) and of parentswas 6 mg/m3.n120-121 males and 71-72 females examined histologically.oNot all animals were exposed for full term, at least 10 males were killed at 3, 6, and 12 mo ofexposure.NS = Not specified.NA = Not applicable.

7-947-95 Table 7-4. Tumor incidences in rats following intratracheal instillation of DE particles (DPM), extracted DPM, carbon black (CB), benzo[ a ]pyrene (B[ a ]P), or particles plus B[ a ]P

Experimental group Number of animals Total dose Animals with tumors (percent) Statistical significance a

Control 47 4.5 mL 0 (0) -
DPM (original) 48 15 mg 8 (17) < 0.01
DPM (extracted) 48 30 mg 10 (21) < 0.001
DPM (extracted) 48 15 mg 2 (4) NS

CB (printex) 48 15 mg 10 (21) < 0.001
CB (lampblack) 48 14 mg 4 (8) NS

B[ a ]P 47 30 mg 43 (90) < 0.001

B[ a ]P 48 15 mg 12 (25) < 0.001

DEP + B[ a ]P 48 15 mg + 170 : g B[ a ]P 4 (8) NS

CB (printex) + B[ a ]P 48 15 mg + 443 : g B[ a ]P 13 (27) < 0.001

Table 7-5. Tumorigenic effects of dermal application of acetone extracts of DPM

Number of animals Strain/sex Sample material Time to first tumor (mo) Survivors at time of first
tumor Total
tumors Duration of experiment (mo)

52 C57BL/40 F C57BL/12 M Extract of DPM obtained during
warmup 13 33 2 22

50 Strain A/M Extract of DPM obtained during full load 15 8 4 23

25 Strain A/F Extract of DPM obtained during full load 13 20 17 17

Source: Kotin et al., 1955.7-96 General Motors Research Laboratories sponsored chronic inhalation studies at the Southwest Research Institute using male Fischer 344 rats, 30 per group, exposed to DPM concentrations of 0.25, 0.75, or 1.5 mg/m 3 (Kaplan et al., 1983; White et al., 1983). The animals were exposed in 12.6 m 3 exposure chambers. Airflow was adjusted to provide 13 changes per hour. Temperature was maintained at 22 ± 2 °C. The exposure protocol was 20 hr/day, 7 days/week for 9 to 15 months. Exposures were halted during normal working hours for servicing. Some animals were sacrificed following completion of exposure, while others were returned to clean air atmospheres for an additional 8 months. Control animals received clean air. Exhaust was generated by 5.7-L Oldsmobile engines (four different engines used throughout the experiment) operated at a steady speed and load simulating a 40-mph driving speed of a full-size

passenger car. Although five instances of bronchoalveolar carcinoma were observed in 90 rats exposed to DE for 15 months and held an additional 8 months in clean air, compared with none among controls, statistical significance was not achieved in any of the exposure groups. These included one tumor in the 0.25 mg/m 3 group, three in the 0.75 mg/m 3 group, and one in the 1.5 mg/m 3 group. Rats kept in clean-air chambers for 23 months did not exhibit any carcinomas. No tumors were observed in any of the 180 rats exposed to DE for 9 or 15 months without a recovery period, or in the respective controls for these groups. Equivocal results may again have been due to less-than-lifetime duration of the study as well as insufficient exposure concentrations. Although the increases in tumor incidences in the groups exposed for 15 months and held an additional 8 months in clean air were not statistically significant, relative to controls, they were slightly greater than the historic background incidence of 3.7% for this specific lesion in this strain of rat (Ward, 1983). The first definitive studies linking inhaled DE to induction of lung cancer in rats were reported by researchers in Germany, Switzerland, Japan, and the UnitedStates in the mid-to-late 1980s. In a study conducted at the Fraunhofer Institute exhaustgenerating system and exposure atmosphere characteristics are presented in Appendix A. The type of engine used (3-cylinder, 43 bhp diesel) is normally used in mining situations and was of Toxicology and Aerosol Research, female Wistar rats were exposed for 19 hr/day, 5 days/week to both filtered and unfiltered (total) DE at an average particulate matter concentration of 4.24 mg/m 3 . Animals were exposed for a maximum of 2.5 years. The exposure system as described by Heinrich et al. (1986a) used a 40 kilowatt 1.6-L diesel engine operated continuously under the U.S. 72 FTP driving cycle. The engines used European Reference Fuel with a sulfurcontent of 0.36%. Filtered exhaust was obtained by passing engine exhaust through a Luwa FP65 HT 610 particle filter heated to 80 °C and a secondary series of filters (Luwa FP-85, Luwa NS-30, and Drager CH 63302) at room temperature. The filtered and unfiltered exhausts were diluted 1:17 with filtered air and passed through respective 12 m 3 exposure chambers. Mass7-97 median aerodynamic diameter of DPM was 0.35 ± 0.10 

### : m (mean ± SD). The gas-phase components of the DE atmospheres are presented in Appendix A. The effects of exposure to either filtered or unfiltered exhaust were described by Heinrich et al. (1986b) and Stöber (1986). Exposure to unfiltered exhaust resulted in 8 bronchoalveolar adenomas and 9 squamous cell tumors in 15 of 95 female Wistar rats examined, for a 15.8% tumor incidence. Although statistical analysis was not provided, the increase appears to be highly significant. In addition to the bronchioalveolar adenomas and squamous cell tumors, there was a high incidence of bronchioalveolar hyperplasia (99%) and metaplasia of the bronchioalveolar epithelium (65%). No tumors were reported among rats exposed to filtered exhaust (n = 92) or clean air (n = 96). Mohr et al. (1986) provided a more detailed description of the lung lesions and tumors identified by Heinrich et al. (1986a,b) and Stöber (1986). Substantial alveolar deposition of carbonaceous particles was noted for rats exposed to unfiltered DE. Squamous metaplasia was observed in 65.3% of the rats breathing unfiltered DE, but not in the control rats. Of nine squamous cell tumors, one was characterized as a Grade I carcinoma (borderline atypia, few to moderate mitoses, and slight evidence of stromal invasion), and the remaining eight were classified as benign keratinizing cystic tumors. Iwai et al. (1986b) examined the long-term effects of DE inhalation on female F344 rats. The exhaust was generated by a 2.4-L displacement truck engine. The exhaust was diluted 10:1 with clean air at 20 °C to 25 °C and 50% relative humidity. The engines were operated at 1,000 rpm with an 80% engine load. These operating conditions were found to produce exhaust with the highest particle concentration and lowest NO 2 and SO 2 content. For those chambers using filtered exhaust, proximally installed high-efficiency particulate air (HEPA) filters were used. Three groups of 24 rats each were exposed to unfiltered DE, filtered DE, or filtered room air for 8 hr/day, 7 days/week for 24 months. Particle concentration was 4.9 mg/m 3 for unfiltered exhaust. Concentrations of gas-phase exhaust components were 30.9 ppm NO
x , 1.8 ppm NO 2 , 13.1 ppm SO 2 , 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 4.5%. No lung tumors were observed in rats exposed to filtered DE. Nineteen of the 24 exposed to unfiltered exhaust survived for 2 years. Of these, 14 were randomly selected for sacrifice at this time. Four of the rats developed lung tumors; two of these were malignant. Five rats of this 2-year exposure group were subsequently placed in clean room air for 3 to 6 months and four eventually (time not specified) exhibited lung tumors (three malignancies). Thus, the lung tumor incidence for total tumors was 42.1% (8/19) and 26.3% (5/19) for malignant tumors in rats exposed to whole DE. The tumor types identified were adenoma (3/19), adenocarcinoma7-98 (1/19), adenosquamous carcinoma (2/19), squamous carcinoma (1/19), and large-cell carcinoma (1/19). The lung tumor incidence in rats exposed to whole DE was significantly greater than that of controls ( p 

### # 0.01). Tumor data are summarized in Table 7-3. Malignant splenic lymphomas were detected in 37.5% of the rats in the filtered exhaust group and in 25.0% of the rats in the unfiltered exhaust group; these values were significantly ( p 

### # 0.05) greater than the 8.2% incidence noted in the control rats. The study demonstrates production of lung cancer in rats following 2-year exposure to unfiltered DE. In addition, splenic malignant lymphomas occurred during exposure to both filtered and unfiltered DE. This is the only report to date of tumor induction at an extrarespiratory site by inhaled DE in animals. A chronic (up to 24 months) inhalation exposure study was conducted by Takemoto et al.(1986), in which female Fischer 344 rats were exposed to DE generated by a 269-cc YANMAR40CE NSA engine operated at an idle state (1,600 rpm). Exposures were 4 hours/day, 4 days/week. The animals were exposed in a 376-L exposure chamber. Air flow was maintained at 120 L/min. Exhaust was diluted to produce a particle concentration of 2-4 mg/m 3 . When not exposed the animals were maintained in an air-conditioned room at a temperature of 24 ± 2°C and a relative humidity of 55 ± 5% with 12 hr of light and darkness. Temperature and humidity in the exposure chambers was not noted. The particle concentration of the DE in the exposure chamber was 2 to 4 mg/m 3 . B[a]P and 1-nitropyrene concentrations were 0.85 and 93 

### : g/g of particles, respectively. No lung tumors were reported in the diesel-exposed animals. It was also noted that the diesel engine employed in this study was originally used as an electrical generatorand that its operating characteristics (not specified) were different from those of a dieselpowered automobile. However, the investigators deemed it suitable for assessing the effects of
diesel emissions. Mauderly et al. (1987) provided data affirming the carcinogenicity of automotive diesel engine exhaust in F344/Crl rats following chronic inhalation exposure. Male and female rats were exposed to diesel engine exhaust at nominal DPM concentrations of 0.35 (n = 366), 3.5 (n = 367), or 7.1 (n = 364) mg/m 3 for 7 hr/day, 5 days/week for up to 30 mo. Sham-exposed (n = 365) controls breathed filtered room air. A total of 230, 223, 221, and 227 of these rats (sham-exposed, low-, medium-, and high-exposure groups, respectively) were examined for lung tumors. These numbers include those animals that died or were euthanized during exposure and those that were terminated following 30 months of exposure. The exhaust was generated by 1980 model 5.7-L Oldsmobile V-8 engines operated through continuously repeating U.S. Federal Test Procedure (FTP) urban certification cycles. The engines were equipped with automatic transmissions connected to eddy-current dynamometers and flywheels simulating resistive and inertial loads of a midsize passenger car. The D-2 diesel control fuel (Phillips Chemical Co.) met U.S. EPA certification standards and contained approximately 30% aromatic7-99 hydrocarbons and 0.3% sulfur. Following passage through a standard automotive muffler and tailpipe, the exhaust was diluted 10:1 with filtered air in a dilution tunnel and serially diluted to the final concentrations. The primary dilution process was such that particle coagulation was retarded. Mokler et al. (1984) provided a detailed description of the exposure system. No exposure-related changes in body weight or lifespan were noted for any of the exposed animals, nor were there any signs of overt toxicity. Collective lung tumor incidence was greater (z statistic, p 

### # 0.05) in the high (7.1 mg/m 3 ) and medium (3.5 mg/m 3 ) exposure groups (12.8% and 3.6%, respectively) versus the control and low (0.35 mg/m 3 ) exposure groups (0.9% and 1.3%, respectively). In the high-dose group the incidences of tumor types reported were adenoma (0.4%), adenocarcinomas plus squamous cell carcinomas (7.5%), and squamous cysts (4.9%). In the medium-dose group adenomas were reported in 2.3% of animals, adenocarcinomas plus squamous cell carcinomas in 0.5%, and squamous cysts in 0.9%. In the low-exposure group adenocarcinomas plus squamous cell carcinomas were detected in 1.3% of the rats. Using the same statistical analysis of specific tumor types, adenocarcinoma plus squamous cell carcinoma and squamous cyst incidence was significantly greater in the high-exposure group, and the incidence of adenomas was significantly greater in the medium-exposure group. A significant ( p <0.001) exposure-response relationship was obtained for tumor incidence relative to exposure concentration and lung burden of DPM. These data are summarized in Table 7-3. A logistic regression model estimating tumor prevalence as a function of time, dose (lung burden of DPM), and sex indicated a sharp increase in tumor prevalence for the high dose level at about 800 days after the commencement of exposure. A less pronounced, but definite, increase in prevalence with time was predicted for the medium-dose level. Significant effects were not detected at the low concentration. DPM (mg per lung) of rats exposed to 0.35, 3.5, or 7.1 mg of DPM/m 3 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
conditions. In summary, this study demonstrated the pulmonary carcinogenicity of high concentrations of whole, diluted DE in rats following chronic inhalation exposure. In addition, increasing lung particle burden resulting from this high-level exposure and decreased clearance was demonstrated. A logistic regression model presented by Mauderly et al. (1987) indicatedthat both lung DPM burden and exposure concentration may be useful for expressing exposureeffect relationships. A long-term inhalation study (Ishinishi et al., 1988a; Takaki et al., 1989) examined the effects of emissions from a light-duty (LD) and a heavy-duty (HD) diesel engine on male and female Fischer 344/Jcl rats. The LD engines were 1.8-L, 4-cylinder, swirl-chamber-type power plants, and the HD engines were 11-L, 6-cylinder, direct-injection-type power plants. The7-100 engines were connected to eddy-current dynamometers and operated at 1,200 rpm (LD engines) and 1,700 rpm (HD engines). Nippon Oil Co. JIS No. 1 or No. 2 diesel fuel was used. The 30- months whole-body exposure protocol (16 h/day, 6 days/week) used DPM concentrations of 0, 0.5, 1, 1.8, or 3.7 mg/m 3 from HD engines and 0, 0.1, 0.4, 1.1, or 2.3 mg/m 3 from LD engines. The animals inhaled the exhaust emissions from 1700 to 0900 h. Sixty-four male rats and 59 to 61 female rats from each exposure group were evaluated for carcinogenicity. For the experiments using the LD series engines, the highest incidence of hyperplastic lesions plus tumors (72.6%) was seen in the highest exposure (2.3 mg/m 3 ) group. However, this high value was the result of the 70% incidence of hyperplastic lesions; the incidence of adenomas was only 0.8% and that of carcinomas 1.6%. Hyperplastic lesion incidence was considerably lower for the lower exposure groups (9.7%, 4.8%, 3.3%, and 3.3% for the 1.1, 0.4, and 0.1 mg/m 3 and control groups, respectively). The incidence of adenomas and carcinomas, combining males and females, was not significantly different among exposure groups (2.4%, 4.0%, 0.8%, 2.4%, and 3.3% for the 2.3, 1.1, 0.4, and 0.1 mg/m 3 groups and the controls, respectively). For the experiments using the HD series engines, the total incidence of hyperplastic lesions, adenomas, and carcinomas was highest (26.6%) in the 3.7 mg/m 3 exposure group. The incidence of adenomas plus carcinomas for males and females combined equaled 6.5%, 3.3%, 0%, 0.8%, and 0.8% at 3.7, 1.8, 1, and 0.4 mg/m 3 and for controls, respectively. A statistically significant difference was reported between the 3.7 mg/m 3 and the control groups for the HD series engines. The carcinomas were identified as adenomas, adenosquamous carcinomas, and squamous cell carcinomas. Although the number of each was not reported, it was noted that the majority were squamous cell carcinomas. A progressive dose-response relationship was not demonstrated. Tumor incidence data for this experiment are presented in Table 7-3. The Ishinishi et al. (1988a) study also included recovery tests in which rats exposed to whole DE (DPM concentration of 0.1 or 1.1 mg/m 3 for the LD engine and 0.5 or 1.8 mg/m 3 for the HD engine) for 12 months were examined for lung tumors following 6-, 12-, or 18-month recovery periods in clean air. The incidences of neoplastic lesions were low, and pulmonary DPM burden was lower than for animals continuously exposed to whole DE and not provided a recovery period. The only carcinoma observed was in a rat examined 12 months following exposure to exhaust (1.8 mg/m 3 ) from the HD engine. Brightwell et al. (1986, 1989) studied the effects of DE on male and female F344 rats. The DE was generated by a 1.5-L Volkswagen engine that was computer-operated according to the U.S. 72 FTP driving cycle. The engine was replaced after 15 mo. The engine emissions were diluted by conditioned air delivered at 800 m 3 /h to produce the high-exposure (6.6 mg/m 3 ) DE atmosphere. Further dilutions of 1:3 and 1:9 produced the medium- (2.2 mg/m 3 ) and low-7-101 (0.7 mg/m 3 ) exposure atmospheres. The CO and NO
x concentrations (mean ± SD) were 32 ± 11 ppm and 8 ± 1 ppm in the high-exposure concentration 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. For males and females combined, a 1.2% (3/260), 0.7% (1/144), 9.7% (14/144), and 38.5% (55/143) incidence of primary lung tumors occurred in F344 rats following exposure to clean air or 0.7, 2.2, and 6.6 mg of DPM/m 3 , respectively (Table 7-3). DE-induced tumor incidence in rats was dose-related and higher in females than in males (Table 7-3). These data included animals sacrificed at the interim periods (6, 12, 18, and 24 mo); therefore, the tumor incidence does not accurately reflect the effects of long-term exposure to the DE atmospheres. When tumor incidence is expressed relative to the specific intervals, a lung tumor incidence of96% (24/25), 76% (19/25) of which were malignant, was reported for female rats in the highdose group exposed for 24 months and held in clean air for the remainder of their lives. For male rats in the same group, the tumor incidence equaled 44% (12/27), of which 37% (10/27) were malignant. It was also noted that many of the animals exhibiting tumors had more than one tumor, often representing multiple histological types. The numbers and types of tumors identified in the rats exposed to DE included adenomas (40), squamous cell carcinomas (35), adenocarcinomas (19), mixed adenoma/adenocarcinomas (9), and 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. Lewis et al. (1989) also examined the effects of inhalation exposure of DE and/or coal dust on tumorigenesis on F344 rats. Groups of 216 male and 72 female rats were exposed to clean air, whole DE (2 mg soot/m 3 ), coal dust (2 mg/m 3 respirable concentration; 5 to 6 mg/m 3 total concentration), or DE plus coal dust (1 mg/m 3 of each respirable concentration; 3.2 mg/m 3 total concentration) for 7 h/day, 5 days/week during daylight hours for up to 24 mo. Groups of 10 or more males were sacrificed at intermediate intervals (3, 6, and 12 mo). The DE was produced by a 7.0-L, 4-cycle, water-cooled Caterpillar Model 3304 engine using No. 2 diesel fuel (<0.5% sulfur by mass). The exhaust was passed through a Wagner water scrubber, which lowered the exhaust temperature and quenched engine backfire. The animals were exposed in 100-cubic-foot chambers. Temperature was controlled at 22 ± 2 °C and relative humidity at 50%±10%. The exhaust was diluted 27-fold with chemically and biologically filtered clean air to achieve the desired particle concentration. Histological examination was performed on 120 to 121 male and 71 to 72 female rats terminated after 24 months of exposure. The exhaust exposure did not significantly affect the tumor incidence beyond what would be expected for aging F344 rats. There was no7-102 postexposure period, which may explain, in part, the lack of significant tumor induction. The particulate matter concentration was also less than the effective dose in several other studies. In a more recent study reported by Heinrich et al. (1995), female Wistar rats were exposed to whole DE (0.8, 2.5, or 7.0 mg/m 3 ) 18 h/day, 5 days/week for up to 24 mo, then held in clean air an additional 6 mo. The animals were exposed in either 6 or 12 m 3 exposure chambers. Temperature and relative humidity were maintained at 23-25 °C and 50%-70%, respectively. DE was generated by two 40-kw 1.6-L diesel engines (Volkswagen). One of them was operated according to the U.S. 72 cycle. The other was operated under constant load conditions. The first engine did not supply sufficient exhaust, which was filled by the second engine. Cumulative exposures for the rats in the various treatment groups were 61.7, 21.8, and 7.4 g/m 3 × h for the high, medium, and low whole-exhaust exposures. Significant increases in tumor incidences were observed in the high (22/100; p <0.001) and mid (11/200; p <0.01) exposure groups relative to clean-air controls (Table 7-3). Only one tumor (1/217), an adenocarcinoma, was observed in clean-air controls. Relative to clean-air controls, significantly increased incidences were observed in the high-exposure rats for benign squamous cell tumors (14/100; p <0.001), adenomas (4/100; p <0.01), and adenocarcinomas (5/100; p <0.05). Only the incidence of benign squamous cell tumors (7/200; p <0.01) was significantly increased in the mid-exposure group relative to the clean-air controls. Particle lung burden and alveolar clearance also were determined in the Heinrich et al. (1995) study. Relative to clean air controls, alveolar clearance was significantly compromised by exposure to mid and high DE. For the high-diesel-exhaust group, 3-mo recovery time in clean air failed to reverse the compromised alveolar clearance. In a study conducted at the Inhalation Toxicology Research Institute (Nikula et al., 1995) F344 rats (114-115 per sex per group) were exposed 16 hr/day, 5 days/week during daylight hours to DE diluted to achieve particle concentrations of 2.5 or 6.5 mg/m 3 for up to 24 mo. Controls (118 males, 114 females) were exposed to clean air. Surviving rats were maintained an additional 6 weeks in clean air, at which time mortality reached 90%. DE was generated with two 1988 Model LH6 General Motors 6.2-L V-8 engines burning D-2 fuel that met EPA certification standards. Chamber air flow was sufficient to provide about 15 exchanges per hour. Relative humidity was 40% to 70% and temperature ranged from 23 to 25
°C. Following low and high DE exposure, the lung burdens were 36.7 and 80.7 mg, respectively, for females and 45.1 and 90.1 mg, respectively, for males. The percentages of susceptible rats (males and females combined) with malignant neoplasms were 0.9 (control), 3.3 (low DE), and 12.3 (high DE). The percentages of rats (males and females combined) with malignant or benign neoplasms were 1.4 (control), 6.2 (low DE), and 17.9 (high DE). All7-103 primary neoplasms were associated with the parenchyma rather than the conducting airways of the lungs. The first lung neoplasm was observed at 15 mo. Among 212 males and females examined in the high-dose group, adenomas were detected in 23 animals, adenocarcinomas in 22 animals, squamous cell carcinomas in 3 animals, and an adenosquamous carcinoma in 1 animal. For further details see Table 7-3. Analysis of the histopathologic data suggested a progressive process from alveolar epithelial hyperplasia to adenomas and adenocarcinomas. Iwai et al. (1997) carried out a series of exposures to both filtered and whole exhaust using a light-duty (2,369 mL) diesel engine. The protocol for engine operation was not stated. Groups of female SPF F344 Fischer rats were exposed for 2 years for 8 hr/day, 7 days/week, 8 hr/day, 6 days/week, or 18 hr/day, 3 days/week to either filtered exhaust or exhaust diluted to a particle concentration of 9.4, 3.2, and 5.1 mg/m 3 , respectively. Cumulative exposure (mg/m 3 × hrs of exposure) equaled 274.4, 153.6, and 258.1 mg/m 3 . The animals were then held for an additional 6 months in clean air. Lung tumors were reported in 5/121 (4%) of controls, 4/108 (4%) of those exposed to filtered exhaust, and 50/153 (35%) among those exposed to whole exhaust. Among rats exposed to whole DE 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 tumors probably corresponded to squamous cysts in another classification.

7.3.1.2. Mouse Studies A series of inhalation studies using strain A mice was conducted by Orthoefer et al. (1981). Strain A mice are usually given a series of intraperitoneal injections with the test agent; they are then sacrificed at about 9 months and examined for lung tumors. In the present series, inhalation exposure was substituted. DE was provided by one of two Nissan CN6-33 diesel engines having a displacement of 3244 cc and run on a Federal Short Cycle. Flow through the exposure chambers was sufficient to provide 15 air changes per hour. Temperature was maintained at 24 °C and relative humidity at 75%. In the first study, groups of 25 male Strong A strain (A/S) mice were exposed to irradiated DE (to simulate chemical reactions induced by sunlight) or nonirradiated DE (6 mg/m 3 ) for 20 h/day, 7 days/week. Additional groups of 40 Jackson A strain (S/J) mice (20 of each sex) were exposed similarly to either clean air or DE, then held in clean air until sacrificed at 9 months of age. No 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 months of age and approximately 100 at 12 months of age). With the exception of those treated with urethan, the number of tumors per mouse did not exceed historical control levels in any of the studies. Exposure to DE,7-104 however, significantly inhibited the tumorigenic effects of the 5-mg urethan treatment. Results
are listed in Table 7-3. Kaplan et al. (1982) also reported the effects of diesel exposure in strain A mice. Groups of male strain A/J mice were exposed for 20 h/day, 7 days/week for 90 days and held until 9 months of age. Briefly, the animals were exposed in inhalation chambers to DE generated by a 5.7-L Oldsmobile engine operated continuously at 40 mph at DPM concentrations of 0, 0.25, 0.75, or 1.5 mg/m 3 . Controls were exposed to clean air. Temperature was maintained at 22 ± 2 °C and relative humidity at 50% ± 10% within the chambers. Among 458 controls and 485 exposed animals, tumors were detected in 31.4% of those breathing clean air versus 34.2% of those exposed to DE. The mean number of tumors per mouse also failed to show significant
differences. In a follow-up study, strain A mice were exposed to DE for 8 months (Kaplan et al., 1983; White et al., 1983). After exposure to the highest exhaust concentration (1.5 mg/m 3 ), the percentage of mice with pulmonary adenomas and the mean number of tumors per mouse were significantly less ( p <0.05) than those for controls (25.0% vs. 33.5% and 0.30 ± 0.02 [S.E.] vs. 0.42 ± 0.03 [S.E.]) (Table 7-3). Pepelko and Peirano (1983) summarized a series of studies on the health effects of diesel emissions in mice. Exhaust was provided by two Nissan CN 6-33, 6-cylinder, 3.24-L diesel engines coupled to a Chrysler A-272 automatic transmission and Eaton model 758-DG dynamometer. Sixty-day pilot studies were conducted at a 1:14 dilution, providing DPM concentrations of 6 mg/m 3. The engines were operated using the Modified California Cycle. These 20-hr/day, 7-days/week pilot studies using rats, cats, guinea pigs, and mice produced decreases in weight gain and food consumption. Therefore, at the beginning of the long-term studies, exposure time was reduced to 8 h/day, 7 days/week at an exhaust DPM concentration of 6 mg/m 3 . During the final 12 months of exposure, however, the DPM concentration was increased to 12 mg/m 3 . For the chronic studies, the engines were operated using the Federal Short Cycle. Chamber temperature was maintained at 24 °C and relative humidity at 50%. Airflow was sufficient for 15 changes per hour. Pepelko and Peirano (1983) described a two-generation study using Sencar mice exposed to DE. Male and female parent-generation mice were exposed to DE at a DPM concentration of 6 mg/m 3 prior to (from weaning to sexual maturity) and throughout mating. The dams continued exposure through gestation, birth, and weaning. Groups of offspring (130 males and 130 females) were exposed to either DE or clean air. The exhaust exposure was increased to a DPM concentration of 12 mg/m 3 when the offspring were 12 weeks of age and was maintained until termination of the experiment when the mice were 15 months old.7-105 The incidence of pulmonary adenomas (16.3%) was significantly increased in the mice exposed to DE compared with 6.3% in clean-air controls. The incidence in males and females combined was 10.2% in 205 animals examined compared with 5.1% in 205 clean-air controls. This difference was also significant. The incidence of carcinomas was not affected by exhaust exposure in either sex. These results provided the earliest evidence for cancer induction following inhalation exposure to DE. The increase in the sensitivity of the study, allowing detection of tumors at 15 mo, may have been the result of exposure from conception. It is likely that Sencar mice are sensitive to induction of lung tumors because they are also sensitive to
induction of skin tumors. These data are summarized in Table 7-3. Takemoto et al. (1986) reported the effects of inhaled DE (2 to 4 mg/m 3 , 4 h/day, 4 days/week, for up to 28 mo) in ICR and C57BL mice exposed from birth. Details of the exposure conditions are presented in Section 7.3.2.1. All numbers reported are for males and females combined. Four adenomas and 1 adenocarcinoma were detected in 34 DE-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 2 adenocarcinomas were detected in 79 C57BL mice autopsied at 13 to 18 mo, compared with none in 19 unexposed animals. Among males and females autopsied at 19 to 28 mo, 8 adenomas and 3 adenocarcinomas were detected in 71 exposed animals, compared with 1 adenoma among 32 controls. No significant increases in adenoma or adenocarcinoma were reported for either strain of exposed mice. However, the significance of the increase in the combined incidence of adenomas and carcinomas was not evaluated statistically. A statistical analysis by Pott and Heinrich (1990a) indicated that the difference in combined benign and malignant tumors between whole DE-exposed C57BL/6N mice and corresponding controls was significant at

p <.05. See Table 7-3 for details of tumor incidence. Heinrich et al. (1986b) and Stöber (1986), as part of a larger study, also evaluated the effects of DE in mice. Details of the exposure conditions reported by Heinrich et al. (1986a) are given in Section 7.3.1.1 and Appendix A. Following lifetime (19 h/day, 5 days/week, for a maximum of 120 weeks) exposure to DE diluted to achieve a particle concentration of 4.2 mg/m 3 , 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 adenomas showed little change, adenocarcinomas increased significantly from 2.4% for controls to 17% for exhaust-exposed mice. In a follow-up study, however, Heinrich et al. (1995) reported a lack of tumorigenic response in either female NMRI or C57BL/6N mice exposed 17 h/day, 57-106 days/week for 13.5 to 23 months to whole DE diluted to produce a particle concentration of 4.5 mg/m 3 . These data are summarized in Table 7-3. The lack of a carcinogenic response in mice was reported by Mauderly et al. (1996). In this study, groups of 540 to 600 CD-1 male and female mice were exposed to whole DE (7.1, 3.5, or 0.35 mg DPM/m 3 ) for 7 hr/day, 5 days/week for up to 24 mo. Controls were exposed to filtered air. DE was provided by 5.7-L Oldsmobile V-8 engines operated continuously on theU.S. Federal Test Procedure urban certification cycle. The chambers were maintained at 25 °C28 °C, relative humidity at 40%-60%, and a flow rate sufficient for 15 air exchanges per hour. Animals were exposed during the light cycle, which ran from 6:00 AM to 6:00 PM. DPM accumulation in the lungs of exposed mice was assessed at 6, 12, and 18 months of exposure and was shown to be progressive; DPM burdens were 0.2 ± 0.02, 3.7 ± 0.16, and 5.6 ± 0.39 mg for the low-, medium-, and high-exposure groups, respectively. The lung burdens in both the medium- and high-exposure groups exceeded that predicted by exposure concentration ratio for the low-exposure group. Contrary to what was observed in rats (Heinrich et al., 1986b; Stöber, 1986; Nikula et al., 1995; Mauderly et al., 1987), an exposure-related increase in primary lung neoplasms was not observed in the CD-1 mice, supporting the contention of a species difference in the pulmonary carcinogenic response to poorly soluble particles. The percentage incidence of mice (males and females combined) with one or more malignant or benign neoplasms was 13.4, 14.6, 9.7, and 7.5 for controls and low-, medium-, and high-exposure groups, respectively.Although earlier studies provided some evidence for tumorigenic responses in dieselexposed mice, no increases were reported in the two most recent studies by Mauderly et al. (1996) and Heinrich et al. (1995), which utilized large group sizes and were well designed and conducted. Overall, the results in mice must therefore be considered to be equivocal.

7.3.1.3. Hamster Studies Heinrich et al. (1982) examined the effects of DE exposure on tumor frequency in female Syrian golden hamsters. Groups of 48 to 72 animals were exposed to clean air or whole DE at a mean DPM concentration of 3.9 mg/m 3 . Inhalation exposures were conducted 7 to 8 hr/day, 5 days/week for 2 years. The exhaust was produced by a 2.4-L Daimler-Benz engine operated under a constant load and a constant speed of 2,400 rpm. Flow rate was sufficient for about 20 exchanges per hour in the 250-L chambers. No lung tumors were reported in either exposure

group. In a subsequent study, Syrian hamsters were exposed 19 hr/day, 5 days/week for a lifetime to DE diluted to a DPM concentration of 4.24 mg/m 3 (Heinrich et al., 1986b; Stöber, 1986). Details of the exposure conditions are reported in Appendix A. Ninety-six animals per7-107 group were exposed to clean air or exhaust. No lung tumors were seen in either the clean-air group or in the DE-exposed group. In a third study (Heinrich et al., 1989b), hamsters were exposed to exhaust from a Daimler-Benz 2.4-L engine operated at a constant load of about 15 kW and at a uniform speed of 2,000 rpm. The exhaust was diluted to an exhaust-clean air ratio of about 1:13, resulting in a mean particle concentration of 3.75 mg/m 3 . Exposures were conducted in chambers maintained at 22 to 24 °C and 40% to 60% relative humidity for up to 18 mo. Surviving hamsters were maintained in clean air for up to an additional 6 mo. The animals were exposed 19 hr/day, 5 days/week beginning at noon each day, under a 12-hr light cycle starting at 7 AM. Forty animals per group were exposed to whole DE or clean air. No lung tumors were detected in either the clean-air or diesel-exposed hamsters. Brightwell et al. (1986, 1989) studied the effects of DE on male and female Syrian golden hamsters. Groups of 52 males and 52 females, 6 to 8 weeks old, were exposed to DE at DPM concentrations of 0.7, 2.2, or 6.6 mg/m 3 . They were exposed 16 hr/day, 5 days/week for a total of 2 years and then sacrificed. Exposure conditions are described in Section 7.3.1.1. No statistically significant ( t test) relationship between tumor incidence and exhaust exposure was reported. In summary, DE alone did not induce an increase in lung tumors in hamsters of either sex in several studies of chronic duration at high exposure concentrations.

7.3.1.4. Monkey Studies Fifteen male cynomolgus monkeys were exposed to DE (2 mg/m 3 ) for 7 hr/day, 5 days/week for 24 months (Lewis et al., 1989). The same numbers of animals were also exposed to coal dust (2 mg/m 3 respirable concentration; 5 to 6 mg/m 3 total concentration), DE plus coal dust (1 mg/m 3 respirable concentration for each component; 3.2 mg/m 3 total concentration), or filtered air. Details of exposure conditions were listed previously in the description of the Lewis et al. (1989) study with rats (Section 7.3.1.1) and are listed in Appendix A. None of the monkeys exposed to DE exhibited a significantly increased incidence of preneoplastic or neoplastic lesions. It should be noted, however, that the 24-mo time frame employed in this study may not have allowed the manifestation of tumors in primates, because this duration is only a small fraction of the monkeys’ expected lifespan. In fact, there have been no near-lifetime exposure studies in nonrodent species.

7.3.2. Inhalation Studies (filtered DE) Several studies have been conducted in which animals were exposed to DE filtered to remove PM. As these studies also included groups exposed to whole exhaust, details 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 al.7-108 (1986b) and Stöber (1986) reported negative results for lung tumor induction in female Wistar rats exposed to filtered exhaust diluted to produce an unfiltered particle concentration of 4.24 mg/m 3 . Negative results were also reported in female Fischer 344 rats exposed to filtered exhaust diluted to produce an unfiltered particle concentration of 4.9 mg/m 3 (Iwai et al., 1986a), in Fischer 344 rats of either sex exposed to filtered exhaust diluted to produce an unfiltered particle concentration of 6.6 mg/m 3 (Brightwell et al., 1989), in female Wistar rats exposed to filtered exhaust diluted to produce an unfiltered particle concentration of 7.0 mg/m 3 (Heinrich et al., 1995), and in female Fischer 344 rats exposed to filtered exhaust diluted to produce unfiltered particle concentrations of 5.1, 3.2, or 9.4 mg/m 3 (Iwai et al., 1997). In the Iwai et al. (1986a) study, splenic lymphomas were detected in 37.5% of the exposed rats compared with
8.2% in controls. In the study reported by Heinrich at al. (1986a) and Stober (1986), primary lung tumors were seen in 29/93 NMRI mice (males and females combined) exposed to filtered exhaust, compared with 11/84 in clean-air controls, a statistically significant increase. In a repeat study by Heinrich et al. (1995), however, significant lung tumor increases were not detected in either female NMRI or C57BL/6N mice exposed to filtered exhaust diluted to produce an unfiltered particle concentration of 4.5 mg/m 3

. Filtered exhaust also failed to induce lung tumor induction in Syrian Golden hamsters (Heinrich et al., 1986a; Brightwell et al., 1989). Although lung tumor increases were reported in one study and lymphomas in another, these results could not be confirmed in subsequent investigations. It is therefore concluded that little direct evidence exists for carcinogenicity of the vapor phase of DE in laboratory animals at
concentrations tested.

7.3.3. Inhalation Studies (DE plus Cocarcinogens) Details of the studies reported here have been described earlier and in Table 7-3. Tumor initiation with urethan (1 mg/kg body weight i.p. at the start of exposure) or promotion with 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 DE up to 12 mg/m 3 (Pepelko and Peirano, 1983). Heinrich et al. (1986b) exposed Syrian hamsters of both sexes to DE diluted to a particle concentration of 4 mg/m 3 . See Section 7.3.1.1 for details of the exposure conditions. At the start of exposure the hamsters received either one dose of 4.5 mg diethylnitrosamine (DEN) subcutaneously per kg body weight or 20 weekly intratracheal instillations of 250 

### : g B[ a ]P. Female NMRI mice received weekly intratracheal instillations of 50 or 100 

### : g B[ a ]P for 10 or 20 weeks, respectively, or 50 

### : g dibenz[ah]anthracene (DBA) for 10 weeks. Additional groups of 96 newborn mice received one s.c. injection of 5 or 10 

### : g DBA between 24 and 48 hr after birth. Female Wistar rats received weekly subcutaneous injections of dipentylnitrosamine7-109 (DPN) at doses of 500 and 250 mg/kg body weight, respectively, during the first 25 weeks of exhaust inhalation exposure. Neither DEN, DBA, or DPN treatment enhanced any tumorigenic responses to DE. Response to B[ a ]P did not differ from that of BaP alone in hamsters, but results were inconsistent in mice. Although 20 B[ a ]P instillations induced a 71% tumor incidence in mice, concomitant diesel exposure resulted in only a 41% incidence. However, neither 10 B[ a ]P instillations nor DBA instillations induced significant effects. Takemoto et al. (1986) exposed Fischer 344 rats for 2 years to DE at particle concentrations of 2 to 4 mg/m 3 . One month after start of inhalation exposure one group of rats received di-isopropyl-nitrosamine (DIPN) administered i.p. at 1 mg/kg weekly for 3 weeks. Among injected animals autopsied at 18 to 24 mo, 10 adenomas and 4 adenocarcinomas were reported in 21 animals exposed to clean air, compared with 12 adenomas and 7 adenocarcinomas in 18 diesel-exposed rats. According to the authors, the incidence of adenocarcinomas was not significantly increased by exposure to DE. Brightwell et al. (1989) investigated the concomitant effects of DE and DEN in Syrian hamsters exposed to DE diluted to produce particle concentrations of 0.7, 2.2, or 6.6 mg/m 3 for 2 years. The animals received a single dose of 4.5 mg DEN s.c. 3 days prior to start of inhalation exposure. DEN did not affect the lack of responsiveness to DE alone. Heinrich et al. (1989b) also exposed Syrian hamsters of both sexes to DE diluted to a particle concentration of 3.75 mg/m 3 for up to 18 mo. After 2 weeks of exposure, groups were treated with either 3 or 6 mg DEN/kg body weight, respectively. Again, DEN did not significantly influence the lack of tumorigenic responses to DE. Heinrich et al. (1989a) investigated the effects of DPN in female Wistar rats exposed to DE diluted to achieve a particle concentration of 4.24 mg/m 3 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 25 weeks of exposure. The tumorigenic responses to DPN were not affected by exposure to DE. For details of exposure conditions of the hamster studies see Section 7.3.1.3. Heinrich et al. (1986a) and Mohr et al. (1986) compared the effects of exposure to particles having only a minimal carbon core but a much greater concentration of PAHs than DPM does. The desired exposure conditions were achieved by mixing coal oven flue gas with pyrolyzed pitch. The concentration of B[a]P and other PAHs per milligram of DPM was aboutthree orders of magnitude greater than that of DE. Female rats were exposed to the flue gaspyrolyzed pitch for 16 hr/day, 5 days/week at particle concentrations of 3 to 7 mg/m 3 for 22 mo, then held in clean air for up to an additional 12 mo. Among 116 animals exposed, 22 tumors were reported in 21 animals, for an incidence of 18.1%. One was a bronchioloalveolar adenoma, one was a bronchioloalveolar carcinoma, and 20 were squamous cell tumors. Among the latter, 16 were classified as benign keratinizing cystic tumors and 4 were classified as carcinomas. No tumors were reported in 115 controls. The tumor incidence in this study was comparable to that reported previously for the DE-exposed animals.7-110 In analyzing the studies of Heinrich et al. (1986a,b), Heinrich (1990b), Mohr et al. (1986), and Stöber (1986), it must be noted that the incidence of lung tumors occurring following exposure to whole DE, coal oven flue gas, or carbon black (15.8%, 18.1%, and 8% to 17%, respectively) was very similar. This occurred despite the fact that the PAH content of the PAH-enriched pyrolyzed pitch was more than three orders of magnitude greater than that of DE;carbon black, on the other hand, had only traces of PAHs. Based on these findings, particleassociated effects appear to be the primary cause of diesel-exhaust-induced lung cancer in rats exposed at high concentrations. This issue is discussed further in Chapter 7.

7.3.4. Lung Implantation or Intratracheal Instillation Studies
7.3.4.1. Rat Studies Grimmer et al. (1987), using female Osborne Mendel rats (35 per treatment group), provided evidence that PAHs in DE that consist of four or more rings have carcinogenic potential. Condensate was obtained from the whole exhaust of a 3.0-L passenger-car diesel engine connected to a dynamometer operated under simulated city traffic driving conditions. This condensate was separated by liquid-liquid distribution into hydrophilic and hydrophobic fractions representing 25% and 75% of the total condensate, respectively. The hydrophilic, hydrophobic, or reconstituted hydrophobic fractions were surgically implanted into the lungs of the rats. Untreated controls, vehicle (beeswax/trioctanoin) controls, and positive (B[a]P) controls were also included in the protocol (Table 7-6). Fraction Ilb (made up of PAHs with four to seven rings), which accounted for only 0.8% of the total weight of DPM condensate, produced the highest incidence of carcinomas following implantation into rat lungs. A carcinoma incidence of 17.1% was observed following implantation of 0.21 mg IIb/rat, whereas the nitro-PAH fraction (IId) at 0.18 mg/rat accounted for only a 2.8% carcinoma incidence. Hydrophilic fractions of the DPM extracts, vehicle (beeswax/trioctanoin) controls, and untreatedcontrols failed to exhibit carcinoma formation. Administration of all hydrophobic fractions (IIad) produced a carcinoma incidence (20%) similar to the summed incidence of fraction IIb (17.1%) and IId (2.8%). The B[a]P positive controls (0.03, 0.1, 0.3 mg/rat) yielded a carcinoma incidence of 8.6%, 31.4%, and 77.1%, respectively. The study showed that the tumorigenic agents were primarily four- to seven-ring PAHs and, to a lesser extent, nitroaromatics. However, these studies demonstrated that simultaneous administration of various PAH compounds resulted in a varying of the tumorigenic effect, thereby implying that the tumorigenic potency of PAH mixtures may not depend on any one individual PAH. This study did not provide any information regarding the bioavailability of the particle-associated PAHs that might be responsible for carcinogenicity. Kawabata et al. (1986) compared the effects of activated carbon and DE on lung tumor formation. One group of 59 F344 rats was intratracheally instilled with DPM (1 mg/week for 107-111Table 7-6. Tumor incidence and survival time of rats treated by surgical lung implantation withfractions from DE condensate (35 rats/group)Material portion by weight (%) Dose (mg)Mediansurvival timein weeks(range)Number ofcarcinomasaNumber ofadenomasbCarcinomaincidence (%)Hydrophilic fraction (I) (25) 6.7 97 (24-139) 0 1 0Hydrophobic fraction (II) (75)Nonaromatics +PACc2 + 3 rings (IIa) (72)PAHd4 to 7 rings (IIb) (0.8)Polar PAC (IIc) (1.1)Nitro-PAH (IId) (0.7)20.0019.220.210.290.1999 (50-139)103 (25-140)102 (50-140)97 (44-138)106 (32-135)50601 1000 14.2017.102.8Reconstituted hydrophobics(Ia, b, c, d) (74.5)Control, unrelatedControl (beeswax/trioctanoin)B[a]P19.910.30.10.0393 (46-136)110 (23-138)103 (51-136)69 (41-135)98 (22-134)97 (32-135)70027113 101000 20.00077.131.48.6aSquamous cell carcinoma.bBronchiolar/alveolar adenoma.cPAC = polycyclic aromatic compounds.dPAH = polycyclic aromatic hydrocarbons.Source: Adapted from Grimmer et al., 1987.7-112 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 surviving 30 months or more postinstillation were sacrificed and evaluated for histopathology. Among 42 animals exposed to DPM surviving 18 months or more, tumors were reported in 31, including 20 malignancies. In the subgroup surviving for 30 mo, tumors were detected in 19 of 20 animals, including 10 malignancies. Among the rats exposed to activated carbon, the incidence of lung tumors equaled 11 of 23 autopsied, with 7 cases of malignancy. Data for those dying between 18 and 30 months and those sacrificed at 30 months were not reported separately. Statistical analysis indicated that activated carbon induced a significant increase in lung tumor incidence compared with no tumors in 50 uninjected controls and 1 tumor in 23 solvent-injected controls. The tumor incidence was significantly greater in the DPM-instilled group and was significantly greater than the increase in the carbon-instilled group. A study reported by Rittinghausen et al. (1997) suggested that organic constituents of diesel particles play a role in the induction of lung tumors in rats. An incidence of 16.7% pulmonary cystic keratinizing squamous cell lesions was noted in rats intratracheally instilled with 15 mg whole DE particles, compared with 2.1% in rats instilled with 15 mg particles extracted to remove all organic constituents, and none among controls. Instillation of 30 mg of extracted particles induced a 14.6% incidence of squamous lesions, indicating the greater effectiveness of particles alone as lung particle overload increased. Iwai et al. (1997) instilled 2, 4, 8, and 10 mg of whole diesel particles over a 2- to 10- week period into female F/344 rats, 50 or more per group. Tumors were reported in 6%, 20%, 43%, and 74% of the rats, with incidence of malignant tumors equal to 2%, 13%, 34%, and 48%, respectively. In a second experiment comparing whole with extracted diesel particles, tumor incidence equaled 1/48 (2%) in uninjected controls, 3/55 (5%) in solvent controls, 12/56 (21%) in extracted diesel particles, and 13/106 (12%) in animals injected with unextracted particles. Although the extracted particles appeared to be more potent, when converted to a lung burden basis (mg/100 mg dry lung) the incidence was only 14% among those exposed to extracted exhaust compared with 31% in those exposed to whole particles. Dasenbrock et al. (1996) conducted a study to determine the relative importance of the organic constituents of diesel particles and particle surface area in the induction of lung cancer in rats. Fifty-two female Wistar rats were intratracheally instilled with 16-17 doses of DPM, extracted DPM, printex carbon black (PR), lampblack (LB), B[ a ]P, DPM + B[ a ]P, or PR + B[ a ]P. The animals were held for a lifetime or sacrificed when moribund. The lungs were necropsied and examined for tumors. Diesel particles were collected from a Volkswagen 1.6-L engine operating on a US FTP-72 driving cycle. The mass median aerodynamic diameter (MMAD) of the diesel particles was 0.25 

### : m and the specific surface area was 12 m 2 /gm.7-113 Following extraction with toluene, specific surface area increased to 138 m 2 /gm. The MMAD for extracted PR was equal to 14 nm, while the specific surface area equaled 271 m 2 /gm. The MMAD for extracted lampblack was equal to 95 nm, with a specific surface area equal to 20

m 2 /gm. The B[ a ]P content of the treated particles was 11.3 mg per gm diesel particles and 29.5 mg B[ a ]P per gm PR. Significant increases in lung tumors were detected in rats instilled with 15 mg unextracted DPM and 30 mg extracted DPM, but not 15 mg extracted DPM. Printex CB was more potent than lampblack CB for induction of lung tumors, whereas B[ a ]P was effective only at high doses. Total dose and tumor responses are shown in Table 7-4. A number of conclusions can be drawn from these results. First of all, particles devoid of organics are capable of inducing lung tumor formation, as indicated by positive results in the groups treated with high-dose extracted diesel particles and printex. Nevertheless, toluene extraction of organics from diesel particles results in a decrease in potency, indicating that the organic fraction does play a role in cancer induction. A relationship between cancer potency and particle surface area was also suggested by the finding that printex with a large specific surface area was more potent than either extracted DPM or lampblack, which have smaller specific areas. Finally, while very large doses of B[ a ]P are very effective in the induction of lung tumors, smaller doses adsorbed to particle surfaces had little detectable effect, suggesting that other organic components of DE may be of greater importance in the induction of lung tumors at low doses pf B[ a ]P (0.2-0.4 mg).

7.3.4.2. Syrian Hamster Studies Kunitake et al. (1986) and Ishinishi et al. (1988b) conducted a study in which total doses of 1.5, 7.5, or 15 mg of a dichloromethane extract of DPM were instilled intratracheally over 15 weeks into male Syrian hamsters that were then held for their lifetimes. The tumor incidences of 2.3% (1/44), 0% (0/56), and 1.7% (1/59) for the high-, medium-, and low-dose groups, respectively, did not differ significantly from the 1.7% (1/56) reported for controls. Addition of 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% and malignant tumor incidence of 88%. B[a]P (7.5 mg over 15 weeks) alone produced a tumor incidence rate of 88.2% (85% of these being malignant), which was not significantly different from the DPM extract + B[a]P group. Intratracheal administration of 0.03 

### : g B[a]P, the equivalent content in 15 mg of DPM extract, failed to cause a significant increase in tumors in rats. This study demonstrated a lack of detectable interaction between DPM extract and B[a]P, 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 composition between DPM extract and DPM. As with all intratracheal instillation protocols, DPM extract lacks the complement of volatile chemicals found in whole DE.7-114 The effects on hamsters of intratracheally instilled DPM suspension, DPM with Fe 2 O 3 , or DPM extract with Fe 2 O 3 as the carrier were studied by Shefner et al. (1982). The DPM component in each of the treatments was administered at concentrations of 1.25, 2.5, or 5.0 mg/week for 15 weeks to groups of 50 male Syrian golden hamsters. The total volume instilled was 3.0 mL (0.2 mL/week for 15 weeks). The DPM and dichloromethane extracts were suspended in physiological saline with gelatin (0.5% w/v), gum arabic (0.5% w/v), and propylene glycol (10% by volume). The Fe 2 O 3 concentration, when used, was 1.25 mg/0.2 mL of suspension. Controls received vehicle and, where appropriate, carrier particles (Fe 2 O 3 ) without the DPM component. Two replicates of the experiments were performed. Adenomatous hyperplasia was reported to be most severe in those animals treated with DPM or DPM plus Fe 2 O 3 particles and least severe in those animals receiving DPM plus Fe 2 O 3 . Of the two lung 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.

7.3.4.3. Mouse Studies Ichinose et al. (1997a) intratracheally instilled 36 four-week-old male ICR mice per group weekly for 10 weeks with sterile saline or 0.05, 0.1, or 0.2 mg DPM. Particles were collected from a 2.74-L four-cylinder Isuzu engine run at a steady speed of 1,500 rpm under a load of 10 torque (kg/m). Twenty-four hours after the last instillation, six animals per group were sacrificed for measurement of lung 8-hydroxydeoxyguanosine (8-OHdG). The remaining animals were sacrificed after 12 months for histopathological analysis. Lung tumor incidence varied from 4/30 (13.3%) for controls to 9/30 (30%), 9/29 (31%), and 7/29 (24.1%) for mice instilled with 0.05, 0.1, and 0.2 mg/week, respectively. The increase in animals with lung tumors compared with controls was statistically significant for the 0.1 mg dose group, the only group analyzed statistically. Increases in 8-OHdG, an indicator of oxidative DNA damage, correlated well with the increase in tumor incidence in the 0.05 mg dose group, although less so with the other two. The correlation coefficients r = 0.916, 0.765, and 0.677 for the 0.05, 0.10, and 0.20 mg DPM groups, respectively. In a similar study, 33 four-week-old male ICR mice per group were intratracheally instilled weekly for 10 weeks with sterile saline, 0.1 mg DPM, or 0.1 mg DPM from which the organic constituents were extracted with hexane (Ichinose et al., 1997b). Exhaust was collected from a 2.74-L four-cylinder Izuzu engine run at a steady speed of 2,000 rpm under a load of 6 torque (kg/m). Twenty-four hours after the last instillation, six animals per group were sacrificed for measurement of 8-OHdG. Surviving animals were sacrificed after 12 mo. The incidence of lung tumors increased from 3/27 (11.1%) among controls to 7/27 (25.9%) among those instilled with extracted diesel particles and 9/26 (34.6%) among those instilled with7-115 unextracted particles. The increase in number of tumor-bearing animals was statistically significant compared with controls ( p <0.05) for the group treated with unextracted particles. The increase in 8-OHdG was highly correlated with lung tumor incidence, r = 0.99.

7.3.5. Subcutaneous and Intraperitoneal Injection Studies
7.3.5.1. Mouse Studies In addition to inhalation studies, Orthoefer et al. (1981) also tested the effects of i.p. injections of DPM on male (A/S) strain mice. Three groups of 30 mice were injected with 0.1 mL of a suspension (particles in distilled water) containing 47, 117, or 235 

### : g of DPM collected from Fluoropore filters in the inhalation exposure chambers. The exposure system and exposure atmosphere are described in Appendix A. Vehicle controls received injections of particle suspension made up of particulate matter from control exposure filters, positive controls received 20 mg of urethan, and negative controls received no injections. Injections were made three times weekly for 8 weeks, resulting in a total DPM dose of 1.1, 2.8, and 5.6 mg for the low-, medium-, and 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 DPM groups, the tumor incidence was 2/30, 10/30, and 8/30, respectively. The incidence among urethan-treated animals (positive controls) was 100% (29/29), with multiple tumors per animal. The tumor incidence for the DPM-treated animals did not differ significantly from that of vehicle controls (8/30) or negative controls (7/28). The number of tumors per mouse was also unaffected by treatment. In further studies conducted by Orthoefer et al. (1981), an attempt was made to compare the potency of DPM with that of other environmental pollutants. Male and female Strain A mice were injected i.p. three times weekly for 8 weeks with DPM, DPM extracts, or various environmental mixtures of known carcinogenicity, including cigarette smoke condensate, coke oven emissions, and roofing tar emissions. Injection of urethan or dimethylsulfoxide (DMSO) served as positive or vehicle controls, respectively. In addition to DPM from the Nissan diesel previously described, an eight-cylinder Oldsmobile engine operated at the equivalent of 40 mph was also used to compare emission effects from different makes and models of diesel engine. 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 mice injected with coke oven extract. Furthermore, the increase in the extract-treated mice was significant only in comparison with uninjected controls (not injected ones) and did not occur when the experiment was repeated. Despite the use of a strain of mouse known to be sensitive to tumor induction, the overall findings of this study were negative. The authors provided several possible explanations for these findings, the most likely of which were (1) the carcinogens that7-116 were present were very weak, or (2) the concentrations of the active components reaching the lungs were insufficient to produce positive results. Kunitake et al. (1986) conducted studies using DPM extract obtained from a 1983 HD MMC—6D22P 11-L V-6 engine. Five s.c. injections of DPM extract (500 mg/kg per injection) resulted in a significant ( p <0.01) increase in subcutaneous tumors for female C57BL mice (5/22 [22.7%] vs. 0/38 among controls). Five s.c. doses of DPM extract of 10, 25, 30, 100, or 200 mg/kg failed to produce a significant increase in tumor incidence. One of 12 female ICR mice (8.3%) and 4 of 12 male ICR mice (33.3%) developed malignant lymphomas following neonatal s.c. administration of 10 mg of DPM extract per mouse. The increase in malignant lymphoma incidence for the male mice was statistically significant at p <0.05 compared with an incidence of 2/14 (14.3%) among controls. Treatment of either sex with 2.5 or 5 mg of DPM extract per mouse did not result in statistically significant increases in tumor incidence. Additional studies using DPM extract from LD (1.8-L, 4-cylinder) as well as HD engines with female ICR and nude mice (BALB/c/cA/JCL-nu) were also reported (Kunitake et al., 1988). Groups of 30 ICR and nude mice each were given a single s.c. injection of 10 mg HD extract, 10 mg HD + 50 

### : g 12-O-tetradecanoylphorbol 13-acetate (TPA), 10 mg LD extract +
50 

### : g TPA, or 50 

### : g TPA. No malignant tumors or papillomas were observed. One papillomatous lesion was observed in an ICR mouse receiving LD extract + TPA, and acanthosis was observed in one nude mouse receiving only TPA. In what appears to be an extension of the Kunitake et al. (1986) s.c. injection studies, Takemoto et al. (1988) presented additional data for subcutaneously administered DPM extract from HD and LD diesel engines. In this report, the extracts were administered to 5-week-old and neonatal (<24 hr old) C57BL mice of both sexes. DPM extract from HD or LD engines was administered weekly to the 5-week-old mice for 5 weeks at doses of 10, 25, 50, 100, 200, or 500 mg/kg, with group sizes ranging from 15 to 54 animals. After 20 weeks, comparison with a control group indicated a significant increase in the incidence of subcutaneous tumors for the 500 mg/kg HD group (5 of 22 mice [22.7%], p <0.01), the 100 mg/kg LD group (6 of 32 [18.8%],

p <0.01), and the 500 mg/kg LD group (7 of 32 [21.9%], p <0.01) in the adult mouse experiments. The tumors were characterized as malignant fibrous histiocytomas. No tumors were observed in other organs. The neonates were given single doses of 2.5, 5, or 10 mg DPM extract subcutaneously within 24 hr of birth. There was a significantly higher incidence of malignant lymphomas in males receiving 10 mg of HD extract and of lung tumors for males given 2.5 mg HD extract and for males given 5 mg and females given 10 mg LD extract. A dose-related trend that was not significant was observed for the incidences of liver tumors for both the HD extract- and LD extract-treated neonatal mice. The incidence of mammary tumors in female mice and multiple-organ tumors in male mice was also greater for some extract-treated7-117 mice, but was not dose related. The report concluded that LD DPM extract showed greater carcinogenicity than did HD DPM extract.

7.3.6. Dermal Studies
7.3.6.1. Mouse Studies In one of the earliest studies of diesel emissions, the effects of dermal application of extract from DPM were examined by Kotin et al. (1955). Acetone extracts were prepared from the DPM of a diesel engine (type and size not provided) operated at warmup mode and under load. These extracts were applied dermally three times weekly to male and female C57BL and 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 engine operated in warmup mode, two papillomas were detected after 13 mo. Four tumors were detected 16 months after the start of treatment in 8 surviving of 50 exposed male strain A mice treated with DPM extract from an engine operated under full load. Among female strain A mice treated with DPM extract from an engine operated under full load, 17 tumors were detected in 20 of 25 mice surviving longer than 13 mo. This provided a significantly increased tumor incidence of 85%. Carcinomas as well as papillomas were seen, but the numbers were not reported. Depass et al. (1982) examined the potential of DPM and dichloromethane extracts ofDPM to act as complete carcinogens, carcinogen initiators, or carcinogen promoters. In skinpainting studies, the DPM was obtained from an Oldsmobile 5.7-L diesel engine operated under constant load at 65 km/h. The DPM was collected at a temperature of 100°C. Groups of 40 C3H/HeJ mice were used because of their low spontaneous tumor incidence. For the complete carcinogenesis experiments, DPM was applied as a 5% or 10% suspension in acetone. Dichloromethane extract was applied as 5%, 10%, 25%, or 50% suspensions. Negative controls received acetone, and positive controls received 0.2% B[a]P. For tumor-promotion experiments, 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-initiation studies, a single initiating dose of 10% diesel particle suspension, 50% diesel particle extract, acetone, or PMA was followed by repeated applications of 0.0001% PMA. Following 8 months of treatment, the PMA dose in the initiation and promotion studies was increased to 0.01%. Animals were treated three times per week in the complete carcinogenesis and initiation experiments and five times per week in promotion experiments. All test compounds were applied to a shaved area on the back of the mouse. In the complete carcinogenesis experiments, one mouse receiving the high-dose (50%) suspension of extract developed a squamous cell carcinoma after 714 days of treatment. Tumor7-118 incidence in the B[a]P group was 100%, and no tumors were observed in any of the other groups. For the promotion studies, squamous cell carcinomas with pulmonary metastases were identified in one mouse of the 50% DPM extract group and in one in the 25% extract group. Another mouse in the 25% extract group developed a grossly diagnosed papilloma. Nineteen positive control mice had tumors (11 papillomas, 8 carcinomas). No tumors were observed for any of the other treatment groups. For the initiation studies, three tumors (two papillomas and one carcinoma) were identified in the group receiving DPM suspension and three tumors (two papillomas and one fibrosarcoma) were found in the DPM extract group. These findings were reported to be statistically insignificant using the Breslow and Mantel-Cox tests. Although these findings were not consistent with those of Kotin et al. (1955), the occurrence of a single carcinoma in a strain known to have an extremely low spontaneous tumor incidence may be of importance. Furthermore, a comparison between studies employing different strains of mice with varying spontaneous tumor incidences may result in erroneous assumptions. Nesnow et al. (1982) studied the formation of dermal papillomas and carcinomas following dermal application of dichloromethane extracts from coke oven emissions, roofing tar, DPM, and gasoline engine exhaust. DPM from five different engines, including a preproduction Nissan 220C, a 5.7-L Oldsmobile, a prototype Volkswagen Turbo Rabbit, a Mercedes 300D, and a HD Caterpillar 3304, was used for various phases of the study. Male and female Sencar mice (40 per group) were used for tumor initiation, tumor promotion, and complete carcinogenesis studies. For the tumor-initiation experiments, the DPM extracts were topically applied in single doses of 100, 500, 1,000, or 2,000 

### : g/mouse. The high dose (10,000 

### : g/mouse) was applied in five daily doses of 2,000 

### : g. One week later, 2 

### : g of the tumor promoter TPA was applied topically twice weekly. The tumor-promotion experiments used mice treated with 50.5 

### : g of B[a]P followed by weekly (twice weekly for high dose) topical applications (at the aforementioned doses) of the extracts. For the complete carcinogenesis experiments, the test extracts were applied weekly (twice weekly for the high doses) for 50 to 52 weeks. Only extracts from the Nissan, Oldsmobile, and Caterpillar engines were used in the complete carcinogenesis experiments. In the tumor-initiation studies, both B[a]P alone and the Nissan engine DPM extract followed by TPA treatment produced a significant increase in tumor (dermal papillomas) incidence at 7 to 8 weeks postapplication. By 15 weeks, the tumor incidence was greater than90% for both groups. No significant carcinoma formation was noted for mice in the tumorinitiation experiments following exposure to DPM extracts of the other diesel engines, although the Oldsmobile engine DPM extract at 2.0 mg/mouse did produce a 40% papilloma incidence in male mice at 6 mo. This effect, however, was not dose dependent.7-119 B[a]P (50.5 

### : g/week), coke oven extract (at 1.0, 2.0, or 4.0 mg/week), and the highest dose of roofing tar extract (4.0 mg/week) all tested positive for complete carcinogenesis activity. DPM extracts from only the Nissan, Oldsmobile, and Caterpillar engines were tested for complete carcinogenic potential, and all three proved to be negative using the Sencar mouse

assay. 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 three times or greater than that of the background value. Based on these criteria, only emissions from the Nissan were considered positive. Tumor initiation and complete carcinogenesis assays required that at least one dose produce a tumor incidence of at least 20%. None of the DPM samples yielded positive results based on this criterion. Kunitake et al. (1986, 1988) evaluated the effects of a dichloromethane extract of DPM obtained from a 1983 MMC M-6D22P 11-L V-6 engine. An acetone solution was applied in 10 doses every other day, followed by promotion with 2.5 

### : g of TPA three times weekly for 25 weeks. Exposure groups received a total dose of 0.5, 5, 15, or 45 mg of extract. Papillomas were reported in 2 of 50 animals examined in the 45 mg exposure group and in 1 of 48 in the 15 mg group compared with 0 of 50 among controls. Differences, however, were not statistically significant.

7.3.7. Summary and Conclusions of Laboratory Animal Carcinogenicity Studies 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 1980s, no chronic studies assessing inhalation of DE, the relevant mode for human exposure, had been reported. Since then long-term inhalation bioassays with DE have been carried out in the United States, Germany, Switzerland, and Japan, testing responses of rats, mice, and Syrian hamsters, and to a limited extent cats and monkeys.7-120Table 7-7. Dermal tumorigenic and carcinogenic effects of various emission extractsTumor initiationCompletecarcinogenesis TumorpromotionSample PapillomasaCarcinomasbCarcinomasbPapillomasaBenzo[a]pyrene +/+c+/+ +/+ +/+Topside coke oven +/+ -/+ NDdNDCoke oven main +/+ +/+ +/+ +/+Roofing tar +/+ +/+ +/+ +/+Nissan +/+ +/+ -/- NDOldsmobile +/+ -/- -/- NDVW Rabbit +/+ -/- IeNDMercedes +/- -/- ND NDCaterpillar -/- -/- -/- NDResidential furnace -/- -/- ND NDMustang +/+ -/+ ND NDaScored at 6 mo.bCumulative score at 1 year.cMale/female.dND = Not determined.eI = Incomplete.Source: Nesnow et al., 1982.7-121 It can be reasonably concluded that with adequate exposure, inhalation of DE is capable of inducing lung cancer in rats. Responses best fit cumulative exposure (concentration × daily exposure duration × days of exposure). Examination of rat data shown in Table 7-8 indicates a trend of increasing tumor incidence at exposures exceeding 1 × 10 4 mg 

### @ hr/m 3 . Exposures greater than approximately this value result in lung particle overload, characterized by slowed particle clearance and lung pathology, as discussed in Chapters 3 and 5, respectively. Tumor induction at high doses may therefore be primarily the result of lung particle overload with associatedinflammatory responses. Although tumorigenic responses could not be detected under nonparticle-overload conditions, the animal experiments lack sensitivity to determine if a threshold exists. However, studies such as those reported by Driscoll et al. (1996) support the existence ofa threshold if it is assumed that inflammation is a prerequisite for lung tumor induction. If lowdose effects do occur, it can be hypothesized that the organic constituents are playing a role. See Chapter 7 for a discussion of this issue. Although rats develop adenomas, adenocarcinomas, and adenosquamous cell carcinomas, they also develop squamous keratinizing lesions. This latter spectrum appears for the most part to be peculiar to the rat. In a recent workshop aimed at classifying these tumors (Boorman et al., 1996), it was concluded that when these lesions occur in rats as part of a carcinogenicity study, they must be evaluated on a case-by-case basis and regarded as a part of the total biologic profile of the test article. If the only evidence of tumorigenicity is the presence of cystic keratinizing epitheliomas, it may not have relevance to human safety evaluation of a substance or particle. Their use in quantifying cancer potency is even more questionable. The evidence for response of common strains of laboratory mice exposed under standard inhalation protocols is equivocal. Inhalation of DE induced significant increases in lung tumors in female NMRI mice (Heinrich et al., 1986b; Stöber, 1986) and in female Sencar mice (Pepelko and Peirano, 1983). An apparent increase was also seen in female C57BL mice (Takemoto et al., 1986). However, in a repeat of their earlier study, Heinrich et al. (1995) failed to detect lung tumor induction in either NMRI or C57BL/6N mice. No increases in lung tumor rates were reported in a series of inhalation studies using strain A mice (Orthoefer et al., 1981; Kaplan et al., 1982, 1983; White et al., 1983). Finally, Mauderly et al. (1996) reported no tumorigenic responses in CD-1 mice exposed under conditions resulting in positive responses in rats. The successful induction of lung tumors in mice by Ichinose et al. (1997a,b) via intratracheal instillation may have been the result of focal deposition of larger doses. Positive effects in Sencar mice may be due to use of a strain sensitive to tumor induction in epidermal tissue by organic agents, as well as exposure from conception, although proof for such a hypothesis is lacking.7-122Table 7-8. Cumulative (concentration × time) exposure data for rats exposed to whole DEStudyExposurerate/duration(hr/week, mo)Total exposuretime (hr)Particleconcentration(mg/m3)Cumulative exposure(mg"hr/m3)Tumor incidence(%)aPer week TotalMauderly et al. (1987) 35, 3035, 3035, 3035, 304.20042004e+15 00.353.57.1012.25122.5248.514701470029820 0.91.33.612.8Nikula et al. (1995) 80, 2380, 2380, 23736073607360 02.56.50200.0520.01840047840 1.07.018.0Heinrich et al. (1986a) 95, 3595, 351330013300 4.24 402.8 56392 17.8Heinrich et al. (1995) 90, 2490, 2490, 2490, 248.64086409e+15 00.82.57.0072.0225.0630.074002180061700 005.522.0Ishinishi et al. (1988a)(Light-duty engine)(Heavy-duty engine)96, 3096, 3096, 3096, 3096, 3096, 3096, 3096, 3096, 3096, 301.15201152e+49 00.10.41.12.300.51.01.83.709.638.4105.6220.8048.096.0172.8355.21.152460813e+373.32.40.84.12.40.80.803.36.5Table 7-8. Cumulative (concentration × time) exposure data for rats exposed to whole DE (continued)StudyExposurerate/duration(hr/week, mo)Total exposuretime (hr)Particleconcentration(mg/m3)Cumulative exposure(mg"hr/m3)Tumor incidence(%)aPer week Total

7-123Brightwell et al. (1989) 80, 2480, 2480, 2480, 247.6807681e+15 00.72.26.6056.0176.0528.053761689650688 1.20.79.738.5Kaplan et al. (1983) 140, 15140, 15140, 15140, 158.4008401e+15 00.250.751.5035.0105.0210.02100630012600 03.310.03.3Iwai et al. (1986b) 56, 2456, 2453765376 4.9 274.4 26342 36.8Takemoto et al. (1986) 16, 18-2416, 18-241,152-1,5361,152-1,53602-4032-6403,456-4,6080Karagianes et al.(1981)30, 2030, 2024002400 8.3 249 19920 16.6Iwai et al. (1997) 56, 2448, 2454, 24537649925616 9.43.25.1526154275 5.47041597e+14 4212427-124 Attempts to induce significant increases in lung tumors in Syrian hamsters by inhalation of whole DE were unsuccessful (Heinrich et al., 1982, 1986b, 1989b; Brightwell et al., 1986). However, hamsters are considered to be relatively insensitive to lung tumor induction. For example, while cigarette smoke, a known human carcinogen, was shown to induce laryngeal cancer in hamsters, the lungs were relatively unaffected (Dontenwill et al., 1973). Neither cats (Pepelko and Peirano, 1983 [see Chapter 7]) nor monkeys (Lewis et al., 1989) developed tumors following 2-year exposure to DE. The duration of these exposures, however, was likely to be inadequate for these two longer-lived species, and group sizes were quite small. Exposure levels were also below the maximum tolerated dose (MTD) in the monkey studies and, in fact, only borderline for detection of lung tumor increases in rats. Long-term exposure to DE filtered to remove particulate matter failed to induce lung tumors in rats (Heinrich et al., 1986b; Iwai et al., 1986b; Brightwell et al., 1989), or in Syrian hamsters (Heinrich et al., 1986b; Brightwell, 1989). A significant increase in lung carcinomas was reported by Heinrich et al. (1986b) in NMRI mice exposed to filtered exhaust. However, in a more recent study the authors were unable to confirm earlier results in either NMRI or C57BL/6N mice (Heinrich et al., 1995). Although filtered exhaust appeared to potentiate the carcinogenic effects of DEN (Heinrich et al., 1982), because of the lack of positive data in rats and equivocal or negative data in mice it can be concluded that filtered exhaust is either not carcinogenic or has a low cancer potency. Kawabata et al. (1986) demonstrated the induction of lung tumors in Fischer 344 rats following intratracheal instillation of DPM. Rittinghausen et al. (1997) reported an increase in cystic keratinizing epitheliomas following intratracheal instillation of rats with either original DPM or DPM extracted to remove the organic fraction, with the unextracted particles inducing a slightly greater effect. Grimmer et al. (1987) showed not only that an extract of DPM was carcinogenic when instilled in the lungs of rats, but also that most of the carcinogenicity resided in the portion containing PAHs with four to seven rings. Intratracheal instillation did not induce 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 significant skin tumor increases in two studies (Kotin et al., 1955; Nesnow et al., 1982). Kunitake et al. (1986) also reported a marginally significant increase in skin papillomas in ICR mice treated with an organic extract from an HD diesel engine. Negative results were reported by Depass et al. (1982) for skin-painting studies using mice and acetone extracts of DPM suspensions. However, in this study the exhaust particles were collected at temperatures of 100 °C, which would minimize the condensation of vapor-phase organics and, therefore, reduce the availability of potentially carcinogenic compounds that might normally be present on DE particles. A significant increase in the incidence of sarcomas in female C57Bl mice was reported by Kunitake et al. (1986) following s.c. administration of LD DPM extract at doses of7-125 500 mg/kg. Takemoto et al. (1988) provided additional data for this study and reported an increased tumor incidence in the mice following injection of LD engine DPM extract at doses of 100 and 500 mg/kg. Results of i.p. injection of DPM or DPM extracts in strain A mice were generally negative (Orthoefer et al., 1981; Pepelko and Peirano, 1983), suggesting that the strain 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 inconclusive regarding the tumor-promoting potential of either filtered or whole DE. A reportby Heinrich et al. (1982), however, indicated that filtered exhaust may promote the tumorinitiating effects of DEN in hamsters. Several reports (Wong et al., 1985; Bond et al., 1990) affirm observations of the potential carcinogenicity of DE by providing evidence for DNA damage in rats. These findings are discussed in more detail in Chapter 3, Section 3.6. Evidence for the mutagenicity of organic agents present in diesel engine emissions is also provided in Chapter 4. Evidence for the importance of the carbon core was initially provided by studies of Kawabata et al. (1986), which showed induction of lung tumors following intratracheal instillation of carbon black that contained no more than traces of organics, and studies of Heinrich (1990b) that indicated that exposure via inhalation to carbon black (Printex 90) particles induced lung tumors at concentrations similar to those effective in DPM studies. Additional studies by Heinrich et al. (1995) and Nikula et al. (1995) confirmed the capability of carbon particles to induce lung tumors. Induction of lung tumors by other particles of low solubility, such as titanium dioxide (Lee et al., 1986), confirmed the capability of particles to induce lung tumors. Pyrolyzed pitch, on the other hand, essentially lacking a carbon core but having much higher PAH concentrations than DPM, also was effective in tumor induction (Heinrich et al., 1986a, 1994). The relative importance of the adsorbed organics, however, remains to be elucidated and is of some concern because of the known carcinogenic capacity of some of these chemicals. These include polycyclic aromatics as well as nitroaromatics, as described in Chapter 2. Organic extracts of particles also have been shown to induce tumors in a variety of injection, intratracheal instillation, and skin-painting studies, and Grimmer et al. (1987) have, in fact, shown that the great majority of the carcinogenic potential following instillation resided in the fraction containing four- to seven-ring PAHs. In summary, based on positive inhalation studies in rats exposed to high concentrations, intratracheal instillation studies in rats and mice exposed to high doses, and supported by positive mutagenicity studies, the evidence for carcinogenicity of DE is considered to be adequate in animals. The contribution of the various fractions of DE to the carcinogenic response is less certain. Exposure to filtered exhaust generally failed to induce lung tumors. The presence of known carcinogens adsorbed to diesel particles and the demonstrated7-126 tumorigenicity of particle extracts in a variety of injection, instillation, and skin-painting studies indicate a carcinogenic potential for the organic fraction. Studies showing that long-term exposure at high concentrations of poorly soluble particles (e.g., carbon black, TiO 2 ) can also induce tumors, on the other hand, have provided definitive evidence that the carbon core of the diesel particle is primarily instrumental in the carcinogenic response observed in rats under sufficient exposure conditions. The ability of DE to induce lung tumors at non-particle-overload conditions, and the relative contribution of the particles’ core versus the particle-associated organics (if effects do occur at low doses) remains to be determined.

7.4. MODE OF ACTION OF DIESEL EXHAUST-INDUCED CARCINOGENESIS As noted in Chapter 2, DE is a complex mixture that includes a vapor phase and a particle phase. The particle phase consists of an insoluble carbon core with a large number of organic compounds, as well as inorganic compounds such as sulfates, adsorbed to the particlesurface. Some of the semivolatile and particle-associated compounds, in particular PAHs, nitroPAHs, oxy-PAHs, and oxy-nitro-PAHs (Scheepers and Bos, 1992), are considered likely to be carcinogenic in humans. The vapor phase also contains a large number of organic compounds, including several known or probable carcinogens such as benzene and 1,3-butadiene. Because exposure to the vapor phase alone, even at high concentrations, failed to induce lung cancer in laboratory animals (Heinrich et al., 1986b), the mode-of-action discussion will focus on the particulate matter phase. Additive or synergistic effects of vapor-phase components, however, cannot be ruled out, as chronic inhalation bioassays involving exposure to diesel particles alone
have not been carried out. Several hypotheses regarding the primary mode of action of DE have been proposed. Initially it was generally believed that cancer was induced by particle-associated organics acting via a genotoxic mechanism. By the late 1980s, however, studies indicated that carbon particles virtually devoid of organics could also induce lung cancer at sufficient inhaled concentrations (Heinrich, 1990b). This finding provided support for a hypothesis originally proposed by Vostal (1986) that induction of lung tumors arising in rats exposed to high concentrations of DE is related to overloading of normal lung clearance mechanisms, accumulation of particles, and cell damage followed by regenerative cell proliferation. The action of particles is therefore mediated by epigenetic mechanisms that can be characterized more by promotional than initiation stages of the carcinogenic process. More recently several studies have focused upon the production of reactive oxygen species generated from particle-associated organics, which may induce oxidative DNA damage at exposure concentrations lower than those required to produce lung particle overload. Because it is likely that more than one of these factors is involved in the carcinogenic process, a key consideration is their likely relative contribution at different exposure levels. The following discussion will therefore consider the possible relationship of the organic components of exhaust, inflammatory responses associated with lung particle overload, reactive oxygen7-127 species, and physical characteristics of diesel particles to cancer induction, followed by a hypothesized mode of action, taking into account the likely contribution of the factors discussed.

7.4.1. Potential Role of Organic Exhaust Components in Lung Cancer Induction More than 100 carcinogenic or potentially carcinogenic components have been specifically identified in diesel emissions, including various PAHs and nitroarenes such as 1-nitropyrene (1-NP) and dinitropyrenes (DNPs). The majority of these compounds are adsorbed to the carbon core of the particulate phase of the exhaust and, if desorbed, may become available for biological processes such as metabolic activation to mutagens. Among such compounds identified from DE are B[ a ]P, dibenz[ a,h ]anthracene, pyrene, chrysene, and nitroarenes such as 1-NP, 1,3-DNP, 1,6-DNP, and 1,8-DNP, all of which are mutagenic, carcinogenic, or implicated as procarcinogens or cocarcinogens (Stenback et al., 1976; Weinstein and Troll, 1977; Thyssen et al., 1981; Pott and Stöber, 1983; Howard et al., 1983; Hirose et al., 1984; Nesnow et al., 1984; El-Bayoumy et al., 1988). More recently Enya et al. (1997) reported isolation of 3-nitrobenzanthrone, one of the most powerful direct-acting mutagens known to date, from the organic extracts of DE. Grimmer et al. (1987) separated DE particle extract into a water- and a lipid-soluble fraction, and the latter was further separated into a PAH-free, a PAH-containing, and a polar fraction by column chromatography. These fractions were then tested in Osborne-Mendel rats by pulmonary implantation at doses corresponding to the composition of the original DE. The water-soluble fraction did not induce tumors; the incidences induced by the lipid-soluble fractions were 0% with the PAH-free fraction, 25% with the PAH and nitro-PAH- containing fractions, and 0% with the polar fraction. The PAH and nitro-PAH-containing fraction, comprising only 1% by weight of the total extract, was thus shown to be responsible for most, if not all, of the carcinogenic activity. Exposure of rats by inhalation to 2.6 mg/m 3 of an aerosol of tar-pitch condensate with no carbon core but containing 50 

### : g/m 3 B[ a ]P along with other PAHs for 10 months induced lung tumors in 39% of the animals. The same amount of tar-pitch vapor condensed onto the surface of carbon black particles at 2 and 6 mg/m 3 resulted in tumor rates that were roughly two times higher (89% and 72%). Because exposure to 6 mg/m 3 carbon black almost devoid of extractable organic material induced a lung tumor rate of 18%, the combination of PAHs and particlesincreases their effectiveness (Heinrich et al., 1994). Although this study shows the tumorinducing capability of PAHs resulting from combustion, it should be noted that the B[ a ]P content in the coal-tar pitch was about three orders of magnitude greater than in diesel soot. Moreover, because organics are present on diesel particles in a thinner layer and the 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 in the carcinogenic process.7-128 Exposure of humans to related combustion emissions provides some evidence for the involvement of organic components. Mumford et al. (1989) reported greatly increased human lung cancer mortality in Chinese communes burning so-called smoky coal, but not wood, in unvented open-pit fires used for heating and cooking. Although particle concentrations were similar, PAH levels were five to six times greater in the air of communes burning smoky coal. Coke oven emissions, containing high concentrations of PAHs but lacking an insoluble carbon core, have also been shown to be carcinogenic in humans (Lloyd, 1971). Adsorption of PAHs to a carrier particle such as hematite, CB, aluminum, or titanium dioxide enhances their carcinogenic potency (Farrell and Davis, 1974). As already noted, adsorption to carbon particles greatly enhanced the tumorigenicity of pyrolyzed pitch condensate containing B[ a ]P and other aromatic carcinogens (Heinrich et al., 1995). The increased effectiveness can be partly explained by more efficient transport to the deep lung. Slow release also enhances residence time in the lungs and prevents overwhelming of activating pathways. As discussed in Chapter 3, free organics are likely to be rapidly absorbed into the bloodstream, which may explain why the vapor-phase component of exhaust is relatively ineffective in the induction of pathologic or carcinogenic effects. Even though the organic constituents may be tightly bound to the particle surface, significant elution is still likely because particle clearance half-times are nearly 1 year in humans (Bohning et al., 1982). Furthermore, Gerde et al. (1991) presented a model demonstrating that large aggregates of inert dust containing crystalline PAHs are unlikely to form at doses typical of human exposure. This allows the particles to deposit and react with the surrounding lung medium, without interference from other particles. Particle-associated PAHs can then be expected to be released more rapidly from the particles. Bond et al. (1984) provided evidence that alveolar macrophages from beagle dogs metabolized B[ a ]P coated on diesel particles to proximate carcinogenic forms. Unless present on the particle surface, B[ a ]P is more likely to pass directly into the bloodstream and escape activation by phagocytic cells. The importance of DE-associated PAHs in the induction of lung cancer in humans may be enhanced because of the possibility that the human lung is more sensitive to these compounds than are rat lungs. Rosenkranz (1996) summarized information indicating that in humans and mice, large proportions of lung cancers contain both mutated p 53 suppressor genes and K- ras genes. Induction of mutations in these genes by genotoxins, however, is much lower in rats than
in humans or mice. B[ a ]P, although only one of many PAHs present in DE, is the one most extensively studied. Bond et al. (1983, 1984) demonstrated metabolism of particle-associated B[ a ]P and free B[ a ]P by alveolar macrophages (AM) and by type II alveolar cells. The respiratory tract cytochrome P-450 systems have an even greater concentration in the nonciliated bronchiolar cells (Boyd, 1984). It is worth noting that bronchiolar adenomas that develop following diesel exposure have been found to resemble both Type II and nonciliated bronchiolar cells. It should7-129 also be noted that any metabolism of procarcinogens by these latter two cell types probably involves the preextraction of carcinogens in the extracellular lining fluid and/or other endocytotic cells, as they are not especially important in phagocytosis of particles. Thus, bioavailability is an important issue in assessing the relative importance of PAHs.Additionally, a report by Borm et al. (1997) indicates that incubating rat lung epithelialderived cells with human polymorphonucleocytes (PMNs) (either unactivated or activated by preexposure to phorbol myristate acetate) increases DNA adduct formation caused by exposure to B[ a ]P; at 0.05 to 0.5 micromolar concentration, addition of more activated PMN in relation to the number of lung cells further increased adduct formation in a dose-dependent manner. The authors suggest that “an inflammatory response in the lung may increase the biologicallyeffective dose of PAHs, and may be relevant to data interpretation and risk assessment of PAHcontaining particles.” These data raise the possibility that DE exposure at low concentrations may result in levels of neutrophil influx that would not necessarily be detectable via histopathological examination as acute inflammation, but that might be effective at amplifying any potential DE genotoxic effect. Nitro-PAHs have also been implicated as potentially involved in diesel-exhaust-induced lung cancer. Although the nitro-PAH fraction of diesel was less effective than PAHs in the induction of lung cancer when implanted into the lungs of rats (Grimmer et al., 1987), in a study of various extracts of DE particles, 30%-40% of the total mutagenicity could be attributed to a group of six nitroarenes (Salmeen et al., 1984). Moreover, Gallagher et al. (1994) reported results suggesting that DNA adducts are formed from nitro-PAHs present in DNA and may play a role in the carcinogenic process. Nitroarenes, however, quantitatively represent a very small percentage of diesel particle extract (Grimmer et al., 1987), making their role in the tumorigenic response uncertain. The induction of DNA adducts in humans occupationally exposed to DE indicates the likelihood that PAHs are participating in the tumorigenic response, and that these effects can occur at exposure levels less than those required to induce lung particle overload. Distinct adduct patterns were found among garage workers occupationally exposed to DE when compared with nonexposed controls (Nielsen and Autrup, 1994). Furthermore, the findings were concordant with the adduct patterns observed in groups exposed to low 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 DE exposure compared with unexposed mechanics. Hou et al. (1995) found elevated adduct levels in bus maintenance workers exposed to DE. Although no difference in mutant frequency was observed between the groups, the adduct levels were significantly different (3.2 vs. 2.3 × 10 -8 ). Nielsen et al. (1996b) measured three biomarkers in DE-exposed bus garage workers: lymphocyte DNA adducts, hydroxyethylvaline adducts in hemoglobin, and 1-hydroxypyrene in urine. Significantly increased levels were reported for all three. Qu et al. (1996) detected increased adduct levels, as7-130 well as increases in some individual adducts, in the blood of underground coal miners exposed to
DE.

7.4.2. Role of Inflammatory Cytokines and Proteolytic Enzymes in the Induction of Lung Cancer in Rats by Diesel Exhaust It is well recognized that the deposition of particles in the lung can result in the efflux of PMNs from the vascular compartment into the alveolar space compartment in addition to expanding the AM population size. Following acute exposures, the influx of the PMNs is transient, lasting only a few days (Adamson and Bowden, 1978; Bowden and Adamson, 1978; Lehnert et al., 1988). During chronic exposure the numbers of PMNs lavaged from the lungs of diesel-exposed rats generally increased with increasing exposure duration and inhaled DPM concentration (Strom, 1984). Strom (1984) also found that PMNs in diesel-exposed lungs remained persistently elevated for at least 4 months after cessation of exposure, a potential mechanism that may be related to an ongoing release of phagocytized particles. Evidence in support of this possibility was reported by Lehnert et al. (1989) in a study in which rats were intratracheally instilled with 0.85, 1.06, or 3.6 mg of polystyrene particles. The PMNs were not found to be abnormally abundant during the clearance of the two lower lung burdens, but they became progressively elevated in the lungs of the animals in which alveolar-phase clearance was inhibited. Moreover, the particle burdens in the PMNs became progressively greater over time. Such findings are consistent with an ongoing particle relapse process, in which particles released by dying phagocytes are ingested by new ones. The inflammatory response, characterized by efflux of PMNs from the vascular compartment, is mediated by inflammatory chemokines. Driscoll et al. (1996) reported that inhalation of high concentrations of carbon black stimulated the release of macrophage inflammatory protein 2 (MIP-2) and monocyte chemotactic protein 1 (MCP-1). They also reported a concomitant increase in hprt mutants. In a following study it was shown that particle exposure stimulates production of tumor necrosis factor TNF- 

### " , an agent capable of activating expression of several proteins that promote both adhesion of leucocytes and chemotaxis (Driscoll et al., 1997a). In addition, alveolar macrophages also have the ability to release several other effector molecules or cytokines that can regulate numerous functions of other lung cells, including their rates of proliferation (Bitterman et al., 1983; Jordana et al., 1988; Driscoll et al., 1996). Another characteristic of AMs and PMNs under particle overload conditions is the release of a variety of potentially destructive hydrolytic enzymes, a process known to occur simultaneously with the phagocytosis of particles (Sandusky et al., 1977). The essentially continual release of such enzymes during chronic particle deposition and phagocytosis in the lung may be detrimental to the alveolar epithelium, especially to Type I cells. Evans et al. (1986) showed that injury to Type I cells is followed shortly thereafter by a proliferation of Type7-131 II cells. Type II cell hyperplasia is a common feature observed in animals that have received high lung burdens of various types of particles, including unreactive polystyrene microspheres. Exaggerated proliferation as a repair or defensive response to DPM deposition may have the effect of amplifying the likelihood of neoplastic transformation in the presence of carcinogens beyond that which would normally occur with lower rates of proliferation, assuming an increase in the cycling of target cells and the probability of a neoplastic-associated genomic disturbance.

7.4.3. Role of Reactive Oxygen Species in Lung Cancer Induction by Diesel Exhaust Phagocytes from a variety of rodent species produce elevated levels of oxidant reactants in response to challenges, with the physiochemical characteristics of a phagocytized particle being a major factor in determining the magnitude of the oxidant-producing response. Active oxygen species released by the macrophages and lymphatic cells can cause lipid peroxidation in the membrane of lung epithelial cells. These lipid peroxidation products can initiate a cascade of oxygen free radicals that progress through the cell to the nucleus, where they damage DNA. If this damage occurs during the epithelial cell’s period of DNA synthesis, there is some probability that the DNA will be replicated unrepaired (Lechner and Mauderly, 1994). The generation of reactive oxygen species by both AMs and PMNs should therefore be considered as one potential factor of what probably is a multistep process that culminates in the development of lung tumors in response to chronic deposition of DPM. Even though products of phagocytic oxidative metabolism, including superoxide anions, hydrogen peroxide, and hydroxyl radicals, can kill tumor cells (Klebanoff and Clark, 1978), and the reactive oxygen species can peroxidize lipids to produce cytotoxic metabolites such as malonyldialdehyde, some products of oxidative metabolism apparently can also interact with DNA to produce mutations. Cellular DNA is damaged by oxygen free radicals generated from a variety of sources (Ames, 1983; Trotter, 1980). Along this line, Weitzman and Stossel (1981) found that human peripheral leukocytes are mutagenic in the Ames assay. This mutagenic 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 a patient with chronic granulomatous diseases, in which neutrophils have a defect in the NADPH oxidase generating system (Klebanoff and Clark, 1978), were less effective in producing mutations than were normal leukocytes. Of related significance, Phillips et al. (1984) demonstrated that the incubation of Chinese hamster ovary cells with xanthine plus xanthine oxidase (a system for enzymatically generating active oxygen species) resulted in genetic damage hallmarked by extensive chromosomal breakage and sister chromatid exchange and produced an increase in the frequency of thioguanidine-resistant cells (HGPRT test). Aside from interactions of oxygen species with DNA, increasing evidence also points to an important role of7-132 phagocyte-derived oxidants and/or oxidant products in the metabolic activation of procarcinogens to their ultimate carcinogenic form (Kensler et al., 1987). Driscoll et al. (1997b) have demonstrated that exposure to doses of particles producing significant neutrophilic inflammation are associated with increased mutation in rat alveolar type II cells. The ability of particle-elicited macrophages and neutrophils to exert a mutagenic effect 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. Hatch and co-workers (1980) have demonstrated that interactions of guinea pig AMs with a wide variety of particles, such as silica, metal oxide-coated fly ash, polymethylmethacrylate beads, chrysotile asbestos, fugitive dusts, polybead carboxylate microspheres, glass and latex beads, uncoated fly ash, and fiberglass increase the production of reactive oxygen species. Similar findings have been reported by numerous investigators for human, rabbit, mouse, and guinea pig AMs (Drath and Karnovsky, 1975; Allen and Loose, 1976; Beall et al., 1977; Lowrie and Aber, 1977; Miles et al., 1977; Rister and Baehner, 1977; Hoidal et al., 1978). PMNs are also known to increase production of superoxide radicals, hydrogen peroxide, and hydroxyl radicals in response to membrane-reactive agents and particles (Goldstein et al., 1975; Weiss et al., 1978; Root and Metcalf, 1977). Although these responses may occur at any concentration, they are likely to be greatly enhanced at high exposure concentrations with slowed clearance and lung particle overload. Reactive oxygen species can also be generated from particle-associated organics. Sagai et al. (1993) reported that DPM can nonenzymatically generate active oxygen species (e.g., superoxide [O 2 - ] and hydroxyl radical [

. OH] in vitro without any biologically activating systems) such as microsomes, macrophages, hydrogen peroxide, or cysteine. Because DPM washed with methanol could no longer produce these radicals, it was concluded that the active components were compounds extractable with organic solvents. However, the nonenzymatic contribution to the DPM-promoted active oxygen production was negligible compared with that generated via an enzymatic route (Ichinose et al., 1997a). They reported that O 2 - and

. OH can be enzymatically generated from DPM by the following process. Soot-associated quinone-like compounds are reduced to the semiquinone radical by cytochrome P-450 reductase. These semiquinone radicals then reduce O 2 to O 2 - , and the produced superoxide reduces ferric ions to ferrous ions, which catalyzes the homobiotic cleavage of H 2 O 2 dismutated from O 2 by superoxide dismutase or spontaneous reactions to produce

. OH. According to Kumagai et al. (1997), while quinones are likely to be the favored substrates for this reaction, the participation
of nitroaromatics cannot be ruled out. One of the critical lesions to DNA bases generated by oxygen free radicals is 8- hydroxydeoxyguanosine (8-OHdG). The accumulation of 8-OhdG as a marker of oxidative DNA damage could be an important factor in enhancing the mutation rate leading to lung cancer7-133 (Ichinose et al., 1997a). For example, formation of 8-OHdG adducts leads to G:C to T:A transversions unless repaired prior to replication. Nagashima et al. (1995) demonstrated that the production of (8-OHdG) is induced in mouse lungs by intratracheal instillation of DPM. Ichinose et al. (1997b) reported further that although intratracheal instillation of DPM in mice induced a significant increase in lung tumor incidence, comparable increases were not reported when mice were instilled with extracted DPM (to remove organics). Lung injury was also less in the mice instilled with extracted DPM. Moreover, increases in 8-OHdG in the mice instilled with unextracted DPM correlated very well with increases in tumor rates. In a related study, Ichinose et al. (1997a) intratracheally instilled small doses of DPM, 0.05, 0.1, or 0.2 mg weekly for 3 weeks, in mice fed standard or high-fat diets either with or without 

### $ -carotene. High dietary fat enhanced DPM-induced lung tumor incidence, whereas 

### $ -carotene, which may act as a free radical scavenger, partially reduced the tumorigenic response. Formation of 8-OHdG was again significantly correlated with lung tumor incidence in these studies, except at the highest dose. Dasenbrock et al. (1996) reported that extracted DPM, intratracheally instilled into rats (15 mg total dose) induced only marginal increases in lung tumor induction, while unextracted DPM was considerably more effective. Although adducts were not measured in this study, it nevertheless provides support for the likelihood that activation of organic metabolites and/or generation of oxygen free radicals from organics are involved in the carcinogenic process. Additional support for the involvement of particle-associated radicals in tissue damage was provided by the finding that pretreatment with superoxide dismutase (SOD), an antioxidant, markedly reduced lung injury and death due to instillation of DPM. Similarly, Hirafuji et al. (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 species resulting from exposure to DPM is involved in the carcinogenic process, it should be noted that 8-OHdG is efficiently repaired and that definitive proof of a causal relationship in humans is still lacking. It is also uncertain whether superoxide or hydroxyl radicals chemically generated by DPM alone promote 8-OHdG production in vivo and induce lung toxicity, because SOD is extensively located in mammalian tissues. Nevertheless, demonstration that oxygen free radicals can be generated from particle-associated organics, that their presence will induce adduct formation and DNA damage unless repaired, that tumor induction in experimental animals correlates with OhdG adducts, and that treatment with antioxidant limits lung damage, provides strong support for the involvement of oxygen free radicals in the toxicologic and carcinogenic response to DE.

7.4.4. Relationship of Physical Characteristics of Particles to Cancer Induction The biological potential of inhaled particles is strongly influenced by surface chemistry 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). A7-134 major factor is specific surface area (surface area/mg). PMNs characteristically are increased abnormally in the lung by DE exposure, but their presence in the lungs does not appear to be excessive following the pulmonary deposition of even high lung burdens of spherical TiO 2 particles in the 1-2 

### : m diameter range (Strom, 1984). In these studies lung tumors were detected only at an inhaled concentration of 250 

### : g/m 3 . In a more recent study in which rats were exposed to TiO 2 in the 15-40 nm size range, inhibition of particle clearance and tumorigenesis were induced at concentrations of 10 mg/m 3 (Heinrich et al., 1995). Comparison of several chronic inhalation studies correlating particle mass and particle surface area retained in the lung with tumor incidence indicated that particle surface area is a much better dosimeter than particle mass (Oberdörster and Yu, 1990; Driscoll et al., 1996). Heinrich et al. (1995) also found that lung tumor rates increased with specific particle surface area following exposure to DE, carbon black, or titanium dioxide, irrespective of particle type. Langer and Nolan (1994) reported that the hemolytic potential of Min-U-Sil15, a silica flour, increased in direct relationship to specific surface area at nominal particle diameters ranging from 0.5 to 20 

### : m.Ultrafine particles appear to be more likely to be taken up by lung epithelial cells. RiebeImre 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 present in tumor cells. Johnson et al. (1993) reported that 20-nm polytetrafluoroethylene particles are taken up by pulmonary epithelial cells as well as polymorphonuclear leucocytes, inducing an approximate 4-, 8-, and 40-fold increase in the release of interleukin-1 alpha and beta, inducible nitric oxide synthetase, and macrophage inflammatory protein, respectively. The carcinogenic potency of diesel particles, therefore, appears to be related, at least to some extent, to their small size and convoluted shape, which results in a large specific particle surface area. Toxicity and carcinogenicity increased with decreasing particle size into the submicron range. For example, Heinrich et al. (1995) have shown that ultrafine titanium dioxide (approximately 0.2 

### : m diameter) is much more toxic than particles with a 10-fold greater diameter of the same composition used in an earlier study by Lee et al. (1986). This increase in toxicity has been noted with even smaller particles. For example, carbon black particles 20 nm in diameter were shown to be significantly more toxic than 50 nm particles (Murphy et al., 1999). The relationship between particle size and toxicity is of concern because, as noted in Chapter 2, approximately 50%-90% of the number of particles in DE are in the size range from 5 to 50 nm. Other than disruption of the cytoskeleton of epithelial cells, there is little information regarding the means by which particle size influences carcinogenicity as well as noncancer toxicity.7-135 7.4.5. Integrative Hypothesis for Diesel-Induced Lung Cancer The induction of lung cancer in rats by large doses of carbon black via inhalation (Heinrich et al., 1995; Mauderly et al., 1991; Nikula et al., 1995) or intratracheal instillation (Kawabata et al., 1994; Pott et al., 1994; Dasenbrock et al., 1996) led to the development of the lung particle overload hypothesis. According to this hypothesis the induction of neoplasia by insoluble low- toxicity particles is associated with an inhibition of lung particle clearance and the involvement of persistent alveolar epithelial hyperplasia. Driscoll (1995), Driscoll et al. (1996), and Oberdörster and Yu (1990) outlined a proposed mechanism for the carcinogenicity of DE at high doses that emphasizes the role of phagocytic cells. Following exposure, phagocytosis of particles acts as a stimulant for oxidant production and inflammatory cytokine release by lung phagocytes. It was hypothesized that at high particle exposure concentrations the quantity of mediators released by particle-stimulated phagocytes exceeds the inflammatory defenses of the lung (e.g., antioxidants, oxidant-metabolizing enzymes, protease inhibitors, cytokine inhibitors), resulting in tissue injury and inflammation. With continued particle exposure and/or the persistence of excessive particle burdens, there then develops an environment of phagocytic activation, excessive mediator release-tissue injury and, consequently, more tissue injury, inflammation, and tissue release. This is accompanied by cell proliferation. As discussed in a review by Cohen and Ellwein (1991), conceptually, cell 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 chronic particle exposures sufficient to elicit inflammation and cell proliferation in the rat lung is an increased probability that the genetic changes necessary for neoplastic transformation will occur. A schematic of this hypothesis has been outlined by McClellan (1997) (see Figure 7-3). In support of this hypothesis, it was reported that concentrations of inhaled CB resulted in increased cytokine expression and inflammatory influx of neutrophils (Oberdörster et al., 1995), increased formation of 8-OhdG (Ichinose et al., 1997b), and increase in the yield of hprt mutants, an effect ameliorated by treatment with antioxidants (Driscoll, 1995; Driscoll et al., 1996). Metabolism of carcinogenic organics to active forms as well as the generation of reactive oxygen species from certain organic species are likely to contribute to the toxic and carcinogenic process. At low exposure concentrations, the lung particle overload condition is not present and the overload-induced inflammatory effects are not present. Note, however, as discussed in Chapters 5 and 6, that other types of inflammation are present in the rat lung at exposures below those inducing lung particle overload. However, at low exposures, activation of organic carcinogens and generation of oxidants from the organic fraction can still be expected. Actual contribution depends upon elution/bioavailability and the effectiveness of antioxidants. Direct effects of ultrafine diesel particles taken up by epithelial 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 quantitative chapter, the relevance of the rat pulmonary response is therefore problematic. The7-136 rat pulmonary noncancer responses to DPM, however, have fairly clear interspecies and human parallels. In response to poorly soluble particles such as DPM, humans and rats both develop an alveolar macrophage response, accumulate particles in the interstitium, and show mild interstitial fibrosis (ILSI, 2000). Other species (mice, hamsters) also have shown similar noncancer pulmonary responses to DPM, but without accompanying cancer response. The rat response for noncancer pulmonary histopathology, however, seems to be more pronounced compared with humans or other species, i.e., rats appear to be more sensitive. Although many critical elements of interspecies comparison, such as the role of airway geometry and patterns of particle deposition, need further elucidation, this basic interspecies similarity and the possible greater sensitivity of pulmonary response seen after longer exposures at high doses make pulmonary histopathology in rats a valid basis for noncancer dose-response assessment.

Figure 7-3. Pathegenesis of lung disease in rats with chronic, high-level exposures to particles.

Source: Modified from McClellan, 1997.7-137 7.4.6. Summary Recent studies have shown rat lung tumor rates resulting from exposures to nearly organic-free carbon black (CB) particles at high concentrations to be similar to those observedfor DE exposures, thus providing strong evidence for a particle overload mechanism for DEinduced pulmonary carcinogenesis in rats. Such a mechanism is also supported by the fact that carbon particles per se cause inflammatory responses and increased epithelial cell proliferation and that AM function may be compromised under conditions of particle overload. The particle overload hypothesis appears sufficient to account for DE-induced lung cancer in rats. However, there is also biological plausibility for lung cancer induction in humans at concentrations insufficient to induce lung particle overload as seen in rats (Chapter 3, Section 3.4 and ILSI, 2000). The uptake of particles by epithelial cells at ambient or occupational exposure levels, DNA damage resulting from oxygen-free radicals generated from organic molecules, and the gradual in situ extraction and activation of procarcinogens associated with the diesel particles may play a role in this response and provide a basis for the plausibility. The slower particle clearance rates in humans (up to a year or more) may result in greater extraction of organics. This is supported by reports of increased DNA adducts in humans occupationally exposed to DE at concentrations unlikely to induce lung particle overload. Although these modes of action can be expected to function at lung overload conditions also, they are likely to be overwhelmed by inflammatory associated effects. The evidence to date indicates that caution must be exercised in extrapolating observations made in animal models to humans when assessing the potential for DE-induced pulmonary carcinogenesis. The carcinogenic response and the formation of DNA adducts in rats exposed to DE and other particles at high exposure concentrations may be species-specific and not DPM specific. The likelihood that different modes of action predominate at high and low doses, such as lung particle overload, also renders high-dose extrapolation to lower ambient
concentrations uncertain.

7.5. WEIGHT-OF-EVIDENCE EVALUATION FOR POTENTIAL HUMAN
CARCINOGENICITY

A carcinogenicity weight-of-evidence evaluation is a synthesis of all pertinent information addressing the question of how likely an agent is to be a human carcinogen. EPA’s 1986 Guidelines for Carcinogen Risk Assessment (U.S. EPA, 1986) provide a classification system for the characterization of the overall weight of evidence for potential human carcinogenicity based on human evidence, animal evidence, and other supportive data. This system includes Group A: Human Carcinogen ; Group B: Probable Human Carcinogen ; Group
C: Possible Human Carcinogen ; Group D: Not Classifiable as to Human Carcinogenicity ; and Group E: Evidence for Noncarcinogenicity to Humans

.7-138 As part of the guidelines development and updating process, the Agency has developed revisions to the 1986 guidelines to take into account knowledge gained in recent years about the carcinogenic processes. With regard to the weight-of-evidence evaluation for potential human 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. To express the weight of evidence for potential human carcinogenicity, EPA’s proposed 1996 and 1999 guidelines utilize a hazard narrative in place of the 1986 A-E classification system. In order to provide some measure of consistency in using the 1996 and 1999 draft guidelines, standard hazard descriptors are used as part of the hazard narrative. The revised guidelines also stress the importance of considering the mode(s) of action information for making an inference about potential cancer hazard beyond the range of observation, typically encountered at levels of exposure in the general environment. “Mode of action” refers to a series of key biological events and processes that are critical to the development of cancer. This is contrasted with “mechanisms of action,” which is defined as a more detailed description of the complete sequence of biological events at the molecular level that must occur to produce a carcinogenic response. The sections to follow evaluate and weigh the individual lines of evidence and combine all evidence to make an informed judgment about the carcinogenicity hazard of DE. A conclusion in accordance with EPA’s 1986 classification system (U.S. EPA, 1986) is provided, as well as a hazard narrative along with appropriate hazard descriptors according to EPA’s Proposed Guidelines (U.S. EPA, 1996b, 1999). These sections draw on information reviewed in Chapters 2, 3, 4, and 7.

7.5.1. Human Evidence Twenty-two epidemiologic studies about the carcinogenicity of workers exposed to DE in various occupations are reviewed in Section 7.2. Exposure to DE has typically been inferred based on job classification within an industry. Increased lung cancer risk, although not always statistically significant, has been observed in 8 out of 10 cohort and 10 of 12 case-control studies within several industries, including railroad workers, truck drivers, heavy equipment operators, and professional drivers. The increased lung cancer relative risks generally range from 1.2 to 1.5, though a few studies show relative risks as high as 2.6. Statistically significant increases in pooled relative risk estimates (1.33 to 1.47) from two independent meta-analyses further support a positive relationship between DE exposure and lung cancer in a variety of DE-exposed occupations. The generally small increased lung cancer relative risk (less than 2) observed in the epidemiologic studies and meta-analyses potentially weakens the evidence of causality. When a relative risk is less than 2, if confounders (e.g., smoking, asbestos exposure) are having an effect on the observed risk increases, it could be enough to account for the increased risk. With the7-139 strongest risk factor for lung cancer being smoking, there is a concern that smoking effects may be influencing the magnitude of the observed increased relative risks. However, in studies for which the effects of smoking were accounted for, increased relative risks for lung cancer prevailed. Though some studies did not have information on smoking, significant confounding by smoking is unlikely because the comparison populations were from the same socioeconomic class. Moreover, when the meta-analysis focused only on the smoking-controlled studies, the
relative risks tended to increase. As evaluated in Section 7.2.4.5, application of the criteria for causality (including the biological plausibility) leads to the conclusion that the increased risks observed in available epidemiologic studies are consistent with a causal association between exposure to DE and occurrence of lung cancer. Overall, the human evidence for potential carcinogenicity for DE is judged to be strong, but less than sufficient for DE to be considered as a human carcinogen because of exposure uncertainties (lack of historical exposure data for workers exposed to DE) and an inability to reach a fully and direct accounting for all possible confounders.

7.5.2. Animal Evidence DE and its organic constituents, both in the gaseous and particle phase, have been extensively tested for carcinogenicity in many experimental studies using several animal species
and with different modes of administration. Several well-conducted lifetime rat inhalation studies have consistently demonstrated that chronic inhalation exposure to sufficiently high concentrations of DE produced dose-related increases in lung tumors (benign and malignant). However, the lung cancer responses in rats from high-concentration exposures appear to be mediated by impairment of lung clearance mechanisms through particle overload, resulting in persistent chronic inflammation and subsequent pathologic and neoplastic changes in the lung. Overload conditions are not expected to occur in humans as a result of environmental or most occupational exposures to DE. Thus, the rat lung tumor response is not considered relevant to an evaluation of the potential for a human environmental exposure-related hazard (Section 7.4). The chronic inhalation studies of DE in mice showed equivocal results, whereas negative findings were consistently seen in hamsters. The gaseous phase of DE (filtered exhaust without particulate fraction) was found not to be carcinogenic in rats, mice, or hamsters. In several intratracheal instillation studies, diesel particulate matter (DPM), carbon black, and the organic DPM extracts which were virtually devoid of PAHs, have been found to produce increased lung tumors in rats. When directly implanted into the rat lung, DPM condensate containing mainly four- to seven-ring PAHs induced increases in lung tumors. In several dermal studies in mice, DPM extracts have also been shown to cause skin tumors and sarcomas in mice following subcutaneous injection.7-140 Available data and hypotheses suggest that both the carbon core and the adsorbed organics have potential roles in inducing lung tumors in the rat, although their relative contribution to the carcinogenic response remains to be determined. The consistent findings of carcinogenic activity by DPM and the organic extracts of DPM in noninhalation studies (intratracheal instillation, lung implantation, skin painting) contribute to the overall evidence for a human hazard potential for DE. The lack of a tumor response from traditional animal inhalation studies in other rodent species is noted. Without understanding the mode(s) of action of DE’s carcinogenicity in humans it is difficult to assess the meaning of nonpositive results from the mouse and hamster inhalation bioassays, and the unusable results from the rat, while having other evidence of carcinogenic potential and plausibility. It should be noted that the animal studies used DE from engines available in the 1980s, and that present-day engine emissions have different characteristics (e.g., higher elemental carbon content and lesser amounts of adsorbed organics on the carbon particles), with uncertain impact on the outcome of the experimental studies. The same point can be made for the occupational epidemiologic studies.

7.5.3. Other Key Data Other key data are judged to be supportive of potential carcinogenicity of DE. As discussed in Chapter 2, DE is a complex mixture of hundreds of constituents in either gaseous phase or particle phase. Although present in small amounts, several organic compounds in the gaseous phase (e.g. PAHs, formaldehyde, acetaldehyde, benzene, 1,3-butadiene) are known toexhibit mutagenic and/or carcinogenic activities. PAHs and PAH derivatives, including nitroPAHs, present on the diesel particle are also known to be mutagenic and carcinogenic. As reviewed in Chapter 4, DPM and DPM organic extracts have been shown to induce gene mutations in a variety of bacteria and mammalian cell test systems. In addition, DE, DPM and DPM extracts have been found to cause chromosomal aberrations, aneuploidy, and sister chromatid exchange in both in vivo and in vitro tests. There is also suggestive evidence for the bioavailability of the organics from DE (Chapter 3, Section3.5). Elevated levels of DNA adducts in lymphocytes have been reported in workers exposed to DE. In addition, animal studies showed that some of the radiolabeled organic compounds are eluted from DE particles following deposition in the lungs.

7.5.4. Mode of Action As discussed in Section 7.4, the modes of action of DE-induced carcinogenicity in humans is not understood. It can be suggested that one or multiple modes of action may be involved. These may include: (a) mutagenic and genotoxic events (e.g., direct and indirect effects on DNA and effects on chromosomes) by organic compounds in the gaseous and particle7-141 phases; (b) indirect DNA damage via the production of reactive oxygen species (ROS) induced by particle-associated organics; and (c) particle-induced chronic inflammatory response leading to oxidative DNA damage through the release of cytokines, ROS, etc., and an increase in cell proliferation. The particulate phase or whole DE exposure, as measured by DPM, appears to have the greatest observable contribution to the carcinogenic effects, and both the particle core and the associated organic compounds have demonstrated carcinogenic properties, although a role for the gas-phase components cannot be ruled out. The carcinogenic activity of DE may also be related to the small size of the particles. Moreover, the relative contribution of the possible mode(s) of action may be different at different exposure levels. For example, available evidence from rat studies indicates the importance of the role of the DPM in mediating lung tumor response at high exposure levels. Thus, the role of the adsorbed organic compounds may take on increasing importance at lower exposure levels.

7.5.5. Characterization of Overall Weight of Evidence: EPA’s 1986 Guidelines for Carcinogen Risk Assessment The totality of evidence supports the conclusion that DE is a probable human carcinogen (Group B1). This conclusion is based on:

• “Limited”evidence (i.e., strong but less than sufficient evidence for “known human carcinogen”), for a causal association between DE exposure and increased risk of lung cancer among workers in different occupations; • Evidence of carcinogenicity of DPM in rats and mice by noninhalation routes of exposure (intratracheal instillation, lung implantation, skin painting, and subcutaneous injection); and

• Extensive supporting data including the demonstrated mutagenic and/or chromosomal effects of DE and its organic constituents, suggestive evidence for the bioavailability of the organics from DE, and knowledge of the known mutagenic and/or carcinogenic activity of a number of individual organic compounds present on the particles (e.g,. PAH and derivatives) and in the DE gases (e.g., benzene, 1,3-butadiene, and aldehydes).7-142 7.5.6. Weight-of-Evidence Hazard Narrative: EPA’s Proposed Guidelines for Carcinogen Risk Assessment (1996b, 1999) The combined evidence supports the conclusion that DE is likely to be carcinogenic to
humans by inhalation and that this hazard applies to environmental exposure conditions. The spectrum of evidence and the inferences drawn provide a substantial case for this hazard potential. The weight of evidence of human carcinogenicity is based on:

• Strong but less than sufficient epidemiologic evidence for a causal association between DE exposure and increased risk of lung cancer among workers in different occupations; • Evidence of carcinogenicity of DPM in rats and mice by noninhalation routes of exposure (intratracheal instillation, lung implantation, skin painting, and subcutaneous injection); and • Extensive supporting data including the demonstrated mutagenic and/or chromosomal effects of DE and its organic constituents, suggestive evidence for the local and systemic bioavailability of the organics from DE, and knowledge of the known mutagenic and/or carcinogenic activity of a number of individual organic compounds present on the particles (e.g., PAH and derivatives) and in the DE gases (e.g., benzene, 1,3 butadiene, and aldehydes).

The weight-of-evidence for the lung cancer hazard is considered strong, even though inferences and uncertainties are involved. Major uncertainties include:

• There is scientific debate about the significance of the occupational epidemiologic evidence for a causal association between occupational exposure and increased lung cancer risk. Some experts view the evidence as weak given that most of the relative risk increases are <2.0, whereas others consider the evidence as more than adequate and compelling. With relatively low relative risks (<2.0), the effects of possible confounding exposures or other factors could play a significant role in the risk increases. For example, there is specific concern about whether the effects of smoking, a known cause of lung cancer, has been adequately or fully accounted for in the key studies. In more general terms, the lack of historical exposure data to retrospectively validate estimated DE exposure levels is also a
limitation. • A lack of knowledge about the mode(s) of action of DE lung cancer in humans results in the use of a number of default risk assessment assumptions which, while justifiable by evidence or policy choice, introduce uncertainty. To date, available evidence for the role of DPM, both the adsorbed organics and the carbon core7-143 particle, has been shown only for high exposure conditions in the rat lung. The tumor inducing mode-of-action in the rat lung appears to depend on particle overloading of the lung and subsequent pathology. This sequence is judged not to be relevant for assessing the hazard to humans exposed in the ambient environment. There is virtually no information about the relative role of DE constituents in mediating the carcinogenic effects at lower experimental exposure levels, though hypotheses exist.

While a major uncertainty relates to the incomplete understanding of DE’s mode(s) of action for the induction of lung cancer in humans, available data and hypotheses suggest that DE-induced lung carcinogenicity may be mediated by mutagenic and nonmutagenic events from both the particles and the associated organic compounds, although a role for the organics in the gaseous phase cannot be ruled out. Given that there is some evidence for a mutagenic mode of action, a cancer hazard is presumed at environmental exposure levels. This is consistent with EPA’s science policy position, which assumes a nonthreshold effect for carcinogens in the absence of definitive data demonstrating a nonlinear or threshold mechanism. It should also be noted that there are not orders of magnitude differences between lower level occupational and higher end environmental exposure levels, in fact, there appears to be exposure overlap. This observation means that an extrapolation of the occupational hazard to lower environmental exposure levels is minimal, and thus, the conclusion of an environmental hazard is supported. Given these circumstances, linear low-dose extrapolation also would be an appropriate default choice in dose-response assessment that is focused on environmental levels of exposure (Chapter 8, Section 8.2). Because of insufficient information, the human carcinogenic potential of DE by oral and dermal exposures cannot be determined.

7.6. EVALUATIONS BY OTHER ORGANIZATIONS Several organizations have reviewed the relevant data and evaluated the potential human carcinogenicity of DE or its particulate component. The conclusions reached by these organizations are generally comparable to the evaluation made in this assessment using EPA’s Carcinogen Risk Assessment Guidelines. A summary of available evaluations conducted by other organizations is provided in Table 7-9.

7.7. CONCLUSION It is concluded that environmental exposure to DE may present a lung cancer hazard to humans. The particulate phase appears to have the greatest contribution to the carcinogenic effect, both the particle core and the associated organic compounds have demonstrated7-144 Table 7-9. Evaluations of DE as to human carcinogenic potential

Organization Human data Animal data Overall evaluation

NIOSH (1988) Limited Confirmatory Potential occupational carcinogen

IARC (1989) Limited Sufficient Probably carcinogenic to humans

IPCS (1996) N/A a N/A Probably carcinogenic to humans

California EPA (1998) “Consistent evidence for a causal association” “Demonstrated carcinogenicity” DPM as a “toxic air contaminant” (California Air Resources Board)

NTP (2000) “Elevated lung cancer in occupationally exposed groups” “Supporting animal and mechanistic data” DPM-Reasonably anticipated to be a carcinogen

a Not applicable.

carcinogenic properties, although a role for the DE gas-phase components cannot be ruled out. Using either EPA’s 1986 Guidelines for Carcinogen Risk Assessment (U.S. EPA, 1986) or the proposed revisions (U.S. EPA, 1996b, 1999), DE is judged to be a probable human carcinogen, or likely to be carcinogenic to humans by inhalation, respectively. The weight of evidence for potential human carcinogenicity for DE is considered strong, even though inferences are involved in the overall assessment. Major uncertainties of the hazard assessment include the following unresolved issues:

• There has been a considerable scientific debate about the significance of the available human evidence for a causal association between occupational exposure and increased lung cancer risk. Some experts view the evidence as weak given
that most of the relative risk increases are <2.0 whereas others consider the evidence as more than adequate and compelling. Additionally, there is debate about whether the effects of smoking have been adequately accounted for in key studies, as well as the lack of historical DE exposure data to retrospectively validate estimated DE exposure levels for the available studies. • A lack of knowledge about the mode(s) of action for DE lung cancer in humans results in the use of a number of default risk assessment assumptions which, while justifiable by evidence or policy choice, introduce uncertainty. To date, available evidence for the role of DPM, both the adsorbed organics and the carbon core particle, has been shown only for high-exposure conditions in the rat lung. The7-145 tumor-inducing mode of action in the rat lung appears to depend on particle overloading of the lung but this is judged to be not relevant for assessing the human hazard of ambient exposures. There is virtually no information about the relative role of DE constituents in mediating carcinogenic effects at lower experimental or environmental exposure levels. Furthermore, there is only a limited understanding regarding the relationship between DPM particle size and carcinogenicity. • DE is present in ambient PM (e.g., PM 2.5 or PM 10 ); however, a cancer hazard for ambient PM has not been identified, as of 1996 (EPA 1996b). An updated evaluation is expected in 2002.

Additional research is needed to address these issues to reduce the uncertainty associated with the potential cancer hazard of exposure to DE.

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8.1. INTRODUCTION Dose-response assessm ent for carcinogenicity defines the relationship between the exposure/dose of an agent and the degree of carcinogenic response, and evaluates potential
cancer risks to hum ans at exposure/dose levels of interest. Most often, the exposure/dose response of interest is well below the range of observation. As a result, dose-response

assessm ent usually entails an extrapolation from the generally high exposures in studies on
hum ans or laboratory anim als to the exposure levels expected from hum an contact with the agent
in the environm ent. It also includes consid erations of the scientif ic validity of these extrapolations based on available knowledge about the underlying m echanism s or m odes of carcinogenic action. The com p lete sequence of biological events that m u st occur to produce an
adverse effect is defined as “m echanism of action.” In cases where only partial inform ation is available, the term “m ode of action” is used to refer to the m echanism s for key events that are judged to be sufficient to inform about the shape of the dose-response curve beyond the range of
observation. This chapter evaluates the available exposure/dose-response data and discusses extrapolation issues in estim ating the cancer risk of environm ental exposure to diesel engine exhaust (DE). It concludes that available data are inadequate to confidently derive a cancer unit
risk estim ate for DE or its com ponent, diesel particulate m a tter (DPM). Unit risk is one possible output from a dose-response assessm ent and is defined as the estim ated upper-bound cancer risk at a specific exposure or dose from a continuous average lifetim e exposure to a carcinogen (in this case, cancer risk per 

### : g/m 3 of DPM). In lieu of unit-risk-based quantitative risk estim ates, this chapter provides a perspective about potential risk at environm ental levels. Subsequent sections of this chapter discuss issues related to dose-response evaluation of hum an cancer risk for DE exposure, including the target tum o r site and underlying m ode of action, suitable

m easures of dose, approaches to low-dose extrapolation, and appropriate data to be used in the dose-response analysis. This is followed by a sim p le analysis of the possible degree and extent
of risk from environm ental exposure to DE. Appendix C provides a sum m a ry review of dose-response assessm ents conducted to date by other organizations and investigators. These risk estim ations were perform e d on the basis of either epidem iologic and/or experim e ntal data

. As concluded in Section 8.5, EPA finds that available epidem iologic data are too uncertain to confidently derive a unit risk estim ate for DEinduced lung cancer, and that rat data are not suitable for estim ating hum an risk. Nevertheless, a review of dose-response evaluations is provided in the appendix for historical context.

8-18.2. MODE OF ACTION AND DOSE-RESPONSE APPROACH According to EPA’s 1996 Proposed Guidelines for Carcinogen Risk Assessm ent (U.S. EPA, 1996), dose-response assessm ent is perform e d in two steps: assessm ent of observed data to derive a point of departure, followed by extrapolation to lower exposures to the extent

necessary. Hum a n data are always preferred over anim al data, if available, as their use obviates the need for extrapolation across species. Mode-of-action inform ation is important to doseresponse evaluation, as it inform s about the relevance of anim al data to assessm ent of hum an hazard and risk, the shape of the dose-response curve at low doses, and the m o st appropriate

m easure(s) of exposure/dose and response.
If

there are suf f i cient quantitative data (hum ans and/or anim als) and adequate understanding of the carcinogenic process, the preferred approach is to use a biologically based

m odel for both the range of observation and extrapolation below that range. Otherwise, as a default procedure, a standard m a them atical m odel is used to curve-fit the observed doseresponse data to obtain a point of departure, which is the lower 95% confidence lim it of the lowest exposure/dose that is associated with a selected m a gnitude of excesses of cancer risk in
hum an or anim al studies. Default approaches for low-dose extrapolation should be consistent with the current understanding of the m ode(s) of action. These include approaches that assum e linearity or nonlinearity, or both. Linear extrapolation is used when there is insufficient understanding of the m odes of action, or the m ode-of-action inform ation indicates that the doseresponse curve at low dose is, or is expected to be, linear. Linear extrapolation involves the calculation of the slope of the line drawn from the point of departure to zero exposure or dose (i.e., above background). W h en there is sufficient evidence for a nonlinear m ode of action but not enough data to construct a biologically based m odel for the relationship, a m a rgin of exposure is used as a default approach. A m a rgin-of-exposure analysis com p ares the point of departure (i.e., the lowest exposure associated with som e cancer risk) with the dose associated
with the environm ental exposure(s) of interest and determ ines whether or not the exposure

m a rgins are adequate. Both default approaches m a y be used for a tum o r response if it is

m e diated by linear and nonlinear m odes of action. The dose-response approaches considered in this chapter follow the principles of EPA’s guidelines for carcinogen risk assessm ent (U.S. EPA, 1986, 1996, 1999). As reviewed in Chapter 7, there is substantial evidence from com b ined hum an and experim e ntal evidence that DE is likely to pose a cancer hazard to hum ans at anticipated levels
of environm ental exposure. The critical target organ is the lung. Evidence exists for a causal relationship between risk for lung cancer and occupational exposure to DE in certain occupational workers such as railroad workers, truck drivers, heavy equipm ent operators, transit workers, etc.

8-2The m echanism ( s) by which DE induces lung cancer in hum ans has not been established. As discussed in Chapter 7, Section 7.4, several m odes of action have been postulated on the
basis of available m echanistic studies, including direct DNA effects (gene m u tations) by the adsorbed organic com pounds and the gaseous fr actions, indirect DNA effects (e.g., chrom o som a l aberrations, sister chrom a tid exchange [SCE], m i cronuclei) by DE and DPM, oxidative DNA
dam a ge by DPM via release of reactive oxygen species (ROS), and particle-induced chronic
inflam m a tory response leading to epithelial cell cytotoxicity and regenerative cell proliferation via release of cytokines, growth factors, and RO S. It is likely that a com b ination of m odes of action contributes to the overall carcinogenic activity of DE, and that the relative contribution of
the various m odes of action m a y vary with different exposure levels. In the absence of a full understanding of the re lative roles of DE constituents in inducing lung cancer in hum ans, and because there is som e evidence for a m u tagenic m ode of action, linear low-dose extrapolation is an appropriate and prudent default choice for m odeling doseresponse, and if needed, risk extrapolation from high to lower exposures (U.S. EPA, 1986, 1996, 1999). It also should be noted that there are not order of m a gnitude differences between lower levels of occupational and higher end environm ental exposure estim ates. In fact, there appears to be exposure overlap. This m eans that an extrapolation of the occupational hazard to lower
environm ental exposure levels is m i nim a l. Other individuals and organizations have used either linear risk extrapolation m odels and/or m echanistically based m odels to estim ate cancer risk
from environm ental exposure to DE (e.g., IPCS , 1996; Cal EPA, 1998; also see Appendix C).
These were exam ined but not found to provide a com p elling basis for unit risk derivation because of database uncertainty and/or recent understandings about the suitability of the rat data
for m odeling a dose-response at environm ental levels of exposure.
For exam ple, there are an observable series of events showing how DE causes lung

tum o rs in the rat under high exposure experim e ntal conditions. Prolonged exposure to high concentrations of a variety of poorly soluble particles including DPM (and its carbon core, devoid of organics) causes lung tum o rs in rats through a m ode of action that involves im pairm e nt of lung clearance m echanism s (referred to as “lung overload response”), leading to persistent
chronic inflam m a tion, cell proliferation, m e taplasia, and ultim ately the developm ent of lung

tum o rs (ILSI, 2000). Because this m ode of action is not expected to be operative at
environm ental exposure conditions, the rat lung tum o r dose-response data are not considered suitable for predicting hum an risk at low environm ental exposure concentrations.

8.3. USE OF EPIDEMIOLOGIC STUDIES FOR QUANTITATIVE RISK ASSESSMENT As discussed above, hum an data are considered m o re appropriate than anim al data in
estim ating environm ental cancer risk f o r DE. Still, there are m a ny uncertainties in using the

8-3available epidem iologic studies that have quantitative exposure data to extrapolate the risk to the general population for am bient-level DE exposure.

8.3.1. Sources of Uncertainty The greatest uncertainty in estim ating DE-induced cancer risk from epidem iologic studies is the lack of knowledge of actual historical exposures for individual workers, particularly for the early years. Reconstructi on of historic exposures is based on job exposure categories, industrial hygiene m easurem ents, and assum p tions m a de about exposure patterns. Another related uncertainty is the choice of m a rkers of exposure to DE. As discussed above, the m odes of action for DE-induced lung cancer in hum ans are not fully understood, and
thus the best m easure of DE exposure is unknown. Various m a rkers of DPM (e.g., respirablesized particles, elem ental carbon [EC]) have been used as dosim eters for DE. Though EC is

m o re sensitive and m o re specif i c than respirable-sized particles, both are considered appropriate
dosim eters. Related to the choice of dosim eter, having a relatively constant relationship between the organics (on the particle) and the particle m a ss would be consistent with a possible m ode-ofaction role for both the particle and organic com ponents. However, evidence of such a constant historic relationship rem a ins unclear. As discussed in Chapter 2 (Section 2.5.2), it appears that

newer m odel on-road engine exhaust has a lesser quantity of organics adsorbed onto the particle

com p ared to older m odel engines. On the other hand, with regard to DE in the am bient air, there is significant variation in the am ounts of DPM organic com ponents em itted because of aged vehicles in the on-road fleet, driving patterns, and the additional presence of nonroad DE (e.g., m a rine vessels and locom o tives, which generally use older technology than on-road engines).
Another m a jor uncertainty associated with m a ny of the DE epidem iologic studies was the inability to f u lly control f o r sm oking ef f ects, resulting in possible errors in estim ating relative
risk increases. Changes in adjustm e nts for sm oking could result in considerable changes in relative risk, because sm oking has a m u ch larger effect on relative lung cancer risk than is likely
for DE. It is difficult to effectively control for a sm oking effect in a statistical analysis because cigarette sm oke contains an array of biologically active com pounds and affects m u ltiple steps of carcinogenesis, thus probably m a king sm okers m o re susceptible to DE-induced lung cancer than

are nonsm okers. A statistical analysis would not be able to adjust for such an effect without having a detailed record of the sm oking history of individuals. A potential uncertainty involves the use of occupational worker data to extrapolate cancer hazard risk to the general population and sensitive subgroups. By sex, age, and general health status, workers are not fully representative of the general population. For exam ple, there is virtually no inform ation to determ ine whether infants and children or people in poor health

8-4respond differently to DE exposure than do workers. Finally, the use of linear low-dose extrapolation m a y contribute to uncertainty in estim ating environm ental risks.

8.3.2. Evaluation of Key Epidemiologic Studies for Potential Use in Quantitative Risk
Estimates
Am ong the available epidem iologic studies, only the railroad worker studies and the
Team ster truck driver studies have reconstr ucted quantitative historical exposure data for possible use in deriving a unit risk estim ate for DE-induced lung cancer. This section evaluates the strengths and lim itations of these data and their suitability for dose-response analysis.

8.3.2.1. Railroad Worker Studies Garshick and colleagues conducted both cohort and case-control studies of lung cancer

m o rtalities am ong U.S. railroad workers registered with the U.S. Railroad Retirem ent Board (RRB). In the cohort study (Garshick et al., 1988), lung cancer m o rtality was ascertained through 1980 in 55,407 railroad workers, age 40 through 64 in 1959, with at least 10 years of work in selected railroad jobs (39 job titles). The cohort was selected on the basis of job titles in 1959. Industrial hygiene evaluations and descriptions of job activities were used to classify jobs as exposed or unexposed to diesel em issions. W o rkers with recognized asbestos exposure were
excluded from the job categories selected for st udy. However, a few jobs with som e potential for asbestos exposure were included in the cohort. Each subject’s work history was determ ined
from a yearly job report filed by his em ployer with the RRB from 1959 until death or retirem ent. The year 1959 was chosen as the effective start of DE exposure for this study because by this
tim e 95% of the locom o tives in the United States were diesel powered. The author reported statistically significant relative risk increases of 1.57 for the 40-44 year age group and 1.34 for the 45-49 year age group, after exclusion of workers exposed to asbestos and controls for

sm oking. Age groups were determ ined by their ages in 1959.
A m a in strength of the cohort study is the large sam p le size (55,407), which allowed sufficient power to detect sm all risks. This study also perm itted the exclusion of workers with potential past exposure to asbestos. The stability of job career paths in the cohort ensured that of the workers 40 to 64 years of age in 1959 classified as DE-exposed, 94% of the cases were still in DE-exposed jobs 20 years later.
The m a in lim itation of the cohort study is the lack of quantitative data on exposure to
DE. The num ber of years exposed to DE was used as a surrogate for dose. The dose, based on
duration of em ploym ent, has inaccuracies because individuals were working on both steam and
diesel locom o tives during the transition period. Itshould be noted that the investigators included

8-5only exposures after 1959; the duration of exposure prior to 1959 was not known. Other
lim itations of this study include its inability to ex am ine the effect of years of exposure prior to 1959 and the less-than-optim al latency period for lung cancer expression. No adjustm e nt for

sm oking was m a de in this study. For a detailed description of this study please refer to Chapter 7, Section 7.2.1.7. Garshick and colleagues also conducted a case-control study of railroad workers who died of lung cancer between 1981 and 1982 (Garshick et al., 1987). The author reported statistically significant increased odds ratios (with asbestos exposure accounted for) of 1.41 (95% confidence interval [CI] = 1.06, 1.88) for the 

### # 64 year age group and 1.64 (95% CI = 1.18, 2.29) for the 

### # 64 year age group with 

### $ 20 years of exposure when com p ared with the 0-4 year

exposure group. The population base for this case-control study was approxim a tely 650,000
active and retired m a le U.S. railroad workers with 10 years or m o re of railroad service who were
born in 1900 or later. The cases were selected from deaths with prim ary lung cancer, which was the underlying cause of death in m o st cases. Each case was m a tched to two deceased controls whose dates of birth were within 2.5 years of th e date of birth of the case and whose dates of death were within 31 days of the date of death noted in the case. Controls were selected
random ly from workers who did not have cancer noted anywhere on their death certificates and
who did not die of suicide or of accidental or unknown causes. A total of 1,256 cases and 2,385 controls were selected for the study. Am ong younger workers, approxim a tely 60% had exposure to DE, whereas am ong older workers, only 47% were exposed to DE. DE exposure surrogates
for workers were sim ilar to those in the cohort study. Asbestos exposure was categorized on the basis of jobs held in 1959, or on the last job held if the subject retired before 1959. Sm oking history inform ation was obtained from the next of kin. The strengths of the case-control study are consideration of confounding factors such as asbestos exposure and sm oking; classification of DE exposures by job titles and industrial hygiene sam p ling; and exploration of interacti ons between sm oking, asbestos exposure, and DE

exposure. Major lim itations of this study include (a) possible overestim ation of cigarette

consum ption by surrogate respondents; (b) use of the Interstate Com m e rce Com m i ssion (ICC) job classification as a surrogate for exposure, which m a y have led to m i sclassification of DE exposure jobs with low intensity and interm ittent exposure, such as railroad police and bus drivers, as unexposed; (c) lack of data on the contribution of unknown occupational or
environm ental exposures and passive sm oking; and (d) a suboptim al latency period of 22 years,
which m a y not be long enough to observe a full expression of lung cancer. For a detailed description of this study, please see Chapter 7, Section 7.2.2.4. As a part of these epidem iologic studies, W o skie et al. (1988a) conducted an industrial hygiene survey in the early 1990s for selected jobs in four sm all northern railroads. DE

8-6exposure was considered as a yes/no variable based on job in 1959 and estim ated years of work in a diesel- exposed job as an index of exposure. Thirty-nine job titles were originally identified and were then collapsed into 13 job categories and, for som e statistical analyses, into 5 categories (clerks, signal m a intainers, engineers/firers, brakers/conductors/hostlers, and shop workers) (W oskie et al., 1988b; Ham m ond et al., 1988). As discussed below, these exposure
estim ations were used by Crum p et al. (1991) and by Cal EPA (1998) for their dose-response analyses.

8.3.2.1.1. Potential for the data to be used for dose-response modeling. Both case-control and cohort studies can be used for dose-response analysis if exposure for each worker is available.
Control of a sm oking effect is im portant when lung cancer is the disease of interest. However, as discussed previously (see Section 8.3.1), one m a y not be able to control sm oking com p letely in a dose-response analysis because of the lack of detailed records of the sm oking history of
individuals. Garshick et al. (1988) reported a positive relationship of relative risk and duration of exposure by m odeling age in 1959 as a covariate in an exposure-response analysis. The positiverelationship disappeared when attained age was used instead of age in 1959, and a negative doseresponse was observed (Crum p et al., 1991). This negative dose-response continued to be upheld in a subsequent reanalysis (Crum p, 1999). Garshick (letter from Garshick, Harvard Medical School, to Chao Chen, U.S. EPA, dated August 15, 1991) perform e d further analysis and reported that the relationship between years of exposure and risk of lung cancer, when adjusted for attained age and calendar year, was flat to negative depending upon which m odel
was used. In contrast, California EPA (Cal EPA, 1998) found a positive dose-response by using age in 1959 but allowing for an interaction term of age and calendar year in the m odel.
Crum p et al. (1991) also found, and Garshick (letter from Garshick, Harvard Medical School, to Chao Chen, U.S. EPA, dated August 15, 1991) confirm e d, that in the years 1977-1980
the death ascertainm e nt was not com p lete. About 20% to 70% of deaths were unaccounted for, depending upon the calendar year. Further analysis, based on job titles in 1959 and lim ited to deaths occurring through 1976, showed that the youngest workers still had the highest risk of dying of lung cancer. Extensive statistical analyses were conducted by a panel convened by HEI (1999) to investigate the utility of the railroad worker cohort for use in dose-response based quantitative
risk assessm ent. Seven m odels were used to test the data, and the m odels were form ed by varying a num ber of covariates in different co m b inations. The covariates included em ploym ent duration, cum u lative exposure with and without correction for background exposure, and three job categories: clerks and signalm en, train work ers (which include engineers/firers/brakers/

8-7conductors), and shop workers. The coefficient for each covariate in a m odel is used to calculate
relative risk for the associated covariate. In sum m a ry, the panel found that effects of exposure as defined by an exposure-response curve were either flat or negative in all of the m odels. In these analyses, relative risk for each job category was assum e d to be constant with respect to age. Further exploration of the data showed that the relative risk for train workers was not constant. The panel’s statistical analyses also revealed the com p lexity of the data and dif f i culties of providing an adequate sum m a ry m easure of effect, probably because calendar year and

cum u lative exposure are highly correlated, which m a kes it especially difficult to sort out their separate effects. The difficulty of providing an adequate m easure of DE effect was further
dem onstrated in Table C.3 of the HEI report, in which negative or positive ef f ects f o r cum u lative exposure (with background exposure adjustm e nt) we re obtained depending on whether or not job category was included in the m odel. A sim ilar review of the divergent views about the railroad worker dose response also can be found in Chapter 7, Section 7.2.1.7. The diverging results about the presence or absence of exposure response for the railroad
worker data have becom e a source of continuing debate about the suitability of these data for
estim ating DE cancer risk. Although it is difficult to identify the exact reason for the diverging findings, the “age effect” appears to be a m a in source of uncertainty because age, calendar year,
and cum u lative exposure are not m u tually independent. Therefore, an ideal dose-response analysis would account for the ages when exposure to DE began and term inated, along with the attained age and other covariates for each person, using age-dependent exposure concentration
rather than cum u lative exposure over lifetim e as a dosim eter. This analysis would be possible
for the railroad workers if inform ation were available on the ages when exposure began and

term inated. Given the equivocal evidence for positive exposure response, EPA has not derived a unit
risk on the basis of the available railroad worker data. This determ ination should not be construed, however, to im ply that the railroad wo rker studies contain no useful inform ation on lung cancer risk from exposure to DE.

8.3.2.2. Teamsters Union Trucking Industry Studies Steenland et al. (1990) conducted a case-control study of lung cancer deaths in the
Central States Team sters Union to determ ine the risk of lung cancer am ong different trucking industry occupations. The study found statistically significant increased odds ratios for lung cancer of 1.89 and 1.64, depending on years of em ploym ent. Cases com p rised all deaths from lung cancer (1,288). The 1,452 controls com p rised every sixth death from the entire file, excluding deaths from lung cancer, bladder cancer , and m o tor vehicle accidents. Individuals were required to have 20 years tenure in the union to be eligible to claim benefits.

8-8Detailed inform ation on work history and potential confounders such as sm oking, diet, and asbestos exposure was obtained by questionnaire. On the basis of interview data and the 1980 census occupation and industry codes, subjects were classified either as nonexposed or as having held other jobs with potential DE exposure. The Team sters Union work history file did
not have inform ation on whether m e n drove diesel or gasoline trucks, and the four principal occupations were long-haul drivers, short-haul or city drivers, truck m echanics, and
dockworkers. Subjects were assigned the job cat egory in which they had worked the longest.
The m a in strengths of the study are the availability of detailed records from the
Team sters Union, a relatively large sam p le size, availability of sm oking data, and m easurem ent of possible asbestos exposures. Som e lim itations of this study include possible

m i sclassifications of exposure and sm oking habits , as inform ation was provided by next-of-kin and lack of sufficient latency to observe lung cancer excess. Steenland et al. (1998) conducted an exposure-response analysis by supplem enting the
data from their earlier case-control study of lung cancer and truck drivers in the Team sters Union with exposure estim ates based on a 1990 industrial hygiene survey of EC exposure (Zaebst et al., 1991), a surrogate for DE in the trucking industry. Available data indicate that exposure to workers in the trucking industry in 1990 averaged 2-27 

### : g/m 3 of EC. The 1990 exposure
inform ation was used by Steenland as a baseline exposure m easurem ent to reconstruct past exposure (in the period of 1949 to 1983) by assum i ng that the exposure for workers in different job categories is a function of highway m ileages traveled by heavy-duty vehicles and efficiency of the engine over the years. The industrial hygiene survey by Zaebst et al. (1991) of EC exposures in the trucking industry provided exposure estim ates for each job category in 1990. The EC m easurem ents were generally consistent with the epidem iologic results, in that m echanics were found to have the highest exposures and relative risk, followed by l ong-haul and short-haul drivers. Dockworkers who had the lowest exposures also had the lowest relative risks. Past exposures were estim ated assum i ng that they were a function of (a) the num ber of heavy-duty trucks on the road, (b) the particulate em issions (gram s /m ile) of diesel engines over
tim e, and (c) leaks from truck exhaust system s for long-haul drivers. Estim ates of past exposure to EC (as a m a rker for DE exposure) were m a de based on the assum p tion that average 1990 levels for a particular job category could be assigned to all subjects in that category, and that levels prior to 1990 were directly proportional to vehicle m iles traveled by heavy-duty trucks
and the estim ated em ission levels of diesel engines. For exam ple, a 1975 exposure level was
estim ated by the following equation: 1975 level = 1990 level × (vehicle m iles 1975/vehicle m iles 1990) × (em i ssions 1975/em issions 1990). Once estim ates of exposure for each year of work history were derived for each subject, analyses were conducted by cum u lative level of estim ated

8-9carbon exposure. As with m o st epidem iologic studies, the endeavors to reconstruct exposures for epidem iologic studies are subject to uncertainties.

8.3.2.2.1. Potential for the data to be used for dose-response modeling. Steenland et al. (1998) analyzed their case-control data and showed a significant positive trend in lung cancer risk with increasing cum u lative exposure to DE. The study by Steenland et al. (1998) could provide a valuable database for calculating unit risk for DE em issions. The strength of this data set is that
the sm oking histories of workers were obtained to the extent possible. Sm oking is especially
im portant in assessing the lung cancer risk due to DE exposure because sm oking has m u ch higher relative risk (or odds ratio) of lung cancer than does DE. In the Steenland et al. (1998) study, the overall (ever-sm okers vs. nonsm okers) odds ratio for developing lung cancer from

sm oking is about 7.2, which is about fivefold larger than the 1.4 relative risk increase from a large synthesis of m a ny DE epidem iologic studies. It is possible that a m odest change of
inform ation on sm oking and diesel exposure m i ght alter the conclusion and risk estim ate. Another strength of the Team ster data for use in environm ental risk assessm ent for the general population is that exposures of Team sters are closer to am bient exposures than are those
of railroad workers. The Team sters Union truck driver case-control workers had cum u lative exposure ranging from 19 to 2,440 

### : g/m 3 -years of EC, with the m e dian and 95 th percentile, respectively, of 358 and 754 

### : g/m 3 -years of EC. The m e dian and 95 th percentile of an environm entally equivalent exposure would be 3 and 6 

### : g/m 3 , respectively. 1 These
environm ental equivalent exposures for the Team sters Union truck drivers are close to the
estim ated am bient exposures of <1.0 

### : g/m 3 to 4.0 

### : g/m 3 (see Table 2-31). Steenland et al. (1998) stated that their risk assessm ent is exploratory because it depends
on estim ates about unknown past exposures. An EPA reanalysis of DE exposure for this study is underway. W ith a revised exposure assessm ent, a reevaluation of the dose-response would be appropriate. In a recent review, HEI (1999) concluded that the Team sters studies m a y be useful
f o r quantitative risk assessm ent, but signif i cant f u rther evaluation and developm ent are needed. Given the ongoing reanalysis of exposure, EPA will not, at this tim e, use the Steenland et al. (1998) occupational risk assessm ent findings to derive equivalent environm ental param e ters and
cancer unit risk estim ates.

1 The conversion assum e s (a) DPM = 40% EC as reported by Steenland et al. (1998), (b) environm ental equivalent exposure is approxim a tely = 0.21 × occupational exposure, and (c) 70 : g/m 3 -y ears is equivalent to a lifetim e of exposure at 1 : g/m 3

.
8-108.3.3. Conclusion Even though available evidence supports a conclusion that DE is likely to be a hum an lung carcinogen, the conclusion of the dose-response evaluation is that the available data are not sufficient to confidently estim ate a cancer unit risk or unit risk range. The absence of such a
cancer unit risk f o r DE lim its the ability to quantif y, with conf idence, the potential im pact of the hazard on exposed populations. Two significant short-term activities are underway to im prove the epidem iologic database for dose-response assessm ent: (1) a follow-up study to correct the undercounting of m o rtality in the Garshick et al. (1988) railroad worker study, and (2) an EPAsponsored effort to im prove the exposure estim ates for Team sters Union truck drivers (Steenland et al., 1998). EPA will m onitor this ongoing research as well as the ongoing NCI-NIOSH study
of nonm etal m i ners and the recently NCI-funded epidem iologic study of truck drivers. As these
studies or other new data becom e publicly available, EPA will reconsider the m e rit of conducting additional dose-response analysis and unit risk derivation.

8.4. PERSPECTIVES ON CANCER RISK Although the available data are considered inadequate to confidently estim ate a cancer unit risk, this does not m ean that there is no inform ation about the possible cancer risk of DE.
To exam ine the significance of the potential cancer hazard from environm ental exposure to DE, all relevant epidem iologic and exposure data as well as sim p le risk assessm ent tools can be used. Such an approach does not produce confident estim ates of cancer unit risk. Rather, these exploratory approaches provide a perspective on the possible m a gnitude of cancer risk and thus insight about the potential significance of the hazard. This section describes approaches and

m e thods that are used to gauge the m a gnitude of possible cancer risk from am bient exposure to
DE. The first approach involves exam ining the differences between the levels of occupational
and am bient environm ental exposures, while assum i ng that cancer risk to DE is linearly proportional with cum u lative lifetim e exposure. Risks to the general public would be low in

com p arison to occupational risk if the differences in exposure are large (e.g., about three orders
of m a gnitude or m o re). On the other hand, if the exposure differences are sm aller (i.e., within
one to two orders of m a gnitude), environm ental risks becom e m o re of a concern as they approach the range of workers’ risk observed in epidem iologic studies of past occupational

exposures. This assum e s that the carcinogenic pot ency of historical and current-day DE is not significantly different, a reasonable a ssum p tion, though not without uncertainty. Table 8-1 shows occupational exposure estim ates for som e of the occupational groups where increased relative risks of lung cancer (e.g., m e ta-analyses) have been analyzed. Given that no statistical properties associated with these exposure estim ates are known, their use here is

8-11------------------------

---------------------

---------------------------------

-----------------

----------------- Table 8-1. DPM exposure margins (ratio of occupational ÷ environmental exposures )

Occupational group E stimated occupational exposure/concentration

Environmental equivalent a Exposure margin ratio - average environmental exposure for 0.8 : g/m 3 of environmental exposure b Exposure margin ratio - high-end environmental exposure for 4.0 : g/m 3 of environmental exposure b Reference c

Public transit workers 15-98 : g/m 3

3-21 : g/m 3 4-26 0.8-5 Birch and Cary, 1996

U.S. railroad workers 39-191 : g/m 3

8- 40 : g/m 3 10-50 2-10 W o skie et al., 1988b

Fork Lift Operators 7 -403 : g/m 3

1- 85 : g/m 3 2-106 0.37-21 Groves and Cain, 2000 d

High end boundary estim ate 1200 : g/m 3

252 : g/m 3 315 63 see text discussion in Section 8.4

a E quivalent envir onm ental exposur e = occ upational exposur e × 0. 21, see Chapter 2, Section 2. 4. 3. 1, som e values ar e r ounded.

b 0. 8 : g/m 3 = aver age 1990 nationwide on- r o ad exposur e estim ate fr om HAPE M m odel; the com p anion r u r a l estim ate is 0. 5 : g/m 3

, and 4 : g/m 3 is a high- end estim ate. T h e 1996 nationwide aver age is 0. 7 : g/m 3 , the com p anion rural estim ate is 0.2 : g/m 3 ; however
, a high- end estim ate is not available for 1996. See Chapter 2, Sections 2. 4. 3. 2. 1 and 2. 4. 3. 2. 2.

c See T a ble 2- 27 for m o r e details about Bir c h and Car y , W o skie.

d 403 : g/m 3 is a 99 percentile estim ate of EC/ : g/m 3

, the DPM equivalent of the E C m easur em ent can be estim ated as DPM = E C x 0. 62 to 1. 31.

not intended to be precise or to m a tch with specific epidem iologic data, but rather to provide a broad range of possible exposures in the workplace. The purpose is to identify a high- and lowend occupational exposure consistent with the occupational groups of interest and then to

com p are these to estim ates of environm ental exposure. Given the special interest in discerning
the lower risk m a gnitude, especially to see if the lower risk m i ght be above or below one in 1

m illion, a high-end exposure estim ate would be used, and as discussed later, the occupational exposure can be arbitrarily increased (e.g., toward an extrem e value) to ensure that a low end of risk is identified, consistent with the reported occupational risk increases. Environm ental exposure data from on-road vehicle em issions are based on the 1990 nationwide exposure
estim ates from the HAPEM m odel (see Chapter 2, Section 2.4.3.2.1). Both average (0.8 

### : g/m 3 ) and high-end exposures (4 

### : g/m 3 ) are used.
In order to com p are differences between occupational and environm ental exposures, it is necessary to convert occupational exposure to continuous exposure (i.e., environm ental equivalent exposure = 0.21 × occupational exposure, see Section 2.4.3.1). Accordingly, Table 8- 1 shows equivalent environm ental levels and the ratios of occupational to environm ental

8-12exposures, referred to as exposure m a rgins (EMs). An EM of 1 or less indicates that
environm ental exposure is com p arable to occupational exposure. An EM >1 m eans that the occupational equivalent exposure is greater than the benchm arked environm ental exposure. Table 8-1 shows that the EMs based on the average nationwide environm ental exposure (0.8 

### : g/m 3 ) approach three orders of m a gnitude. EM’s that range from 1 to 10 also can be viewed as showing that adjusted occupational exposures are relatively close to the am bient
environm ental levels that were chosen as benchm arks. This closeness sets the stage for less uncertainty in hazard and risk extrapolation from the occupational to environm ental setting. It also raises a concern that risks to the general public could approach worker risks that were observed in the occupational epidem iologic studies. Table 8-1 is based upon DE exposure
estim ates from on-road sources only. W ith the addition of exposure from nonroad sources, the average nationwide-based EM ratios would be lowe r

. For exam ple, using 1996 exposure data for urban populations (Table 2-30), the exposure from on-road sources is 0.5 

### : g/m 3 , whereas
nonroad sources contribute 0.9 

### : g/m 3 , for a total of 1.4 

### : g/m 3

. Using this exposure value in place of the EM calculation of Table 8-1 (1990 estim ate of 0.8) produces a nearly 43% reduction
in the EM ratio. A com p arison of EM changes for the high-end on-road plus nonroad exposure is not possible at the present tim e because the 1996 data have not yet been m odeled to obtain a high-end value sim ilar to the 1990 value of 4.0 

### : g/m 3

. A second approach to explore the possible cancer risk to the general population from
environm ental exposure to DE is m o re quantitative. One begins by exam ining the risk observed in DE-exposed workers and then m a king reasoned assum p tions as to how these risks can be
translated to environm ental exposure conditions. Such an approach involves three m a jor

assum p tions: (1) the excess lung cancer risk as shown in num erous epidem iologic studies and in

two m e ta-analyses is indeed due to DE exposure, (2) the increased lung cancer risk over background is linearly proportional to the lifetim e exposure to DE, and (3) the past DE exposure
for workers has the sam e cancer-inducing potential as the current DE in am bient air. Any of
these assum p tions could have an im pact on the possible environm ental risk by either increasing or decreasing the risk estim ates, including the possibility of a lower or zero risk at environm ental
levels. As reviewed in Chapter 7, Section 7.2, num erous epidem iologic studies have shown increased lung cancer risks (i.e., som e are deaths, som e are cases) am ong workers in certain occupations. The relative risks or odds ratios range from 1.2 to 2.6. Two independent m e taanalyses show sm oking-adjusted relative risk increases of 1.35 (Bhatia et al., 1997) and 1.47 (Lipsett and Cam p lem a n, 1999). For this analysis, a relative risk of 1.4 is selected as a
reasonable estim ate. This risk m eans that the workers faced an extra risk 40% higher than the

8-135% background lifetim e lung cancer risk in the U.S. population. 2 Thus, using the relationship [excess risk = (relative risk-1) × 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 - 1) × (0.05) = 0.02]. The validity of this interpretation depends on an im portant assum p tion: that
the observed increm ental risk of 40% was due to DE exposure alone and not to other unknown
extraneous factors. It should be noted, however , that the conclusion about the risk perspective would not be changed even if the increm ental risk of 40% were greatly reduced (e.g., to 20%); the conclusion would be changed only if alm o st all of the increm ental risk were due to nondiesel
factors. Next, one would consider the EM (i.e., the EM ratio) between the occupational exposures and general-population environm ental exposures. DPM concentrations in the workplace, used as a surrogate for worker exposure, are shown for three occupational worker groups in Table 8-1. These range from 7-403 

### : g/m 3 (with an equivalent continuous exposure of 1-85 

### : g/m 3 ). These worker groups are consistent with m a ny of those cited in the m e ta-analyses. For this exploratory risk estim ation approach, we wa nt to intentionally adopt a high-end boundary exposure estim ate that is unlikely to be exceeded, so that a lower bounding of the risk would be
identified. An occupational boundary exposure of 1,200 

### : g/m 3 with its environm ental equivalent estim ated value of 252 

### : g/m 3 has been purposefully adopted to represent a high-end boundary estim ate. This happens to be about three tim es the forklift operator value shown in Table 8-1, and clearly a high-end estim ate when Table 2-27 is exam ined, exclusive of the
estim ates for m i ners which are not included in the m e ta-analyses. It should be noted once again
that none of these estim ates are intended to be precise. Table 8-1 shows that the DPM exposure m a rgin ratio between occupational and
environm ental exposure, using the nationwide average exposure value of 0.8 

### : g/m 3 , m a y range from 2 to 315 when the boundary estim ate is used, and range from 0.37-63 when 4.0 

### : g/m 3 is used as a high-end environm ental benchm ark exposure. Som e of these extrem e values will be used, as discussed in the next paragraph.
Risks from environm ental exposure depend on the shape of the dose-response curve in the range between occupational and environm ental exposures. If lifetim e risks in this range were

2 The background rate of 0.05 is an approxim a ted lifetim e risk calculated by the m e thod of lifetable analysis using age-specific lung cancer m o rtality data and probab ility of death in the age group taken from the National Health Statistics (HRS) m onographs of Vital Statistics of the U.S. (Vol. 2, Part A, 1992). Sim ilar values based on two rather crude approaches also can be obtained: (1) 59.8 × 10 -5 / 8.8 × 10 -3 = 6.8 × 10 -2 where 59.8 × 10 -5 and 8.8 × 10 -3 are, respectively, the crude estim ates of lung cancer d eaths (including intrathoracic organs, estim ated to be less than 105 of the total cases) and total deaths for 1996 reporte d in Statistical Abstract of the U.S. (Bureau of the Census, 1998, 118 th 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 Statis tics J of the Am erican Cancer Society, Jan/Feb 2000), 270,000,000 is the current U.S. population, and 76 is the expected lifespan.

8-14to fall proportionately with reduced exposure, and if one assum e s that past occupational exposures were at the high-boundary end, then the risk from average environm ental exposure
could be between 10 -5 and 10 -4 (0.02 ÷ 315 = 6 × 10 -5 ). On the other hand, if occupational exposures for different groups were lower, risk s from environm ental exposure would be higher.
For exam ple, if occupational concentrations or exposures were closer to 100 

### : g/m 3 , a value that is represented in several data sets shown in Table 8-1 (with an equivalent environm ental exposure of 21 

### : g/m 3 and a corresponding EM of 26), then risks from environm ental exposure would approach 10 -3 (0.02 ÷ 26 = 8 × 10 -4 ). If lif etim e risks were to f a ll m o re than proportionately, then risks would be lower. The latter two sources of dose-response uncertainty (i.e., the actual occupational exposures and the shape of the dose-response curve at low exposures) cannot be further defined with currently available inform ation. Use of higher
environm ental exposures (>0.8 up to 4.0 

### : g/m 3 ) lowers the EM value and increases the
estim ated risk.
The m a gnitude of the estim ated lifetim e cancer risk derived from using a very high-end occupational-to-environm ental exposure difference, establishes a reasonable basis to believe that the general population could face possible risks high enough to be of concern. This does not directly address the segm ents of the population that m a y be at highest risk: those who are additionally exposed to nonroad sources of DE and children who m a y be m o re sensitive to earlylife DE exposure, if not in fact, a longe r period of potential lifetim e exposure. The analyses presented above are not intended to be precise but are useful in gauging the possible risk range using sim p le risk explorati on m e thods. Best scientific judgm ent guided the
selection of assum p tions and other elem ents of this analysis which are deem ed reasonable and appropriate for identifying possible risks based on the inform ation currently available. These analyses provide a sense of where an upper lim it (or “upper bound”) of the cancer risk m a y be.
The sim p le m e thodologies used are generic and are valid for any increased relative risk data, however, they are not unique to the DE data. These analyses are subject to num erous uncertainties, particularly the lack of actual exposure inform ation in the epidem iologic studies
and uncertainties related to the three m a jor underlying assum p tions m e ntioned earlier. Any of
these uncertainties could have an im pact on the possible risk levels discussed above. Lower risks are possible and one cannot rule out zero risk. The risks could be zero because (a) som e individuals within the population m a y have a high tolerance to exposure from DE and therefore not be susceptible to the cancer risk from environm ental exposure, and (b) although evidence of this has not been seen, there could be a threshold of exposure below which there is no cancer
risk.
The estim ated possible risk ranges (10 -5 to 10 -3 as well as lower and zero risk) provide a perspective of the potential significance of the lung cancer hazard. This perspective should not

8-15be viewed as a def i nitive quantitative characterization of risk. The developm ent of risk estim ates does not constitute endorsem e nt of their validity as surrogates for cancer unit risk or their suitability for estim ating num bers of cancer cases. Further research is needed to m o re accurately
assess and characterize environm ental cancer risks of DE.

8.5. SUMMARY AND DISCUSSION
There are num erous published quantitative dose-response assessm ents to estim ate hum an cancer risk from exposure to DE that use epidem iologic and/or experim e ntal anim al data (see Appendix C). These dose-response assessm ents were considered but failed to present a fully
sufficient basis for a confident derivation of a cancer unit risk. This is because of epidem iologic exposure-related database uncertainties and the recent understanding about the lack of relevance of the rat lung cancer response to environm entally exposed hum ans. The dose-response analysis in this chapter has f o cused on the f easibility of using the occupational epidem iologic data to develop dose-response relationships and extrapolating them to the presum ably lower levels of
environm ental exposure. Typically, this would result in the derivation of an exposure/dosespecific cancer unit risk. In the absence of an understanding about the m ode(s) of action for DEinduced lung cancer in hum ans, coupled with the consideration that DE contains m a ny

m

u tagenic and carcinogenic constituents, linear low-dose extrapolation is judged to be an appropriate default choice for dose-response analysis, should there be satisfactory data to perform such an analysis. This chapter specif i cally evaluated the suitability of using the railroad worker studies (Garshick et al., 1987, 1988) and the Team ster Uni on truck driver studies (Steenland et al., 1990, 1998) for dose-response analysis. However, because of uncertainties about the exposureresponse for the railroad workers and exposure uncertainties for the truck drivers, a doseresponse-based cancer unit risk estim ate for DE is not developed from these data sets at this
tim e. In the absence of a unit risk to assess environm ental cancer risk, som e insight about the possible significance of the hazard can be drawn from the available epidem iologic data and exploratory risk evaluation techniques. A discussi on of possible risk is presented in the form of a perspective on the possible m a gnitude of risk from environm ental exposure. The perspective
discussion notes the sm all exposure m a rgins and possible overlap between som e occupational
and environm ental exposure levels. This lessens the uncertainty of extrapolating the occupational hazard and observed risk into the environm ental setting. Furtherm ore, based on a

m o re quantitative approach involving the observed lung cancer from occupational exposures and
the m a gnitude of occupational and environm en tal exposure differences, an exploratory risk analysis shows that environm ental cancer risks possibly range from 10 -5 to nearly 10 -3 , while a

8-16consideration of the num erous uncertainties and assum p tions also indicates that lower risk is possible and zero risk cannot be ruled out. These risk findings are only general indicators of the potential significance of the lung cancer hazard and should not be viewed as a definitive quantitative characterization of risk or be used to estim ate an exposure-specific population
im pact, i.e., estim ating num bers of cancer deaths. Best scientific judgm ent guided the selection
of assum p tions and other elem ents of the analysis which are deem ed reasonable and appropriate for identifying possible risks based on the inform ation currently available. Further research is
needed to m o re accurately assess and characterize environm ental cancer risks of DE. This exploratory risk analysis uses data collected f r om engines built prior to the m i d1990s. W h ile engine exhaust em issions have been decreasing and exhaust com position has been changing in recent years, particularly for on-road engines, EPA believes that the insight gained
from the risk perspective is pertinent to current on-road and nonroad engine use. New and cleaner engines will becom e available in response to environm ental concerns and strict new regulations. As cleaner engines replace a substantial num ber of existing engines, the risk perspective will need to be reevaluated.

REFERENCES FOR CHAPTER 8

Bhatia, R; Lopipero, P; Sm ith, A. (1997) Diesel e xhaust exposure and lung cancer. Epidem iology 9(1):84-91.

Birch, ME; Cary, RA. (1996) Elem ental carbon-based m e thod for m onitoring occupational exposures to particulate diesel exhaust. Aerosol Sci Technol 25:221-241.

California Environm ental Protection Agency (Cal EPA). (1998) Health risk assessm ent for DE. Public and Scientific Review Draft.

Crum p, KS. (1999) Lung cancer m o rtality and DE: reanalysis of a retrospective cohort study of U.S. railroad workers. Inhal Toxicol 11:1-17.

Crum p, KS; Lam b ert, T; Chen, C. (1991) Assessm ent of risk from exposure to diesel engine em issions. Clem ent International Corporation. Prepared for U.S. EPA under contract no. 68-02-4601.

Garshick, E; Schenker, MB; Munoz, A; et al. (1987) A case-control study of lung cancer and DE exposure in railroad workers. Am Rev Respir Dis 135:1242-1248.

Garshick, E; Schenker, MB; Munoz, A; et al. (1988) A re trospective cohort study of lung cancer and DE exposure in railroad workers. Am Rev Respir Dis 137:820-825.

Ham m ond, SK; Sm ith, TJ; W o skie, SR; et al. (1988) Markers of exposure to diesel exhaust and cigarette sm oke in railroad workers. Am Ind Hyg Assoc J 49:516-522.

Health Effects Institute (HEI). (1999) Diesel em i ssions and lung cancer: epidem iology and quantitative risk

assessm ent. A special report of the Institute’s Diesel Epidem iology Expert Panel. Cam b ridge, MA: Health Effects Institute.

8-17International Life Sciences Institute (ILSI). (2000) ILSI Risk Science Institu te workshop: The relevance of the rat lung response to particle overload for hum an risk a ssessm ent. Gardner, DE, ed

. Inhal Toxicol:12(1-2).

International Program m e on Chem ical Safety: W o rld Health Organization (IPCS). (1996) Diesel fuel and exhaust

em issions. Environm ental Health Criteria 171. Geneva: W o rld Health Organization.

Lipsett, M; Cam p lem a n, S. (1999) Occupational exposure to DE and lung cancer: a m e ta-analysis. Am J Publ Health 89(7):1009-1017.

Steenland, NK; Silverm a n, DT; Hornung, RW . (1990) Case-control study of lung cancer and truck driving in the Team sters Union. Am J Publ Health 80:670-674.

Steenland, K; Deddens, J; Stayner, L. (1998) DE and lung cancer in the trucking industry: exposure-response analysis and risk assessm ent. Am J Ind Med 34:220-228.

U.S. Environm ental Protection Agency (EPA). (1986) Guid elines for carcinogen risk assessm ent. Federal Register 51(185):33992-34003.

U.S. EPA. (1996) Proposed guidelines for carcinogen risk assessm ent. Federal Register 61(79):17960-18011.

U.S. EPA. (1999) Guidelines for Carcinogen Risk Assessm ent. Review Draft. NCEA-F -0644, July. Risk Assessm ent Forum
, W a shington, DC. http://www.epa.gov/ncea/raf/cancer.htm

.

W o skie, SR; Sm ith, TJ; Ham m ond, SK; et al. (1988a) Estim ation of the DE exposures of railroad workers: II. National and historical exposures. Am J Ind Med 13:395-404.

W o skie, SR; Sm ith, TJ; Ham m ond, SK; et al. (1988b) Estim ati on of the DE exposures of railroad workers: I. Current exposures. Am J Ind Med 13:381-394.

Zaebst, D; Clapp, D; Blade, L; et al. (1991) Quantitative determ ination of trucking industry workers’ exposures to diesel particles. Am Ind Hyg Assoc J 52:529-541.

8-189. CHARACTERIZ ATION OF POTENTIAL HUMAN HEALTH EFFECTS OF DIESEL EXHAUST: HAZ ARD AND DOSE-RESPONSE ASSESSMENTS

9.1. INTRODUCTION
Hum a n health risk assessm ent entails the evaluation of all pertinent inf o rm ation on the
hazardous nature of environm ental agents, on the extent of hum an exposure to them , and on the characterization of the potential risk to an exposed population. The inform ation is typically organized into four com ponents: hazard assessm ent, dose-response assessm ent, exposure

assessm ent, and risk characterization. This health assessm ent docum ent focuses only on the hazard and dose-response assessm ent com ponents. The overall objectives of this diesel engine exhaust (DE) assessm ent are:

• t o identify and characterize the hum an health effects, i.e., hazards that m a y result
from environm ental exposure to DE;

• t o determ ine whether there is a quantitative exposure/dose-response relationship for DE exposure and the health effect in the range of observation, and if sufficient data are available (1) for noncancer effects to derive estim ates of exposure that are believed to be without appreciable risk, and (2) for carcinogenicity to derive an exposure/dose-specific cancer unit risk; and

• t o sum m a rize and integrate the findings of the assessm ent into a characterization and
discuss the uncertainties.

This chapter sum m a rizes and integrates the key findings about the nature and
characteristics of environm ental exposure to DE (Chapter 2), health hazard inform ation (Chapters 3, 4, 5, and 7), and dose-response analyses (Chapters 6 and 8) that are relevant to the potential hum an health effects associated with current-day environm ental exposure to DE. It also discusses the uncertainties associated with the key f i ndings, including critical data and knowledge gaps, key assum p tions, and EPA’s science policy choices that are used to bridge the data and knowledge gaps.
This assessm ent is the Agency’s first health assessm ent for DE em issions and was developed to provide inform ation about the potential for DE-related environm ental health hazards that could be used in evaluating regulatory initiatives under provisions of the Clean Air
Act.

9-19.2. PHYSICAL AND CHEMICAL CO MPOSITION OF DIESEL EXHAUST As reviewed in Chapter 2, DE is a com p lex m i xture of hundreds of constituents in gas or particle phases. Gaseous com ponents of DE include carbon dioxide, oxygen, nitrogen, water vapor, carbon m onoxide, nitrogen com pounds, sulfur com pounds, and low m o lecular-weight hydrocarbons and their derivatives. The particulate m a tter of DE, diesel particulate m a tter (DPM), is com posed of elem ental carbon (EC), adsorbed organic com pounds, and sm all am ounts of sulfate, nitrate, m e tals, trace elem ents, water, and unidentified com pounds. Incom p lete

com bustion of fuel hydrocarbons as well as engine oil and other fuel com ponents such as sulfur
leads to the f o rm ation of DPM. DPM is either directly em itted f r om diesel-powered engines (prim a ry particulate m a tter) or is form ed from the gaseous com pounds em itted by a diesel engine (secondary particulate m a tter).
DE em issions vary in chem ical com position and particle sizes am ong different engine types, fuel form ulations, and within engine types according to operating conditions. As the

em issions age in the environm ent they also change. There also have been changes in DE

em issions over tim e as a result of changes in engine technology and fuel reform ulation. The following sections identify and characterize the key com ponents of DE that are of concern in possible health outcom e s, and discuss the changes in the com position of DE over tim e. The
latter inform ation is critical for m a king a scientific judgm ent about the appropriateness of using epidem iologic and toxicologic findings from past DE exposures to assess hazard and risk from current-day environm ental exposures. It should be noted that available anim al studies are based on exhaust exposures from various m odel year on-road diesel engines since 1980, whereas m a ny of the epidem iologic studies refer to exposures from on-road and nonroad diesel engines in use
from the 1950s through the m i d-1990s.
After em ission from the tailpipe, DE undergoes dilution, chem ical and physical
transform a tions, and dispersion and transport into the atm o sphere. After a day or so in the

am bient environm ent, the exhaust m i xture is said to be aging, a recognition of the atm o spheric
transform a tion processes, m o stly focused on the organics present, though som e particle size changes also m a y occur. The public health im pact of DE m i xture transform a tions is not clear, as

som e atm o spheric processes alter chem ical form s to a less toxic form whereas others seem to produce a chem ical form with increased toxicity (Chapter 2, Section 2.3).

9.2.1. Diesel Exhaust Components of Possible Health Concern
The com ponents of DE that are of health concern for this assessm ent are the particles (elem e ntal carbon core), the organic com pounds adsorbed to the particles, and the organic

com pounds present in the gas phase. The gaseous oxides of carbon, nitrogen, and sulfur are also

9-2of public health interest and the relevant health considerations for these are reviewed separately
in EPA’s Am bient Air Quality Criteria Docum e nts.

9.2.1.1. Diesel Particles Approxim a tely 80%-95% of DPM m a ss is in the fine particle size range ( 

### # 2.5

m i crom eters, am bient particulate m a tter [PM]), with a m ean particle aerodynam i c diam eter of
about 0.2 m i crom eters. Ultrafine particles (<0.1 m i crom eters), a sm aller size com ponent of the fine particles, average about 0.02 m i crom eters in aerodynam i c diam eter and account for about
1%-20% of the DPM m a ss and 50%-90% of the to tal num ber of particles in DPM (Chapter 2, Section 2.2.8.3).
Particle size is im portant for a num ber of reasons. Particles with aerodynam i c diam eters
>2.5 m i crom eters (i.e., >PM 2.5 ) tend to be retained in the upper portions of the respiratory tract, whereas particles with diam eters <2.5 m i crom eters (i.e., PM 2.5 ) are deposited in all areas, especially into the lower portions of the respiratory tract, including the deep lung. These fine and ultrafine particles have a very large surface area per gram m a ss (Chapter 2, Section 2.2.2),
which enables them to adsorb and transport inorganic and organic com pounds into the lung (Chapter 3, Section 3.3). DPM is part of am bient particulate m a tter (PM). The EPA Em issions Trends Report (U.S. EPA, 2000) indicates that annual nationwide em issions of diesel PM 2.5 (on-road and nonroad) in 1998 were 77% of all m obile-source em issions in 1998, 23% of the total PM 2.5 inventory excluding natural and m i scellaneous sources, and 6% if the natural and m i scellaneous
sources are included. Som e geographic areas have a higher percentage of DPM in am bient PM 2.5 because of differences in the number and types of diesel engines present in the area (e.g., onroad engines as well as nonroad engines). For instance, in Manhattan, New York, on-road diesel PM was reported to contribute about 53% of am bient PM 10 during 3 days in 1993, whereas 1996-
1997 studies in the Phoenix and Denver areas showed diesel PM to be 10%-15% of total PM 2.5

m a ss (Chapter 2, Section 2.4.2.1). DPM generally contains a high percentage of EC per unit m a ss, which can be used as a distinguishing feature from noncom bustion sources of PM 2.5 and, to an extent, other com bustion

sources. The DPM EC content can range from m o re than 50% to approxim a tely 75% of the
DPM m a ss depending on age of engine, type of engine (heavy-duty versus light-duty), fuel characteristics, and driving conditions. The organic carbon portion of DPM can range approxim a tely from 19% to 43%, though higher and lower values also have been reported. In

com p arison, gasoline engine exhaust generally has a reverse pattern of low EC content and a high percentage of organics on the particle m a ss (see Chapter 2, Table 2-13).

9-39.2.1.2. Organic Compounds The organic com pounds present in the gases and adsorbed onto the particles include a wide spectrum of com pounds related to unburned diesel fuel, lube oil, low levels of partial

com bustion, and pyrolysis products (see Chapter 2, Table 2-19). The organic com pounds present in the gaseous phase include alkanes, alkenes, aldehydes, m onocyclic arom atic

com pounds, and polycyclic arom atic hydrocarbons (PAHs). Am ong the gaseous com ponents of DE, the aldehydes are particularly im portant because of their potential carcinogenic effects and because they m a ke up an im portant fraction of the gaseous em issions. Form aldehyde accounts
for a m a jority of the aldehyde em issions (65%- 80%) from diesel engines. Acetaldehyde and
acrolein are the next m o st abundant aldehydes. Other gaseous com ponents of DE that are notable for their carcinogenic effects include benzene, 1,3-butadiene, PAHs, and nitro-PAHs (including those with 

### # 4 rings and nitro-PAHs with 2 and 3 rings). A num ber of the gaseous

com pounds (e.g., aldehydes, alkanes, alkenes, NO

x , SO
x ) also are known to induce respiratory tract irritation given sufficient exposure (see Chapter 2, Table 2-21). Very sm all am ounts of
dioxins have been m easured in heavy-duty diesel truck exhaust. These em issions are estim ated to represent about 1.2% of the 1995 national di oxin inventory; dioxin em issions from nonroad
exhausts have not been estim ated (Chapter 2, Section 2.2.7.2). Organic substances adsorbed onto DPM include C 14-35 hydrocarbon com pounds, PAHs
with 

### $ 4 rings, and nitro-PAHs. PAHs and their derivatives com p rise <1% of the DPM m a ss (Chapter 2, Section 2.2.8). Many of these hydrocarbons are known to have m u tagenic and carcinogenic properties. California EPA (Cal EPA, 1998) identified at least 19 hydrocarbons present in DE that are known or suspected carcinogens, according to evaluations by the International Agency for Research on Cancer (IARC).

9.2.2. “Fresh” Versus “Aged” Diesel Exhaust Newly em itted exhaust is term ed “f resh,” whereas exhaust that is m o re than 1 or 2 days old is referred to as “aged” because of alterations caused by sunlight and other chem ical physical
reactions that occur in the atm o sphere. The overall toxicological consequence of DE aging is unclear because during aging som e com pounds in the DE m i xture are altered to m o re toxic form s while others are m a de less toxic. For exam ple, PAHs present in fresh em issions m a y be nitrated by atm o spheric NO 3 to form nitro-PAHs, thus adding to the existing burden of toxic nitro-PAHs present in fresh exhaust. On the other hand, PAHs present in the gas phase can react with hydroxyl radicals present in the am bient air, leading to a reduced atm o spheric lifetim e of the original PAHs. Alkanes and alkenes m a y be c onverted to aldehydes, and oxides of nitrogen to nitric acid (Chapter 2, Section 2.3).

9-49.2.3. Changes in Diesel Exhaust Emissions and Composition Over Time Chapter 2, with its Sum m a ry in Section 2.5, provides a full review of em issions trends
and a com p lete characterization of the physical and chem ical changes in DE over the years, taking into consideration the lack of consistent analytical and m easurem ent techniques and the variability in em issions based on vehicle m i x, driving cycles, engine deterioration, and other
factors. Key findings and inferences relevant to the potential health effects of DE are discussed
below. As discussed in Chapter 2, Section 2.2.3, the EPA Em issions Trend Report estim ates that
DPM 10 on-road em issions decreased 27% between 1980 and 1998. DPM em ission factors (g/m ile by m odel year) from new on-road diesel vehicles decreased on average by a factor of six
from the m i d-1970s to the m i d-1990s. These significant reductions are largely attributable to
reductions in three PM com ponents: EC, organic carbon, and sulfate. Lim ited data are available to assess the changes in em ission rates from locom o tive, m a rine, or other nonroad diesel sources
over tim e, although it is estim ated that DPM 10 ( 

### # 10 

### : m ) em issions from nonroad diesel engines increased 17% between 1980 and 1998 (Chapter 2, Section 2.2.5). Because of changes in engine technology and fuel com position, the chem ical

com position of DPM from on-road vehicles has also changed over tim e. The percentage of soluble organic m a terial associated with DPM decreased by m odel year from the 1980s to the 1990s, and the proportion of EC is correspondingly hi gher. PAHs and nitro-PAHs are present in
DPM from both new and older diesel engines. There are insufficient data to provide clear insight into the potential f o r changes in total PAH em issions over tim e or specif ic PAHs such as benzo[a]pyrene and 1-nitropyrene. It should be noted that the chem ical com position of am bient DPM to which people are currently exposed is determ ined by a com b ination of exhaust from older and newer engines as well as on-road and nonroad applications of those engines. Consequently, the decrease in the soluble organic fraction of DPM by m odel year for on-road engines does not directly translate into a proportional decrease in DPM-associated organic

m a terial to which people are exposed. In a ddition, the contributions from high-em itting and/or

sm oking diesel engines have not been quantified (Chapter 2, Section 2.5.2). Because of these uncertainties, the exposure im pact of changes in DPM com position over
tim e cannot be confidently characterized. Availa ble data clearly indicate that toxicologically significant organic com ponents of DE (e.g., PAHs , PAH derivatives, nitro-PAHs) were present in DPM and DE in the 1970s and are still present. Even though a significant fraction of am bient DPM (possibly m o re than 50%) is em itted by nonroad equipm ent, data are currently inadequate to characterize changes in the chem ical com position of DPM from nonroad equipm ent over tim e. Given the variation in fuel, engine technology, and in-use operational factors over the years, caution should be exercised in presum ing that a decrease in the am ount of em issions or em ission

9-5constituents f r om older engines to present day in-use engines will result in a decrease in
hazard/risk. In m eeting the 2007 federal regulations for heavy-duty DE, the exhaust com position
will be m a rkedly changed with a consequence that health hazards are expected to be significantly
reduced.

9.3. AMBIENT CONCENTRATIONS AND EXPOSURE TO DIESEL EXHAUST Chapter 2, Section 2.4 provides inform ation on occupational and environm ental exposures to DE in order to provide a context for the hazard assessm ent and dose-response analysis. Highlights of the available inform ation are discussed below.
DE is em itted from a variety of sources, both on-road (e.g., m o tor vehicles, construction equipm ent) and nonroad (e.g., farm equipm ent, railway locom o tives, or m a rine uses).
Environm ental exposure to DE is generally higher in urban areas than in rural areas. The
concentration of DE in the air will vary within any geographic area depending on the num ber and types of diesel engines in the area and the atm o spheric patterns of dispersal. Som e im portant factors that determ ine the difference be tween the am bient concentration of DE and the resultant exposure to an individual include the proxim ity of a person to the DE source and his/her pattern of activity which, for exam ple, includes outdoor versus indoor activities as well as related breathing rates. Certain occupational populations (e.g., transportation and garage workers, heavy-equipm ent operators, and others who spend considerable tim e outdoors) can be exposed to m u ch higher levels of DE than the general population. The am ount or num ber of particles delivered and retained in the lung is one factor that could contribute to differential
hum an susceptibility to DPM. For exam ple, children have sm aller lungs than adults and thus could have a higher lung burden of inhaled DPM per lung surface area if their activity pattern results in a high breathing frequency.
As DE is a com p lex m i xture of m a ny constituents, environm ental concentration

m easurem ents and related hum an exposure is difficult to precisely m easure. Even though levels
of a num ber of DE constituents are generally known, it is difficult to quantify the portion that

com e s f r om DE since other types of em ission sources also m a y em it the sam e constituent. Moreover, there is still incom p lete knowledge about the relative roles of the relevant DE
constituents in m e diating the potential health effects of DE. Historically, exposure levels to DPM have been used as a surrogate m a rker/dosim eter for whole DE. Although uncertainty
exists as to whether DPM m a ss (expressed as 

### : g/m 3 of DPM ) is the m o st appropriate dosim eter
f o r health ef f ect purposes, it is considered to be a reasonable choice until m o re def i nitive
inform ation is available about the m echanism s or m ode(s) of toxicity action of DE. Several techniques exist for estim ating am bient concentrations of DPM, including
chem ical m a ss balance (CMB) source apportionm ent, dispersion m odeling, and using EC as a

9-6surrogate for DPM. DPM concentrations repor ted from CMB and dispersion m odeling studies in the 1980s suggest that in urban and suburba n areas (Phoenix and Southern California), the annual average DPM concentration ranged from 2 to 13 

### : g/m 3

. In the 1990s, annual or seasonal average DPM concentrations in suburban or urban locations have ranged from 1.2 to 4.5 

### : g/m 3

. DPM concentrations at a m a jor bus stop in downtown Manhattan ranged from 13.2 to 46.7


### : g/m 3 over a 3-day period in 1993. In nonurban and rural areas in the 1980s, DPM concentrations were reported to range from 1.4 to 5 

### : g/m 3

. In the 1990s, nonurban air basins in California were reported to have DPM concentrations ranging from 0.2 to 2.6 

### : g/m 3 (Chapter 2, Section 2.4.2).
A com p rehensive exposure assessm ent is not presented in this assessm ent, though EPA is developing this in an analysis called the National Air Toxics Assessm ent. Interim exposure estim ation based on EPA’s Hazardous Air Pollutant Exposure Model (HAPEM-MS3 m odel), for on-road sources only, suggests that in 1996 annual average DPM exposure in urban areas from only on-road engines was 0.7 

### : g/m 3 , while in rural areas exposure was 0.3 

### : g/m 3

. Am ong 10 urban areas, the 1996 annual average estim ated exposure ranged from 0.5 to 1.2 

### : g/m 3

. A highend exposure estim ate for 1996 is not yet available. Com p arable 1990 exposure estim ates for on-road sources ranged from 0.9 

### : g/m 3 for urban areas to 0.5 

### : g/m 3 for rural areas. In 1990 exposure estim ates for the m o st highly exposed individuals (e.g., outdoor workers and children who spend large am ounts of tim e outdoors) were estim ated to be up to 4.0 

### : g/m 3 (Chapter 2, Section 2.4.3.2, Table 2-29). Nationwide level nonroad em ission exposures are estim ated to be nearly double those from on-road sources.
Estim ates for occupational exposures to DE as DPM m a ss are generally higher than
environm ental exposures. Tables 2-27 and 2-28 pr ovide historic exposure estim ates for specific worker categories. For exam ple, historic DPM exposure estim ates range from 39–191 

### : g/m 3 for railroad workers, 4–748 

### : g/m 3 for firefighters, 7–98 

### : g/m 3 for public transit workers and airport crews, 5–61 

### : g/m 3 for m echanics and dock workers, and 2–7 

### : g/m 3 for long- and short-haul
truck drivers. For a direct com p arison of lifetim e exposures between an occupational setting (8 hours per day, 5 days per week, for 45 years) and environm ental exposure (continuous exposure for 70 years), the occupational estim ates are converted to an equivalent environm ental lifetim e estim ate, 1 which is also shown in Table 2-28. A conversion of EC-based m easurem ents to total
DPM also m a y be needed for som e estim ates. The estim ated 70-year lifetim e exposures equivalent to those for the occupational groups discussed above range from about 0.4–157


### : g/m 3

. These data indicate that som e lower-end occupational estim ates of DPM, when
converted to environm ental equivalents, overlap the range of estim ated environm ental exposures

1 Environm ental equivalent occupational exposure = 0.21 × occupational exposure.

9-7to DPM from on-road em issions (national average in 1990 of 0.8 

### : g/m 3 , with high-end exposures up to about 4 

### : g/m 3 ). The addition of nonroad em ission exposures, when appropriate,

m a kes the case for overlap of occupational and environm ental exposure m o re prevalent.

9.4. HAZ ARD CHARACTERIZ ATION The prim ary health effects of concern from environm ental exposure to DE include
effects associated with both acute and short-term exposures as well as chronic exposures. It is recognized that acute exposures m a y produce transitory physiological sym p tom s of varied severity as well as exacerbation of allergenic effects from acute and repeated exposures. On the
basis of com b ined hum an and experim e ntal evidence from chronic exposure studies, noncancer respiratory effects and lung cancer are observed.
The health effects data are based on DE from a variety of engines existing before the

m i d-1990s. There have been changes in the physical and chem ical com position of som e DE

em issions (on-road vehicle em issions) over tim e, though there is no definitive inform ation to
show that the em ission changes portend significant toxicological changes. The m ode(s) of action for DE toxicity in hum ans is not understood, and hence knowledge is lacking about the role of
exhaust m i xture com ponents in m odulating the toxicity. Taken together, these considerations have lead to a judgm ent that the hazards identified from older technology-based exposures are applicable to current-day exposures. As new and cleaner diesel engines replace a substantial

num ber of existing engines, the general applicability of the older data will need to be
reevaluated. As discussed in Chapter 6 (Section 6.4), it is also reasonable to expect that DPM, being a
constituent of am bient f i ne PM (PM 2.5 ), would contribute to the wider spectrum of effects that
have been associated with am bient PM 2.5 . Com m unity epidem iologic studies have shown that

am bient PM 2.5 exposure is statistically associated with increased m o rtality (especially am ong people over 65 years of age with preexisting cardiopulm onary conditions) and m o rbidity as

m easured by increases in hospital adm i ssions, respiratory sym p tom rates, decrem ents in lung function, and exacerbation of asthm a , and possibl y im m unological effects in the respiratory system

. There continues to be little epidem iologic evidence for an effect of am bient exposure to PM on cancer rates (U.S. EPA, 1996a,b), though U.S. EPA’s Criteria Docum e nt for Am bient PM (expected to be released in 2002) will exam ine the question further.

9.4.1. Acute and Short-Term Exposures
The com b ined hum an and anim al evidence indicates that DE can induce irritation to the eye, nose, and throat, as well as inflam m a tory responses in the airways and the lung following

9-8acute and/or short-term exposure to high concentrations. There also is suggestive evidence for possible im m unological and allergenic effects of DE.

9.4.1.1. Acute Irritation DE contains various respiratory irritants in the gas phase and in the particulate phase (e.g., SO

x , NO
x , aldehydes). Acute exposure to DE has been associated with irritation of the eye, nose, and throat, respiratory sym p tom s (cough and phlegm ) , and neurophysiological sym p tom s such as headache, lightheadedness, nausea, vom iting, and num bness or tingling of the

extrem ities. Such sym p tom s have been described m a inly in reports of individuals exposed to DE in the workplace, or in clinical studies in hum ans exposed acutely to high concentrations of DE. Because of the general lack of validating exposur e inform ation in the reports, the role of DE in causing these effects is unknown. An exposure-response relationship for these acute irritation and respiratory sym p tom s has not been dem onstrated (Chapter 5, Section 5.1.1.1).

9.4.1.2. Respiratory Effects Available studies of occupational exposure to DE have not provided evidence for significant decrem ents of lung function in work ers over a work shift or after a short-term exposure period. Short-term and subchronic inhalation studies of DE in anim als (rats, m i ce, ham s ters, cats, guinea pigs) showed inflam m a tion of the airways and m i nim a l or no lung function changes. These effects were associated with high DE exposures (up to 6 m g /m 3 ). Exposure-response relationships have not been established for these responses (Chapter 5, Sections 5.1.1, 5.1.2, and 5.1.3).

9.4.1.3. Immunological Effects
Recent hum an and anim al studies show that acute DE exposure episodes can exacerbate
im m unological reactions to other allergens or initiate a DE-specific allergenic reaction. The
effects seem to be associated with both the organic and carbon core fraction of DPM. In hum an subjects, intranasal adm i nistration of DPM has resulted in m easurable increases of IgE antibody production and increased nasal m R NA for som e proinflam m a tory cytokines. These types of responses also are m a rkers typical of asthm a , though for DE, evidence has not been produced in
hum ans that DE exposure results in asthm a

. The ability of DPM to act as an adjuvant to other allergens also has been dem onstrated in hum an subjects. For exam ple, co-exposure to DPM and ragweed pollen was reported to significantly enhance the IgE antibody response and cytokine expression relative to ragweed pollen alone. Available anim al studies also dem onstrate the potential adjuvant effects of DPM with m odel allergens, e.g., in m ouse studies the allergenic
reaction to ovalbum in and Japanese cedar pollen (Chapter 5, Sections 5.1.1.1.3 and 5.1.1.1.4).

9-9Additional research is needed to further characterize im m unological effects of DE and to
determ ine whether or not the im m unological effects constitute a low-exposure hazard. This health endpoint is of considerable public health concern, given the increases in allergic hypersensitivity in the U.S. population (Chapter 5, Section 5.6.2.6).

9.4.2. Chronic Exposure
9.4.2.1. Noncancer Effects Available long-term and cross-sectional hum an studies have provided evidence for an association between respiratory sym p tom s (cough and phlegm ) and DE exposure, but there was no consistent effect on lung function. DE has b een shown in m a ny anim al studies of several species to induce lung injury (chronic infla m m a tion and histopathologic changes) following long-term inhalation exposure. DE also has been tested in laboratory anim als for other health effects, but no significant effects have been found. Overall, available data lead to the conclusion of a potential chronic respiratory hazard to hum ans from long-term exposure to DE.

9.4.2.1.1. Respiratory effects. A few hum an studies in various diesel occupational settings suggest that DE exposure m a y im pair pulm onary function, as evidenced by increases in respiratory sym p tom s and som e reductions in baseline pulm onary function consistent with restrictive airway disease. Other studies found no particular effects. The m e thodologic
lim itations in available hum an studies lim it their usefulness in drawing any firm conclusions about DE exposure and noncancer respirat ory effects (Chapter 5, Section 5.1.1.2).
Available studies in anim als, however, provide a large body of evidence dem onstrating that prolonged inhalation exposure to high concentrations of DE can result in pulm onary injury.
A num ber of long-term laboratory studies in rats, m i ce, ham s ters, cats, and m onkeys found varying degrees of adverse lung pathology incl uding focal thickening of the alveolar walls, replacem ent of Type I alveolar cells by type II cells, and fibrosis. The rat is the m o st sensitive
anim al species to DE-induced pulm onary toxicity (Chapter 5, Sections 5.1.3 and 5.4).
Available m echanistic data, m a inly in rats, indicates that the DPM f r action of DE is a controlling factor in the etiology of pulm onary toxicity, although a role for the adsorbed organic

com pounds on the particles and in the gaseous phase cannot be ruled out. The lung injury appears to be m e diated by an invasion of alveol ar m acrophages that release chem otactic factors that attract neutrophils and additional alveolar m acrophages, which in turn release m e diators (e.g., cytokines, growth factors) and oxygen radicals. These m e diators result in persistent
inflam m a tion, cytotoxicity, im paired phagocytosis, clearance of particles, and eventually deposition of collagen by activated fibroblasts. This m ode of action seem s to be operative for a variety of poorly soluble particles in addition to DPM (ILSI, 2000). Because long-term exposure

9-10to DE has been shown to induce exposure-depende nt chronic respiratory effects in a wide range
of anim al species, and the m ode of action is deem ed relevant to hum ans, there is a sufficient scientific basis to support a conclusion that hum ans also could be at hazard for these effects under a chronic exposure condition.

9.4.2.1.2. Other noncancer effects

. The negative results from available studies in several
anim al species (rats, m i ce, ham s ters, rabbits, m onkeys) indicate that DE is not likely to pose a reproductive or developm ental hazard to hum ans. There has been som e evidence from anim al studies indicating possible neurological and behavior al effects, as well as liver effects. These effects, however, are seen at exposures higher than the respiratory effects. Overall, there is insufficient evidence to conclude that a low-exposure hazard exists for these endpoints (Chapter 5, Section 5.1.3.3).

9.4.2.2. Carcinogenic Effects Many epidem iologic and toxicologic studies have been conducted to exam ine the potential for DE to cause or contribute to the developm ent of cancer in hum ans and anim als, respectively. In addition, there are som e m ode-of-action studies that seek to provide an
im proved understanding about the underlying carci nogenic process and thus contribute to a better understanding of the likelihood of hazard to hum ans. The available evidence indicates that chronic inhalation of DE is likely to pose a lung cancer hazard to hum ans. There is
insufficient inform ation for an evaluation of the potential cancer hazard of DE by oral and
derm al routes of exposure.

9.4.2.2.1. Epidemiologic studies. Twenty-two epidem iologic studies about the carcinogenicity of workers exposed to DE in various occupations are reviewed in Chapter 7, Section 7.2. Exposure to DE has typically been inferred on the basis of job classification within an industry,
with cum u lative exposure based on duration of em ploym ent or age. Increased lung cancer risk, although not always statistically significant, has been observed in 8 out of 10 cohort studies and 10 of 12 case-control studies within several industries, including railroad workers, truck drivers, heavy-equipm ent operators, farm tractor operators, and professional diesel vehicle drivers. The increased lung cancer relative risks generally range from 1.2 to 1.5, although a few studies show relative risks as high as 2.6. Statistically significant increases in relative risk, 1.33 to 1.47, are also shown in two independent m e ta-analyses. The m e ta-analyses dem onstrate the effect of pooling m a ny studies and in this case show the positive relationship between DE exposure and lung cancer across a variety of DE-exposed occupations.

9-11The generally sm all increases in lung cancer relative risk (1.2 to 1.5, i.e., less than 2) observed in the epidem iologic studies potentially weakens the evidence of causality. This is because with a relative risk of less than 2, if confounders (e.g., sm oking, asbestos exposure) were having an effect on the observed risk increases, then it could be enough to account for the
increased risk. W ith the strongest risk factor for lung cancer being sm oking, there is a lingering uncertainty as to whether sm oking effects m a y be influencing the m a gnitude of the observed increased relative risks, in spite of the f act that in key studies the investigating epidem iologists assert that they have effectively controlled for sm oking. In studies in which the effects of

sm oking were controlled, increased relative risks for lung cancer prevailed. W h ile som e studies
did not have inform ation on sm oking, confounding by sm oking is judged unlikely to be significant if the com p arison populations were from the sam e socioeconom ic class. As evaluated in Chapter 7 (Section 7.2.4.5), application of the criteria for causality provides a rational basis to conclude that the increased risks observed in available epidem iologic studies are consistent with a causal association between exposure to DE and occurrence of lung

cancer. Overall, the hum an evidence for potential carcinogenicity for DE is judged to be strong but less than sufficient to satisfy the criteria for a “known” hum an carcinogen because of exposure uncertainties (lack of historical exposure of workers to DE) and residual uncertainty as to whether all confounders have been satisfactorily accounted for. The epidem iologic evidence is inconclusive for DE being associated with other form s of cancer.

9.4.2.2.2. Animal studies. DE and its organic constituents, both in the gaseous and particle phase, have been extensively tested for carcinogenicity in m a ny experim e ntal studies using
several anim al species and with different m odes of adm i nistration. Several well-conducted studies have consistently dem onstrated that chronic inhalation exposure to sufficiently high concentrations of DE produced dose-related increases in lung tum o rs (benign and m a lignant) in

rats. In contrast, chronic inhalation studies of DE in m i ce showed equivocal results, whereas negative f i ndings were consistently seen in ham s ters. The gaseous phase of DE (f iltered exhaust without particulate fraction), was found not to be carcinogenic in rats, m i ce, or ham s ters. The
available data indicate that am ong the traditional anim al test species, the rat is the m o st sensitive species to DE. As reviewed in Chapter 7, Section 7.4, the lung cancer response in rats from high-concentration exposures to DE appears to be m e diated by im pairm e nt of lung clearance

m echanism s owing to particle overload, resulting in persistent chronic inflam m a tion and subsequent pathologic and neoplastic changes (i.e., cancer) in the rat lung. Particle overload
conditions in the hum an lung are not expected to occur as a result of environm ental or m o st occupational exposures to DE. Thus, the increased lung tum o rs in the rat are not an appropriate

9-12basis from which to judge the potential for a hum an hazard or perform a dose-response analysis
to derive a cancer unit risk for hum ans. In several intratracheal instillation studies, DPM, DPM organic extracts, and carbon black, which is virtually devoid of PAHs, have been found to produce increased lung tum o rs in

rats. W h en directly im planted into the rat lung, DPM condensate containing m a inly four- to seven-ring PAHs induced increases in lung tum o rs. DPM extracts also have been shown to
cause skin tum o rs in several derm al studies in m i ce and sarcom as in m i ce following subcutaneous injection. Overall there are consistent findings of carcinogenic activity by the organic extracts of DPM in noninhalation studies (i.e., intratracheal instillation, lung
im plantation, skin painting). This contributes to the evidence for a potential hum an hazard.

9.4.2.2.3. Other key data. W h ile not as extensive as the hum an and anim al carcinogenicity data, other types of data are judged to be supportive of DE’s potential carcinogenicity in hum ans. As

m e ntioned previously, DE is a com p lex m i xture of hundreds of constituents in either the gaseous phase or particle phase. Although present in sm all am ounts, several organic com pounds in the gaseous phase (e.g., PAHs, form aldehyde, acetaldehyde, benzene, 1,3-butadiene) are known to
exhibit m u tagenic and/or carcinogenic activities. PAHs and PAH derivatives, including nitroPAHs present on the diesel particle, also are known to be m u tagenic and carcinogenic. As reviewed in Chapter 4, DPM and DPM organic extracts have been shown to induce gene

m

u tations in a variety of high-dose bacteria and m a m m a lian cell test system s. DPM and DPM organic extracts also have been shown to induce chrom o som a l aberrations, aneuploidy, and sister
chrom a tid exchange in both rodent and hum an in vitro tests. There also is suggestive evidence for the bioavailability of organic com pounds from the
DE m i xture. Elevated levels of DNA adducts in lym phocytes have been reported in workers exposed to DE. In addition, inhalation studies of anim als using radio-labeled m a terials indicate

som e elution of organic com pounds from DPM after deposition in the lung as m easured by their presence in biological tissue and fluids (Chapter 3, Section 3.5).

9.4.2.2.4. Modes of carcinogenic action. The term “m ode of action” refers to a series of key biological events and processes that are critical to the developm ent of cancer. As discussed in Section 9.4.2.2.2, there is an understanding of the m odes of action for the DE-induced lung

tum o rs in the rat. However, the m odes of action by which DE increases lung cancer risks in
hum ans are unknown, and the evidence in rats is not applicable to environm entally exposed
hum ans. As discussed in Chapter 7, Section 7.4, it is hypothesized that m u ltiple m odes of action
could be involved in m e diating the carcinogenic effect of DE. These m odes of action m a y

9-13include: (a) m u tagenic events (e.g., direct effects on DNA and effects on chrom o som e s) by organic com pounds in the gas and particle phase, (b) indirect DNA dam a ge via the production of reactive oxygen species (ROS) induced by particle-associated organics, and (c) particle-induced
chronic inflam m a tory response leading to oxidative DNA dam a ge through the release of cytokines, ROS, etc., and an increase in cell proliferation. In rats, the particulate phase appears to have the greatest contribution to the carcinogenic effects, and both the particle core and the associated organic com pounds have dem onstrated carcinogenic properties in one or m o re test system s. W h ile lim ited rat data and com p arative potency calculations suggest that gas-phase com ponents are not the prim ary factors in the developm ent of lung cancer, a contributory role of the recognized toxic com ponents cannot be
dism issed. The relative im portance of the various m odes of action m a y be different at different exposure levels. Evidence from rat studies indicates the im portance of the EC com ponent of the DE particle in m e diating lung tum o r response at high exposure levels. As for the particleabsorbed organics, their inherent toxicity potentia l gives rise to a hypothesis that they m a y play a role in low or high exposures to DE.

9.4.2.2.5. Weight-of-evidence evaluation. Chapter 7, Section 7.5, provides an evaluation of the overall weight of evidence for hum an carcinogenicity in accordance with EPA’s Guidelines for Carcinogen Risk Assessm ent (U.S. EPA, 1986, 1996a, 1999). The totality of evidence supports
the conclusion that DE is a probable human carcinogen (Group B1) by inhalation exposure using the criteria in the 1986 guidelines. A cancer hazard narrative for DE also is provided in accordance with the revised draft 1996/1999 guidelines, which concludes that DE is likely to be carcinogenic to humans by inhalation from environm ental exposures. The com m on bases for either conclusion include the following lines of evidence:

• s trong but less than sufficient evidence for a causal association between DE exposure and increased lung cancer risk am ong workers in varied occupations where exposure to DE occurs;

• e xtensive supporting data including the dem onstrated m u tagenic and/or chrom o som a l effects of DE and its organic constituents, and knowledge of the known m u tagenic and/or carcinogenic activity of a num ber of individual organic com pounds present with particles and in the DE gases;

• e vidence of carcinogenicity of DPM and the associated organic com pounds in rats
and m i ce by noninhalation routes of exposure; and

• s uggestive evidence for the bioavailability of DE organics from DE in hum ans and
anim als.

9-14A notable uncertainty in the characterization of the potential cancer hazard of DE at low
levels of environm ental exposure is the incom p lete understanding of about its m ode(s) of action for the induction of lung cancer in hum ans. Available data suggest that DE-induced lung carcinogenicity m a y be m e diated by m u tagenic and nonm utagenic events by both the particles and the associated organic com pounds, and that a role for the organics in the gaseous phase
cannot be ruled out. Given that there is som e evidence for a m u tagenic m ode of action, a cancer hazard is presum ed possible at environm ental levels of exposure. This is consistent with EPA’s science policy position that assum e s a nonthreshold effect for carcinogens with a m u tagenic

com ponent in the absence of definitive data dem onstrating a threshold or nonlinear m echanism

. Additional support for an environm ental hazard also com e s from a com p arison of the estim ated
environm ental levels to the estim ated occupational exposure levels where risk is seen. Given that there is only a m i nim a l m a rgin between environm ental and occupational exposure ranges, if not an overlap, the extrapolation of observable hazard from the occupational setting to the

am bient environm ent is relatively confident. Because of insufficient inform ation, the hum an carcinogenic potential of DE by oral and derm al exposures cannot be determ ined. Several organizations previously have reviewed available relevant data and evaluated the potential hum an carcinogenicity of DE or the particulate com ponent (DPM) of DE. Sim ilar conclusions were reached by various organizations (see Table 7-9). For exam ple, som e organizations have concluded that DE is probably carcinogenic to hum ans (IARC, 1989; IPCS, 1996), or reasonably is anticipated to be a carcinogen (NTP, 2000). Overall, the weight of evidence for potential hum an carcinogenicity for DE is considered strong, even though inferences are involved. Uncertainties are present, however, and include the following unresolved issues. First, there has been a considerable scientific debate about the significance of the
available hum an evidence for a causal association between occupational exposure and increased lung cancer risk. Som e experts view the evidence as weak and/or inconsistent while others
consider the evidence com p elling. This is due to a lack of consensus about whether the effects
of sm oking and other potential confounders have been adequately accounted for in key studies, and the lack of agreed-upon historical DE exposure data for the available studies. Second, while the m ode of action for DE-induced lung tum o rs in rats from high exposure is sufficiently understood, the m ode of action for the DE lung cancer risk in hum ans is not
known. To date, available evidence for the role of both the adsorbed organics and the carbon core particle has only been shown under high-exposure experim e ntal anim al test conditions. There is virtually no inform ation about the relative role of DE constituents in m e diating carcinogenic effects at the low-exposure levels.

9-15Additional research is needed to address these issues to reduce the uncertainty associated with the potential cancer hazard of exposure to DE.
The relevance of this hazard characterization to current am bient DE exposures hinges on recognizing that the health effects data are de rived from engine technologies and fuels that existed in the past, and that som e changes in the DE exhaust m i xture have occurred and can be expected in the future. Although decreases in am ount and changes in com position of DE

em issions have occurred over tim e for on-road engines, a change is slow to m a nifest in the
environm ent because, for exam ple, vehicular fleet turnover is slow and the change is slow to
dom inate across an engine fleet. Available studies have not focused on the potential toxicological effect of the em ission changes. Th ere is no com p elling evidence at present to show that past and present exhaust characteristics are so toxicologically dissim ilar as to render the
current use of the assessm ent’s findings outdated.

9.5. DOSE-RESPONSE ASSESSMENT
In assessm ents of estim ated hum an health risks, hum an data from environm ental exposures are always preferred over anim al data, if available, as their use obviates the need for extrapolation across species, e.g., from anim als to hum ans. However, for m o st environm ental agents, available health effects inform ation is generally lim ited to occupational exposures in
studies of hum ans (e.g., workers) or high experim e nt al exposures to laboratory anim als. For the agents with high-exposure data com p ared to environm ental exposure levels of interest, doseresponse assessm ent is perform e d in two steps: assessm ent of data in the observable range to derive a point of departure (which usually is the lowest exposure or dose that induces som e ,

m i nim a l, or no apparent effects), followed by extrapolation to lower exposures to the extent

necessary. Extrapolation to low exposures is ideally based on the understanding of m ode(s) of toxic action of the agent which allows the developm ent and use of a m ode of action specific

exposure-response m odel. In the absence of sufficient data, default m e thods and m odels are used to extrapolate to the lower exposure levels. For DE, there is suf f i cient evidence to conclude that acute or short-term inhalation exposure at relatively high levels can cause irritant effects to the eye, nose, and throat, respiratory sym p tom s , and neurophysiological sym p tom s such as headache, nausea, etc., however, no quantitative data are available to derive an estim ate of hum an exposure that is not likely to elicit irritant and inf l am m a tory ef f ects in hum ans. There is also sufficient evidence to support the conclusion that DE has the potential to cause cancer and noncancer effects of the lung from long-term inhalation exposure. Chapters 6 and 8 provide dose-response analyses related to the noncancer and cancer hazards to hum ans, respectively, from lifetim e exposure to DE. A dose-response analysis to estim ate the expected

9-16response at environm ental exposure levels has less uncertainty the closer the anim al test or
estim ated hum an epidem iologic-related exposures are to the environm ental levels of interest.
W ith increasing exposure m a rgins (EM), and thus a greater range of extrapolation, the uncertainty about the shape of the dose-response curve in the region of low-dose extrapolation increases and the possibility of a zero risk cannot be ruled out.

9.5.1. Evaluation of Risk for Noncancer Health Effects As discussed previously (Section 9.4.2.1), the evidence for potential chronic noncancer health effects of DE is based prim arily on findings from chronic anim al inhalation studies showing a spectrum of dose-dependent chronic inflam m a tion and histopathological changes in the lung in several anim al species including rats, m i ce, ham s ters, and m onkeys. A lim ited

num ber of epidem iologic investigations of workers exposed to DE have not provided consistently clear evidence of significant chronic respiratory effects associated with DE

exposure. On the other hand, the relatively large epidem iologic database for am bient PM shows a clear relationship between respiratory effects and am bient fine PM that is partially com posed
of

DPM. The specif i c role of DPM or any other source-related constituent of am bient PM in causing the observed respiratory effects has not been defined. The approach taken in this assessm ent to estim ate a level of DE in the air to which
hum ans m a y be exposed throughout their lifetim e without an appreciable risk of deleterious effects is to derive a reference concentration (RfC) for DE based on the consistent data for respiratory inflam m a tion in the rat studies. This approach assum e s that hum ans would respond
to DE sim ilarly to the tested anim als under sim ilar exposure conditions. An uncertainty of this approach stem s from the circum stance that anim al studies have used high DE exposures, and the
anim al results m u st be translated to hum ans as well as to lower exposure levels since the potential chronic health effects of DE in hum ans at environm ental exposure levels cannot be
ascertained from the available DE hum an data. It also is relevant to recognize that DPM is a com ponent of am bient fine PM and that
there is a relative wealth of hum an effects data for am bient PM showing a sim ilarity of certain
adverse health effects for DPM and am bient fine PM. This allows one to reasonably expect that
the PM 2.5 National Am bient Air Quality Standard (NAAQS) would provide a m easure of protection from DPM, reflecting DPM’s current and approxim a te proportion to PM 2.5 . Am bient
PM 2.5 has been shown to be statistically associated with increased m o rtality (especially am ong people over 65 years of age with preexisting cardiopulm onary conditions) and m o rbidity, as

m easured by increases in hospital adm i ssions, respiratory sym p tom rates, and decrem ents in lung function with both long- and short-term changes in am bient PM 2.5 concentrations.

9-179.5.1.1. Chronic Reference Concentrations for Diesel Exhaust An inhalation Reference Concentration (RfC) is based upon long-term data, i.e., chronic exposure, and can be derived from either hum an or anim al data. An RfC is correctly defined as
“an estim ate of a continuous inhalation exposure to the hum an population, including sensitive subgroups, with uncertainty spanning perhaps an order of m a gnitude, that is likely to be without appreciable risks of deleterious noncancer effects during a lifetim e.” The RfC m e thodology

assum e s that there is an exposure threshold below which effects will not occur. The RfC is not a bright line; rather, as the long-term hum an exposure increases above the RfC, the m a rgin of protection decreases.
W ith the absence of DE exposure-response data in hum ans, this assessm ent derives an RfC for DE based on dose-response data from four chronic inhalation studies in rats (Mauderly et al., 1987; Ishinishi et al., 1988; Heinrich et al., 1995; Nikula et al., 1995). All of these studies used DPM (expressed as 

### : g/m 3 ) as a m easure of DE exposure. The pulm onary effects, including
inflam ation and histopathologic lesions, were consid ered to be the critical noncancer effects. As shown in Table 6-2, the no-observable-adverse-e ffects levels (NOAELs), the lowest-observableadverse-effects levels (LOAELs), and the adverse effects levels (AELs) for lung inflam m a tion and histopathologic changes were identified for th e first three studies. For the Nikula et al. study, lower 95% confidence estim ates of the c oncentrations of DPM associated with a 10% incidence (BMCL 10 ) of chronic pulm onary inflam m a tion and fibrosis were derived since
NOAEL’s were not observed. For all four studi es, hum an equivalent concentrations (HECs) corresponding to the anim al NOAEL, LOAEL, AEL, and BMCL 10 were then com puted using a
dosim etry m odel developed by Yu et al. (1991) as described in Chapter 6, Section 6.5.2, and Appendix A. The dosim etry m odel accounts for species differences (rat to hum an) in respiratory exchange rates, particle deposition efficiency, di fferences in particle clearance rates at high and low doses, and transport of particles to lym ph nodes. The purpose is to identify the highest HEC value with no apparent effect, i.e., NOAEL HE C . The highest NOAEL HE C associated with no apparent effect is 144 

### : g/m 3 from the Ishinishi et al. (1988) study; this then becom e s the point of departure for deriving an RfC. To obtain the RfC, this point of departure was divi ded by two types of uncertainty factors (UF): a factor of 3 recognizes interspecies (i.e., rat to hum an) extrapolation uncertainties, and a factor of
10 reflects uncertainties about interindividual hum an variation in sensitivity. An evaluation of the interspecies extrapolation issues for dosim eteric and pharm acodynam i c equivalence between
rats and hum ans showed that although som e adjustm e nts could be accounted for, there rem a ined a residual uncertainty, and thus an uncertainty of 3 out of a possible factor of 10 is used. In the
absence of m echanistic or specific data, a default value of 10 is considered appropriate to account for possible hum an variability in sensitivity, particularly for children and people with

9-18preexisting respiratory conditions. The spectrum of the population that m a y have a greater susceptibility cannot be better characterized until there is additional knowledge about m ode of
action. The resulting RfC for DE is 5 

### : g/m 3 of DPM. Overall, the confidence level in the RfC is considered m e dium in a range of low to high
conf idence. A principal uncertainty of the Rf C analysis is the reliance on anim al data to predict
hum an risk. The critical effects, chronic infl am m a tion, and pathologic changes, which are well
characterized in four anim al species, are considered relevant to hum ans. Collective evidence for all poorly soluble particles, including DPM, indicates that the rat is the m o st sensitive laboratory
anim al species tested to date. Although in general the rat is thought to be m o re sensitive to lung injury than hum ans to poorly soluble particles (ILSI, 2000), it is not clear that this is the case specifically for diesel. W e m u st recall that DE is a m i xture of not just carbon particles but also various organics, both on the particles and in gases. In addition, differences in particle deposition, retention, and clearance m echanism s have been largely but perhaps not com p letely addressed by the use of the rat-to-hum an dosim etry m odel. The use of rat data is not likely to grossly underestim ate the hum an risk for pulm onary noncancer health effects. In term s of the potential for other critical health effects, there is growing evidence suggesting that DE can exacerbate allergenic effects to known sensitizers, while also evoking production of biochem i cal

m a rkers typically associated with asthm a

. Som e work in this area indicates that hum ans m a y be
as sensitive as rats and m i ce to the im m unologic effects (Chapter 6, Section 6.3.4). This database is currently lacking key exposure-response data, but m a y in the future provide an
alternative basis for RfC derivation. It also should be noted that the am bient PM health effects data show a broader array of adverse hum an health concerns (e.g., cardiovascular effects, as well as acute exposure effects). W ith DPM being a ubiquitous com ponent of am bient PM, there is an uncertainty about the adequacy of the existing DE noncancer database to identify all of the pertinent DE-caused noncancer health hazards.

9.5.1.2. Risks Based on Ambient PM 2.5 As discussed in Chapter 6 (Section 6.4), in 1997 EPA established an annual NAAQS for
PM 2.5 , at a level of 15 

### : g/m 3 to provide protection against adverse health effects associated with both long- and short-term exposures to am bient fine PM. DPM is a typical constituent of

am bient fine PM (generally about 10% of PM 2.5 with som e exam ples up to 36%). 2 Given the

2 “A qualitative com p arison of adverse effects of exposure to DPM and am bient fine PM shows that the respiratory system is adversely affected in both cases, though a wider spectrum of adverse effects has been identified for am bient fine PM. In contrast to the diesel PM database, there is a wealth of hum an data for fine PM noncancer effects which indicates that the health effects from fine PM do not have a discernable threshold at this tim e.”

9-19sim ilarity of health concerns for respirator y inflam m a tion and pulm onary health effects from both DPM and fine particles, it is reasonable to expect that DPM contributes to som e of the
health effects associated with PM 2.5 . Current knowledge is insufficient, however, to describe the relative potencies of DPM and the other com ponents of PM 2.5. As long as the percentage of
DPM to total am bient PM 2.5 rem a ins in sim ilar proportion, protective levels for PM 2.5 would be expected to offer a m easure of protection from effects associated with DPM.

9.5.1.3. Conclusions
This assessm ent estim ates an exposure air level of DE (as m easured by DPM) to which
hum ans m a y be exposed throughout their lifetim e without experiencing any adverse noncancer
health effects. The approach taken applies the RfC m e thod using data specific to DE to produce
an RfC of 5 

### : g/m 3 of DPM on the basis of four chronic inhalation studies of DE in rats and a

com posite uncertainty factor of 30. In addition, this assessm ent also recognizes the relative wealth of data regarding health effects associat ed with am bient PM and presum es that a health protective level f o r PM 2.5 also would be expected to provide a m easure of protection from DPM, a constituent part of PM 2.5 . The PM 2.5 standard of 15 

### : g/m 3 as an annual average thus is expected to provide a m easure of protection from DPM noncancer health effects, reflecting DPM’s current approxim a te proportion to PM 2.5 .

9.5.2. Evaluation of Cancer Risks As discussed in Section 9.4.3, the com b ined weight of evidence indicates that DE has the potential to pose a cancer hazard to hum ans at anticipated levels of environm ental exposure. The target organ of DE-induced carcinogenicity is the lung. Strong evidence exists for a causal relationship between risk for lung cancer and occupational exposure to DE in certain occupational workers such as railroad workers, truck drivers, heavy-equipm ent operators (e.g., shipyard, diesel farm equipm ent, and construction), and transit workers. The evidence, however, was less than sufficient to confidently characterize DE as carcinogenic to hum ans, and instead
the assessm ent concludes that DE is likely to be a hum an carcinogen. It also has been shown unequivocally in several studies that DE can cause benign and m a lignant lung tum o rs in rats in a
dose-related m a nner following chronic inhalation exposure to high concentrations; however, this response is not thought applicable to predict a hazard to hum ans exposed at lower environm ental
levels. The m echanism ( s) by which DE would induce lung cancer in hum ans has not been established, but available data suggest that m u tagenic and nonm utagenic m odes of action are possible. Hence, for estim ating DE cancer risk at low environm ental exposures, linear low-dose extrapolation would be considered an appropriate default assum p tion, which is consistent with EPA’s science policy position that in the absence of an understanding of m odes of carcinogenic

9-20action, a nonthreshold effect is to be presum ed (U.S. EPA, 1986, 1996a). This sam e assum p tion has been used by other organizations/risk assessors who have previously used either linear risk extrapolation m odels or m echanistically based m odels to estim ate cancer risk from
environm ental exposure to DE (e.g., W HO-IPCS, 1996; Cal EPA, 1998; also see Appendix C). Dose-response assessm ent is based on either hum an or anim al data, although hum an data are always preferred if available. Several quantitative assessm ents have been conducted by organizations and investigators on the basis of both occupational data and rat data (see Appendix C). However, m o re recent evidence indicates that DE causes tum o rs in the rat lung via a m ode
of action that involves im pairm e nt of lung clearance m echanism s (referred to as “lung overload response”) associated with high exposures. This lung overload response is not expected in
hum ans exposed to environm ental levels (nor is it expected to occur in m a ny occupational exposures), and thus the rat lung tum o r dose-response data are not considered suitable for predicting hum an risk at low environm ental exposures. Given that the rat data are not appropriate for estim ating cancer risk to hum ans, this assessm ent focuses on using the occupational epidem iologic data for estim ating environm ental risk of DE to hum ans. Even though occupational data are considered m o st relevant for use in dose-response

assessm ent, uncertainties exist, including the f o llowing issues:

• t he use of DPM (expressed as 

### : g/m 3 ) as a surrogate dosim eter for DE exposure, given that the relative roles of various constituents in m e diating carcinogenic effects
and the m ode of carcinogenic action are still unknown;

• t he representativeness of occupational populations for the general population and vulnerable subgroups, including infants and children and individuals with preexisting diseases, particularly respiratory conditions;

• t he lack of actual DE exposure data for workers in the available epidem iologic studies;

• possible confounders (sm oking and asbestos exposure) that could contribute to the observed lung cancer risk in occupational studies of DE if the control for these confounders is not adequate; and

• w hether or not an exposure-response relationship for occupational lung cancer risk
can be estim ated for DE.

Chapter 8, Section 8.3, provides a discussion of these uncertainties, along with an evaluation of the suitability of available occupational studies f o r a derivation of a cancer unit risk estim ate f o r DE. Unit risk is defined as the estim ated upper-bound cancer risk at a specific exposure or dose

9-21from a continuous average lifetim e exposure of 70 years (in this case, cancer risk per 

### : g/m 3 of DPM).
Am ong the occupational studies, the railroad worker studies (Garshick et al., 1987, 1988)
and the Team sters Union truck driver studies (Steenland et al., 1990, 1998) are considered to have the best available exposure data for possible use in establishing exposure-response relationships and deriving a cancer unit risk. There have been dif f e rent views on the suitability
of these studies for estim ating environm ental cancer risks (e.g., Cal EPA, 1998; HEI, 1995, 1999). Given the equivocal evidence for the presence or absence of an exposure-response relationship for the study of railroad workers, and exposure uncertainties for the study of truck drivers, it is judged that available data are too uncertain at this tim e for the developm ent of a confident quantitative dose-response analysis and subsequent derivation of cancer unit risk for
DE. In the absence of a cancer unit risk to assess population cancer risk, this assessm ent provides a “perspective” about the possible m a gnitude of risk in the population from
environm ental exposure to DE. One approach to estim ating the possible m a gnitude of risk
involves sim p ly noting that risks to the general public would be low in com p arison with occupational risk if the differences in the lower environm ental exposures com p ared to the higher occupational exposures are large. If the differences are sm all, the environm ental risks would approach the workers’ risk observed in studies of past occupational exposures. A com p arison of
environm ental equivalent occupational and am bient environm ental exposures showed that for certain occupations, there is a potential for overlap between environm ental exposure and the
estim ated environm ental equivalent occupational exposure, while in other cases the
environm ental exposures could be up to about 100-fold lower than the occupational levels (see Table 8-1). For the exposure overlap case, one can infer that the environm ental risk could be the

sam e as, or approach, the risk m a gnitudes observed in the occupational studies. In the 100-fold lower case, the environm ental risk could be about 100-fold lower than the observed risk

m a gnitudes in the occupational studies. Risks to the general public are of potential concern when a significant risk is seen in the occupational setting and the difference between occupational and am bient exposure m a y overlap or is relatively sm all (within one to two orders
of m a gnitude). A second approach, which is related to the first approach but m o re quantitative, is to
estim ate possible ranges of lung cancer risk from occupational exposures to DE, and then use a proportional relationship of exposure differences (e.g., EMs) to scale the occupational risk to the
environm ental exposure setting. Given the range of observed relative risks or odds ratios of lung
cancer in a num ber of occupational studies, a relative risk increase of 1.4 was selected as a
reasonable estim ate of occupational risk for the purpose of this analysis. The relative risk of 1.4

9-22m eans that the workers faced an extra risk that is 40% higher than the approxim a te 5% background lifetim e lung cancer risk in the U.S. population. Using the relationship [ excess risk = (relative risk-1) × background risk ], 2% of these DE-exposed workers (i.e., 10 -2 risk) would have been at risk (and developed lung cancer) attributable to occupational exposure to DE. Using a nationwide average environm ental exposure (0.8 

### : g/m 3 DPM), and assum i ng (a) the excess lung cancer risk from occupational exposure is about 10 -2 ; and (b) the past occupational exposures were no higher than about 1,200 

### : g/m 3 (equivalent to an environm ental equivalent EM of 315, connoting a relatively large EM), the environm ental cancer risk would
f a ll between 10 -4 and 10 -5

. The selection of 1,200 

### : g/m 3 is a very high value intentionally selected to illustrate a high-end exposure boundary and thus a lower bounding of risk calculated by this exploratory approach. On the other hand, if occupational exposures for som e groups were lower, for exam ple, closer to 100 

### : g/m 3 (equivalent to an environm ental equivalent EM of 26, connoting a sm aller EM), the environm ental risk would be higher and approach 10 -3

. The
selection of 100 

### : g/m 3 is purposefully toward the lower end of the reported occupational exposure range which spans 7–403 

### : g/m 3 in Table 8-1. The risk estim ates are attended by

num erous uncertainties; their inclusion in this docum ent does not constitute Agency
endorsem e nt of their validity as a surrogate for cancer unit risk; the range of values is not useful
for estim ating num bers of cancer cases; and the range of possible risk from environm ental exposures also could be lower and a zero risk cannot be ruled out. These types of exploratory analyses are not intended to be precise or provide a def i nitive
characterization of cancer risk but are usef ul in illustrating and gauging the possible range of risk based on applying reasonable judgm ent. The analyses provides a sense of where an upper lim it (or “upper bound”) of the risk m a y be. These analyses are subject to uncertainties, particularly the lack of actual exposure inform ation for the occupational epidem iologic studies and the use of public-health-conservative risk assessm ent assum p tions. The possible risks also could be lower and a zero risk cannot be ruled out because (a) som e individuals in the population m a y have a high tolerance to exposure from DE and therefore not be susceptible to cancer from
environm ental exposure, and (b) although not reported, there could be a threshold of exposure
below which there is no cancer risk. Given these circum stances, we refer to this risk analysis as a “perspective” on possible risks. Best scientific judgm ent guided the selection of assum p tions
and other elem ents of this analysis which are deem ed reasonable and appropriate for identifying possible risks based on the inform ation currently available. Further research is needed to m o re accurately assess and characterize environm ental cancer risks from DE.

9-239.6. SUMMARY AND CONCLUSIONS The available health effects data show that acute (short-term episodic exposure) and chronic (long-term ) exposure to DE can pose hazards to hum ans and that environm ental exposures, in som e cases, m a y have a risk. At relatively high acute exposures, DE can cause acute irritation to the eye and upper respiratory airways and sym p tom s of respiratory irritation which m a y be tem porarily debilitating. Evidence also shows that DE has im m unological toxicity that can induce allergic responses (som e of which are also typical of asthm a ) and/or exacerbate existing respiratory allergies. W h ile the hazard potential is im portant for these acute and short exposure-related effects, quantitative dose-response estim ates for these effects could not be developed because of the lack of exposure-response inform ation. It is concluded that long-term exposure to low levels of DE poses a hazard for chronic
inflam m a tion and pathological changes in the hum an lung. A level of hum an lifetim e exposure thought to be without appreciable risk for lung dam a ge is estim ated to be 5 

### : g/m 3 of DPM, this being a calculated RfC value for DE. Because DPM is a constituent of am bient PM 2.5 and there
is som e sim ilarity in potential adverse effects from DE and PM 2.5 , it is expected that a m easure of protection from health effects associated with DE is provided by the 1997 annual PM 2.5 NAAQS,
set at a level of 15 

### : g/m 3

. DE is considered to pose a hum an lung carcinogenicity hazard, which is expressed in a weight-of-evidence conclusion that DE is judged to be a “probable” hum an carcinogen, or is “likely to be carcinogenic in hum ans by inhalation” at environm ental or higher exposure
conditions. Because of uncertainty in the available exposure-response data, a cancer unit risk/cancer potency for DE has not been derived. One should note that the closeness of the highend environm ental exposures and low-end estim ates of occupational exposure suggest less uncertainty in the extrapolation of hazard and possible risk to the environm ental setting. Exploratory analyses using public health conservative assum p tions provides a perspective on the possible range of lung cancer risk from environm en tal exposure to DE. Best scientific judgm ent guided the selection of assum p tions and other elem ents of this analysis which are deem ed reasonable and appropriate for identifying possible risks based on the inform ation currently
available. These analyses indicate that lifetim e cancer risk m a y exceed 10 -5 and could be as high
as 10 -3 or nearly so, though considering the assum p tions used and the uncertainties, lower risk is possible and a zero risk cannot be ruled out. This range of values is attended by num erous uncertainties, the inclusion of the range in this assessm ent does not constitute Agency
endorsem e nt of their validity as surrogates for cancer unit risk values, and the range is not
suitable for estim ating num bers of cancer cases. These risk findings should not be viewed as a
def i nitive characterization of risk.

9-24Even though the evidence for potential hum an health hazards for DE is convincing and persuasive, uncertainties exist because of the use of assum p tions to bridge data and knowledge gaps about hum an exposures to DE and the underlying m echanism s by which DE m a y cause the
observed toxicities in hum ans and anim als. A notable uncertainty of this assessm ent is how the physical and chem ical nature of DE em issions has changed over the years because the toxicological and epidem iologic observations are based on older engines and their em issions, yet the desire is to focus on the potential health hazards related to exposure from present-day or
future em issions. There have been changes in the physical and chem ical com position of som e DE em issions (on-road vehicle em issions) over tim e, though there is no definitive inform ation to
show that the em ission changes portend significant toxicological changes. The m ode(s) of action for DE toxicity in hum ans is not understood, and hence knowledge is lacking about the role of
exhaust m i xture com ponents in m odulating the toxicity. Taken together, these considerations have lead to a judgm ent that the hazards identified from older technology-based exposures are applicable to current-day exposures. As new and cleaner diesel engines replace a substantial

num ber of existing engines, the general applicability of the conclusions in this assessm ent will
need to be reevaluated.
Other uncertainties include the assum p tions that health effects observed at high doses

m a y be applicable to low doses, and that toxicologic findings in laboratory anim als are predictive of hum an responses. Also, the available data are not sufficient to dem onstrate the absence or presence of an exposure/dose-response threshold in hum ans for DE toxicity at
environm ental exposures. Again, this is due in part to the lack of understanding of how DE m a y cause adverse health effects in exposed hum ans and laboratory anim als. Although there are hypotheses about the specific m echanism s by which DE m i ght cause cancer and other toxicities, no specif i c biological pathways or specif i c constituents of DE have been f i rm ly established as responsible for low-dose effects. The assum p tions used in this assessm ent, i.e., the presence of a biological threshold for chronic respiratory effects based on cum u lative dosage and the absence of a threshold for lung cancer stem m i ng from subtle and irreversible effects, are considered prudent and reasonable default choices.
The characterization of health hazards and risks contained in this docum ent assum e s that the potential DE health hazards are relevant for long-term exposures, up to and including
lifetim e exposures, and would apply to a wide sp ectrum of individuals but not necessary those that would have signif i cant dif f e rential susceptibility. There is no DE-specif i c inf o rm ation that provides direct insight into the question of dif f e rential susceptibility within the general hum an population or vulnerable subgroups, for exam ple, children or the elderly. Although default approaches to account for interindividual variation have been included in the derivation of the noncancer effects RfC (i.e., use of an uncertainty factor of 10), this m a y or m a y not adequately

9-25protect certain subgroups that could be m o re vulnerable. Differential susceptibility to DPM

am ong individuals in the population would be due to differences in dosim etry (i.e., differences in retained particle m a ss or num ber in the lung) and/or differences in respiratory system tissue response sensitivity. From the dosim etry perspective, we understand that age, gender, and disease status can influence deposition in the lung and other areas of the respiratory tract (U.S. EPA, 1996b, Section 10.7.7). For exam ple, given that DE chronic toxicity is focused on the respiratory system , vulnerable subgroups m i ght in clude those individuals who predispose their lungs to increased particle retention (e.g., sm oking, high particulate burdens from nondiesel sources) or those having existing respiratory or lung inflam m a tion, repeated respiratory infections, or chronic bronchitis or asthm a

. For children, there is also the hypothesis of possible increased sensitivity to exposure, given the ongoing processes of developm ent from birth to

m a turation, of the respiratory and im m une system s. Despite the uncertainties regarding intraspecies variability, the def a ult approach of using an uncertainty factor of 10 in the derivation of the noncancer effects RfC to account for possible interindividual variation in the toxic response to DE exposure is appropriate and reasonable given the lack of DE-specific data. Variation in DE exposure is another source of uncertainty. Because of variation in
hum an activity patterns and their proxim ity to DE sources of em issions, different population subgroups could potentially receive higher or lower exposure to DE. The highest exposed are clearly occupational subgroups whose job brings them very close to DE sources, such as diesel engine vehicle drivers and workers, diesel powered m achinery operators, som e underground

m i ners, etc. High exposures in the general population would be to those living very near or having tim e outdoors in proxim ity to DE sources as well as those engaged in activities that cause high breathing rates where DE is present. Accordingly, where appropriate, analyses in this

assessm ent have included possible high-end DE exposures in addition to the lower nationwide average exposure estim ates. Lastly, this assessm ent considers only potential heath effects from exposures to DE alone. DE exposure could be additive or synergistic to concurrent exposures to other air pollutants. For exam ple, there is evidence that DPM that has been altered by being in the presence of am bient ozone significantly increases the rat lung inflam m a tory effect com p ared to DPM that was not subjected to ozone (Madde n et al., 2000). This observation suggests a hypothesis that inflam m a tion-related noncancer hazards of airborne DPM m a y be worsened by the increasing presence of ozone in the am bient air. Other concerns include the possible
im pacts for children and adults on the exacerbation of existing allergens resulting from DE

exposure. However, in the absence of m o re definitive data dem onstrating interactive effects

9-26from com b ined exposures to DE and other pollutants, it is not possible to further address these
issues at this tim e.

REFERENCES FOR CHAPTER 9

California Environm ental Protection Agency-OEHHA (Cal EP A). (1998) Part B: Health risk assessm ent for diesel exhaust, Public and Scientific Review Draft. February 1998.

Garshick, E; Schenker, MB; Munoz, A; et al. (1987) A case-control study of lung cancer and diesel exhaust exposure in railroad workers. Am Rev Respir Dis 135:1242-1248.

Garshick, E; Schenker, MB; Munoz, A; et al. (1988) A re trospective cohort study of lung cancer and diesel exhaust exposure in railroad workers. Am Rev Respir Dis 137:820-825.

Health Effects Institute (HEI). (1995) Diesel exhaust: a cr itical analysis of em issions, exposure, and health effects. Cam b ridge, MA: Health Effects Institute.

HEI. (1999) Diesel em issions and lung cancer: epidem iol ogy and quantitative risk assessm ent. A special report of the Institute’s Diesel Epidem iology Expert Pane l. Cam b ridge, MA: Health Effects Institute.

Heinrich, U; Fuhst, R; Rittinghausen, S; et al. (1995) Chronic inhalation expos ure of W i star rats and two different strains of m i ce to diesel engine exhaust, car bon black, and titanium dioxide. Inhal Toxicol 7:553-556.

International Agency for Research on Cancer (IARC). (1989) Diesel and gasoline engine exhausts and som e nitroarenes. IARC m onographs on the evaluation of carcinoge nic risks to hum ans: v. 46. Lyon, France: W o rld Health Organization; pp. 41-185.

International Life Sciences Institute (IL SI). (2000) ILSI Risk Science Institu te workshop: the relevance of the rat lung response to particle overload for hum an risk a ssessm ent. Gardner, DE, ed . Inhal Toxicol 12(1-2).

International Program m e on Chem ical Safety (IPCS), W o rl d Health Organization. (1996) Diesel fuel and exhaust

em issions. Environm ental Health Criteria 171. Geneva: W o rld Health Organization.

Ishinishi, N; Kuwabara, N; Takaki, Y; et al. (1988) Long-term inhalation expe rim e nts on diesel exhaust. In: Diesel exhaust and health risks. Results of the HERP studies. Ibaraki, Japan: Research Com m ittee for HERP Studies; pp. 11-84.

Madden, M; Richards, J; Dailey, L; et al. (2000) Effect of ozone on diesel exha ust toxicity in rat lung. Toxicol Appl Pharm acol 168:140-148.

Mauderly, JL; Jones, RK; Griffith, W C ; et al. (1987) Di esel exhaust is a pulm onary carcinogen in rats exposed chronically by inhalation. Fundam Appl Toxicol 9:208-221.

National Toxicology Program (NTP). (2000) 9 th report on carcinogens. Public Health Service, U.S. Departm e nt of Health and Hum a n Services, Research Triangle Park, NC. Available from : http://ntp-server.niehs.nih.gov.

Nikula, KJ; Snipes, MB; Barr, EB; et al. (1995) Co m p arative pulm onary toxicities and carcinogenicities of chronically inhaled diesel exhaust and carbon bl ack in F344 rats. Fundam Appl Toxicol 25:80-94.

Steenland, K; Silverm a n, DT; Hornung, RW . (1990) Case-control study of lung cancer and truck driving in the Team sters Union. Am J Public Health 80:670-674.

9-27Steenland, K; Deddens, J; Stayner, L. (1998) Diesel e xhaust and lung cancer in the trucking industry: exposureresponse analyses and risk assessm ent. Am J Ind Med 34:220-228.

U.S. Environm ental Protection Agency (EPA). (1986) Guid elines for carcinogen risk assessm ent. Federal Register 51(185):33992-34003.

U.S. EPA. (1996a) Proposed guidelines for carcinogen risk assessm ent. Office of Research and Developm ent. Federal Register 61(79):17960-18011. EPA/600/P-92/003C.

U.S. EPA. (1996b) Air quality criteria for particulate m a tter. National Center for Environm ental Assessm ent; Research Triangle Park, NC: EPA/600/P-95/001aF.

U.S. EPA. (1999) Guidelines for Carcinogen Risk A ssessm ent. Review Draft. NCEA-F-0644, July. Risk Assessm ent Forum
, W a shington, DC. http://www.epa.gov/ncea/raf/cancer.htm

. U.S. EPA. (2000) National air pollutant em ission trends, 1900-1998. EPA-454/R-00-002, March.

W o rld Health Organization (W HO-IPCS). (1996) Diesel fuel and exhaust em issions. Environm ental Health Criteria 171, W HO-IPCS, Geneva, Switzerland.

Yu, CP; Yoon, KJ; Chen, YK. (1991) Retention m odeling of DE particles in rats and hum ans. J Aerosol Med 4:79-115.

9-28

## Appendix A

Calculation of Human Equivalent Continuous Exposure Concentrations (HECs)

A-1

### A.1. INTRODUCTION As discussed in Chapter 3, the lung burden of diesel particulate m a tter (DPM) during exposure is determ ined by both the am ount and site of particle deposition in the lung and, subsequently, by rates of translocation and clearance from the deposition sites. Mathem atical

m odels have often been used to com p lem e nt experim e ntal studies in estim ating the lung burdens of inhaled particles in different species unde r different exposure conditions. This appendix presents a m a them atical m odel that sim u lates the deposition and clearance of DPM in the lungs
of rats and hum ans of Yu et al.(1991) also published as Yu and Yoon (1990). Diesel particles are aggregates form ed from prim ary spheres 15-30 nm in diam eter. The aggregates are irregularly shaped and range in size from a few m o lecular diam eters to tens of

m i crons. The m a ss m e dian aerodynam i c diam eter (MMAD) of the aggregates is typically 0.2

μm and is polydisperse with a geom etric st andard deviation of around 2.3. The organics adsorbed onto the aggregates norm a lly account for 10% to 30% of the particle m a ss. However, the exact size distribution of DPM and the specific com position of the adsorbed organics depend

upon m a ny factors, including engine design, fuel s used, engine operating conditions, and the
therm odynam i c process of exhaust. The physical and chem ical characteristics of DPM have been reviewed extensively by Am ann and Siegla (1982) and Schuetzle (1983).
Four m echanism s deposit DPM within the respiratory tract during exposure: im paction,
sedim e ntation, interception, and diffusion. The contribution from each m echanism to deposition, however, depends upon lung structure and size, the breathing condition of the subject, and particle size distribution. Under norm a l breat hing conditions, diffusion is the m o st dom inant

m echanism and the other three m echanism s play m i nor roles. Once DPM is deposited in the respiratory tract, both the carbonaceous core and the adsorbed organics will be rem oved f r om the deposition sites by m echanical clearance, provided by m u cociliary transport in the ciliated conducting airways as well as m acrophage phagocytosis
and m i gration in the nonciliated airways, and dissolution. As the carbonaceous core or soot of DPM is insoluble, it is rem oved from the lung prim arily by m echanical clearance, whereas the adsorbed organics are rem oved principally by dissolution (Chapter 3).



### A.2. PARTICLE MODEL To develop a m a them atical m odel that sim u lates the deposition and clearance of DPM in the lung, an appropriate m odel for diesel particles m u st be introduced. For the deposition study, an equivalent sphere m odel developed by Yu and Xu (1987) was used to sim u late the dynam i cs and deposition of DPM in the respiratory tract by various m echanism s

. For the clearance study, a diesel particle is assum e d to be com posed of three different m a terial com ponents according to
their characteristic clearance rates: (1) a carbonaceous core of approxim a tely 80% of the particle

A-2m a ss; (2) absorbed organics of about 10% of pa rticle m a ss, which are slowly cleared from the lung; and (3) adsorbed organics quickly cleared from the lung, accounting for the rem a ining 10% of particle m a ss. The presence of two discrete organic phases in the particle m odel is suggested by observations that the rem oval of particle-associated organics from the lung exhibits a biphasic clearance curve (Sun et al., 1984; Bond et al., 1986), as discussed in Chapter 3. This curve represents two m a jor kinetic clearance phenom ena: a fast- phase organic washout with a halftim e of a few hours, and a slow phase with a half-tim e that is a few hundred tim es longer. The
detailed com ponents involved in each phase are not known. It is possible that the fast phase consists of organics that are leached out prim arily by diffusion m echanism s while the slow phase

m i ght include any or all of the following com ponents: (a) organics that are “loosened” before they are released, (b) organics that have becom e intercalated in the carbon core and whose
release is thus im peded, (c) organics that are associated for longer periods of tim e because of hydrophobic interaction with other organic-phase m a te rials, (d) organics that have been ingested by m acrophages and as a result effectively rem a in in the lung for a longer period of tim e because
of m e tabolism by the m acrophage (m etabolites form ed m a y interact with other cellular

com ponents), and (e) organics that have directly acted on cellular com ponents, such as the
form ation of covalent bonds with DNA and other biological m acrom olecules to form adducts. The above distinction of the organic com ponents is general and m a de to account for the biphasic clearance of DPM; it does not specifically im ply the actual nature of the adsorbed organics. For aerosols m a de of pure organics, such as benzo( a )pyrene (BaP) and nitropyrene (NP) in the sam e size range of DPM, Sun et al. (1984) and Bond et al. (1986) observed a nearly

m onophasic clearance curve. This m i ght be explained by the absence of intercalative phenom ena (a) and of hydrophobic interaction im posed by a heterogeneous m i xture of organics (b). The m easurem ent of a pure organic m i ght also neglect that quantity which has becom e intracellularly (c) or covalently bound (d).



### A.3. COMPARTMENTAL LUNG MODEL

The m odel of Yu et al. (1991) com p rises three principal com p artm ents involved in deposition and clearance: tracheobronchial (T or TB), alveolar (A), and lung-associated lym p h node (L), as shown in Figure A-1. The outside com p artm ents blood (B) and GI tract (G) and nasopharyngeal or head (H) are also represented. The alveolar com p artm ent in the m odel is obviously the m o st im portant for long-term retention studies. However, for short-term consideration, retentions in other lung com p artm ents m a y also be significant. The presence of these lung com p artm ents and the two outside com p artm ents in the m odel therefore provides a

com p lete description of all clearance processes involved.

A-3In Figure A-1, r (i ) T , and r (i ) H

, r (i ) A are, respectively, the m a ss deposition rates of DE m a terial

com ponent i (i=1 [core], 2 [slowly cleared organi cs], and 3 [rapidly cleared organics]) in the head, tracheobronchial, and alveolar com p artm ents; and λ (i ) XY

represents the transport rate of

m a terial com ponent i from any com p artm ent X to any com p artm ent Y. Let the m a ss fraction of

m a terial com ponent i of a diesel particle be ƒ i . Then

r ( i )

' f i r H , (A-1) H

r T

( i )

' f i r T , (A-2)

r

A ( i )

' f i r A , (A-3)

where r H , r T , and r A are, respectively, the total m a ss deposition rates of DPM in the H, T, and
A com p artm ents, determ ined from the equations:

r

H ' c ( TV )( RF )(

DF ) H , (A-4)

r T ' c ( TV )( RF )(

DF ) T , (A-5)

r

A ' c ( TV )( RF )(

DF ) A . (A-6)

In Equations A-4 to A-6, c is the m a ss concentration of DPM in the air, TV is the tidal
volum e, RF is the respiratory frequency, and (DF) H , (DF) T , and (DF) A are, respectively, the deposition fractions of DPM in the H, T, and A com p artm ents over a respiratory cycle. The

A-4values of (DF) H , (DF) T , and (DF) A , which vary with the particle size, breathing conditions, and lung architecture, were determ ined from the deposition m odel of Yu and Xu (1987). The differential equations for m (i ) XY

, the m a ss of m a terial com ponent i in com p artm ent X as a function of exposure tim e t, can be written as

Head (H)

Tracheobronchial (T)

Alveolar (A)

Lym ph nodes (L) dm T ( i )

dt

dm A ( i )

dt dm ( i ) H

' r ( i )

& λ ( i ) H HB m ( i )

, (A-7) dt H HG m ( i )

& λ ( i ) H

' r T ( i ) % λ ( i ) TG m T ( i )

& λ ( i ) Α T m A ( i )

& λ ( i ) TB m T ( i )

, (A-8)

' r A ( i )

& λ ( i ) AL m A ( i )

& λ ( i ) AT

m A ( i )

& λ ( i ) AB m A ( i )

, (A-9)

dm L ( i )

' λ ( i ) LB m L ( i )

. (A-10) AL m A ( i )

& λ ( i ) dt

Equation A-9 m a y also be written as

A m A ( i )

, (A-11) dm A ( i )

' r A ( i )

& λ ( i ) dt

where

λ ( i )

' λ ( i ) % λ ( i ) % λ ( i )

. (A-12) A AT AL AB

is the total clearance rate of m a terial com ponent i from the alveolar com p artm ent. In Equations A-7 to A-10, we have assum e d vanishing m a terial concentration in the blood com p artm ent to calculate diffusion transport.
The total m a ss of the particle-associated organics in com p artm ent X is the sum of m (2 ) X and m (3 ) X the total m a ss of DPM in com p artm ent X is equal to

m X ' m X (1 ) % m X (2 ) % m X (3 ) (A-13)

A-5< The lung burdens of diesel soot (core) and organics are defined, respectively, as

m (1 )

Lung ' m T (1 ) % m A (1 )

, (A-14)

and

m (2 ) % (3 )

' m T (2 ) % m A (2 ) % m T (3 ) % m A (3 )

. (A-15) Lung

Because the clearance of diesel soot from com p artm ent T is m u ch faster than from com p artm ent A, m (1 ) A T < m (1 ) a short tim e after exposure, Equation A-14 leads to

m (1 )

Lung – m A (1 )

. (A-16)

Solution to Equations A-7 to A-10 can be obtained once all the transport rates λ (i ) XY

are known. W h en λ (i ) are constant, which is the case in linear kinetics, Equations A-7 to A-10 will XY have a solution that increases with tim e at the beginning of exposure but eventually saturates and reaches a steady-state value. This is the classical retention m odel developed by the International
Com m i ssion of Radiological Protection (ICRP, 1979). However, as discussed in Chapter 3, data have shown that when rats are exposed to DPM at high concentration for a prolonged period, long-term ed clearance is im paired. This is the so-called overload effect, observed also for other insoluble particles. The overload effect cannot be predicted by the classical ICRP m odel.
Soderholm (1981) and Strom et al. (1987, 1988) have proposed a m odel to sim u late this effect by adding a separate sequestering com p artm ent in the alveolar region. In the present approach, a single com p artm ent for the alveolar region of the lung is used and the overload effect is accounted for by a set of variable transport rates λ (i ) AL , and λ A (i ) which are functions of m A . The AT

, λ (i )
transport rates λ A (i ) and λ (i ) AL in Equations A-7 to A-10 can be determ ined directly from experim e ntal data on lung and lym ph node burdens, and λ (i ) AB from Equation A-12. AT

and λ (i )



### A.4. SOLUTIONS TO KINETIC EQUATIONS Equation A-11 is a nonlinear differential equation of m A (i ) with known function of λ A (i

. ) For diesel soot, this equation becom e s

A-6dm A (1 )

' r (1 )

& λ (1 ) A A ( m A ) m A (1 )

. (A-17) dt

Because clearance of the particle-associated organics is m u ch faster than diesel soot, m (2 ) and m (3 ) A A constitute only a very sm all f r action of the total particle m a ss (less than 1%) af ter a long exposure, and we m a y consider λ (1 ) A alone. Equation A-17 is then reduced to a A as a function of m (1 )
dif f e rential equation with m (1 ) A the only dependent variable. The general solution to Equation A-17 for constant r (1 ) A at any tim e, t, can be obtained by the separation of variables to give

0 m (1 ) dm (1 )
A

r (1 ) A

A m A (1 ) ' t

. (A-18)

A & λ (1 )

If r (1 ) A is an arbitrary function of t, Equation A-17 needs to be solved num erically such as by a Runge-Kutta m e thod. Once m (1 ) A is found, the other kinetic equations A-7 to A-10 for both diesel soot and the particle-associated organics can be solved readily, as they are linear equations. The solutions to these equations for constant r (i ) T , and r A (i ) are given below: H , r (i ) Head (H)

m ( i )

' r ( i ) H H0 H )/ λ ( i ) H H H / λ ( i ) % ( m ( i )

& r ( i ) H ) e xp (

& λ ( i ) t ) (A-19)

where λ ( i )

' λ ( i ) HB (A-20) H HG % λ ( i )

A-7Tracheobronchial (T)

t m T ( i )

' exp (

& λ ( i ) t )

0 ( r T ( i ) % λ ( i ) m A ( i ) ) e xp ( λ ( i ) t ) dt % m ( i ) (A-21) T AT T T0

where λ ( i )

' λ ( i ) % λ ( i ) (A-22) T TG TB

Lym ph nodes (L)

t

LB 0 AL m A ( i ) exp( λ ( i ) t ) dt % m ( i ) m L ( i )

' exp(

& λ ( i ) t ) λ ( i ) LB L0 ) (A-23)

In Equations A-19 to A-23, m (i ) XO represents the value of m X (i ) at t = 0. In the sections to follow, the m e thods of determ ining r (i ) T , and r (i ) H , r (i ) A , or (DF) H , (DF) T , and

H , r (D F ) XY in the com p artm ental lung m odel are (DF) A r (D F ) T , and r (D A F ) as well as the values of λ (i ) presented.



### A.5. DETERMINATION OF DEPOSITION FRACTIONS

The m a them atical m odels for determ ining the deposition fractions of DPM in various regions of the respiratory tract have been developed by Yu and Xu (1986, 1987) and are adopted in this report. Yu and Xu consider DPM as a polydisperse aerosol with a specified m a ss m e dian aerodynam i c diam eter (MMAD) and geom etrical standard deviation σ g . Each diesel particle is represented by a cluster-shaped aggregate within a spherical envelope of diam eter d
e . The envelope diam eter d
e is related to the aerodynam i c diam eter of the particle by the relation

d

e ' φ & 1/2 ( C
a ) 1/2) ( ζ ) 1/2 (A-24) d

a C
e ζ

o

where 

### ζ is the bulk density of the particle in g/cm 3

, ζ 0 = 1 g/cm 3 ; φ is the packing density, which is the ratio of the space actually occupied by prim ary particles in the envelope to the overall envelope volum e; and C

x is the slip factor given by the expression:

A-8C
x ' 1 % 2 λ [1.257 % 0.4 exp

& ( 0.55 d
x )] (A-25) d

x λ

in which λ – 8 × 10 -6 cm 3 is the m ean free path of air m o lecules at standard conditions. In the diesel particle m odel of Yu and Xu (1986), ζ has a value of 1.5 g/cm 3 and a φ value of 0.3 is chosen based upon the best experim e ntal estim ates. As a result, Equation A-24 gives d
e /d
a = 1.35. In determ ining the deposition fraction of DPM, d
e is used for diffusion and interception according to the particle m odel.

A.5.1. Deposition in the Head Particle deposition in the naso- or oropharyngeal region is referred to as head or extrathoracic deposition. The am ount of particles that enters the lung depends upon the breathing m ode. Norm ally, m o re particles are collected via the nasal route than by the oral route
because of the nasal hairs and the m o re com p lex air passages of the nose. Since the residence
tim e of diesel particles in the head region during inhalation is very sm all (about 0.1 s for hum an adults at norm a l breathing), dif f u sional deposition is insignif i cant and the m a jor deposition

m echanism is im paction. The following em pirical form ulas derived by Yu et al. (1981) for
hum an adults are adopted for deposition prediction of DPM:
For m outh breathing:

( DF ) H

, in ' 0, for d
a 2 # 3000

( DF ) H

, in ' & 1.117 % 0.324 log( d
a 2 Q ), for d
a 2 Q > 3000

( DF ) H

, ex ' 0,

and for nose breathing:

( DF ) H

, in ' & 0.014 % 0.023 log( d
a 2 Q ), for d
a 2 Q # 337 (A-26)

(A-27)

(A-28)

(A-29)

A-9( DF ) H

, in ' & 0.959 % 0.397 log( d
a 2 Q ), for d
a 2 Q > 337 (A-30)

( DF ) H

, ex ' 0.003 % 0.033 log( d
a 2 Q ), for d
a 2 Q # 215 (A-31)

( DF ) H

, ex ' & 0.851 % 0.399 log( d
a 2 Q ), for d
a 2 Q > 215 (A-32)

where (DF) H is the deposition efficiency in the head, the subscripts in and ex denote inspiration and expiration, respectively, d
a is the particle aerodynam i c diam eter in μm , and Q is the air
f l owrate in cm 3 /sec.
Form ulas to calculate deposition of diesel particles in the head region of children are
derived from those for adults using the theory of sim ilarity, which assum e s that the air passage in the head region is geom etrically sim ilar for all ages and that the deposition process is characterized by the Stokes num ber of the particle. Thus, the set of em pirical equations from A-26 through A-32 are transform e d into the following form : For m outh breathing:

( DF ) H

, in ' 0, for d
a 2 Q # 3000 (A-33) and for nose breathing:

( DF ) H

, in ' & 1.117 % 0.972 logK % 0.324 log( d
a 2 Q ), for d
a 2 Q > 3000 (A-34)

( DF ) H

, ex ' 0. (A-35)

( DF ) H

, in ' & 0.014 % 0.690 log K % 0.023 log( d
a 2 Q ), for d
a 2 Q # 337 (A-36)

( DF ) H

, in ' & 0.959 % 1.191 log K % 0.397 log ( d
a 2 Q ), for d
a 2 Q > 337 (A-37)

( DF ) H

, ex ' 0.003 % 0.099 log K % 0.033 log( d
a 2 Q ), for d
a 2 Q # 215 (A-38)

A-10where K is the ratio of the linear dim e nsion of the air passages in the head region of adults to that
of

children, which is assum e d to be the sam e as the ratio of adult/child tracheal diam eters.

( DF ) H

, ex ' 0.851 % 1.197 log K % 0.399 log( d
a 2 Q ), for d
a 2 Q >215 (A-39)

For rats, the following em pirical equations are used for deposition prediction of DPM in
the nose:

( DF ) H

, in ' ( DF ) H

, ex ' 0.046 % 0.009 log( d
a 2 Q ), for d
a 2 Q # 13.33 (A-40)

A.5.2. Deposition in the Tracheob ronchial and Alveolar Regions The deposition m odel adopted for DPM is the one previously developed for

m onodisperse (Yu, 1978) and polydisperse spherical aerosols (Diu and Yu, 1983). In the m odel,

( DF ) H

, in ' ( DF ) H

, ex ' & 0.522 % 0.514 log( d
a 2 Q ), for d
a 2 Q > 13.33 (A-41)

the branching airways are viewed as a cham ber m odel shaped like a trum pet (Figure A-2). The
cross-sectional area of the cham ber varies with airway depth, x, m easured from the beginning of
the trachea. At the last portion of the trum pet, additional cross-sectional area is present to
account for the alveolar volum e per unit length of the airways. Inhaled diesel particles that escape capture in the head during inspiration will enter the trachea and subsequently the bronchial airways (com partm e nt T) a nd alveolar spaces (com partm e nt A).
Assum i ng that the airways expand and contract uniform ly during breathing, the equation for the conservation of particles takes the form :

β (

A 1 % A 2 ) M c % Q M c

' & Qc η (A-42) M

x M x

where c is the m ean particle concentration at a given x and tim e t; A 1 and A 2 are, respectively,
the sum m e d cross-sectional area (or volum e per unit length) of the airways and alveoli at rest; η is the particle uptake efficiency per unit length of the airway; β is an expansion factor, given by:

β ' 1 % V t (A-43) V

l

A-11A-12 (A-44)

(A-45)

(A-46)

(A-47)

(A-48)

(A-49) and Q is the air flow rate, varying with x and t according to the relation

where Q 0 is the air flow rate at x = 0. In Equations A-43 and A-44, V t is the volume of new air in the lungs and V

x and V R are, respectively, the accumulated airway volume from x = 0 to x, and total airway volume at 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 1 and x 2 from time t 1 to t 2 can then be found from the expression

For diesel particles, 

### 0 is the sum of those due to the individual deposition mechanisms described above, i.e.,
where 

### 0 I , 

### 0 S , 

### 0 P , and 

### 0 D are, respectively, the deposition efficiencies per unit length of the

airway due to impaction, sedimentation, intercep tion, and diffusion. On the basis of the particle model described above, the expressions for 

### 0 I , 

### 0 S , 

### 0 P , and 

### 0 D are obtained in the following form:A-13 (A-50)

(A-51)

(A-52)

(A-53)

(A-54)

(A-55) for Reynolds numbers of the flow smaller than 2000, and for Reynolds numbers greater than or equal to 2000, where ST=d 2

a u/(18 

### : R) is the particle Stokes number, 

### 2 = L/(8R), 

### 0 = 3 

### : u s

L/(32uR), 

### ' = d
e /R, and 

### ) = DL/(4R 2 u) . In the above definitions u is the air velocity in the airway; 

### : is the air viscosity; L and R are, respectively, the length and radius of the airway; u s = C

a d

2

a

/(18 

### : ) is the particle settling velocity; and D = C
e kT(3 

### B: d

e ) is the diffusion coefficient with k denoting the Boltzmann constant and T the absolute temperature. In the deposition model, it is also assumed that 

### 0 I and 

### 0 P = 0 for expiration, while 

### 0 D and 

### 0 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:

where E D and E S are, respectively, the deposition efficiencies due to the individual mechanisms of diffusion and sedimentation over the pause period. The expression for E D and E S are given by

where 

### J D = D 

### J /R 2 in which 

### J is the pause time and 

### " 1 , 

### " 2 , and 

### " 3 are the first three roots of the equation:

in which J
o is the Bessel function of the zero th order, and:and

E S ' 1

& 0.0069 τ & 1

& 0.0859 τ & 2

& 0.0582 τ & 3 S S S , (A-56) for τ S > 1
,

where τ S = u S τ /2R. The values of (DF) T and (DF) A over a breathing cycle are calculated by superim posing DF for inspiration, deposition efficiency E dur ing pause, and DF for expiration in the tracheobronchial airways and alveolar space. It is assum e d that the breathing cycle consists of a constant flow inspiration, a pause, and a constant flow expiration, each with a respective duration fraction of 0.435, 0.05, and 0.515 of a breathing period.

A.5.3. Lung Models Lung architecture affects particle deposition in several ways: the linear dim e nsion of the airway is related to the distance the particle tr avels before it contacts the airway surface; the air flow velocity by which the particles are transported is determ ined by the cross-section of the airway for a given volum etric flowrate; and flow characteristics in the airways are influenced by the airway diam eter and branching patterns. Thus, theoretical prediction of particle deposition depends, to a large extent, on the lung m odel chosen.

A.5.3.1. Lung Model for Rats Morphom etric data on the lung airways of rats were reported by Schum and Yeh (1979). Table A-1 shows the lung m odel data for Long Evans rats with a total lung capacity of
13.784 cm 3

. Application of this m odel to Fischer rats is accom p lished by assum i ng that the rat
has the sam e lung structure regardless of its strain and that the total lung capacity is proportional to the body weight. In addition, it is also assum e d that the lung volum e at rest is about 40% of the total lung capacity and that any linear dim e nsion of the lung is proportional to the cubic root of the lung volum e.

A.5.3.2. Lung Model for Human Adults The lung m odel of m a ture hum an adults used in the deposition calculation of DPM is the

sym m e tric lung m odel developed by W e ibel (1963). In W e ibel’s m odel, the airways are

assum e d to be a dichotom ous branching system with 24 generations. Beginning with the 18th generation, increasing num bers of alveoli are present on the wall of the airways, and the last three generations are com p letely aleveolated. Thus, the alveolar region in this m odel consists of

A-14all the airways in the last seven generations. Table A-2 presents the m o rphom etric data of the airways of W eibel’s m odel adjusted to a total lung volum e of 3000 cm 3

.

A.5.3.3. Lung Model for Children The lung m odel for children in the diesel study was developed by Yu and Xu (1987) on
the basis of available m o rphom etric m easurem ents. The m odel assum es a lung structure with
dichotom ous branching of airways, and it m atches W eibel’s m odel for a subject when evaluated at the age of 25 years, the age at which the lung is considered to be m ature. The num ber and size of airways as functions of age t (years) are determ ined by the following equations.

A.5.3.3.1. Number of airways and alveoli. The num ber of airways N i (t) at generation i for age t is given by

N i ( t ) ' 2 i

,

N

21 ( t ) ' N
r ( t ),
N

22 ( t ) ' N 23 ( t ) ' 0.

N

21 ( t ) ' 2 21

, N

22 ( t ) ' N
r ( t )

& 2 21

, N

23 ( t ) ' 0,

N

21 ( t ) ' 2 21

, N

22 ( t ) ' 2 22

, N

23 ( t ) ' N
r ( t )

& 2 21

& 2 22 f or 0 # i # 20 (A-57)

for N
r ( t ) # 2 21 (A-58)

f or 2 21 < N
r ( t ) # 2 22 (A-59)

for N
r ( t )>2 21 % 2 22

, (A-60)

where N
r (t) is the total num ber of airways in the last three airway generations. The em pirical equation for N
r which best f its the available data is

Thus, N
r (t) increases from approxim ately 1.5 m illion at birth to 15 m illion at 8 years of age and

N

r ( t ) ' 2.036 × 10 7 (1

& 0.926 e & 0. 15 t ), t # 8 (A-61) 1.468 × 10 7

, t >8

rem ains nearly constant thereafter. Equations A-58 to A-60 also im ply that in the last three

A-15generations, the airways in the subsequent generation begin to appear only when those in the preceding generation have com p leted developm ent. The num ber of alveoli as a function of age can be represented by the following equation according to the observed data:

N

A ( t ) ' 2.985 × 1 0 8 (1

& 0.919 e & 0. 45 t ) (A-62)

The num ber of alveoli distributed in the unciliated airways at the airway generation level
is determ ined by assum i ng that alveolization of airways takes place sequentially in a proxim a l
direction. For each generation, alveolization is considered to be com p lete when the num ber of alveoli in that generation reaches the num ber determ ined by W e ibel’s m odel.

A.5.3.3.2. Airway size

. Four sets of data are used to determ ine airway size during postnatal growth: (a) total lung volum e as a function of age; (b) airway size as given by W e ibel’s m odel; (c) the growth pattern of the bronchial airways; a nd (d) variation in alveolar size with age. From these data, it is found that the lung volum e, LV(t) at age t, norm a lized to W e ibel’s m odel at 4800

cm 3 for an adult (25 years old), follows the equation

LV ( t ) ' 0.959 × 1 0 5 (1

& 0.998 e & 0. 002 t ) ( cm 3 ). (A-63)

The growth patterns of the bronchial airways are determ ined by the following equations

D i ( t )

& D iw ' α i [ H ( t )

& H (25)], (A-64)

L i ( t )

& L iw ' β i [ H ( t )

& H (25)], (A-65)

where D i (t) and L i (t) are, respectively, the airway diam eter and length at generation i and age t,
D iw and L iw the corresponding values for W e ibel’s m odel, α i and β i are coefficients given by

α i ' 3.26 × 1 0

& 2 exp[

& 1.183 ( i % 1) 0. 5 ] (A-66)

β i ' 1.05 × 1 0

& 6 exp [ 10.1] ( i % 1)

& 0. 2 ] (A-67)

A-16and H(t) is the body height, which varies with age t in the form

H

( t ) ' 1.82 × 1 0 2 (1

& 0.725 e & 0. 14 t ) ( cm ). (A-68)

For the growth patterns of the airways in the alveolar region, it is assum e d that

D i

' L i

' D
a ' f ( t ), for 17 # i # 23 (A-69) D iw L iw D
aw

where D
a is the diam eter of an alveolus at age t, D
aw = 0.0288 cm is the alveolar diam eter for
adults in accordance with W e ibel’s m odel, and f(t) is a function determ ined from

3 16 { LV ( t )

& j π D i 2 ( t ) L i ( t ) N i ( t )} f

( t ) ' 23 i ' 0 4 (A-70)
{ j π D 2

aw N

A ( t )} iw

L iw N

i ( t ) % 5 π D 3
i ' 17 4 36



### A.6. TRANSPORT RATES The values of transport rates λ (i) XY for rats have been derived from the experim e ntal data of clearance for diesel soot (Chan et al., 1981; Stromet al., 1987, 1988) and for the particleassociated organics (Sun et al., 1984; Bond et al ., 1986; Yu et al., 1991). These values are used in the present m odel of lung burden calculation and are listed below:

λ ( i )

HG ' 1.73 ( i ' 1,2,3) (A-71)

λ (1 ) λ (1 ) λ (1 ) AB

HB ' TB ' LB ' λ ( i )

' 0.00018 (A-72)

λ (2 ) λ (2 ) λ (2 ) λ (2 )

HB ' TB ' LB ' AB ' 0.0129 (A-73)

λ (3 ) λ (3 ) λ (3 ) λ (3 )

HB ' TB ' LB ' AB ' 12.55 (A-74)

A-17λ ( i )

' 0.693 ( i ' 1,2,3) (A-75) TG

λ (1 )

AL ' 0.00068 [1

& exp(

& 0.046 m A 1. 62 )] (A-76)

λ ( i )

' 1 λ ( i ) ( i ' 2,3) (A-77)

AL 4 AB

λ ( i )

' 0.012 exp(

& 0.11 m A 1. 76 ) % AT (A-78) 0.00068 exp(

& 0.046 m A 1. 62 ) ( i ' 1,2,3)

λ (1 )

' AL % λ (1 ) % λ (1 )

'

A λ (1 ) AT AB (A-79) 0.012 exp(

& 0.11 m A 1. 76 ) % 0.00086

λ (2 )

' AL % λ (2 ) % λ (2 )

' 0.012 exp(

& 0.11 m A 1. 76 ) % A λ (2 ) AT AB (A-80) 0.00068 exp(

& 0.046 m a 1. 62 ) % 0.0161

λ (3 )

' AL % λ (3 ) % λ (3 )

' 0.012 exp(

& 0.11 m A 1. 76 ) % A λ (3 ) AT AB (A-81) 0.00068 exp(

& 0.046 m A 1. 62 ) % 15.7

XY

is the unit of day -1 , and m A 

### – m (1 ) is the particle burden (in m g ) in the alveolar where λ (i ) A

com p artm ent. Experim e ntal data on the deposition and clearance of DPM in hum ans are not available.
To estim ate the lung burden of DPM for hum an exposure, it is necessary to extrapolate the transport rates λ (i ) XY

from rats to hum ans. For organics, it is assum e d that the transport rates are the

sam e for rats and hum ans. This assum p tion is based upon the observation of Schanker et al. (1986) that the lung clearance of inhaled lipophilic com pounds appears to depend only on their lipid/water partition coef f i cients and is independent of species. In contrast, the transport rates of
diesel soot in hum ans should be different from t hose of rats, since the alveolar clearance rate, λ A ,

A-18of insoluble particles at low lung burdens for hum an adults is approxim a tely seven tim es that of rats (Bailey et al., 1982). No data are available on the change of the alveolar clearance rate of insoluble particles in
hum ans due to excessive lung burdens. It is seen from Equation A-79 that λ (1 ) A for rats can be
written in the f o rm

λ (1 )

' a exp(

& bm A c ) % d (A-82) A

where a, b, c, and d are constants. The right-hand side of Equation A-82 consists of two term s, representing, respectively, m acrophage-m ediated m echanical clearance and clearance by
dissolution. The first term depends upon the lung burden, whereas the second term does not. To extrapolate this relationship to hum ans, we assum e that the dissolution clearance term is independent of species and that the m echanical clearance term for hum ans varies in the sam e proportion as in rats under the sam e unit surface particulate dose. This assum p tion results in the following expression for λ (1 ) A in hum ans

λ (1 ) a

A ' P exp[

& b ( m A / S ) c ] % d (A-83)

where P is a constant derived from the hum an/rat ratio of the alveolar clearance rate at low lung burdens and S is the ratio of the pulm onary surface area between hum ans and rats. Equation
A-83 im plies that rats and hum ans have equivalent am ounts of biological response in the lung to
the sam e specific surface dose of inhaled DPM.
From the data of Bailey et al. (1982), a value of λ (1 ) = 0.00169 day -1 is obtained for A hum ans at low lung burdens leading to P = 14.4. A value for S of 148 is reported from the data
of the anatom ical lung m odel of Schum and Yeh (1979) for rats and W e ibel’s m odel for hum an adults. For hum ans less than 25 years old, the m odel assum e s the sam e value for P, but S is

com puted from the data of the lung m odel for young hum ans (Yu and Xu 1987). The value of S for different ages is shown in Table A-3. The equations for other transport rates that have a lung-burden-dependent com ponent are extrapolated from rats to hum ans in a sim ilar m a nner. The following lists the values of λ (i ) XY (in day -1 ) for hum ans used in the present m odel calculation:

λ (1 )

HG ' 1.73 ( i ' 1,2,3) (A-84)

A-19λ (1 ) λ (1 ) λ (1 ) λ (1 ) HB ' TB ' LB ' AB ' 0.00018 (A-85)

λ (2 ) λ (2 ) λ (2 ) λ (2 )

HB ' TB ' LB ' AB ' 0.0129 (A-86)

λ (3 ) λ (3 ) λ (3 ) λ (3 )

HB ' TB ' LB ' AB ' 12.55 (A-87)

λ ( i )

' 0.693 ( i ' 1,2,3) (A-88)

TG

λ (1 )

AL ' 0.00068 {1

& 0.0694 exp[

& 0.046( m A / S ) 1. 62 ]} (A-89)

λ ( i )

' 1 λ ( i ) ( i ' 2, 3) (A-90)

AL 4 AB λ ( i )

' 0.0694 {0.012 exp[

& 0.11( m A / S ) 1. 76 ] %

AT (A-91) 0.00068 exp[

& 0.046( m A / S ) 1. 76 ]} ( i ' 1, 2, 3)

λ (1 )

' AL % λ (1 ) λ (1 ) AB % λ (1 )

' A AT (A-92) 0.0694 {0.012 exp[

& 0.11( m A / S ) 1. 76 ]} % 0.00086

λ (2 )

' AL % λ (2 ) % λ (2 )

' (A-93)

A λ (2 ) AT AB 0.0694{0.012 exp[

& 0.11( m A / A ) 1. 76 ] %

0.00068 exp[

& 0.046( m A / S ) 1. 76 ]} % 0.016 (A-94)

### A.7. RESULTS

A.7.1. Simulation of Rat Experiments To test the accuracy of the m odel, sim u lation results are obtained on the retention of DPM in the rat lung and com p ared with the data of lung burden and lym ph node burden obtained by Strom et al. (1988). A particle size of 0.19 μm MMAD and a standard geom etric deviation,

σ g , of 2.3 (as used in Strom ’ s experim ent) are used in the calculation. The respiratory param eters for rats are based on their weight and calculated using the following correlations of m i nute volum e, respiratory frequency, and growth curve data.

Minute volum e = 0.9W (cm 3 /m in) ( A-95)

Respiratory frequency = 475W -0.3 (1/m in) ( A-96)

where W is the body weight (in gram s) as determ ined from the equation



### W = 5+537T/(100+T), for T $ 56 days (A-97) in which T is the age of the rat m easured in days. Equation A-95 was obtained from the data of Mauderly (1986) for rats ranging in age
from 3 m o to 2 years old; Equation A-96 was obtai ned from the data of Strom et al. (1988); and Equation A-97 was determ ined from the best fit of the experim e ntal deposition data. Figures A3 and A-4 show the calculated lung burden of diesel soot ( m (1 ) T ) and lym ph node burden, A + m (1 ) respectively, for the experim ent by Strom et al. (1988) using anim als exposed to DPM at
6 m g /m 3 for 1, 3, 6, and 12 weeks; exposure in all cases was 7 days/week and 20 h daily. The solid lines represent the calculated accum u lation of particles during the continuous exposure phase and the dashed lines indicate calculated post-exposure retention. The agreem ent between the calculated and the experim ental data for both lung and lym ph node burdens during and after the exposure periods was very good.
Com p arison of the m odel calculation and the retention data of particle-associated BaP in rats obtained by Sun et al. (1984) is shown in Figure A-5. The calculated retention is shown by
the solid line. The experim ent of Sun et al. consisted of a 30-m i n exposure to diesel particles coated with [ 3 H ] benzo[a]pyrene ([ 3 H] - BaP ) at a concentration of 4 to 6 μg/m 3 of air and followed by a post-exposure period of over 25 days. The fast and slow phase of ([ 3 H] - BaP )
clearance half-tim es were found to be 0.03 day and 18 days, respectively. These correspond to

AO = 0.0385 day -1 and λ (3 ) AO is the value of λ (i) λ 2 AO = 23.1 day -1 in our m odel, where λ (i ) XY at m A 6 0. Figure A-5 shows that the calculated retention is in excellent agreem ent with the experim ental data obtained by Sun et al. (1984).

A-21A.7.2. Predicted Burdens in Humans Selected results of lung burden predictions in hum ans are shown in Figures A-6 to A-9. The particle conditions used in the calculation are 0.2 μm MMAD with σ g = 2.3, and the m a ss fractions of the rapidly and slowly cleared organics are each 10% (ƒ 1 = ƒ 2 = 0.1)

. Figures A-6 and A-7 show, respectively, the lung burdens per unit concentration of diesel soot and the associated organics in hum an adults for differe nt exposure patterns at two soot concentrations,
0.1 and 1 m g /m 3

. The exposure patterns used in the calculation 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, sim u lating environm ental and occupational exposure conditions. The results show that the lung burdens of both diesel soot and the associated organics reached a steady-stat e value during exposure. Because of differences
in the am ount of particle intake, the steady-state lung burdens per unit concentration were highest for exposure pattern (a) and lowest for exposure pattern (b). Also, increasing soot
concentration from 0.1 to 1 m g /m 3 increased the lung burden per unit concentration. However, the increase was not noticeable for exposure pattern (c). The dependence of lung burden on the soot concentration is caused by the reduction of the alveolar clearance rate at high lung burdens
discussed above. Figures A-8 and A-9 show the effect of age on lung burden, where the lung burdens per unit concentration per unit weight are plotted vers us age. The data of lung weight at different ages are those reported by Snyder (1975). The exposure pattern used in the calculation is 24 h/day and 7 days/week for a period of 1 year at the two soot concentrations, 0.1 and 1 m g /m 3

. The results show that, on a unit lung weight basis, the lung burdens of both soot and organics are functions of age, and the m a xim u m lung burdens occur at approxim a tely 5 years of age. Again, for any given age, the lung burden per unit concentration is slightly higher at 1 m g /m 3 than at
0.1 m g /m 3

.



### A.8. PARAMETRIC STUDY OF THE MODEL The deposition and clearance m odel of DPM in hum ans, presented above, consists of a large num ber of param eters that characterize the size and com position of diesel particles, the
structure and dim ension of the respiratory tract, the ventilation conditions of the subject, and the
clearance half-tim es of the diesel soot and the particle-associated organics. Any single or

com b ined changes of these param eters from their norm al values in the m odel would result in a change in the predicted lung burden. A param etric study has been conducted to investigate the effects of each individual param eter on calculated lung burden in hum an adults. The exposure pattern chosen for this study is 24 h/day and 7 days/week for a period of 10 years at a constant
soot concentration of 0.1 m g /m 3

. The following presents two im portant results from the

param etric study.

A-22A.8.1. Effect of Ventilation Conditions The changes in lung burden due to variations in tidal volum e and respiratory frequency are depicted in Figures A-10 and A-11. Increasing any one of these ventilation param eters increased the lung burden, but the increase was m u ch sm aller with respect to respiratory frequency than to tidal volum e. This sm all increase in lung burden was a result of the decrease in deposition efficiency as respiratory frequency increased, despite a higher total am ount of
DPM inhaled. The m ode of breathing has only a m i nor effect on lung burden because switching from nose breathing does not produce any appreciable change in the am ount of particle intake into the lung (Yu and Xu, 1987). All lung burden results presented in this report are for nose breathing.

A.8.2. Effect of Transport Rates Transport rates have an obvious effect on the retention of DPM in the lung after deposition. Because we are m ainly concerned with the long-term clearance of diesel soot and the associated organics, only the effects of two transport rates, λ (1 ) A , are studied. A and λ (2 ) Experim ental data of λ (1 ) from various diesel studies in rats have shown that λ (1 ) can vary by a A A factor of two or higher. W e use a m u ltiple of 0.5 to 2 for the uncertainty in λ (1 ) and λ (2 ) to A A

exam ine the effect on lung burden. Figures A- 12 and A-13 show respectively, the lung burden
results f o r diesel soot and the associated organics versus the m u ltiples of λ (1 ) and λ (2 ) used in the A A calculation. As expected, increasing the m u ltiple of λ (1 ) reduced the lung burden of diesel soot A with practically no change in the organics burden (Figure A-12), while just the opposite occurred
when the m u ltiple of λ (2 ) was increased (Figure A-13). A



### A.9. OPERATIONAL DERIVATION OF HUMAN EQUIVALENT CONCENTRATIONS (HECs)

The m odel of Yu et al. (1991) is ordered into two parts; one part param eterized on the physiology and anatom y of a 300 g rat and the other part param eterized on the physiology and

anatom y of a 25 year old hum an m a le. The sequence of steps taken to calculate the hum an equivalent continuous concentrations (the HECs), outlined in Table A-4, were as follows:

• T he exposure scenario of the rats was entered into the rat portion of the m odel and the

m odel ran to obtain the output of lung burden in m g DPM/ rat lung at the tim e of the
sacrifice of the rats.

• T he output of m g DPM/ rat lung was norm a lized to m g DPM/ cm 2 of rat lung tissue based on a total pulm onary surface area of 4090 cm 2.

.

A-23• T he norm a lized rat lung burdens were used to calculate the corresponding lung burden based on the pulm onary surface area of 627,000 cm 2

. This operation yielded

m g DPM / lung of a 25 year old hum an m a le.

• V arious air concentrations were run in an iterative fashion with the hum an portion of
the m odel under a continuous exposure scenario of 24 h/day, 7d/wk for 70 years with ventilatory param e ters set at 0.926 L for tidal volum e and 15 breaths per m i nute as the respiratory frequency to yield a total daily pulm onary volum e of 20 m 3

. This was continued until the output (m g DPM/lung) was m a tched to the m g DPM /hum an lung obtained from the norm a lized rat lung bur den; the concentration from the m odel
that m a tched this lung burden was term ed the hum an equivalent continuous concentration, the HEC. The hum an m odeling runs did not consider the preadult status of airway and alveoli num ber discussed above but rather were ran for 1 to 70 years with adult (25 years of age) param e ters m e ntioned above.

These HEC values address kinetic issues of DPM deposition and retention in the lung by
hum ans. As noted above, these values do not reflect the kinetic variability that m a y exist in the
hum an population exposed to DPM which includes m e n and wom e n, young and old. However,
the lim ited param e tric analysis of the m odel clearly shows variability of those param e ters m o st determ inative in hum ans (e.g., tidal volum e, respiration rate, and rates of clearance of particles
from the airways) were m i rrored in the corresponding output of the m odel (lung burden of DPM). One interpretation of this parallel in param e ter-output is that the variability in the physiological characteristics of hum ans reflects the variability in the m odel such that, for

exam ple, a sm all tidal volum e would be reflected with a decreased lung burden of DPM. Variability am ong hum ans of these key param e ters such as tidal volum e do vary but within an
order of m a gnitude. This would m ean that the DPM dose received by different individuals in the population from the sam e concentration would indeed vary within the extrem es of these
determ inative param e ters.

A-241

2

3

4

5

6

7

8

9

10

11

12

13

14

15

16

17

18

19 Table A-1. Lung model for rats at total lung capacity

Generation Number of Accumulative

number airw ays Length (cm) Diameter (cm) volume a (cm)

1 2.680 0.340 0.243

2 0.715 0.290 0.338

3 0.400 0.263 0.403

5 0.176 0.203 0.431

8 0.208 0.163 0.466

1 4 0.117 0.134 0.486

2 3 0.114 0.123 0.520

3 8 0.130 0.112 0.569

6 5 0.099 0.095 0.615

109 0.091 0.087 0.674

184 0.096 0.078 0.758

309 0.073 0.070 0.845

521 0.075 0.058 0.948

877 0.060 0.049 1.047
1,477 0.055 0.036 1.414

b 2,487 0.035 0.020 1.185
4,974 0.029 0.017 1.254
9,948 0.025 0.016 1.375
19,896 0.022 0.015 1.595

21 39,792 0.020 0.014 2.003

22 79,584 0.019 0.014 2.607

25 318,336 0.017 0.014 7.554

24 636,672 0.017 0.014 13.784

a Incl udi ng t h e at t ached al veol i vol um e (num ber of al veol i = 3 × 10 7 , al veol ar di am et er = 0.0086 cm ).

b Term i n al bronchi ol es.

A-25Table A-2. Lung model by Weibel (1963) adjusted to 3000 cm 3 lung volume

Generation Number of Accumulative

number airw ays Length (cm) Diameter (cm) volume a (cm)

0 1 10.260 1.539 19.06

2 2 4.070 1.043 25.63

2 4 1.624 0.710 28.63

3 8 0.650 0.479 29.50

4 1 6 1.086 0.385 31.69

5 3 2 0.915 0.299 33.75

6 6 4 0.769 0.239 35.94

7 128 0.650 0.197 38.38

8 256 0.547 0.159 41.13

9 512 0.462 0.132 44.38

10 1,024 0.393 0.111 48.25

11 2,048 0.333 0.093 53.00

12 4,096 0.282 0.081 59.13

13 8,192 0.231 0.070 66.25

14 16,384 0.197 0.063 77.13

15 32,768 0.171 0.056 90.69

16 b 65,536 0.141 0.051 109.25

17 131,072 0.121 0.046 139.31

18 262,144 0.100 0.043 190.60

19 524,283 0.085 0.040 288.16

20 1,048,579 0.071 0.038 512.94

21 2,097,152 0.060 0.037 925.04

22 4,194,304 0.050 0.035 1,694.16

23 8,388,608 0.043 0.035 3,000.00

A-26Table A-3. Ratio of pulmonary surface areas betw een humansand rats as a function of human age
Age (year) Surface area

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 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

A-27Table A-4.Human equivalent continuous concentrations (HECs)calculated with the model of Yu et al. (1991) from long-termrepeated exposure rat studies of DPM exposureRat exposure concsmg DPM/ rat lungmg DPM/cm2mg DPM/HECStudyExposureconditionsa(mg/m3)(modeled)brat&human lungb,chuman lungc(mg/m3)cMauderlyet al., 1987a7 h/day,5 days/wk, 130 wkd0.350.286.85E-5430.038Mauderlyet al., 1987a7 h/day,5 days/wk, 130 wk3.4720.234.95E-331011.375Mauderlyet al., 1987a7 h/day,5 days/wk, 130 wk7.0844.521.09E-268253.05Ishinishi et al., 1988 (LDc)16 h/day,6 days/wk,130 wk0.110.245.87E-5370.032Ishinishi et al., 1988 (LD)16 h/day,6days/wk, 130 wk0.411.002.45E-41530.128Ishinishi et al., 1988 (LD)16 h/day,6days/wk, 130 wk1.1818.454.51E-328281.25Ishinishi et al., 1988 (LD)16 h/day,6days/wk, 130 wk2.3239.899.75E-361152.75Ishinishi et al., 1988 (HD)16 h/day,6days/wk, 130 wk0.461.152.81E-41760.144Ishinishi et al., 1988 (HD)16 h/day,6days/wk, 130 wk0.9612.943.16E-319840.883Ishinishi et al., 1988 (HD)16 h/day,6days/wk, 130 wk1.8431.227.63E-347862.15Ishinishi et al., 1988 (HD)16 h/day,6days/wk, 130 wk3.7264.671.58E-299144.4Nikula et al., 199516 h/day,5 days/wk, 100 wk2.4428.647.00E-343911.95Nikula et al., 199516 h/day,5 days/wk, 100 wk6.376.151.86E-2116745.1Heinrich et al., 199518 h/day,5 days/wk, 104 wk0.843.839.4E-45870.33Heinrich et al., 199518 h/day,5 days/wk, 104 wk2.534.48.4E-352742.35Heinrich et al., 199518 h/day,5 days/wk, 104 wk6.9897.82.4E-2149936.7aThese are entered into the programas hrs/day,days/week forthe total number of weeks exposedand the last week of exposurebefore evaluation (as this would affectclearance). The parameters for the rat were based on abodyweight which was set in the programat 300g.bThese values were obtained with the ratportion of the model and are noted as lung burden, in mg DPM /lung of a 300 g rat, atthe final week of the exposure scenario.Theseoutputs were then normalized to cm2of the rat lung, at 4090 cm2total (Xu and Yu, 1987).cPreparatoryto using the human portion of the model, the mg DPM/cm2value from above was used to project the mg DPM that would be present in the adult human lungbased on a total lung surface area of 627,000 cm2(Xu and Yu, 1987).Various air concentrations were thenentered into the human modelas 70 years continuous exposurescenarios and ran iterativelyuntil the output (in mgDPM / lung atage 70) matched this mgDPM/human lung, i.e., the total lung burden.This matching air concentration is, bydefinition, the human equivalentcontinuous concentration (HEC).dweeks = (months of exposure) × 4.33.

A-28Figure A-1. Compartmental model of DPM retention.

A-29Figure A-2. Trumpet model of lung airw ays.

A-30Figure A-3. The experimental and predicted lung burdens of rats to DPM at a solid and dashed concentration of 0.6 mg/m 3 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).

A-31Figure A-4. Experimental and predicted lymph no de burdens of rats exposed to CEPs at a concentration of 6.0 mg/m 3 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).

A-32Figure A-5. Comparison betw een the calculat ed lung retention (solid line) and the experimental data obtained by Sun et al. (1984) for the particle-associated BaP in rats.

A-33Figure A-6. Calculated lung burdens of dies el soot per unit exposure concentration in human adults exposed continuously to DPM at tw o different concentrations of 0.1 and 1.0 mg/m 3

. Exposure patterns are (a) 24 h/day and 7 days/w eek, (b) 12 h/day and 7 days/w eek, and (c) 8 h/day and 5 days/w eek.

A-34Figure A-7. Calculated lung burdens of the particle-associated organics per unit exposure concentration in human adults exposed continuously to DPM at tw o different concentrations of 0.1 and 1.0 mg/m 3

. Exposure patterns are (a) 24 h/day and 7 days/w eek, (b) 12 h/day and 7 days/w eek, and (c) 8 h/day and 5 days/w eek.

A-35Figure 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 tw o different concentrations of 0.1 and 1.0 mg/m 3 for 7 days/w eek and 24 h daily.

A-36Figure 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 tw o different concentrations of 0.1 and 1.0 mg/m 3 for 7 days/w eek and 24 h daily.

A-37Figure A-10. Calculated lung burdens in human adults versus tidal volume in liters for exposure to DPM at 0.1 mg/m 3 for 10 years at 7 days/w eek and 24 h daily. Parameters used in the calculation are: (a) MMAD=0.2 μm, σ g =2.3, ƒ 2 =0.1, ƒ 3 =0.1; (b) respiratory frequency = 14 min -1 ; and (c) lung volume = 3000 cm 3

.

A-38Figure A-11. Calculated lung burdens in human adults versus respiratory frequency in bpm for exposure to DPM at 0.1 mg/m 3 for 10 years at 7 days/w eek and 24 h daily. Parameters used in the calculation are: (a) MMAD=0.2 μm, σ g =2.3, ƒ 2 =0.1, ƒ 3 =0.1; (b) tidal volume = 500 cm 3 , and (c) lung volume = 3200 cm 3

.

A-39Figure A-12. Calculated lung burdens in human adults versus multiple of λ (1) for A exposure to DPM at 0.1 mg/m 3 for 10 years at 7 days/w eek and 24 h daily. Parameters used in the calculation are: (a) MMAD=0.2 μm, σ g =2.3, ƒ 2 =0.1, ƒ 3 =0.1; (b) tidal volume = 500 cm 3 , respiratory frequency = 14 min -1 ; and (c) lung volume = 3200 cm 3

.

A-40Figure A-13. Calculated lung burdens in human adults versus multiple of λ (1) for exposure A to DPM at 0.1 mg/m 3 for 10 years at 7 days/w eek and 24 h daily. Parameters used in the calculation are: (a) MMAD=0.2 μm σ g =2.3, ƒ 2 =0.1, ƒ 3 =0.1; (b) tidal volume = 500 cm 3 , respiratory frequency = 14 min -1; and (c) lung volume = 3200 cm 3

.

A-41

### REFERENCES FOR APPENDIX A

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B a i l e y , M R ; Fry
, FA; Jam e s, AC . (1982) The l ong-t e rm cl earance ki net i c s of i n sol ubl e part i c l e s from t h e hum an l ung. Ann Occup Hy g 26: 273-289.

Bond, JA; Sun, JD; Medinsky, MA; et al. (1986) Deposition, m e tabolism and excretion of 1-[ 14 C ] ni t r opy rene and 1-[ 14 C ] ni t r opy rene coat ed on di esel exhaust part i c l e s as i n fl uenced by exposure concent r at i on. Toxi col Appl Pharm acol 85: 102-117.

Chan, TL; Lee, PS; Hering, W E . (1981) Deposition and clearance of inhaled diesel exhaust particles in the respi r at ory t r act of Fi sher rat s . J Appl Toxi col 1: 77-82.

Diu, CK; Yu, CP. (1983) Respiratory tract deposition of polydisperse aerosols in hum ans. Am Ind Hyg Assoc J 44: 62-65.

Heinrich, U; Fuhst, R; Rittinghausen, S; et al. (1995) Chronic i nhalation exposure of W i star rats and two different strains of m i ce to diesel engine exhaust, carbon black, and titanium dioxide. Inhal Toxicol 7:533-556.

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Ishi ni shi
, N; Kuwabara, N; Takaki , Y; et al . (1988) Long- t e rm i nhal a t i on experi m e nt s on di esel exhaust

. In: Di esel exhaust and heal t h ri sks: resul t s of t h e HER P st udi es. Tsukuba, Ibaraki
, Japan: Japan Aut o m obi l e R e search In stitu te, In c., Research Co m m ittee fo r HERP Stu d i es; p p

. 1 1

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M a uderl y , JL. 1986. R e spi r at i on of F344 rat s i n nose- onl y i nhal a t i on exposure t ubes. J Appl Toxi col 6: 25-30.

M a uderl y , JL; Jones, R K ; Gri ffi t h , W C ; et al . (1987) Di esel exhaust i s a pul m onary carci nogen i n rat s exposed chroni cal l y by i nhal a t i on. Fundam Appl Toxi col 9: 208-221.

Nikula, KJ; Snipes, MB; Barr, EB; et al. (1995) Co m p arative pulm onary toxic ities and carcinogenicities of chroni cal l y i nhal e d di esel exhaust and carbon bl ack i n F344 rat s . Fundam Appl Toxi col 25: 80-94.

Schanker, LS; M i t c hel l
, EW ; B r own, R A . (1986) Speci es com p ari s on of drug absorpt i on from t h e l ung aft e r aerosol i nhal a t i on or i n t r at racheal i n ject i on. Drug M e t a b Di spos 14(1): 79-88.

Scheut zl e, D. (1983) Sam p l i ng of vehi cl e em i ssi ons for chem i cal anal y s i s and bi ol ogi cal t e st i ng. Envi ron Heal t h Perspect 47: 65-80.

Schum , M; Yeh, HC. (1979) Theoretical evaluation of aerosol deposition in anatom ical m odels of m a m m a lian lung ai rway s. B u l l M a t h B i ol 42: 1-15.

Sny d er, W S . (1975) R e port of t a sk group on reference m a n. Oxford, London: Pergam on Press, pp. 151-173.

Sol d erhol m , SC . (1981) C o m p art m ent a l anal y s i s of di esel part i c l e ki net i c s i n t h e respi r at ory sy st em of exposed ani m al s. Oral present a t i on at EPA Di esel Em i ssi ons Sy m posi u m , R a l e i gh, NC , Oct ober 5-7. In: Toxi col ogi cal effect s of em i ssi ons from di esel engi nes (Lewt a s J, ed.). New York: El sevi er, pp. 143-159.

St rom , KA; C h an, TL; Johnson, JT. (1987) Pul m onary ret e nt i on of i nhal e d subm i c ron part i c l e s i n rat s : di esel exhaust exposures and l ung ret e nt i on m odel

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- 5718. W a rren, M I : General M o t o rs Research Laboratories.

St rom , KA; C h an, TL; Johnson, JT. (1988) Inhal e d part i c l e s VI. Dodgson, J; M c C a l l u m , R I ; B a i l e y , M R ; et al ., eds. London: Pergam on Press, pp. 645-658.

A-42Sun, JD; W o ff, RK; Kanapilly, GM; et al. (1984) Lung re tention and m e tabolic fate of inhaled benzo(a)pyrene associ at ed wi t h di esel exhaust part i c l e s. Toxi col Appl Pharm acol 73: 48-59.

W e i b el , ER . (1963) M o rphom et ry of t h e hum an l ung. B e rl i n : Spri nger-Verl a g.
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Yu, CP. (1978) Exact analysis of aerosol depositi on during steady breathing. Powder Technol 21:55-62.

Yu, CP; Xu, GB. (1986) Predictive m odels for deposition of diesel exhaust participates in hum an and rat lungs. Aerosol Sci Technol 5: 337-347.

Yu, CP; Xu, GB. (1987) Predicted deposition of diesel particles in young hum ans. J Aerosol Sci 18:419-429.

Yu, C P ; Yoon, KJ. (1990) R e t e nt i on m odel i ng of di esel exha ust part i c l e s i n rat s and hum ans. R e s R e p Heal t h Eff Inst 40: 1-33.

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Yu, C P ; Yoon, KJ; C h en,YK. (1991) R e t e nt i on m odel i ng of di esel exhaust part i c l e s i n rat s and hum ans. J. Aerosol M e d. 4(2): 79-115.

A-43B-1 

## Appendix B

Benchmark Concentration Analysis of Diesel Data1 For increases in incidence “extra risk” is used which is response incidence (inc) normalized to the background (BG) incidence; response – BG 

## / 1-BG.

B-2 

### B-1. INTRODUCTION TO BENCHMARK The benchmark dose or benchmark concentration approach, hereafter referred to as the BMC approach, is an alternate to the N/LOAEL option for deriving effect levels. The BMC is currently undergoing extensive consideration by the Agency with promulgation of software and guidelines for application of this methodology (U.S. EPA, 2000). The BMC approach involves fitting a dose-response function to dose and effect information from a single study to derive the best fit of those data. This “best fit”is statistically termed the maximum likelihood estimate but is referred to in the benchmark terminology as the BMC curve. The curve defining the corresponding lower 95% confidence limit of this “b est fit”estimate is termed the BMCL curve. This BMCL curve is used to predict the dose that will result in a level of response that is defined a priori as the benchmark response “x”, BMCL

x . In the analyses below, for example, the benchmark response for a 10% increase in incidence 1 of chronic inflammation is defined as a
BMCL 10 ; the corresponding 10% increase as determined from the BMC curve would be termed
the BMC 10 . This BMCL 10 would be derived by first using the data and the programs to determine the BMC and BMCL curves. The concentration corresponding to a 10% increase in incidence would then be determined directly from the BMCL. The BMCL 10 then would be used as the representative value for the effect level or point of departure in the dose-response assessment. The latest version of the Agency Benchm ark Dose Software (BMDS Version 1.2; U.S. EPA, 2000) was used to analyze data on chr onic inflammation and pulmonary histopathology present in the chronic studies that were amenable to benchmark analysis. At this time, the Agency BMDS offers sixteen different models total that are appropriate for the analysis of dichotomous data (gamma, logistic, probit, Weibull, log-logistic, multistage, log-probit, quantal-linear, quantal-quadratic), continuous da ta (linear, polynomial, power, Hill) and nested developmental toxicology data (NLogistic, NCTR, Rai & Van Ryzin). Results from all models include a reiteration of the model formula and model run options chosen by the user, goodness-of-fit information, a graphical presentation for visual inspection and the concentration estimate for the response at the designated BMCL

x , as well as the corresponding BMC

x . More details on the modeling results are described and presented in the analysis on dichotomous data following. The U.S. EPA benchmark dose (BMD/C) methods guidance has not been finalized at this time to provide definitive procedures and criteria (U.S. EPA 1995). Therefore, in this document provisional criteria for minimum data to perform a benchmark analysis are designated such that (1) complete quantitative information on the response of interest should be available (e.g.,B-3 incidence as number affected / total, means with variability) and that (2) at least two exposure levels with responses that differ from those of the controls are provided, and (3) a benchmark response of 10% is employed such that outcomes are BMCL 10 s. A response of 10% is at or near the limit of sensitivity in most long-term bioassays as determined from both the typical number of animals used in bioassays and a low spontaneous background rate (e.g., 0.1%) for a given effect (Haseman, 1984; Haseman et al., 1989).



### B-2. DIESEL DATA FOR BENCHMARK ANALYSIS Using the criteria set forth in Section B-1 and the information about the critical effects that have been identified (pulmonary inflammati on, pulmonary histopathology including indicators of fibrotic changes such as increases in alveolar-capillary wall thickness) the following rat chronic studies identified in Chapter 6 were analyzed for information suitable for BMC analysis: Ishinishi et al. (1986, 1988), Mauderly et al. (1987a,b; 1988); Heinrich et al. (1986, 1995), and Nikula et al. (1995). Results from this analysis yielded only a few data sets from a single study, that of Nikula et al. (1995), that could be used for BMC analys is. The basis for not including data from the other studies varied. Information on pulmonary hi stopathology in the studies of Ishinishi et al. (1986, 1988), for example, was supplied only in narrative form with no quantitative information given. A similar situation was found for those reports of the ITRI study; Wolff et al. (1987) reports on clearance alterations due to DPM exposure; Henderson et al. (1988) does give information on hydroxyproline but only in graphical form; the 1988 study of Mauderly et al. deals with pulmonary function as a function of DPM lung loading; the 1987a reference of Mauderly et al. discusses tumor prevalence only and the Mauderly 1987b reference reports ondiesel exhaust in developing lung to a single exposure concentration of DPM with no doseresponse information available. Those reports on the General Motor study contain extensive information relating not to the critical effects, but mostly to precursors of inflammation such as levels of polymorphonuclear neutrophils and lym phocytes in bronchoalveolar lavage from DPM exposed rats (Strom, 1984) and guinea pigs (Barnhart et al., 1981) as well as information on collagen biosynthesis (Misiorowski et al., 1980) all of which is presented in graphical rather than tabular form amenable for benchmark analysis. The information on noncancer histopathology reported by Heinrich et al. (1995) is in text form only and this author’s 1986 study deals primarily with clearance and mortality. Nikula et al. (1995), however, do present extensive quantitative dose-response information (incidence / dichotomous data) on several measures of the critical effect including chronic inflamation (presence of focal aggregates of neutrophils), focal fibrosis with epithelial hyperplasia (nodul ar fibrosis rimmed by hyperplasia), and septal fibrosis (interstitial fibrosis within alveol ar septa) although the study had but 2 exposureB-4 concentrations both of which are different from the controls, a minimal number on which benchmark analysis should be performed.



### B-3. BENCHMARK ANALYSIS OF DIESEL DATA These data from Nikula et al. (1995) were ex tracted, HEC concentrations calculated using the model of Yu et al. (1991; Appendix A), and analyzed using all 9 applicable models for dichotomous data. Because the benchmark models were ran with the HEC, general from the model of Yu et al. (1991), the BMCL 10s are also HECs. The results and data are presented in Table B-1. Results were evaluated based on the nature of the data set, visual inspection of the graphical output, and on the goodness-of-fit parameters, including p values and the AIC. When p values were generated for model fits, values for p that were less than 0.1 were considered to reflect a minimal fit to the data and were disqualified from further consideration. However, the small set of only 3 data points was often matched by the number of parameters fitted in several of the models such that the outcome of the model exactly fit the data and thus no p value is generated; these model fits are often referred to as being overparameterized, and are indicated as “NA” in Table B-1. Values for p that were less than 0.1 were considered to reflect a minimal fit to the data. The AIC (Akaike Information Coefficient; Akaike, 1973; Stone, 1998) is a parameter generated for the models in U.S. EPA (2000) that allows for a general comparison among models run on the same data set. The AIC is defined as -2 log L + 2 p where log L is the log likelihood of the fitted model, and p is the number of parameters estimated; smaller values
indicate better fits. The overall results of this mathematical analysis is reasonable in a biologically mechanistic sense in that chronic inflammation is more prevalent and apparently occurs at lower concentrations (i.e., has lower BMCL 10 values) than does focal fibrosis. The information on septal fibrosis were not interpretable as the data were not amenable (no or zero background and then total incidence) to any meaningful benchmark or other dose-response analysis. The most sensitive endpoint, chronic inflammation, is therefore the most sensitive benchmark concentration followed by focal fibrosis. The choice for the most appropriate BMCL 10 from among the various modeled values for chronic inflammation requires analysis of both th e statistical and graphical outputs of the data. 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/m 3 . The shape of the dose-response curves generated by several of the models, including gamma-hit, Weibull, multistage, and quantal linear were all a uniformly upward sloping arc from the origin (graphs not shown) with minimal evidence of any “S” character, a shape not concordant with the data array in Table 6.2. Models that did generate curves with “S”B-5 character included log-logistic, logistic, probit, quantal-quadratic, and log-probit. Because of their concordance with this independent data array on dose-response, the latter outputs are further analyzes.
The results for both chronic inflammation and fo cal fibrosis for those models with outputs having appreciable “S” character suggest that females may be more sensitive than males for these endpoints as the incidences are higher and the BMCL 10 values are generally lower for females than for males. However, the model fits of the BMCL 10 s to the chronic inflammation data segregated by sex were generally inadequate as judged from the p values (most being far less than 0.1) or from visual inspection of the fits to the data, several of which (e.g., log-logistic and log-probit) were lacking any appreciable “S ” character. However, combining female and male data improved data fitting as judged by the increased p values to where nearly all were>0.1 and to where the visual fits were concordant with the independent information on doseresponse. Too, most of the combined BMCL 10 s were either intermediate between the female and
male values or somewhat closer to the female values such that the combined BMCL 10 values
were not much different from the females BMCL 10 s. From among the combined male and female model outputs in Table B-1, the logistic, probit, and quantal quadratic results were all excluded based on the high AIC value relative to the log-logistic and log-probit results. The log-logistic results were excluded based on the shape of the lower portion of the dose-response curve which was upward sloping near the origin (graph not shown) and not as concordant with the i ndependent dose-response information in Table 6-2 as was the fit of the log-probit model (Figure B-1) . This leaves the fit of the log-probit model as being most reflective of the information in Table 6-2. The BMCL 10 of the log-probit curve at 0.37 mg/m 3 remains and, by elimination, appears to be the most defensible choice from among
the BMCL 10 s arrayed in Table B-1. Figure B-1 shows the graphical representation of the logprobit model fit to the data and the origin of the BMCL 10 . This graph also shows the relationship
of the BMCL 10 of 0.37 mg/m 3 to the variability that exists around the control value and that the value of 0.37 mg/m 3 is not far removed from the outer range of this variability. The log-probit
BMCL 10 for focal fibrosis (combined) of 1.3 mg/m 3 noted as being representative of this lesion from the BMC analysis in Table B-1. Characterization of this benchmark value indicates that it may not be a suitable candidate for use as a point of departure for development of a dose-response assessment such as the RfC. An attribute of the benchmark method is that the response (such as the 10% as used here) is near the range of the actual experimental values, such that extrapolation is not far below the observed experimental range. However, due to the paucity of data points overall and lack of any values below an HEC of nearly 2 mg/m 3 in the Nikula et al. (1995) study, the extrapolation of this BMC to the 10% response level is considerable, the BMLC 10 of 0.37 mg/m 3 being > 5-fold below theB-6 nearest observed value of 1.95 mg/m 3 . Also, the high experimental exposures used in this study are in the range of those resulting in pulmonary overload conditions in rats and therefore in the range of the model assumptions of Yu et al. (1991) about this phenomenon in humans for calculation of the HECs (Chapter 3). The BMCL 10 of 0.37 mg/m 3 is considerably greater than other NOAELs in the DPM data base of 0.144 mg/m 3 and 0.128 mg/m 3 (Table 6-2 in Chapter 6), possibly indicating that these NOAELs represent actual incidence levels that are considerably less than10%; from the same log-probit model the corresponding BMCL 05 was 0.21 mg/m 3 (near the range of these NOAELs) and the corresponding BMCL 01 was 0.07 mg/m 3 (below the range of these NOAELs). These limitations on this BMCL 10 make it a less than optimal candidate for consideration as a point of departure in the development of dose-response

assessments.



### B-4. SUMMARY The recently developed EPA Benchmark dose software (U.S. EPA, 2000) and preliminary guidance was utilized to analyze diesel data by the benchmark approach. Data from only one of the array of principal studies identified else where (Chapter 6) was found to contain data amenable to benchmark analysis. The data fr om this study, that of Nikula et al. (1995) on pulmonary inflammation and histopathology, was ex tracted and analyzed as dichotomous data using all available models and designating a 10% response level such that BMCL 10 s were calculated; as the models were ran with HECs, the BMCL 10s were also HECs. The analysis resulted in an array of BMCL 10 s from 3 different effects in two sexes (both separate and combined) with 9 different models. These BMCL 10 s were each considered from a perspective of biological relevance, known dose -response character, and from the individual fit to the data by the models from statistical parameters and visual judgments. The BMCL 10 that emerged after the above considerations was 0.37 mg/m 3 for the combined male plus female incidence of chronic active pulmonary inflammation. A BMCL 10 of 1.3 mg/m 3 for pulmonary focal fibrosis was also noted in this analys is. Characterization of these benchmark values indicates that neither may be a suitable candidate for use as a point of departure in development of a dose-response assessment such as the RfC but that they are concordant with other quantitative dose-response aspects of the DPM database.B-7Table B-1. BMC analysis of pathology incidence data in maleand female F344 rats from the study of Nikula et al. (1995)using the different models available from U. S. EPA benchmarkdose project (U.S. EPA, 2000) for dichotomous data based on10% extra risk (i.e., a 10% increase relative to a total that has been adjusted for background) and no threshold term. Theconcentrations used in the analysis are human continuous equivalent concentrations (HECs) obtained from the interspeciesextrapolation model of Yu et al. (1991). The table listings include the BMCL10(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 notavailable, usually due to the lack of degrees of freedom.AIC = Akaike Information Coefficient (see U.S. EPA, 2000 andbelow) which may be used for model comparison on the same data set.Effect (from Table 5and 6, p 86, Nikulaet al., 1995)Inc @0 mg/m3Inc @1.95 mg/m3HECInc @5.1 mg/m3HECBMCL10(BMC10)log-logisticBMCL10(BMC10)log-probitBMCL10(BMC10)multi-stageBMCL10(BMC10) -WeibullBMCL10(BMC10) -gammaBMCL10(BMC10) -quantallinearBMCL10(BMC10) -probitBMCL10(BMC10) -logisticBMCL10(BMC10)quantalquadraticChronic activeinflammation >18 mos,grades 1-3, male +female combined5/177 59/162 118/174 0.32(0.64)P= NAAIC= 4830.37(.70)P=NAAIC = 4830.43(.49)P= 0.982AIC= 4810.43(.49)P= 0.982AIC= 4810.43(.49)P=0.98AIC= 4800.43(.49)P= .982AIC= 4811.06(1.19)P= 0.000AIC= 4991.12(1.26)P=0.000AIC= 5021.34(1.45)P= 0.000AIC = 505Chronic activeinflammation >18 mos,grades 1-3 in males1/86 19/81 54/85 0.67(1.16)P= NAAIC= 2170.74(1.22)P = NAAIC = 2170.56(.95)undefinedAIC= 217.56(1.04)P= NAAIC= 216.56(1.09)P= NAAIC= 2170.50(.61)P= 0.15AIC= 2161.31(1.55)P= 0.05AIC= 2190.67(1.16)P= NAAIC= 2171.42(1.57)P= 0.055AIC = 218Chronic activeinflammation >18 mos,grades 1-3 in females4/91 40/81 64/89 0.18(0.26)P= NAAIC= 257.016(.30)P = NAAIC = 2570.33(.40)P= 0.173AIC= 2570.33(.40)P= 0.173AIC= 2570.33(.40)P= 0.17AIC= 2570.33(.40)P= 0.173AIC= 2570.83(.96)P= 0.0001AIC= 2720.85(1.0)P= 0.000AIC= 2731.21(1.35)P= 0.000AIC = 279Focal fibrosis withepithelial hyperplasia,grades 1-4 in males andfemales combined0/177 18/162 63/174 1.25(1.8)P= 1.000AIC= 3451.3(1.8)P = 1.000AIC = 3451.21(1.8)P= 1.000AIC= 3451.21(1.8)P= 1.000AIC= 3451.21(1.8)P= 1.0AIC= 3451.1(1.3)P= 0.363AIC= 3452.32(2.61)P= 0.013AIC= 3532.50(2.8)P= 0.006AIC= 3562.14(2.34)P= 0.091AIC = 347Focal fibrosis withepithelial hyperplasia,grades 1-4 in males0/86 5/81 19/85 1.72(2.7)P= 1.00AIC= 1321.6(2.7)P = 1.000AIC = 1321.79(2.8)undefinedAIC= 1341.79(2.8)P= 1.00AIC= 1321.79(2.75P= 1.0AIC= 1321.7(2.4)P= 0.70AIC= 1312.98(3.5)P= 0.199AIC= 1343.17(3.69)P= 0.153AIC= 1352.68(3.1)P=0.552AIC = 131Focal fibrosis withepithelial hyperplasia,grades 1-4 in females0/91 13/81 44/89 0.80(1.4)P= 1.00AIC= 1990.87(1.47)P = 1.000AIC = 1990.77P= 0.99AIC= 1990.77(1.4)P=1.0AIC=1990.71(1.4)P= 1.00AIC= 1990.71(.88)P= 0.445AIC= 1981.76P= 0.037AIC= 2051.89(2.2)P= 0.02AIC= 2071.7(1.9)P= 0.21AIC = 200Septal fibrosis,>18 mos, grades 1-4 inmales1/86 79/81 83/85 .003(.008)P= 0.35AIC= 53(failed) 0.07(.08)P= 0.000AIC= 650.07(.08)P= 0.000AIC= 650.07(.08)P= 0.000AIC= 650.07(.08)P= 0.000AIC= 650.29(.37)P= 0.000AIC= 1140.32(.44)P= 0.000AIC= 860.42(0.47)P= 0.000AIC = 100Septal fibrosis,>18 mos, grades 1- 4 infemales2/91 75/81 87/89 0.009 (.05)P= NAAIC= 87(failed) 0.08(.10)P= 0.003AIC= 910.08(.10)P= 0.000AIC= 910.08(.10)P= 0.003AIC= 910.08(.10)P= 0.003AIC= 910.32(.40)P= 0.000AIC= 1310.34(.45)P= 0.000AIC= 1090.46(.51)P= 0.000AIC = 119B-8 Figure B-1. Benchmark concentration analysis (log-probit) of chronic pulmonary inflammation in rats exposed to DPM from Nikula et al. (1995). BMCL 10 , the lower confidence estimate of the concentration of DPM associated with a 10% incidence (extra risk); BMC 10 , the corresponding estimate from the best (log-probit) fit. (  ) data with 95% error bounds.B-9 

### REFERENCES FOR APPENDIX B

Akaike, H. (1973) Information theory a nd an extension of the maximum likeli hood principle. In: Proceedings of the Second International Symposium on Information Theory, B.N. Petrov and F. Csaki, eds. Akademiai Kiado, Budapest, pp. 267-281

Barnhart, MI; Chen, S-T; Salley, SO; et al. (1981) Ultrastructure and morphometry of the alveolar lung of guinea pigs chronically exposed to diesel engine exha ust: six months’ experience. J Appl Toxicol 1:88-103.

Haseman, JK. (1984) Statistical issues in the design, analysis, analysis and interpretation of animal carcinogenicity studies. Environ Health Persp 58: 385-392.

Haseman, JK; Huff, JE; Rao, GN and Eustis, SL. (1989) S ources of variability in rode nt carcinogenicity studies. Fund Appl Toxicol 12: 793-804.

Heinrich, U; Muhle, H; Takenaka, S; et al. (1986) Chronic effects on the respiratory tract of hamsters, mice, and rats after long-term inhalation of high concentrations of filtered and unfiltered diesel engine emissi ons. J Appl Toxicol 6:383-395.

Heinrich, U; Fuhst, R; Rittinghausen, S; et al. (1995) Chronic i nhalation exposure of Wistar rats and two strains of mice to diesel engine exhaust, carbon black , and titanium dioxide. Inhal Toxicol 7:553-556.

Henderson, RF; Pickrell, JA; Jones, RK; et al. (1988) Re sponse of rodents to inhaled diluted diesel exhaust: biochemical and cytological changes in bronchoalveolar lavage fluid and in lung tissue. Fundam Appl Toxicol 11:546-567.

Ishinishi, N; Kuwabara, N; Nagase, S; et al. (1986) L ong-term inhalation studies on effects of exhaust from heavy and light duty diesel engines on F344 rats. In: Ishinishi, N; Koizumi, A; McClellan, RO; et al., eds. Carcinogenic and mutagenic effects of diesel e ngine exhaust: proceedings of the international satellite symposium on toxicological effects of emissions from diesel engines; July; Tsukuba Science City, Japan. (Developments in toxicology and environmental science: v. 13.) Amsterdam: Elsevier Science Publishers BV; pp. 329-348

Ishinishi, N; Kuwabara, N; Takaki, Y; et al. (1988) Long- term inhalation experiments on diesel exhaust. In: Diesel exhaust and health risks: results of the HERP studi es. Tsukuba, Ibaraki, Japan: Japan Automobile Research Institute, Inc., Research Committee for HERP Studies; pp. 11-84.

Mauderly, JL; Jones, RK; Griffith, WC; et al. (1987a) Di esel exhaust is a pulmonary carcinogen in rats exposed chronically by inhalation. Fundam Appl Toxicol 9:208-221.

Mauderly, JL; Bice, DE; Carpenter, RL; et al. (1987b) E ffects of inhaled nitrogen dioxide and diesel exhaust on developing lung. Cambridge, MA: Health E ffects Institute; research report no. 8.

Misiorowski, RL; Strom, KA; Vostal, JJ; et al. (1980) Lung biochemistry of rats chronically exposed to diesel particulates. In: Pepelko, WE; Danner, RM; Clarke, NA, eds. Health effects of diesel engine em issions: proceedings of an international symposium; D ecember 1979. Cincinnati, OH: U.S. Envi ronmental Protection Agency, Health Effects Research Laboratory; pp. 465-480; EPA report no. EPA-600/9-80-057a. Available from: NTIS, Springfield, VA; PB81-173809.

Nikula, KJ; Snipes, MB; Barr, EB; et al. (1995) Co mparative pulmonary toxic ities and carcinogenicities of chronically inhaled diesel exhaust and carbon bl ack in F344 rats. Fundam Appl Toxicol 25:80-94.

Stone, M. (1998) Akaike’s Criteria. In: Encyclopedia of Biostatistics, Armitage, P. and Colton, T., eds. Wiley, New York.B-10 Strom, KA. (1984) Response of pulmonary cellular defenses to the inhalation of high concentrations of diesel exhaust. J Toxicol Environ Health 13:919-944.

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 a nd Development, Risk Assessment Forum United States Environmental Protection Agency.

U.S. EPA. (2000) Benchmark dose software version 1.2. Washington, DC: National Center for Environmental Assessment, United States Environmental Protecti on Agency. Available: h ttp://www.epa.gov/ncea/bmds.htm [2000, June 7].

Wolff, RK; Henderson, RF; Snipes, MB; et al. (1987) Alterations in particle accumulation and clearance in lungs of rats chronically exposed to diesel exhaust. Fundam Appl Toxicol 9:154-166.

Yu, CP; Yoon, K.J.; Chen, Y.K. (1991) Retention modeli ng of diesel exhaust particles in rats and humans. J. Aerosol Res. 4 (2): 79-115.C-1 

## Appendix C

A Summary Review of Cancer Dose-response Analyses on Diesel ExhaustC-2 C.1. INTRODUCTION Several individuals and organizations have previously conducted dose-response assessments to estimate quantitatively the cancer risk from exposures to DE. Estimations were performed on the basis of either epidemiologic and/or experimental data. As concluded in Section 8.5, EPA finds that available epidemiologic data are too uncertain to confidently derive a unit risk estimate for DE-induced lung cancer, and that rat data are not suitable for estimating human risk. Nevertheless, a review of histori cal dose-response evaluations is provided here as background information. This information is not intended to constitute endorsement or a recommendation for use in quantitative risk assessment. Early analyses to quantitatively assess the carcinogenicity of DE were hindered by a lack of positive epidemiologic studies and long-term animal studies. One means of overcoming these obstacles was the use of comparative potency methods based on combined epidemiologic and experimental data. By the late 1980s, the availability of dose-response data from animal bioassays and epidemiologic studies provided an opportunity for the derivation of both animal and human data-based estimates, although considerable uncertainties were generally acknowledged by the authors of these assessments.

C.2. COMPARATIVE POTENCY METHODS In this method, the potency of diesel particulate matter (DPM) extract is compared with other combustion or pyrolysis products for which epidemiology-based unit risk estimates have been developed. Comparisons are made using short-term tests such as skin painting, mutations,
and mammalian cell transformation. The ratio of the potency of DPM extract to each of these agents is then multiplied by their individual unit risk estimates to obtain the unit risk for DE. If epidemiology-based estimates from more than one pollutant are used, the derived potencies are generally averaged to obtain an overall mean. Major uncertainties of this method include the assumptions that (1) the cancer potency of DE can be determined on the basis of the relative effectiveness of the organic fraction alone; (2) the relative potency in short-term tests is an accurate predictor of lung cancer potency; and (3) DPM extracts are similar in chemical 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 were collected from three light-duty engine s (a Nissan 220 C, an Oldsmobile 350, and a Volkswagen turbocharged Rabbit), all run on a highway fuel economy test cycle, and from a heavy-duty engine (Caterpillar 3304) run under st eady-state, low-load conditions. The DPM extracts were tested in a variety of assays. Dose/concentration-dependent increases in response were obtained for the four assays listed below:C-3 C Ames Salmonella typhimurium (TA98) reverse mutation,

C Gene mutation in L5178Y mouse lymphoma cells,

C Sencar mouse skin tumor initiation test, and

C Viral enhancement of chemical transformation in Syrian hamster embryo cells. Only the first three assays were used to develop comparative potency estimates because of variability of responses in the enhancement of the viral transformation assay. The in vitro studies were carried out both in the presence and absence of metabolic activators. The potency,defined as the slope of the dose-response curve, was measured for each sample in each shortterm assay. The skin tumor initiation test was positive for all the engines tested except the Caterpillar engine. Only the Nissan engine, however, resulted in strong dose-response data. Because skin tumor initiation was considered to be the most biologically relevant test, it was used to derive potency estimates for the Nissan engine. An estimate for the Nissan engine was then derived by multiplying the epidemiology-based potency estimates for each of the three agents (coke oven emissions, roofing tar, and cigarette smoke condensate) by the ratios of their potencies in the
skin tumor initiation test to that of the Nissan diesel engine. According to this method, three 95% upper-bound estimates of lifetime cancer risk per microgram per cubic meter of extractable organic matter were derived for the Nissan diesel, based on potency comparisons with each of the three agents. These values are: coke oven emissions, 2.6 × 10 -4 ; roofing tar, 5.2 × 10 -4 ; and cigarette smoke condensate, 5.4 × 10 -4 . The average of the three equals 4.4 × 10 -4

. The potency of the other diesel emission samples was not estimated directly because of the weak response in the skin tumor initiation test. Instead, their potency relative to the Nissan engine was estimated as the arithmetic mean of their potency relative to the Nissan in the Salmonella assay in strain TA98, the sister chromatid exchange assay in Chinese hamster ovary cells, and the mutation assay in mouse lymphoma cells. The estimated lifetime cancer risk per microgram per cubic meter of extractable organic matter for extracts from these engines are as follows: Volkswagen, 1.3 × 10 -4 ; Oldsmobile, 1.2 × 10 -4 ; and Caterpillar, 6.6 × 10 -6

. Harris (1983) developed comparative potency estimates for the same four engines used by Albert et al. (1983) but used only two epidem iology-based potency estimates: those for coke oven emissions and for roofing tar. He employed preliminary data from three of the same assays used by Albert et al. (1983): the Sencar mouse skin tumor initiation assay, enhancement of viral transformation in Syrian hamster embryo cells, and the L5178 mouse lymphoma test. The DE cancer potency estimates were then derived by multiplying the epidemiology-based cancer potency estimates for both coke oven emissions and roofing tar by the ratio of their potencies compared with DPM extract in each of the three bioassays. Harris (1983) derived an overallC-4 mean relative risk value of 3.5 × 10 -5 per μg/m 3 for the three light-duty engines with a 95% upper
confidence limit of 2.5 × 10 -4 . Individual mean values for each engine were not reported. McClellan (1986), Cuddihy et al. (1981, 1984), and Cuddihy and McClellan (1983)
estimated a risk of about 7.0 × 10 -5 per μg/m 3 DPM using a comparative potency method similar to those reported in the preceding paragraph. The database was similar to that used by Albert et al. (1983) and Harris (1983).

C.3. EPIDEMIOLOGY-BASED ESTIMATION OF CANCER RISK The first lung cancer risk estimates based on epidemiologic data were derived by Harris (1983). He assessed the risk of exposure to DE using data from the London Transport Worker Study reported by Waller (1981). Five groups of employees from the London Transport Authority (LTA) were used: bus garage engineers, bus drivers, bus conductors, engineers in central works, and motormen and guards. The first group was considered to have received the highest exposure; the next two, intermediate; and the last two groups, none. When cancer death rates for the high-exposure group were compared with those of London males, there was no increase in the observed-to-expected (O/E) ratios. The author, in fact, considered the results to be negative. However, because the low rate of lung cancer in all the LTA exposure groups may have been the result of a “healthy worker” effect, Harris (1983) compared the exposed groups with internal controls. He merged the three exposed groups and compared them with the two groups considered to be unexposed. An adjustment was made for the estimated greater exposure levels of garage engineers compared with bus drivers and conductors. Using this method, the relative risk of the exposed groups was greater than 1 but was statistically significant only for garage engineers exposed from 1950 to 1960. In that case, the O/E ratio was 29% greater than the presumed unexposed controls. Harris (1983) identified a variety of uncertainties relative to potency assessment based on this study. These included:

C small unobserved differences in smoking incidences among groups, which could have a significant effect on lung cancer rates;

C uncertainty about the magnitude of exposure in the exposed groups;

C uncertainty regarding the extent of change in exposure conditions over time;

C random effects arising from the stochastic nature of the cancer incidence; and

C uncertainty in the mathematical specification of the model. Taking the uncertainties into account, he derived a maximum likelihood excess relative
risk estimate of 1.23 × 10 -4 , with a 95% upper confidence limit of 5 × 10 -4 per μg/m 3 DPM per

year.C-5McClellan et al. (1989) reported risk estimates based on the Garshick et al. (1987) casecontrol study in which lung cancer in railroad workers was evaluated. Using a logistic regression, the expected relative risk of lung cancer death was estimated to rise 0.016 per year of exposure to DE. Adjustments were made to convert to continuous exposure (168 vs. 40 hours) for 70 years. Because exposure levels could not be defined exactly, two sets of calculations were made, assuming inhaled DPM concentrations of either 500 or 125 μg/m 3 DPM. The number of excess cancer deaths per year in the United States was estimated to be 3,800 (95% C.I. 400-7400 when an exposure of 125 μg/m 3 was used, and 950 (95% C.I. 100-1,900) when 500 μg/m 3 DPM
was used. The California EPA (Cal-EPA, 1998) derived unit risk estimates for lung cancer based upon the Garshick et al. (1987) case-control study and the Garshick et al. (1988) cohort study of U.S. railroad workers. A variety of exposure patterns were considered, characterized by two components: the average exposure concentration for the workers as measured by Woskie et al. (1988) and the extent of change in exposure from 1959 to 1980. The lowest lifetime risk
estimate derived was 1.3 × 10 -4 per μg/m 3 and the highest was 2.4 × 10 -3 per μg/m 3 . The geometric mean was 6 × 10 -4 per μg/m 3

.Steenland et al. (1998) estimated lung cancer risk of truck drivers on the basis of a casecontrol study of decedents in the Teamsters Union (Steenland et al., 1990). Retrospective exposure estimates were made starting with a set of 1990 exposure measurements for different job categories and then retrospectively estima ting from 1982 to about 1950 using various factors, including diesel vehicle miles traveled and engine emission rates per mile. The 1990 job category estimates came from an extensive industrial hygiene survey of elemental carbon (EC) exposures in the trucking industry by Zaebst et al. (1991). Lifetime (through age 75) excess risk of lung cancer death for male truck drivers was calculated with the aid of a cumulative exposure model. Assuming a most likely emissions scenario of 4.5 g/mile in 1970, and a 45-year exposure to 5 μg/m 3 of EC beginning at age 20 and ending at age 65, the estimated excess lung cancer risk was determined to be 1.6% (95% CI 0.4%-3.1%). Using the same data base, Stayner et al (1998) presented an estimate of excess lifetime risk of 4.5E-4 for a worker exposed to 1 μg/m 3 of DE for 45 years.

C.4. ANIMAL BIOASSAY-BASED CANCER POTENCY ESTIMATES With the availability of chronic cancer bioassays, a considerable number of potency estimates were derived using lung tumor induction in rats. A high degree of uncertainty exists in the use of the rat data to predict human risk. Major uncertainties include: (1) differences in particle deposition patterns between rats and humans, (2) differences in sensitivity between rats and humans to the carcinogenic action of DE, and (3) extrapolation of rat lung tumor responsesC-6 at high concentrations to ambient concentrations without a clear understanding of the mode of action of DE. It is now widely recognized that the rat lung tumor response associated with any insoluble particles at high concentrations is mediated by a particle-overload mechanism (ILSI, 2000), suggesting that rat data for DE are not suitable for estimating human risk at low
environmental concentrations. The first risk estimate was reported by Albert and Chen (1986), based on the chronic rat bioassay conducted by Mauderly et al. (1987). Using a multistage model and assuming equivalent deposition efficiency in humans and rats, they derived a 95% upper confidence limit
of 1.6 × 10 -5 for lifetime risk of exposure to 1 μg/m 3 . Pott and Heinrich (1987) also used a linear model and data reported by Brightwell et al. (1989) , Heinrich et al. (1986), and Mauderly et al. (1987). They reported risk estimates ranging from 6 × 10 -5 to 12 × 10 -5 per μg/m 3 . Smith and Stayner (1990), using time-to-tumor models based on the data of Mauderly et al. (1987), derived point (MLE) estimates ranging from 1.0 × 10 -4 to 2.1 × 10 -4 per μg/m 3 after converting from occupational to environmental exposure scenario. Pepelko and Chen (1993) developed unit risk estimates based on the data of Brightwell et al. (1989), Ishinishi et al. (1986), and Mauderly et al. (1987) using a detailed dosimetry model to extrapolate dose to humans and a linearized multistage (LMS) model. Taking the geometric mean of individual estimates from the three bioassays, they derived unit risk estimates of 1.4 ×
10 -5 per μg/m 3 when dose was based on carbon particulate matter per unit lung surface area rather than whole DPM, and 1.2 × 10 -4 per μg/m 3 when based on lung burden per unit body weight. Hattis and Silver (1994) derived a maximum likelihood estimate for occupational exposure of 5.2 × 10 -5 per μg/m 3 based on lung burden and bioassay data reported by Mauderly et al. (1987) and use of a five-stage Armitage -Doll low-dose extrapolation model. California EPA (CAL-EPA, 1998) derived a geometric mean estimate of 6 × 10 -5 per μg/m 3 from five bioassays using an LMS model. To demonstrate the possible influence of particle effects as well as particle-associated organics, an additional modeling approach was conducted by Chen and Oberdorster (1996). Employing a biologically based two-stage model a nd using malignant tumor data from Mauderly et al. (1987), the upper-bound risk estimate for exposure to 1 μg/m 3 was estimated to be
1.7 × 10 -5 . This estimate is virtually identical to that using the LMS model, assuming nonthreshold effect of particles. If a threshol d of particle effect is assumed, however, the
estimated risk decreases about fivefold. The results also show that the mechanism of DEinduced lung tumor at high exposure concentra tions may differ from that at low exposure concentrations, with the organics and particles playing primary roles of tumorigenesis, respectively, at low and high concentrations. Overall, the potency estimates on the basis ofC-7 animal bioassays are in the range of 10 -6 to 10 -4 per 1 μg/m 3 of DPM. Valberg and Crouch (1999) conducted a meta-analysis of rat bioassays by pooling together data of low-dose groups from different bi oassays. There are eight bioassays used in the meta-analysis; half of them had duration of 24 months, and the remaining studies had duration of 30 months or more. Animals with continuous lifetime exposure of less than 600 μg/m 3 of DE were included in the analysis. Continuous lifetime exposure is calculated by protracting actual DE exposure to 30 months (24 hours per day, 7 days per week). The researchers concluded that exposure of rats to DE at concentrations not associated with lung overload is consistent with no tumorigenic effect.

REFERENCES FOR APPENDIX C

Albert, RE; Chen, C. (1986) U.S. EPA diesel studies on inhalati on hazards. In: Carcinogenic ity and mutagenicity of diesel engine exhaust. Ishinishi, N; Koizumi, A; McClellan, R; et al., eds. Amsterdam: Elsevier, pp. 411-419.

Albert, RE; Lewtas, J; Nesnow, S; et al. (1983) Comparative potency method for cancer risk assessment: application to diesel particulate emissions. Risk Anal 3:101-117.

Brightwell, J; Fouillet, X; Cassano-Zoppi , AL; et al. (1989) Tumours of the re spiratory tract in rats and hamsters following chronic inhalation of engine exhaust emissions. J Appl Toxicol 9:23-31.

California Environmental Protection Agency. (CAL-EPA, OEHHA) (1998) Health risk assessment for DE. Public and Scientific Review Draft.

Chen, CW; Oberdörster, G. (1996) Selection of models for assessing dose-response re lationships for particleinduced lung cancer. Inhal Toxicol 8(suppl):259-278.

Cuddihy, RG; McClellan, RO. (1983) Evaluating lung cancer risks from exposures to diesel engine exhaust. Risk Anal 3:119-124.

Cuddihy, RG; Griffith, WC; Clark, CR; et al. (1981) Potential health and environmental effects of light-duty diesel vehicles. Report No. LMF-89. Albuquerque, NM: Inhalation Toxi cology Research Institute.

Cuddihy, RG; Griffith, WC; McClellan, RO. (1984) Health risks from light-duty diesel vehicles. Environ Sci Technol 18:14-21a.

Garshick, E; Schenker, MB; Munoz, A; et al. (1987) A case-control study of lung cancer and DE exposure in railroad workers. Am Rev Respir Dis 135:1242-1248.

Garshick, E; Schenker, MB; Munoz, A; et al. (1988) A re trospective cohort study of lung cancer and DE exposure in railroad workers. Am Rev Respir Dis 137:820-825.

Harris, JE. (1983) Diesel emissions and lung cancer. Risk Anal 3:83-100.

Hattis, D; Silver, K. (1994) Use of mechanistic data in occupational health risk asse ssment: the example of diesel particulates. In: Chemical risk assessment and occupational health. Smith, MC; Christiani, DC; Kelsey, KT, eds. Westport, CT: Auburn House, pp. 167-178.C-8 Heinrich, U; Muhle, H; Takenaka, S; et al. (1986) Chroni c effects on the respiratory tract of hamsters, mice and rats after long-term inhalation of high concentration of filtered and unfiltered diesel engine emissi ons. J Appl Physiol 6:383-395.

Ishinishi, N; Kuwabara, N; Nagase, S; et al. (1986) L ong-term inhalation studies on effects of exhaust from heavy and light duty diesel engines on F344 rats. In: Carcinoge nic and mutagenic effects of diesel engine exhaust. Ishinishi, N; Koizumi, A; McClellan, RO; et al., eds. Amsterdam: Elsevier, pp. 329-348.

Mauderly, JL; Jones, RK; Griffith, WC; et al. (1987) DE is a pulmonary carcinogen in rats exposed chronically by inhalation. Fundam Appl Toxicol 9:208-221.

McClellan, RO. (1986) Health effects of DE: a case study in risk assessment. Am Ind Hyg Assoc J 47:1-13.

McClellan, RO; Cuddihy, RG; Griffith, WC; et al. (1989) Integrating diverse data sets to assess the risks of airborne pollutants. In: Assessment of inhalation hazards. Mohr, U, ed. New York: Springer Verlag, pp. 3-22.

Pepelko, WE; Chen, C. (1993) Quantita tive assessment of cancer risk from e xposure to diesel engine emissions. Regul Toxicol Pharmacol 17:52-65.

Pott, F; Heinrich, U. (1987) Dieselmotorabgas und Lunge nk auf die Gefährdung des Menschen. In: Umwelthygiene, vol. 19. Med. Institut. f. Umwelthygiene, A nnual Report 1986/87. Dusseldorf, F.R.G., pp. 130-167.

Smith, RA; Stayner, L. (1990) An exploratory assessment of the risk of lung cancer associated with exposure to DE based on a study with rats. Final report. Division of St andards Development and Tec hnology Transfer; Cincinnati, OH: NIOSH.

Stayner, L; Ddankovic, D; Smith, R; et al. (1998) Predicted lung cancer risk among miners exposed to diesel exhaust particles. Am J Ind Med 34:207-219.

Steenland, NK; Silverman, DT; Hornung, RW. (1990) Case-control study of lung cancer and truck driving in the Teamsters Union. Am J Publ Health 80:670-674.

Steenland, K; Deddens, J; Stayner, L. (1998) DE and lung cancer in the trucking industry: exposure-response analysis and risk assessment. Am J Ind Med 34:220-228.

Valberg, PA; Crouch, EA. (1999) Meta-analysis of rat l ung tumors from lifetime inhalation of diesel exhaust. Environ Health Perspect 107(9):693-699.

Waller, RE. (1981) Trends in lung cancer in London in relation to exposure to diesel fumes. Environ Int 5:479-483.

Woskie, SR; Smith, TJ; Hammond, SK; et al. (1988) Estimation of the DE exposures of railroad workers: II. National and historical exposures. Am J Ind Med 13:395-404.

Zaebst, D; Clapp, D; Blade, L; et al. (1991) Quantitative determination of trucking industry workers’ exposures to diesel particles. Am Ind Hyg Assoc J 52:529-541.United States Environmental Protection Agency/ORD National Center for Environmental Assessment Washington, DC 20460

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